U.S. Geological Survey 

Toxic Substances Hydrology ProgramProceedings of the Technical Meeting, 

Colorado Springs, Colorado, 

September 20-24, 1993

By David W. Morganwalp and David A. Aronson

U.S. GEOLOGICAL SURVEY

WATER-RESOURCES INVESTIGATIONS REPORT 94-4015

Volume 1 of 2

Tallahassee, Florida 

1996
U.S. DEPARTMENT OF THE INTERIOR 

BRUCE BABBITT, Secretary

U.S. GEOLOGICAL SURVEY 

Gordon P. Eaton, Director

r

For additional information write to:

Assistant Coordinator

Toxic Substances Hydrology Program

U.S. Geological Survey

412 National Center

Reston, Virginia 22092

Copies of this report may be 

purchased from:

U.S. Geological Survey

Earth Science Information Center

Open-File Reports Section

Federal Center, MS 517

Box 25286

Denver, CO 80225
PREFACE

The Toxic Substances Hydrology (Toxics) Program of the U.S. Geological Survey (USGS) provides 

earth-science information needed to prevent or mitigate contamination of the Nation's ground-water and 

surface-water resources. The Toxics Program's objective is to understand the movement and fate of toxic 

substances in the Nation's ground water and surface 

water. The program's research activities include 

(1) field and laboratory investigation of the physical, 

chemical, and biological processes that control the 

transport and fate of contaminants in ground water and 

surface water; (2) the development of new analytical 

methods and sampling techniques to measure the concentrations of toxic substances in water; and (3) the 

investigation of the relations among human activities, 

environmental characteristics, and regional patterns of 

water quality. Most of the research activities take 

place at contaminated sites. The major research 

objective is not characterization of contaminated sites, 

but is obtaining knowledge about contaminant-transport processes transferable to other contaminated sites. 

For example, the knowledge obtained at the research 

sites could be used to predict contaminant transport at 

other sites, improve methods for remediating contaminated ground water, design water-quality assessments, 

and identify potential causes of contamination.

Recent accomplishments of the Toxics Program 

have been presented at a series of technical meetings 

held in Tucson, Arizona (1984), Cape Cod, Massachusetts (1985), Denver, Colorado (1987), Pensacola, 

Florida (1987), Phoenix, Arizona (1988), and 

Monterey, California (1991). This report documents 

the oral and poster presentations given at the most 

recent meeting, held in Colorado Springs, Colorado, 

in September 1993. The sections in this report correspond to the topical organization of the meeting's oral 

sessions. Each session focused on research at a specific contaminated site or on a topical subject. Papers 

associated with poster sessions were integrated into 

the appropriate section. The lead paper in most of the 

sections gives an overview of the research conducted 

at the site discussed.

Section A addresses the biodegradation of gasoline in ground water that resulted from a leaking 

underground storage tank in Galloway Township, 

New Jersey. The papers describe methods to determine in situ biodegradation rates of gasoline on the 

water table, the geochemistry of gasoline constituents

dissolved in ground water, and air-flow modeling of 

gas transport in the unsaturated zone.

Section B describes fluid flow and solute transport in fractured rock at the Mirror Lake Site in New 

Hampshire, where methods to characterize contaminated fractured rock are being developed. The methods are interdisciplinary and draw on techniques from 

hydrology, geology, geochemistry, and geophysics. 

Specific examples of papers are (1) discussions on the 

use of well logs to help design tracer tests and (2) comparisons of subsurface fracture mapping with electromagnetic and seismic tomography and with borehole 

acoustic televiewer and borehole flowmeter logs.

Section C contains papers on the migration of 

sewage-contaminated ground water on Cape Cod, 

Massachusetts, and on the transport and fate of metals, 

microbes, and organic chemicals in ground water. The 

papers describe the use of a large array of multilevel 

samplers to conduct large- and small-scale (hundreds 

of meters to meters) tracer tests designed to study a 

variety of transport processes in ground water. Examples of topics of additional papers in this section 

include: phosphate contamination in ground water, the 

role of protozoa in an organically contaminated aquifer, and numerical simulation of contaminant transport.

Section D is on the biodegradation and transport 

of chlorinated solvents, such as trichloroethylene 

(TCE), in ground water at Picatinny Arsenal, New 

Jersey. The papers describe the desorption of TCE 

from aquifer sediments, the microbial degradation and 

transport of TCE in ground water, the volatilization 

and transport of TCE in the unsaturated zone, and the 

surfactant-enhanced remediation of ground water contaminated with TCE.

Section E presents the development of methods 

to analyze and collect water samples, and to investigate transport processes. These papers discuss, for 

example, the development of a new method to collect 

sediment cores without loss of sample, new analytical 

methods to measure the concentrations of pesticides 

and pesticide metabolites in water, and the use of 

nitrogen isotopes to study the subsurface transport of 

nitrogen.

Section F contains papers on the environmental 

fate of agricultural chemicals in the Midwest. The 

papers describe the occurrence, distribution, and transport of agricultural chemicals in ground water, 

streams, large rivers, reservoirs, and precipitation at

iii
regional to test plot scales. Examples of specific topics in this section are papers on the relation of nitrate 

concentrations in surface water to land use, the persistence of herbicides in reservoirs, and the results of a 

reconnaissance of pesticides in near-surface aquifers. 

The final eight papers in this section present results 

from the Management Systems Evaluation Areas 

(MSEA) program, which is part of an interagency 

research program to evaluate the effects of agricultural-management (farming) systems on water quality.

Section G contains papers on the natural biodegradation and transport of crude oil in the subsurface at Bemidji, Minnesota. The papers describe 

crude-oil contamination that resulted from a pipeline 

failure. Because the time the contamination occurred 

and the amount of crude oil in the subsurface are 

known, this site has provided USGS researchers and 

others a unique opportunity to study contaminant 

hydrology; information on the source of contamination usually is not available for most sites of groundwater contamination. Examples of specific topics in 

this section are papers on the use of simulation to 

study field-scale solute transport and biodegradation, 

and on small-scale chemical heterogeneities in the 

contaminated aquifer.

Section H focuses on the fate of anthropogenic 

contaminants in the San Francisco Bay Estuary and 

tributaries in California. The papers describe the 

transport of pesticides from the San Joaquin and 

Sacramento Rivers, California, through the Sacramento-San Joaquin Delta, and into the Bay. In addition, the papers in this section describe the fate of 

organochlorine compounds, the results of a survey of 

molecular marker compounds, and the distribution 

and flux of sulfides in the Bay.

Section I contains papers on the transport and 

fate of toxic metals in acidic mine drainage in the 

headwaters of the Upper Arkansas River in Colorado. Results of tracer and pH-modification experiments in St. Kevin Gulch, Leadville, Colorado a 

small mountain watershed-are discussed. These 

experiments were designed to isolate hydraulic transport processes so that geochemical processes can be 

studied.

Section J addresses a variety of topics relating 

to the hydrology of toxic substances in ground water 

and surface water. The papers are grouped by topic, 

and discuss, for example, the fate of organic compounds in surface water, the fate of hydrocarbons in 

the subsurface, the microbial degradation of explosive

wastes, and the measurement of unsaturated-zone 

properties.

Section K reports on the transport of agricultural chemicals in ground water at an agricultural 

research plot near Plains, Georgia, where USGS and 

U.S. Department of Agriculture researchers collaborated on a tracer test involving bromide, atrazine, and 

other agricultural chemicals beneath a plot of irrigated 

corn. The papers discuss, for example, the twodimensional distribution of a bromide tracer in the 

unsaturated zone, and the spatial variability of soil 

properties affecting the transport of agricultural chemicals.

Section L contains papers on nonpoint sources 

of ground-water contamination. These papers, which 

report results of studies on Long Island, New York, 

and the New Jersey coastal plain, describe the relation 

between land use and ground-water contamination.

Section M presents papers on acidic groundwater and surface-water contamination from copper 

mining near Final Creek, Arizona. The papers 

discuss, for example, the use of ground-water age dating to confirm estimates of traveltimes, the interaction 

of the contaminant plume with Final Creek, and the 

relations among algal species and chemical gradients 

in Final Creek.

Compiling and editing a proceedings of this size 

is a complex and difficult task. The task was made 

easier by the work and dedication of David Aronson, 

who died on April 7, 1996. David coedited two proceedings of technical meetings for the Toxic Substances Hydrology Program, and received awards for 

his work. This proceedings is a testament to his many 

contributions to the U.S. Geological Survey.

David W. Morganwalp

Acting Coordinator

Toxic Substances Hydrology Program

iv
CONTENTS 

Volume 1

Page 

SECTION A Gasoline in Ground Water in Galloway Township, New Jersey

Overview of Research on the Transport, Microbial Degradation, and Remediation of Hydrocarbons at a 

Subsurface Gasoline-Spill Site in Galloway Township, New Jersey

By Arthur L. Baehr and Jeffrey M. Fischer.............................................................................................................. 1

Hydrogeology, Contaminant Distribution, and Biodegradation Processes at a Gasoline-Spill Research Site 

in Oalioway Township, New Jersey

By Jeffrey M. Fischer, Nicholas P. Smith, and Arthur L. Baehr.............................................................................. 7

Biogeochemical Fate of Organic Acid Metabolites Produced During the Microbial Degradation of 

Hydrocarbons in Ground Water

By Isabelle M. Cozzarelli and Mary Jo Baedecker.................................................................................................. 17

Method for Estimating the Rate of Microbial Degradation of Hydrocarbons Associated with Bioventing 

at a Gasoline-Spill Site in Galloway Township, New Jersey

By Matthew A. Lahvis, Arthur L. Baehr, and Jeffrey M. Fischer............................................................................ 27

Determination of Vapor-Phase Diffusion Coefficients for Unsaturated-Zone Sediments at a 

Gasoline-Spill Site in Galloway Township, New Jersey

By Jeffrey M. Fischer, Ronald J. Baker, Matthew A. Lahvis, and Arthur L. Baehr................................................ 35

Use of Column Studies and a Reactive Transport Model to Measure Biodegradation Rates of Hydrocarbons

By Ronald J. Baker and Arthur L. Baehr................................................................................................................ 43

Method for Monitoring Biodegradation of Hydrocarbons in Unsaturated Porous Media

By John G. Nolan, Ronald J. Baker, and Arthur L. Baehr ...................................................................................... 49

Optimizing the Design of Vapor-Extraction Remediation Systems for Removal of Organic 

Contaminants From the Unsaturated Zone

By Claire Welty, Craig J. Joss, and Arthur L. Baehr............................................................................................... 55

A Multispecies Transport Model to Simulate Vapor Extraction and Bioventing Remediation of Unsaturated 

Zones Contaminated with Gasoline

By C.J. Joss and A.L. Baehr..................................................................................................................................... 61

A Comparison of Two Analytical Techniques For Identifying and Measuring the Concentrations of 

Volatile Hydrocarbons in Ground Water

By Curtis S. Phinney and Isabelle M. Cozzarelli..................................................................................................... 65

SECTION B Research on Flow and Transport in Fractured Rock at Mirror Lake, New Hampshire

Overview of Research on Use of Hydrologic, Geophysical, and Geochemical Methods to

Characterize Flow and Chemical Transport in Fractured Rock at the Mirror Lake Site, New Hampshire

By Alien M. Shapiro and Paul A. Hsieh .................................................................................................................. 71

Characterizing Bedrock Fractures in Outcrop for Studies of Ground-Water Hydrology: An Example 

from Mirror Lake, Grafton County, New Hampshire

By Christopher C. Barton......................................................................................................................................... 81

Use of a Borehole Color Video Camera to Identify Lithologies, Fractures, and Borehole Conditions in 

Bedrock Wells in the Mirror Lake Area, Grafton County, New Hampshire

By Carole D.Johnson.............................................................................................................................................. 89

Fracture Detection in Crystalline Rocks, Mirror Lake Area, Grafton County, New Hampshire

By P.P. Haeni, John W. Lane, Jr., C.C. Barton, and David A. Lieblich.................................................................. 95

Use of Well Logs to Prepare the Way for Packer Strings and Tracer Tests: Lessons from the 

Mirror Lake Study

By Frederick L. Paillet............................................................................................................................................. 103

Velocity, Attenuation, and Dispersion Electromagnetic Tomography in Fractured Rock

By David L. Wright, Gary R. Olhoeft, and Thomas P. Grover................................................................................ 111
Electromagnetic and Seismic Tomography Compared to Borehole Acoustic Televiewer and Flowmeter Logs 

for Subsurface Fracture Mapping at the Mirror Lake Site, New Hampshire

By David L. Wright, Gary R. Olhoeft, Paul A. Hsieh, Ernest L. Majer, Frederick L. Paillet, and

John W. Lane, Jr................................................................................................................................................... 119

Hydraulic Characteristics of Fractured Bedrock Underlying the FSE Well Field at the Mirror Lake Site, 

Grafton County, New Hampshire

By Paul A. Hsieh and Alien M. Shapiro .................................................................................................................. 127

A New Method of Performing Controlled Injection of Traced Fluid in Fractured Crystalline Rock

By Alien M. Shapiro and Paul A. Hsieh.................................................................................................................. 131

Field Instrumentation for Multilevel Monitoring of Hydraulic Head in Fractured Bedrock at the Mirror 

Lake Site, Grafton County, New Hampshire

By Paul A. Hsieh, Richard L. Perkins, and Donald O. Rosenberry......................................................................... 137

Factors Affecting Recharge to Crystalline Rock in the Mirror Lake Area, Grafton County, New Hampshire

By Philip T. Harte and Thomas C. Winter............................................................................................................... 141

Concentrations of Chlorofluorocarbons and Other Gases in Ground Water at Mirror Lake, New Hampshire

By Eurybiades Busenberg and L. Niel Plummer...................................................................................................... 151

Helium Isotope Analysis and Tritium-Helium Age Dating in the Mirror Lake Basin, Grafton County, 

New Hampshire

By S. Drenkard, T. Torgersen, R. Weppernig, K. Parley, P. Schlosser, R.L. Michel, A.M. Shapiro and

W.W. Wood ......................................................................................................................................................... 159

Observational, Experimental and Inferred Evidence for Solute Diffusion in Fractured Granite Aquifers: 

Examples from the Mirror Lake Watershed, Grafton County, New Hampshire

By Warren W. Wood, Alien M. Shapiro, Paul A. Hsieh, and Terry B. Councell.................................................... 167

A Conceptual Model for Estimating Regional Ground-Water Velocity in Bedrock of the Mirror Lake Area, 

Grafton County, New Hampshire

By Alien M. Shapiro, Warren W. Wood, Eurybiades Busenberg, Stefan Drenkard, L. Niel Plummer,

Thomas Torgersen, and Peter Schlosser.............................................................................................................. 171

SECTION C Contaminant Transport Research on Cape Cod, Massachusetts

Overview of Research at the Cape Cod Site: Field and Laboratory Studies of Physical, Chemical, and 

Microbiological Processes Affecting Transport in a Sewage-Contaminated Aquifer

By Denis R. LeBlanc................................................................................................................................................ 179

The Transport of Inorganic Contaminants in a Sewage Plume in the Cape Cod Aquifer, Massachusetts

By Brigid A. Rea, Douglas B. Kent, Linda C. D. Anderson, James A. Davis, and Denis R. LeBlanc.................... 191

Potential Long-Term Effects of Phosphate Contamination in a Sand and Gravel Aquifer, 

Cape Cod, Massachusetts

By Kenneth G. Stollenwerk...................................................................................................................................... 199

Background Aqueous Chemistry and Effects of Carbon Dioxide Variations in Recharge in Shallow 

Ground Water in a Glacial Outwash Aquifer, Cape Cod, Massachusetts

By Roger W. Lee...................................................................................................................................................... 207

Heavy-Metal Transport in a Sand and Gravel Aquifer with Variable Chemical Conditions, Cape Cod, 

Massachusetts

By James A. Davis, Douglas B. Kent, Jennifer A. Coston, andKathrynM. Hess.................................................. 219

Evaluation of Geochemical Indicators of Metal Adsorption in a Sand and Gravel Aquifer, Cape Cod, 

Massachusetts

By Christopher C. Fuller, Jennifer A. Coston, James A. Davis, and Eleanor Dixon............................................... 227

Spatial Variability of Metal-Ion Adsorption and Hydraulic Conductivity in a Sand and Gravel Aquifer, 

Cape Cod, Massachusetts

By Kathryn M. Hess, James A. Davis, Christopher C. Fuller, and Jennifer A. Coston........................................... 235

Small-Scale Tracer Tests Applied to the Measurement of in situ Denitrification Rates in a SewageContaminated Aquifer, Cape Cod, Massachusetts

By Myron H. Brooks, Richard L. Smith, and Stephen P. Garabedian..................................................................... 243

vi
Overview of Research on the Distribution and Role of Protozoa in an Organically Contaminated Aquifer at 

Cape Cod, Massachusetts

By Nancy E. Kinner and Ronald W. Harvey............................................................................................................ 249

Transport of Indigenous Protozoa in a Sandy Aquifer, Cape Cod, Massachusetts

By Ronald W. Harvey, Nancy E. Kinner, Dan MacDonald, Amoret L. Bunn, and David W. Metge..................... 253

Use of Static Column Experiments to Identify Factors Affecting Bacterial Attachment in Contaminated 

Aquifer Sediments From Cape Cod, Massachusetts

By David W. Metge, Ronald W. Harvey, George R. Aiken, and Larry B. Barber, II............................................. 259

An Experiment to Quantify Temporal Variability of Water Samples Obtained from Screened Wells, 

Cape Cod, Massachusetts

By Thomas E. Reilly and Denis R. LeBlanc............................................................................................................ 265

Preliminary Type-Curve Analysis of an Aquifer Test in an Unconfined Sand and Gravel Aquifer, 

Cape Cod, Massachusetts

By Allan F. Moench, Denis R. LeBlanc, and Stephen P. Garabedian..................................................................... 273

Numerical Simulation of Downward Movement of Solutes During a Natural-Gradient Tracer Test in Sand and 

Gravel, Cape Cod, Massachusetts

By Denis R. LeBlanc and Michael A. Celia............................................................................................................. 283

Effects of Changes in Aquifer Properties on Simulated Fluid-Particle Pathlines, Cape Cod, Massachusetts

By John P. Masterson and Donald A. Walter........................................................................................................... 289

SECTION D Chlorinated Solvents in Ground Water at Picatinny Arsenal, New Jersey

Overview of Research Activities on the Transport and Fate of Chlorinated Solvents in Ground Water at 

Picatinny Arsenal, New Jersey, 1991-93

By Thomas E. Imbrigiotta and Mary Martin............................................................................................................ 297

Simulation of Transport, Desorption, Volatilization, and Microbial Degradation of Trichloroethylene in 

Ground Water at Picatinny Arsenal, New Jersey

By Mary Martin........................................................................................................................................................ 307

Simulation of Trichloroethylene Volatilization in the Unsaturated Zone during a Field Infiltration Experiment at 

Picatinny Arsenal, New Jersey

By H. Jean Cho, James A. Smith, and Peter R. Jaffe............................................................................................... 315

Cometabolic Biotransformation of Trichloroethylene and cis-1,2-Dichloroethylene in Unsaturated-Zone 

Soil at Picatinny Arsenal, New Jersey

By Theodore A. Ehlke and Thomas E. Imbrigiotta.................................................................................................. 321

Desorption of Trichloroethylene From Aquifer Sediments at Picatinny Arsenal, New Jersey

By David Koller, Thomas E. Imbrigiotta, Arthur L. Baehr, and James A. Smith ................................................... 329

Effect of Air Drying on Solvent Extraction of Trichloroethylene-Contaminated Soil, Picatinny Arsenal, 

New Jersey

By David Koller, Thomas E. Imbrigiotta, and James A. Smith............................................................................... 339

In Situ Biotransformation of Trichloroethylene and cis-1,2-dichloroethylene at Picatinny Arsenal, 

New Jersey

Ry Theodore A. Ehlke. Barbara H. Wilson, John T. Wilson, and Thomas E. Imbrigiotta...................................... 347

Surfactant-Enhanced Remediation of Ground Water at Picatinny Arsenal, New Jersey

By James J. Deitsch and James A. Smith................................................................................................................. 355

SECTION E Methods Development

Sampling Borehole Flow to Quantify Aquifer Cross-Contamination by Volatile Organic Compounds

By Ronald A. Sloto................................................................... 361

A Sample-Freezing Drive Shoe for a Wireline-Piston Core Sampler

By Fred Murphy and W.N. Herkelrath..................................................................................................................... 369

Determination of Surface Area of Soil Components by Sorption of Nitrogen and Ethylene Glycol Monoethyl 

Ether Vapors and Evaluation of Adsorption, Partition and Cation Solvation Processes

By Cary T. Chiou and David W. Rutherford............................................................................................................ 375

vii
Isolation of Lipophilic Organic Contaminants Along the Upper Mississippi River Using a 

Semipermeable Membrane Device

By Geoffrey S. Ellis and Colleen E. Rostad............................................................................................................. 381

Measurement of Concentrations of Alachlor and Its Ethanesulfonic Acid Metabolite in Water by Solid-Phase 

Extraction and Enzyme-Linked Immunosorbent Assay

By Diana S. Aga, E. Michael Thurman, and Michael L. Pomes.............................................................................. 385

Isocratic Separation of Alachlor Ethanesulfonic Acid, Alachlor Oxoacetic Acid, and Hydroxyatrazine by 

Reversed-Phase Liquid Chromatography

By Michael L. Pomes, Douglas F. Holub, Diana S. Aga, and E.M. Thurman......................................................... 393

Use of 15N to Trace Movement of Nitrogen Fertilizer at a Field Plot

By Dale W. Blevins, Donald H. Wilkison, Steven R. Silva, and Brian P. Kelly..................................................... 401

SECTION F Occurrence, Distribution, and Transport of Agricultural Chemicals in the Midwest

Occurrence, Deposition, and Long Range Transport of Herbicides in Precipitation in the Midwestern and 

Northeastern United States

By Donald A. Goolsby, E. Michael Thurman, Michael L. Pomes, Michael Meyer, and William A. Battaglin...... 411

Atrazine Transport and Degradation in a Pristine Watershed: The Fate of Atrazine Deposited by Precipitation

By Aron E. Cromwell and E. Michael Thurman...................................................................................................... 423

Annual Use and Transport of Agricultural Chemicals in the Mississippi River, 1991-92

By William A. Battaglin, Donald A. Goolsby, and Richard H. Coupe.................................................................... 429

Relation of Nitrate Concentrations in Surface Water to Land Use in the Upper-Midwestern United States, 

1989-90 

By David K. Mueller, Barbara C. Ruddy, and William A. Battaglin....................................................................... 441

Reconnaissance Data for Selected Herbicides and Two Atrazine Metabolites in Surface Water of the Midwestern 

United States: Chemical Analysis by Immunoassay and Gas Chromatography/Mass Spectrometry

By Elisabeth A. Scribner, E. Michael Thurman, and Donald A. Goolsby............................................................... 449

Chemistry, Degradation, and Transport of Triazine Herbicide Metabolites in Surface Water

By E.M. Thurman, M.S. Mills, and M.T. Meyer..................................................................................................... 455

The Transport and Degradation of Cyanazine Metabolites in Surface Water of the Midwestern United States

By M.T. Meyer and E.M. Thurman.......................................................................................................................... 463

Transport and Concentrations of Selected Herbicides and Nitrate in the Sangamon River, Illinois, 

April 1991 -March 1992

By G.P. Johnson and R.H. Coupe ............................................................................................................................ 471

Temporal Distributions of Herbicide and Nitrite-Nitrate Concentrations and Loads in the West Fork of the 

Big Blue River, Nebraska, 1990 and 1991-92

By A.D. Druliner, D.A.Goolsby, and M. Meyer...................................................................................................... 479

Persistence of Herbicides in Selected Reservoirs in the Midwestern United States: Some Preliminary Results 

By Donald A. Goolsby, William A. Battaglin, James D. Fallen, Diana S. Aga, Dona W. Koplin, and

E. Michael Thunnan.............................................................................................................................^ 487

Determining the Relative Age, Transport, and Three-Dimensional Distribution of Atrazine in a Reservoir 

Using Immunoassay

By James D. Fallen and E.M. Thurman................................................................................................................... 499

Pesticides in Near-Surface Aquifers: Results of the Midcontinental United States Ground-Water 

Reconnaissance, 1991 -92

By Dana W. Koplin, Donald A. Goolsby, Dana S. Aga, Jana L. Iverson, and E. Michael Thurman...................... 505

Predicting Nitrate-Nitrogen and Atrazine Contamination in the High Plains Aquifer in Nebraska

By A.D. Druliner and T.S. McGrath........................................................................................................................ 515

Comparison of Simulated and Observed Movement of Bank-Storage Water Adjacent to the Cedar River, Iowa

By Paul J. Squillace and David W. Pollock............................................................................................................. 525

Fate of Alachlor, Atrazine, and Bromide in a Corn Plot near Topeka, Kansas

By D.A. Eckhardt, R.J. Wagenet, E.M. Thurman, and P.L. Barnes......................................................................... 535

viii
Effects of Farming Systems on Ground-Water Quality at the Princeton, Minnesota, Management 

Systems Evaluation Area, 1991

By Matthew K. Landon, Geoffrey N. Delin, John A. Lamb, Robert H. Dowdy, and James L. Anderson.............. 543

Effects of Topography on the Transport of Agricultural Chemicals near Princeton, Minnesota, 1992

By Geoffrey N. Delin and Matthew K. Landon....................................................................................................... 555

Spatial Variability of Unsaturated-Zone Properties in Relation to Topography in a Sand-Plain Setting near 

Princeton, Minnesota

By G.N. Delin, M.K. Landon, R.W. Healy, and H.W. Olsen.................................................................................. 565

Nitrogen Uptake and Soil Water Variability Across a Sand Plain Landscape

By J.A. Lamb, M.D. Tomer, J.L. Anderson, and R.H. Dowdy................................................................................ 575

Geophysical Investigations of Heterogeneity and Scale at the Princeton, Minnesota, Management 

Systems Evaluation Area

By Jeffrey E. Lucius and Gary R. Olhoeft............................................................................................................... 581

Gcuclicmi&lry of Nitrogen in a Farmed Watershed Near Princeton, Minnesota

By Richard B. Wanty, Michele L. Turtle, Matthew K. Landon, Geoffrey N. Delin, and J.K. Bohlke.................... 591

Chemical Loads of Nitrate, Atrazine, and Metolachlor in Walnut Creek Watershed near Ames, Iowa, 1991-92

By P.J. Soenksen, J.L. Hatfield, and D.J. Schmitz................................................................................................... 597

Atrazine Transport Through Preferential Pathways at the Missouri MSEA

By James A. Tindall and William K. Vencill........................................................................................................... 607

Volume 2

SECTION G Crude Oil in the Subsurface Environment at Bemidji, Minnesota

Evolution of the Contaminant Plume in an Aquifer Contaminated with Crude Oil, Bemidji, Minnesota

By Mary Jo Baedecker, Isabelle M. Cozzarelli, Philip C. Bennett, Robert P. Eganhouse, and Marc F. Hull......... 613

Use of Simulation to Study Field-Scale Solute Transport and Biodegradation at the Bemidji, Minnesota, 

Crude-Oil Spill Site

By H.I. Essaid, M.J. Baedecker, and I.M. Cozzarelli............................................................................................... 621

Source Mass Balance Calculated from Changes in Composition of Spilled Crude Oil in the Subsurface near 

Bemidji, Minnesota

By Matthew K. Landon and Marc F. Hult................................................................................................................ 631

Studies of the Molecular Size of Dissolved Organic Carbon Fractions Downgradient from the Oil Body at 

Bemidji, Minnesota

By George Aiken...................................................................................................................................................... 641

Small-Scale Chemical Heterogeneities in a Crude-Oil-Contaminated Aquifer, Bemidji, Minnesota

By Isabelle M. Cozzarelli, Mary Jo Baedecker, George Aiken, and Curtis Phinney............................................... 647

Use of Carbon and Hydrogen Stable Isotopes to Investigate the Production and Fate of Methane at a 

Toxic Waste Site, Bemidji, Minnesota

By Kinga Revesz, Tyler Coplen, Mary J. Baedecker, and Marc Hult..................................................................... 653

Sorption of Polycyclic Aromatic Hydrocarbons to Aquifer Materials Containing a Low-Organic Carbon 

Content from Bemidji, Minnesota

By Joseph J. Piatt, Steven J. Eisenriech, Debera A. Backhus, and Paul D. Capel................................................... 663

Investigations of Microbial Weathering of Minerals

By Philip C. Bennett and Franz K. Hiebert.............................................................................................................. 671

Multiphase-Flow Modeling of the Bemidji, Minnesota Crude-Oil-Spill Site with Geostatistical Simulation of 

Hydraulic Properties

By L.A. Dillard, H.I. Essaid, and W.N. Herkelrath.................................................................................................. 679

IX
SECTION H Anthropogenic Contaminants in the San Francisco Bay Estuary and Tributaries, 

California

Challenges in Detecting Contaminant Effects on an Estuarine Ecosystem Affected by Many Different 

Disturbances: San Francisco Bay

By Samuel N. Luoma, and Frederic H. Nichols....................................................................................................... 689

Hydrodynamics of California's Sacramento-San Joaquin Delta, and the Collection of Flow Data for the 

Tracking of Pesticides

By Richard N.Oltmann...................................^ 695

Diazinon Concentrations and Transport in the Sacramento River and San Francisco Bay, California, 

February 1993

By Kathryn M. Kuivila and Dwight D. Copeland.................................................................................................... 699

Distribution and Possible Biological Effects of Diazinon in the San Joaquin River and the Sacramento-San 

Joaquin Delta, California, February 1993

By Kathryn M. Kuivila, Robert C. Sheipline, and Christopher G. Foe.................................................................... 705

Concentrations of Dissolved Rice Pesticides in the Colusa Basin Drain and Sacramento River, 

California, 1990-92

By Kathryn L. Crepeau, Kathryn M. Kuivila, and Joseph L. Domagalski.............................................................. 711

Method for Estimating the Activity of Carbofuran-Degrading Microorganisms in Environmental Samples

By William J. Meyers and Kathryn M. Kuivila....................................................................................................... 719

Distributions of Organochlorine Compounds in Sediments and Livers of Striped Bass (Morone Saxatilis) in 

San Francisco Bay-Delta Estuary

By Wilfred E. Pereira, Frances D. Hostettler, John R. Cashman and Richard S. Nishioka..................................... 725

A Survey of Molecular Marker Compounds in Sediments of San Francisco Bay, California

By Frances D. Hostettler, John B. Rapp, Wilfred E. Pereira, and Keith A. Kvenvolden........................................ 735

Distributions and Benthic Flux of Dissolved Sulfides in the Toxic Water Column of San Francisco Bay, California

By James S. Kuwabara, Yvonne R. Hunter, and Cecily C.Y. Chang...................................................................... 747

SECTION I Metals in the Headwaters of the Upper Arkansas River, Colorado

Past and Present Research on Metal Transport In St. Kevin Gulch, Colorado

By Briant A. Kimball............................................................................................................................................... 753

Iron and Zinc Budgets in Surface Water for a Natural Wetland Affected by Acidic Mine Drainage, 

St. Kevin Gulch, Lake County, Colorado

ByKathcrineWalton-Day.........................................................................^ 759

Tracer Injections in Small Streams Why and How We Do Them

By Gary W. Zellweger............................................................................................................................................. 765

Processes Controlling Dissolved Copper Concentrations during an Instream pH-Modification Experiment

By Kathleens. Smith............................................................................................................................................... 769

An Equilibrium-Based Simulation Model for Reactive Solute Transport in Small Streams

By Robert L. Runkel, Kenneth E. Bcncala and Robert E. Broshears....................................................................... 775

Simulation of Reactive Transport During a pH Modification Experiment in a Mountain Stream Affected by Acid 

Mine Drainage

By Robert E. Broshears, Briant A. Kimball and Robert L. Runkel......................................................................... 781

Difficulties with Intermediate-Scale Experiments for Studies of Iron Chemistry in Streams Affected by Acidic 

Mine Drainage

By Diane M. McKnight, Kenneth E. Bencala, Richard A. Harnish, and Robert L. Runkel.................................... 789

Metal Concentrations in Sedges in a Wetland Receiving Acidic Mine Drainage from St. Kevin Gulch, 

Leadville, Colorado

By B.M. Erickson, P.H. Briggs, and T.R. Peacock.................................................................................................. 797

Uraniferous Waters of Southeastern Colorado: A Function of Geology, Climate and Land Use

By Robert A. Zielinski and Sigrid Asher-Bolinder.................................................................................................. 805
SECTION J Research on the Fate and Transport of Contaminants in Ground Water and Surface 

Water

Dissolved Organic Compounds in the Mississippi River from Minneapolis-St. Paul, Minnesota 

to New Orleans, Louisiana

By Larry B. Barber, II, Jerry A. Leenheer, Wilfred E. Pereira, and Mary A. Noriega............................................ 813

Linear Alkylbenzene Sulfonate in the Mississippi River

By Charles F. Tabor, Jr., and Larry B. Barber, II..................................................................................................... 821

Organic Carbon Associated With Suspended Silt and Colloid From the Mississippi River and Some of 

Its Tributaries, July 1991-May 1992

By Colleen E. Rostad, Stephanie G. Monsterleet, LaDonna M. Bishop, and Geoffrey S. Ellis.............................. 825

Accumulation of Organochlorine Compounds in Hydrilla Verticillata Relative to Sediments in the Tidal 

Potomac River

By Jessica A. Hopple and Gregory D. Foster........................................................................................................... 831

Pathways of Methanogenic Biodegradation of Creosote-Derived Aromatic Compounds

By E. Michael Godsy, Donald F. Goerlitz, and Dunja Grbic'-Galic'....................................................................... 835

Modeling Steady-State Methanogenic Degradation of Phenols in Ground Water at Pensacola, Florida

By Barbara A. Bekins, E. Michael Godsy, and Donald F. Goerlitz........................................................................ 843

Modeling Breakthrough of Nitrogen Heterocyclic Compounds in Laboratory Columns Containing 

Creosote-Contaminated Aquifer Material

By Ean Warren and E. Michael Godsy.................................................................................................................... 851

Bioremediation of Petroleum and Metal Contamination with Dissimilatory Metal Reduction

By Derek R. Lovley.................................................................................................................................................. 859

Effects of Pb and Terminal-Electron-Accepting Processes on Organic Acid Concentrations in Contaminated 

Aquifer Sediments

By Paul M. Bradley, Francis H. Chapelle, and Don A. Vroblesky.......................................................................... 865

Shifting of Terminal-Electron-Accepting Processes in a Petroleum-Hydrocarbon-Contaminated Aquifer

By Don A. Vroblesky and Francis H. Chapelle ....................................................................................................... 873

Preliminary Assessment of Jet Fuel Contamination of an Atoll Aquifer

By Gordon W. Tribble, Roger W. Lee, Duane S. Wydoski, Jill D. Torikai, and Charles D. Hunt, Jr..................... 881

Microbial Degradation of Nitrotoluenes in Surface Soils and Aquifer Sediments, Weldon Spring, Missouri

By Francis H. Chapelle and Paul M. Bradley .......................................................................................................... 889

Depletion of Nitrogen-Bearing Explosives Wastes in a Shallow Ground-water Plume near Hawthorne, Nevada

By A.S. Van Denburgh, D.F. Goerlitz, and E.M. Godsy......................................................................................... 895

Reaction of Arsenic Minerals and Gold Ores with Alkaline Cyanide Solutions

By Walter H. Ficklin, Geoffrey S. Plumlee, and John B. McHugh......................................................................... 905

Anaerobic Growth of Bacterial Strain Ses3 with Selenate as the Electron Acceptor

By Jodi SwitzerBlum, Charles W. Culbertson, and Ronald S. Oremland............................................................... 911

Chemical, Isotopic, and Microbiological Evidence for Denitrification During Transport of Domestic Wastewater 

Through a Thick Unsaturated Zone in the Mojave Desert

By Roy A. Schroeder, Peter Martin, and J.K. Bohlke.............................................................................................. 917

Potential Anhydrite Precipitation Associated With Deep Injection of Ground-Water Brine from 

Paradox Valley, Colorado

By Yousif K. Kharaka, William C. Evans, Gil Ambats, and James J. Thordsen..................................................... 927

Apparent Solute Dispersion in an Alluvial/Aquifer Stream System

By J. Jeffrey Starn .................................................................................................................................................... 935

Measurements of Hydraulic Conductivity and Capillary Pressure under Unsaturated Conditions in a Laboratory 

Triaxial System

By Harold W. Olsen................................................................................................................................................. 941

Unsaturated Zone Properties at a Hazardous-Waste-Disposal Site at the Idaho National Engineering 

Laboratory

By Stephanie Shakofsky and John R. Nimmo.......................................................................................................... 949

XI
SECTION K Agricultural Chemicals in the Unsaturated Zone and Ground Water Near Plains, Georgia

Fate and Transport of Atrazine at the Plains, Georgia, Ground Water Study Site

By R. A. Leonard, L.R. Marti, D.W. Hicks, and J.B. McConnell............................................................................ 955

Unsaturated Zone Transport at the Plains, Georgia Field Site

By D.D. Bosch, R.A. Leonard, and R.K. Hubbard.................................................................................................. 965

Two-Dimensional Distribution of a Bromide Tracer in the Unsaturated Zone at the Plains, Georgia, 

Research Site

By D.W. Hicks, J.B. McConncll, H.H. Persinger, J.D. Scholz, and R.K. Hubbard................................................. 971

Spatial Variability of Soil Properties Affecting Chemical Transport at the Plains, Georgia, Research Site

By R. K. Hubbard, D. D. Bosch, R. A. Leonard, and D. W. Hicks ........................................................................ 983

SECTION L Nonpoint-Source Contamination Investigations

Modeling Statistical Relations Among Shallow Ground-Water Quality, Human Activities, Land Use, 

and Thickness of the Unsaturated Zone on Long Island, New York

By Paul E. Stackelberg and David A.V. Eckhardt................................................................................................... 993

Evaluating the Effect of Land Use and Sampling Depth on Ground-Water Quality, Long Island, New York

By Paul E. Stackelberg............................................................................................................................................. 1001

Use of Flowpath Simulation to Determine Contributing Areas and Traveltimes of Nonpoint-Source 

Ground-Water Contamination, Gloucester County, New Jersey

By Anthony S. Navoy............................................................................................................................................... 1011

Vulnerability of Water from Public Supply Wells to Contamination by Pesticides, Potomac-Raritan-Magothy 

Aquifer System, New Jersey Coastal Plain

By Eric F. Vowinkel................................................................................................................................................. 1021

A Review of Methods for Assessing the Potential for Regional Ground-Water Contamination 

By David R. Seller ......................................................................^^^

SECTION M Ground-Water and Surface-Water Contamination From Copper Mining Near 

Final Creek, Arizona

Hydrologic and Geochemical Factors affecting Metal Contaminant Transport in Pinal Creek Basin near 

Globe, Arizona

By James G. Brown and Judson W. Harvey ............................................................................................................ 1035

Dissolved Gas and Chlorofluorocarbon Content of Ground Waters in the Pinal Creek Basin, Arizona

By Pierre D. Glynn and Eurybiades Busenberg....................................................................................................... 1043

Unsaturated Zone Diffusion of Carbon Dioxide and Oxygen in the Pinal Creek Basin, Arizona

By Pierre D. Glynn and Eurybiades Busenberg....................................................................................................... 1055

Interactions between Shallow Ground Water and Surface Water that Affect Metal Transport in 

Pinal Creek, Arizona

By Judson W. Harvey, Christopher C. Fuller, and Brian J. Wagner....................................................................... 1065

Association of Selected Metals with Colloidal and Suspended Particulate Material in Shallow Ground Water 

and Surface Water at Pinal Creek, Arizona

By Judson W. Harvey and Christopher C. Fuller..................................................................................................... 1073

Solute-Transport Parameter Estimation for an Injection Experiment at Pinal Creek, Arizona

By Brian J. Wagner and Judson W. Harvey............................................................................................................. 1081

Alteration of Alluvium by Acidic Ground Water Resulting from Copper Mining at Pinal Creek, Arizona

By Carol J. Lind and Kenneth G. Stollenwerk......................................................................................................... 1089

Chemical Processes in Manganese Oxide and Carbonate Precipitation in Pinal Creek, Arizona

By John D. Hem and Carol J. Lind .......................................................................................................................... 1095

Experimental Studies of Trace-Metal Partitioning in Perennial Reaches of Pinal Creek, Arizona

By Christopher C. Fuller.......................................................................................................................................... 1101

Relations Among Algal Species and Chemical Gradients in Pinal Creek, near Globe, Arizona

By Patrice H. Spindler and Milton R. Sommerfeld.................................................................................................. 1107

xii
Overview of Research on the Transport, Microbial Degradation, 

and Remediation of Hydrocarbons at a Subsurface GasolineSpill Site in Galloway Township, New Jersey

By Arthur L. Baehr1 and Jeffrey M. Fischer1

Abstract

Since 1988, the U.S. Geological Survey has 

conducted research on the fate and transport of 

hydrocarbons at a site of subsurface gasoline contamination in Galloway Township, New Jersey, 

Work has focused on three interrelated areas of 

research: ground-water contaminant geochemistry, estimation of microbial degradation rates of 

hydrocarbons on the basis of rates of gas transport 

in the unsaturated /one, and vapor-extraction 

remediation design. Contaminant-geochemistry 

work has produced a water-quality data base that 

indicates zones of aerobic degradation adjacent to 

zones of anaerobic degradation with sharp chemical-concentration gradients in narrow interfacial 

/ones. The analysis of gas transport in the unsaturated zone has provided a method for quantifying 

rates of aerobic hydrocarbon biodegradation in the 

capillary zone and shallow ground water. Mathematical models have been developed to simulate 

vapor-extraction remediation and to optimize system design. Research activity in each area is 

detailed and the project bibliography is updated.

INTRODUCTION

In 1988, the U.S. Geological Survey (USGS) 

began a field-oriented research project at a site of 

subsurface gasoline contamination from a leaky underground storage tank on a farm in Galloway Township, 

New Jersey (fig. 1). The objective of the original 

project, which was conducted in cooperation with the 

New Jersey Department of Environmental Protection 

and Energy, was to determine factors controlling the 

performance of vapor-extraction remediation. During 

the installation and sampling of the initial monitoring

U.S. Geological Survey. West Trenton. N.J.

network in the fall of 1988, it became clear that the site 

afforded a unique opportunity to conduct field and 

laboratory research on the fate and transport of 

hydrocarbons in the subsurface.

The scope of the project was expanded to allow for 

more extensive characterization of contamination in 

the unsaturated zone and shallow ground water before 

site remediation. The project data base provides for 

evaluation of the natural attenuation and microbial 

degradation of hydrocarbons. In 1989, the site was 

selected as a USGS Toxic Substances Hydrology 

program research project. The project workplan in its 

formative stage was presented at the technical meeting 

held in Monterey, Calif., in 1991. The purpose of this 

paper is to summarize research activity and update the 

project bibliography to include results reported since 

the Monterey meeting.

The Galloway Township study site is located in the 

New Jersey Coastal Plain about 30 mi northwest of 

Atlantic City (fig. 1). Unconsolidated sediments 

beneath the site are mostly medium-grained sands, 

with some clay layers. The unsaturated zone is typically about 9 ft thick. Its lower boundary is a perched 

water table which, depending on precipitation history, 

may be coincident with the regional water table. 

Ground water in the regional aquifer usually flows to 

the southeast. Fischer, Smith, and Baehr (1996) 

describe the hydrogeology of the site. The monitoring 

network (fig. 1) consists of 27 nests of vapor probes, 13 

ground-water monitoring wells, 2 multilevel groundwater samplers, 5 vapor-extraction wells, 2 neutronprobe access pipes, and a nest of thermistors (Fischer 

and others, 1991). The composition of unsaturatedzone gases is determined by analysis of gas samples 

collected from the vapor-probe network (Baker and 

others, 1991). Ground-water contamination and inorganic chemistry is determined by analyzing samples 

collected from wells and submerged vapor probes 

(Baedecker and others, 1991; Cozzarelli and others, 

1991). Gibs and others (1993) report on the construction and sampling of the multilevel samplers.
.

40'

\

75' 74" 73*

GALLOWAY 

TOWNSHIP

o ao MILES

^^

0 20 KILOMETERS

\.

h10

I

20 30 40 50 FEET

I I I

I 

10 15 METERS

\

HOUSE

EXPLANATION

 Nested vapor probes

*» U.S. Geological Survey ground-water

observation well 

<§> New Jersey Department of Environmental

Protection ground-water observation well

__________Z______^____________

e Multilevel ground-water 

O Neutron probe access

O Thermistor nest

IS Vapor-extraction well

Figure 1. Location of subsurface gasoline-spill study site in Galloway Township, N.J.
HYDROCARBON CONTAMINATION 

GEOCHEMISTRY

Results of initial studies at the site revealed that 

gasoline was present as a separate-phase liquid on the 

regional water table in the immediate vicinity of well 

DEP-1 and on the perched water table in the vicinity of 

the exhumed tank (fig. 1). Dissolved hydrocarbons 

identified include benzene, toluene, C2 -Cr, C4-alkylbenzenes, and napthalene (Baedecker and others, 1991; 

Phinney and Cozzarelli, 1996). The areawide distribution of contaminants was described by Baedecker and 

others (1991) and Fischer, Smith, and Baehr (1996). 

Contamination in perched water has persisted at high 

concentrations in the vicinity of the exhumed tank. 

Hydrocarbon contamination on the regional water table 

is downgradient from the source and concentrations 

vary with time, depending on the direction of groundwater flow. The concentration of total aromatic hydrocarbons decreases downgradient, as a result of dilution, 

volatilization, and biodegradation.

Ground water was sampled and analyzed to identify biogeochemical processes occurring at the site and 

to identify degradation pathways. Baedecker and others (1991) and Cozzarelli and others (1991), provide 

descriptions of the sampling methods. Results of analyses for inorganic and organic chemical compounds 

indicate active microbial degradation of hydrocarbons 

in ground water. These data are reported by Baedecker 

and others (1991), Cozzarelli and others (1991), Cozzarelli (1993), and Cozzarelli and Baedecker (1996).

Aerobic respiration and reduction of nitrate, sulfate, and iron oxides by microbes have been identified 

as important processes for degrading hydrocarbons at 

the site (Baedecker and others, 1991; Cozzarelli and 

others, 1991). The perched water has been depleted of 

dissolved oxygen and nitrate, leaving sulfate and iron 

reduction as the main biodegradation processes. In the 

regional aquifer, aerobic respiration and nitrate reduction appear to be the main hydrocarbon-degradation 

pathways as evidenced by results of multilevel sampling beneath the water table. Results of microbial 

counting showed higher numbers of hydrocarbondegrading microorganisms in contaminated ground 

water and sediment than in uncontaminated ground 

water and sediment (Mills and Randall, 1991). The fate 

of organic acids, which are intermediate hydrocarbondegradation products, depends on the availability of 

electron acceptors for anaerobic degradation processes 

(Cozzarelli, 1993).

ESIMATION OF MICROBIAL DEGRADATION 

RATES OF HYDROCARBONS ON THE 

BASIS OF GAS TRANSPORT RATES IN 

THE UNSATURATED ZONE

Environmental regulators concerned with the 

effect of gasoline spills on ground-water quality have 

recently acknowledged that microbial degradation of 

hydrocarbons is a significant mechanism for attenuating subsurface gasoline contamination. Methods for 

quantifying degradation rates at a given site, however, 

have not been established. The approach developed for 

the Galloway Township site is to identify stoichiometric relations among oxygen consumption, carbon dioxide production, and hydrocarbon degradation. Once 

these relations are established, fluxes of oxygen and 

carbon dioxide through the unsaturated zone are used 

to estimate in situ biodegradation rates.

Baehr and others (1991) provide an overview of 

this method. Oxygen diffuses toward the water table 

where it is used by microbes to aerobically degrade 

hydrocarbons. Carbon dioxide, which is an end product 

of this degradation, diffuses toward the surface. At the 

Galloway Township site, hydrocarbons are not 

detected in the gaseous phase in most parts of the unsaturated zone, whereas the underlying perched water is 

virtually anoxic and contains high concentrations of 

dissolved hydrocarbons (Cozzarelli and others, 1991; 

Cozzarelli, 1993). These data, combined with the presence of an abundance of hydrocarbon-degrading bacteria at the water table (Mills and Randall, 1991), 

indicate that aerobic hydrocarbon degradation occurs 

near the top of the perched water. Thus, the rates of 

oxygen and carbon dioxide transport are related to the 

rate of aerobic degradation of hydrocarbons.

The relation between the rates of oxygen consumption and carbon dioxide production by hydrocarbondegrading bacteria is complicated by the oxidation of 

ferrous iron. Also, the presence of intermediate metabolites, such as organic acids, detected in the ground 

water (Cozzarelli, 1993) indicates that aerobic degradation is not always complete. If these pathways can be 

neglected for the purpose of determining an approximate stoichiometry, then complete hydrocarbon mineralization (for example, for toluene: C7H8 + 9O2 » 

7CO2 + 6H2O) may provide a working approximation.
This working approximation was substantiated in the 

laboratory by conducting respiration experiments with 

hydrocarbon-contaminated sediment (Noland, 1993; 

Noland and others, 1996). Furthermore, Lahvis (1993) 

determined that the ratio of oxygen and carbon dioxide 

fluxes in the unsaturated zone above contaminated 

perched water at the Galloway Township site was close 

to that predicted by using the mineralization working 

approximation.

A value for the effective diffusion coefficient for a 

constituent in unsaturated porous media is required for 

calculating diffusion rates through the unsaturated zone. 

Cores of sediment were extracted from the unsaturated 

zone at the Galloway Township site and a technique was 

developed for estimating effective diffusion coefficients 

(Fischer, Baker, and Baehr, 1996). The method involves 

measuring changes in tracer concentration along the 

length of the core. A mathematical model developed by 

Baehr and Bruell (1990) is used to calculate the effective 

diffusion coefficient.

Two types of field experiments were conducted to 

evaluate hydrocarbon-biodegradation rates. The first 

experiment involves measuring oxygen and carbon 

dioxide profiles at the site under natural conditions prior 

to remediation. The second experiment involves monitoring the redistribution of gases after a period of vapor 

extraction has caused oxygen concentrations to increase 

and carbon dioxide concentrations to decrease in the 

unsaturated zone to near-atmospheric levels. A vaportransport model developed by Lahvis (1993) is used to 

compute rates of gas movement from gas-phase concentration data. Lahvis and others (1996) conclude that the 

rates of gas movement at the Galloway Township site 

indicate that significant rates of hydrocarbon degradation may be limiting the spread of the contaminant 

plume in ground water.

Water-quality data collected by sampling near the 

perched water table indicate that the aerobic degradation 

rate is highest in a small vertical interval between the top 

of the capillary zone and 1 ft below the water table. This 

information leads to the formulation of the hypothesis 

that coexisting conditions of high moisture content and 

high concentrations of oxygen and hydrocarbons, which 

prevail only within this small vertical interval, are more 

favorable for the aerobic degradation of hydrocarbons 

than those in the relatively dry overlying unsaturated 

zone and underlying anoxic ground water. Construction 

of a field experiment to study these conditions over the 

indicated spatial scale was not possible as a result of 

perched water-table altitude variations. Therefore, a laboratory method to emulate biochemical conditions near 

the water table was developed (Baker, 1993). Sediment

collected just above the water table was packed in glass 

columns. The upper part of the column was unsaturated 

and the column water level was maintained at a constant 

altitude. Hydrocarbon concentration was maintained as 

a constant level in the saturated lower part of the column. 

Concentrations of hydrocarbons, oxygen, and other 

constituents can be controlled by manipulating the end 

reservoirs. Analysis of these experiments with a onedimensional vapor-transport model is reported by 

Baker (1993) and Baker and Baehr (1996). Hydrocarbon-degradation rates within sediments maintained 

near capillary-zone moisture conditions were an order 

of magnitude higher than those in the drier overlying 

unsaturated sediments. Neither hydrocarbon concentrations nor oxygen concentrations were limiting 

factors in these experiments.

VAPOR-EXTRACTION-REMEDIATION 

DESIGN

Remedial action for spills of petroleum products 

generally include an effort to physically recover product 

accumulated on the water table by pumping or bailing. 

Substantial portions of the spill typically remain trapped 

in residual saturations in the unsaturated and capillary 

zones as well as in a smear zone in shallow ground water 

caused by fluctuating water-table altitudes. This part of 

the spill can pose a long-term threat to ground-water 

quality, because it is a source of hydrocarbons that can 

be transported in aqueous and gaseous phases. Vaporextraction technology has been applied successfully to 

remove additional hydrocarbons from the residual saturation source. The success of this technology is based on 

the significant volatility of gasoline hydrocarbons and 

the ability to induce an air-flow field in the unsaturated 

zone with combinations of dry wells and trenches.

Mathematical models of the air-flow field can be 

used to help design a vapor-extraction system. The 

ground-water-flow simulator MODFLOW (McDonald 

and Harbaugh, 1988) has been adapted for this purpose. 

This code, called AIR3D, and a users' manual currently 

are in preparation for distribution in the public domain. 

Application of the air-flow code results in the definition 

of the air-flow field corresponding to a chosen configuration of extraction and injection locations. Welty and 

others (1992; 1996) extended the modeling tool by 

coupling the three-dimensional air-flow model with 

mathematical programming to define an optimal configuration of extraction and injection locations. Publication 

of the computer code that implements this coupling, 

called OPTAIR, and a users' guide in the public domain 

is planned.
The permeability of unsaturated porous media to 

air is required information for air-flow simulations. 

Baehr and Huh (1991) developed analytical solutions 

to the equation defining air flow to or from a single well 

screened in an unsaturated zone to determine air-phase 

permeability from results of pneumatic tests conducted 

at a research site near Bemidji, Minn. A computer code 

that implements the analytical solutions has been 

developed. The code, called AIR2D, and a users' manual is currently in preparation for publication in the 

public domain. Joss and others (1992) report on applications of this parameter-estimation technique at the 

Galloway Township site.

A mathematical model for simulating chemical 

transport associated with vapor extraction and bioventing has been developed by Joss (1993). The model can 

be used to simulate three-dimensional, coupled transport of multiple chemical species and is expected to be 

a useful tool in analyzing and designing vapor-extraction, bioventing, and air-sparging remediation systems. 

This model is summarized by Joss and Baehr (1996).

REFERENCES

Bacdcckcr, M.J., Cozzarclli, I.M., and Phinncy, C.S., 1991, 

Distribution of organic and inorganic constituents in 

ground water at Galloway Township, New Jersey, in 

Mallard, G.E., and Aronson, D.A., cds., U.S. Geological Survey Toxic Substances Hydrology Program-- 

Proceedings of the technical meeting, Montcrcy, California, March 11-15, 1991: U.S. Geological Survey 

Water-Resources Investigations Report 91-4034, 

p. 287-293.

Bachr, A.L., and Brucll, C.J., 1990, Application of the 

Stcfan-Maxwcll equations to determine limitations of 

Pick's law when modeling organic vapor transport in 

sand columns: Water Resources Research, v. 26, no. 6, 

p. 1155-1163.

Bachr, A.L., Fischcr, J.M., Lahvis, M.A., Baker, R.J., and 

Smith, N.P., 1991, Method for estimating rates of 

microbial degradation of hydrocarbons based on gas 

transport in the unsaturated zone at a gasoline-spill site 

in Galloway Township, New Jersey, in Mallard, G.E., 

and Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Montcrcy, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 250-255.

Baehr, A.L., and Hull, M.F., 1991, Evaluation of unsaturated zone air permeability through pneumatic tests: 

Water Resources Research, v. 27, no. 10, p. 2605-2617.

Baker, R.J., 1993, Hydrocarbon microbial degradation rate 

determination in columns of unsaturated porous media 

using vapor transport modeling: Philadelphia, Pa., 

Drcxcl University, Environmental Studies Institute, 

unpublished Ph.D. dissertation, 237 p.

Baker, R.J., and Bachr, A.L., 1996, Use of column studies 

and a reactive transport model to measure biodcgradation rates of hydrocarbons, /'/; Morganwalp, D.W., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program--Proceed ings of the 

technical meeting, Colorado Springs, Colo., September 

20-24. 1993: U.S. Geological Survey Water-Resources 

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Baker, R.J., Fischcr, J.S., Smith, N.P.. Kochnlcin, S.A., and 

Bachr, A.L., 1991, Gas-chromatographic methods for 

determining the composition of the vapor phase in gasoline-contaminate soils, /'/; Mallard, G.E., and Aronson. 

D.A., cds., U.S. Geological Survey Toxic Substances 

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Cozzarclli, I.M., 1993, The biogcochcmical fate of organic 

acids in a shallow aquifer contaminated with gasoline: 

Charlottcsvillc, University of Virginia, Department of 

Environmental Sciences, unpublished Ph.D. dissertation, 240 p.

Cozzarclli, I.M., and Bacdcckcr, M.J., 1996, Biogcochcmical fate of organic acid metabolites produced during the 

microbial degradation of hydrocarbons in ground 

water, in Morganwalp, D.W., and Aronson, D.A., cds., 

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Cozzarclli, I.M., Bacdcckcr, M.J., and Hopple, J.A., 1991, 

Gcochcmical gradients in shallow ground water caused 

by the microbial degradation of hydrocarbons at Galloway Township, New Jersey, /'/; Mallard, G.E., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

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technical meeting, Montcrcy, California, March 11-15, 

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Fischcr, J.M., Bachr, A.L., and Smith, N.P., 1991, An investigation of the transport and microbial degradation of 

hydrocarbons in the subsurface at a gasoline-spill site 

in Galloway Township, New Jersey, //; Mallard, G.E., 

and Aronson, D.A., cds., U.S. Geological Survey Toxic 

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technical meeting, Montcrey, California, March 11-15, 

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1996, Detenuination of vapor-phase diffusion coefficients for unsaturatcd zone sediments at a gasoline-spill 

site in Galloway Township, New Jersey, in Morganwalp, D.W., and Aronson, D.A., cds., U.S. Geological 

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Gibs, J., Brown, G.A., Turner, K.S., Maclcod, C.L., Jclinski, J.C., and Kochntcin, S.A., 1993, Effects of smallscale vertical variations in well-screen inflow rates and 

concentrations of organic compounds on the collection 

of representative ground-watcr-quality samples: 

Ground Water, v. 31, no. 2, p. 201-208.

Joss, C.J., 1993, Multispccics vapor transport and reactions 

in unsaturatcd porous media under conditions of 

induced airflow: Philadelphia, Pa., Drcxcl University, 

Department of Civil Engineering, unpublished Ph.D. 

dissertation, 103 p.

Joss, C.J., and Bachr, A.L., 1996, A multispccics transport 

model to simulate vapor extraction and biovcnting 

remediation of unsaturatcd zones contaminated with 

gasoline, in Morganwalp, D.W., and Aronson, D.A., 

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technique for determining unsaturatcd zone air-permeability, in DiGiulio, D., cd., Proceedings of the symposium on soil venting, April 29-May I, 1991, Houston, 

Texas, EPA/500/R-92/174, 334 p.

Lahvis, M.A., 1993, Estimating rates of microbial degradation of hydrocarbons at a gasoline spill site based on 

vapor diffusion in the unsaturated zone: Philadelphia, 

Pa., Drcxcl University, Department of Civil Engineering, unpublished Ph.D. dissertation, 204 p.

Lahvis, M.A., Bachr, A.L., and Fischcr, J.M., 1996, Method 

for estimating the rate of microbial degradation of 

hydrocarbons associated with biovcnting at a gasolinespill site in Galloway Township, New Jersey, //; Morganwalp, D.W.Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

95-4015.

McDonald, M.G., and Harbaugh, A.W., 1988, A modular 

three-dimensional finite difference ground-water flow 

model: U.S. Geological Survey Techniques of WaterResources Investigations, chap. Al, book 6, 586 p.

Mills, A.L., and Randall, S.E., 1991, Abundance and distribution of microbial guilds in a subsurface plume of 

gasoline contamination at Galloway Township, New 

Jersey, in Mallard, G.E., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, Monterey, 

California, March 11 -15, 1991: U.S. Geological Survey 

Water-Resources Investigations Report 91-4034, 

p. 263-267.

Nolan, J.G., 1993, Relationship between hydrocarbon biodcgradation, oxygen consumption and carbon dioxide 

production in unsaturatcd porous media: Philadelphia, 

Pa., Drcxcl University, Department of Civil Engineering, unpublished master's thesis, 93 p.

Nolan, J.G., Baker, R.J., and Bachr, A.L., 1996, Method for 

monitoring biodcgradation of hydrocarbons in unsaturatcd porous media, in Morganwalp, D.W., and Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 95-4015.

Phinncy, C.S., and Cozzarclli, I.M., 1996, A comparison of 

two analytical techniques for identifying and measuring 

the concentrations of volatile hydrocarbons in ground 

water, in Morganwalp, D.W., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

95-4015.

Wclty, C., Joss, C.J., and Baehr, A.L., 1992, OPTAIR-A 

three dimensional airflow model coupled with mathematical programming to optimize vapor extraction of 

the unsaturatcd zone: unpublished final report submitted to American Petroleum Institute, October 26, 1992, 

57 p.

Wclty, C. Joss, C.J., and Bachr, A.L., 1996, Optimizing the 

design of vapor-extraction remediation systems for 

removal of organic contaminants from the unsaturatcd 

zone, in Morganwalp, D.W., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

95-4015.
Hydrogeology, Contaminant Distribution, and

Biodegradation Processes at a Gasoline-Spill

Research Site in Galloway Township, New Jersey

By Jeffrey M. Fischer1 , Nicholas P. Smith 1 , and Arthur L. Baehr 1

Abstract

The hydrogeology of the Toxic Substances 

Hydrology Program research site in Galloway 

Township, New Jersey, was characterized to determine factors controlling hydrocarbon-contaminant distribution and biodegradation. Sediments 

beneath the site are predominantly sand and clay. 

A localized zone of perched water forms above the 

regional water table. The direction of regional 

ground-water flow in the unconfined aquifer has 

varied from slightly north of east to east-southeast. 

Horizontal hydraulic gradients have been less than 

0.005 since 1988 and estimates of average horizontal linear velocities range from 0.04 to 0.36 feet 

per day. A monitoring well provides a conduit for 

water to flow down from the perched water table 

zone to the unconfined part of the regional aquifer. 

The direction of ground-water flow near this well 

differs from regional flow directions.

Gasoline as a separate-phase liquid is floating on the regional water table in the leaky monitoring well and in the perched water zone near the 

location from which the tank was exhumed. Dissolved hydrocarbons are present in the perchedwater zone within a stable elliptical plume about 

80 feet long and 40 feet wide in the 11 to 14 foot 

depth interval. Hydrocarbon contamination in the 

unconfined aquifer was detected downgradient 

from the leaky well, but contaminant location varied with changes in the direction of ground-water 

flow. The concentration of total aromatic hydrocarbons decreased downgradient as a result of 

dilution, dispersion, and biodegradation. Groundwater flow affects the types of hydrocarbondegradation processes and, thus, the locations in 

which hydrocarbon contaminants are found.

'U.S. Geological Survey, West Trenton, N.J.

INTRODUCTION

Contamination of ground water with gasoline 

hydrocarbons is a serious problem throughout the 

United States. In the subsurface the constituents of 

gasoline move as solutes, as vapors, or as an 

immiscible phase, and can undergo microbial 

degradation. An understanding of the processes that 

control the transport and biodegradation of gasoline is 

needed to provide a strong scientific basis for 

evaluating the threat to ground-water supplies from 

gasoline spills.

In 1984, gasoline was found to have leaked from 

an underground storage tank at a farm in Galloway 

Township, New Jersey (fig. 1). Investigations by State 

agencies revealed a localized area of shallow ground 

water contaminated with gasoline hydrocarbons. No 

contamination was detected in offsite shallow domestic 

wells or in onsite deep irrigation wells. The leaking 

tank was removed in May 1986, and contaminated soils 

were excavated to a depth of 8 ft from an area about 

80 ft long and 25 ft wide (fig. 1). The excavation was 

backfilled with uncontaminated sediments.

In 1988, the U.S. Geological Survey (USGS) 

began investigations at the site. Objectives of these 

investigations are described by Fischer and others 

(1991) and Baehr and Fischer (1996). This paper 

describes the hydrogeology and geochemistry of the 

research site, and discusses the relations among the 

hydrogeology, contaminant distribution, and 

biodegradation processes at the site.

Since the Galloway Township farm was chosen 

as a USGS research site in July 1988, an extensive 

network of multilevel probes and wells has been 

installed for observing small-scale vertical chemical 

gradients immediately above and below the water 

table. The ground-water-monitoring network (fig. 1) 

consists of 13 monitoring wells, 2 multilevel groundwater samplers, 27 nets of submerged vapor probes, 

and 5 vapor-extraction wells. To identify dissolved 

hydrocarbons and inorganic constituents that are
X

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EXPLANATION

 Nested vapor probes

^ U.S. Geological Survey ground-wster

observation well 

<S New Jersey Department of Environmental

Protection ground-water observation well

__ / \ ___________

9 Multilevel ground-wa

« Neutron probe acces

O Thermistor nest

8) Vapor-extraction we

Figure 1. Location of Galloway Township research site showing former tank location, area of excavation, and ephemeral draina

(Section A-A' shown in fig. 2.)
indicators of hydrocarbon biodegradation, groundwater samples were collected in June 1989, January 

1990, July 1990, and April 1991. The monitoring 

network is described in detail by Fischer and others 

(1991). Ground-water sampling and analytical 

procedures, as well as a detailed discussion of the 

ground-water chemistry, can be found in Baedecker 

and others (1991) and Cozzarelli and others (1991).

HYDROGEOLOGY

The research site is located in a sparsely 

populate agricultural section of Galloway Township, in 

Atlantic County, New Jersey. The area is relatively flat 

and no perennial streams or surface impoundments are 

present within 0.25 mi of the site. During periods of 

heavy rain, runoff is channeled across the site in an

ephemeral drainage channel (fig. 1). Measured 

precipitation at the research site was 40,45, and 39 in. 

in 1990, 1991, 1992, respectively.

The site is situated in the Coastal Plain 

physiographic province and is underlain by seawarddipping wedges of unconsolidated sediments 

composed of sand, silt, and clay. Near-surface 

sediments consist of fluvial sands and clays. The sands 

are predominantly quartz with minor microcline and 

ilmenite. Clays near the surface are mostly gibbsite, 

whereas deeper clays are kaolinitic (Owens and 

Minard, 1979). Core samples and geologic logs 

obtained during installation of the vapor probes and 

wells were used to describe the lithology within 25 ft of 

the land surface (fig. 2). Records of domestic and 

irrigation wells at the site and on adjacent properties 

were used to examine the lithology at greater depths 

and to assess the areal extent of the lithologic units.

---10

Clayey-organic soil

Brown-clayey sand

Medium well-sorted

of sand and clay Alternating layers 3/89 v ?__- - _

~~~~

Fine clayey sand

V 1/89

EXPLANATION 

? Geological contact:

Queried where unknown

 Perched water table

 Regional water table

1 Depth of borehole

0123456 METERS

-30

-25 ±

r.-15

t-20

DATUM IS SEA LEVEL VERTICAL EXAGGERATION x 4

Figure 2. Vertical section through the research site illustrating the generalized geohydrology within 25 feet of 

land surface and altitudes of regional and perched water tables in march 1989 and October 1989. (Location of 

section shown in fig. 1.)
The research site overlies the Cohansey aquifer. 

The ground-water-flow system beneath the site 

consists of two hydrologic units that were studied in 

detail: a zone of perched water of local extent and an 

unconfined aquifer of regional extent (hereafter 

referred to as the "regional aquifer). The perched-water 

zone forms above the dense clay stringers that are 

present at depths of 14 to 19 ft (fig. 2). Measurements 

of moisture contents less than saturation beneath the 

clay stringers and the ability to draw vapors from 

probes situated between the base of the clay stringers 

and the top of the regional aquifer confirm that the 

clays are responsible for the presence of the perched 

water. Records of offsite wells indicate that the upper 

part of the regional aquifer is separated from the lower 

part by an areally extensive, 20-ft-thick clay layer, the 

upper surface of which is at a depth of about 40 ft. 

Because initial study results from the site indicated that 

no hydrocarbon contaminants were present in the lower 

part of the regional aquifer, USGS studies were limited 

to the perched-water zone and the upper part of the 

regional aquifer.

The depth of the regional water table has varied 

from 13.8 to 20.9 ft below land surface since waterlevel measurements were first made at the site in March 

1988. The horizontal hydraulic gradients has usually 

been less than 0.005. Directions of ground-water flow 

vary in response to precipitation and irrigation 

withdrawals. Since May 1989, the usual ground-water 

flow direction has been to the east, and normally has 

ranged from slightly north of east to east-southeast 

(fig. 3). Occasionally, flow has been to the eastnortheast or southeast. Results of slug tests conducted 

at monitoring wells in the regional aquifer, and 

analyzed according to the method of Hvorslev (1951), 

indicate that horizontal hydraulic conductivities range 

from about 2 to 30 ft/d. Hydraulic conductivities at two 

adjacent wells, DEP-1 and EVW8D, were lower than 

those measured at other wells (about 0.2 ft/d). Well 

records indicate that these two wells may be screened 

in a clayey part of the regional aquifer. On the basis of 

hydraulic-conductivity measurements, hydraulicgradient data, and an average porosity of 0.41, the 

average linear flow velocity in the regional aquifer was 

estimated to range from about 0.04 to 0.36 ft/d.

The depth of the perched water table varied from 

8.0 to 12.4 ft below land surface from May 1990 to 

December 1992. The perched water table is highest 

near the ephemeral drainage channel (fig. 1). Because 

runoff infiltrates along this channel, the perched water

table slopes away from the channel and ground water 

flows perpendicularly away from it (to the northwest 

and southeast; fig. 2). Slug tests conducted in the 

perched-water zone indicate the horizontal 

conductivities of the sands that are present at depths of 

4 to 11 ft range from 8 to 30 ft/d. Deeper interbedded 

sands and clays, at a depth of 11 to 14 ft, are presumed 

to have lower horizontal hydraulic conductivities. The 

perched-water zone varies in areal extent and thickness 

with precipitation, but it usually is present within a 70- 

ft radius of the location of the exhumed gas tank. 

During dry periods, the southeastern part of the 

perched-water zone, in the vicinity of well USGS-5 

(fig. 1), disappears entirely. Several factors may 

account for this observation. The slope of the clay 

stringers may cause ground water to flow toward the 

northwest during dry periods (fig. 2); sand units with 

relatively high hydraulic conductivity may grade into 

clayey sands of lower hydraulic conductivity; or the 

characteristics of the clay stringers, such as their 

thickness or hydraulic conductivity, may vary and 

cause changes in percolation rates.

Perched water eventually leaks down to the 

regional aquifer; however, the degree of connection 

between the perched-water zone and the regional 

aquifer is uncertain. On a seasonal basis the regional 

water table rises and completely saturates the zone 

beneath the clay stringers. Baedecker and others (1991) 

concluded, on the basis of chemical gradients, that the 

clay stringers hinder mixing between the perchedwater zone and the regional aquifer. Measured 

hydraulic conductivities of cores indicate that the clays 

are relatively impermeable (10-5 to 10-7 ft/d). During 

the study period, water levels in the perched-water 

zone were 5.3 to 9.6 higher than those measured in the 

regional aquifer. This difference produced very high 

vertical gradients; however, the low permeability of the 

clay stringers limits contact between the two water 

zones. Estimated vertical fluxes through the clay 

stringers were less than 1 in/yr and flow velocities were 

less than 2 in/yr.

The annular space around well DEP-1, a 

monitoring well installed prior to USGS investigations, 

appears to provide a conduit for the downward flow of 

water and gasoline from the perched-water zone to the 

regional aquifer. Evidence for this connection includes 

the following: The record of well DEP-1 indicates that 

the top of the screen is in the clay stringers and that the 

sand pack extends up into the perched-water zone; 

water-level altitudes in well DEP-1 are consistently

10
EXPLANATION

Water-level contour-Shows 

in feet. Dashed where app

is variable. Datum is sea l

Location of monitoring wel

measurement

0 5 10 15 

Figure 3. Water-level altitudes at the Gajloway Township research site in January 1990, July 1990, and April 1991. (Water levels in 

0.75, 0.70, and 1.10 feet higher than regional water levels in January 1990, July 1990, and April 1991, respectively.)
higher (from 0.3 to 1.5 ft, depending on time of year) 

than those in nearby wells (fig. 3), and water-level data 

from vapor probe borings indicate that the perched 

water table is depressed in the vicinity of well DEP-1. 

Elevated regional water levels near this well indicate 

that local directions of ground-water flow are radially 

away from the well and vary significantly from 

regional flow directions (fig. 3). In addition, groundwater flow rates must be greater in the vicinity of this 

well than elsewhere in the regional aquifer, and vertical 

How must be significant. The amount of leakage varies 

over time as vertical gradients changes.

CONTAMINANT DISTRIBUTION AND 

DEGRADATION PROCESSES

From the beginning of the field study in 

September 1989 until December 1991, gasoline as a 

separate-phase liquid was detected only on the regional 

water table in well DEP-1 (fig. 1). From September 

1989 to March 1990, the thickness of gasoline in the 

well decreased from about 1.2 ft to less than 0.05 ft. 

From April 1990 through January 1991, only a slight 

sheen of gasoline was detectable in the well; minor 

amounts (less than 0.03 ft) reappeared in the well from 

February 1991 through October 1991. No separatephase gasoline has been detected in well DEP-1 since 

January 1992.

Vapor sampling, begun in July 1989, indicated 

the presence of separate-phase gasoline on the perched 

water table in the area where the tank had been (fig. 1). 

Much of the separate-phase gasoline has been removed 

from the perched-water zone as a result of vaporextraction experiments begun in October 1990. (See 

Baehr and Fischer (1996) for a description of vaporextraction experiments at the site.) Hydrocarbon 

vapors have not been detected above the perched-water 

zone since August 1992; it is presumed that very little 

separate-phase gasoline has been present since then.

Aqueous-phase-hydrocarbon contamination in 

the perched-water zone consists of a small elliptical 

area about 80 ft long by 40 ft wide at a depth of 1 1 to 

14 ft (fig. 4). Studies by Baedeckerand others (1991) 

indicate that the total concentration of benzene, 

toluene, ethylbenzene, o-xylene, m-xylene, and pxylene (BTEX) within this area usually exceeded 2 mg/ 

L. Hydrocarbon concentrations in the center of this 

plume have remained fairly stable over time, possible 

as a result of low rates of ground-water flow in the

interbedded sand and clay layers at the base of the 

perched-water zone. Hydrocarbon concentrations 

decrease toward the sides and top of the plume, partly 

as a result of dilution. Hydrocarbon concentrations in 

the southeastern limb of the plume have decreased 

considerably over time (tig. 4).

In the regional aquifer, concentrations of 

aqueous-phase hydrocarbons and their areal 

distribution have varied substantially during the study 

period as a result of changes in ground-water flow 

directions. Generally, hydrocarbon contaminants were 

detected downgradient from areas where ground water 

leaks through the clay stringers. These areas usually 

were downgradient from well DEP-1. Thus the areal 

distribution of BTEX was east of well DEP-1 in 

January 1990, southeast of the well in July 1990, and 

east of the well in April 1991 (fig. 5). In July 1990, the 

plume was narrow, possibly because the direction of 

ground-water flow was relatively constant for the 4 

months prior to sampling. The plume was much 

broader in July 1990 than it was during the other two 

sampling periods possibly because the direction of 

ground-water flow was highly variable during the 4 

months prior to sampling. The plume location is 

consistent with flow directions if solute-transport times 

are considered. Total BTEX concentrations in the 

center of the plume in the regional aquifer often exceed 

10 mg/L (fig. 5). The concentration of total aromatic 

hydrocarbons decreases downgradient, in part as a 

result of dilution and dispersion. Compositional 

changes from lighter weight to heavier weight 

aromatics are indicators of volatilization and 

biodegradation along the flow path. The contaminant 

plume flows 5 to 10 ft beneath the top of the regional 

water table and is usually less than 10 ft thick 

(Baedecker and others, 1991; Gibs and others, 1993). 

Except in areas where leakage occurs, contaminant 

flow is predominantly horizontal with a slight 

downward component, possibly in response to 

recharge. The downgradient extent of the plume is 

unknown because it extends beyond the monitoring 

wells (a maximum distance of 80 ft from well DEP-1).

Although the plume in the regional aquifer 

forms downgradient from the contaminant source, it 

has variegated appearance. Zones of highly 

contaminated water and slightly contaminated water 

alternate across the plume. The submerged vapor 

probes and wells used for ground-water sampling are 

screened at different depths and were designed for 

sampling water from discrete 6-in. to 24-in. intervals of

12
\

Formej>tank 

location

Maximum extent 

of BTEX detection

\

HOUSE

\

«

>

10

i

20

I

30

I

40

I

I 

10

50 FEET

I ~T 

15 METERS

EXPLANATION

2.0 Line of equal concentration of 

benzene, toluene, ethylbenzene, 

o-xylene, m-xylene, and p-xylene 

(BTEX) in milligrams per liter.

O Monitoring well

 Submerged vapor probe

Figure 4. Change in location of 2-milligram-per-liter concentration contour for total benzene, toluene, ethylbenzene, O-xylene, m-xyl

depth of 11 to 15 ft below land surface in the perched-water zone at the Galloway Township research site from July 1990 to April 19

in milligrams per liter. 1990 data from Baedecker and others, 1991; 1991 data from M.J. Baedecker and I.M. Cozzarelli, U.S. Geolog

1993.)
Former tank location

\\

January 1990

Former tank location

\\

Former tank location

\\

\

\

oO.O

April 1991

EXPLANATIO

 2.0 Line of equal concentration 

ethylbenzene, o-xylene, m

in milligrams per liter. D

-> Approximate local directio

well DEP-1

Monitoring well 

Submerged vapor probe 

Multi-level sampler

0 10 2O 30 40 S

0 5 10 15 

Figure 5. Concentrations of total benzene, toluene, ethylbenzene, o-xylene, m-xylene, and p-xylene (BTEX) at an altitude of 33 t

Galloway Township research site in January 1990, July 1990, and April 1991. (Concentrations shown are in milligrams per liter. 1

1991; 1991 data from M.J. Baedecker and I.M. Cozzarelli, U.S. Geological Survey, written commun., 1993.)
the aquifer. Cozzarelli (1993) has demonstrated 

significant vertical heterogeneity in hydrocarbon 

concentration over short vertical distances. Radial flow 

around well DEP-1 could account for some of the 

observed contaminant distribution; however, 

determination of flow rates and flow directions is 

difficult as a result of steeper vertical gradients and 

lower hydraulic conductivities in the vicinity of this 

well than elsewhere at the site. The observed 

contaminant distribution also could be attributed to 

natural or borehole-induced leakage in the vicinity of 

vapor well VW-9. Another possibility is that separatephase gasoline is trapped beneath the clay stringers and 

moves in response to changing flow directions. All 

these possibilities are currently being investigated.

Hydrologic conditions at the site also control the 

distribution of inorganic constituents used in 

microbially mediated oxidation-reduction reactions. 

Microbially mediated reactions identified at the site 

include aerobic respiration, denitrification, nitrate 

reduction, sulfate reduction, and iron reduction 

(Baedecker and others, 1991; Cozzarelli and others, 

1991; Cozzarelli, 1993). Oxygen, nitrate, and sulfate 

are supplied to the plume in the perched-water zone by 

vertical percolation of recharge and reach the plume in 

the regional aquifer by horizontal flow (Cozzarelli and 

others, 1991). At the base of the perched-water plume, 

in an area of low rates of ground-water flow and high 

concentrations of hydrocarbons, oxygen and nitrate 

have been depleted and sulfate and iron reduction are 

the main biodegradation processes (Cozzarelli, 1993). 

At the edges of this plume, where recharge mixes with 

contaminated water, aerobic respiration, nitrate 

reduction, and denitrification occur. In the regionalaquifer plume, lower concentrations of oxygen and 

nitrate and higher concentrations of ammonia than 

those in uncontaminated water indicate that aerobic 

respiration, denitrification, and, possibly, nitrate 

reduction are the main biodegradation processes. On 

occasion, anoxic zones have formed in the regionalaquifer plume. This may be related to the rate of 

contaminant input and the length of time that one 

ground-water flow direction has been maintained. 

Changes in flow direction facilitate mixing of 

contaminants and oxygenated water and may preclude 

the formation of anoxic zones.

Water-chemistry data show that the 

geochemical environment in the regional aquifer is less 

stable than that in the perched-water zone (Baedecker 

and others, 1991). Reduced chemical species, such as

reduced iron and hydrogen sulfide, are produced in 

anoxic environments and typically are not stable in 

oxygenated water. Leakage of anoxic perched water to 

the regional aquifer, either at well DEP-1 or other 

locations, may explain some of the instability. Also, 

anoxic zones may form at different locations in the 

regional aquifer at various times, depending on the 

recent stability of the regional flow direction.

SUMMARY

The USGS began investigations in 1988 at a site 

in Galloway Township, New Jersey, at which ground 

water had been contaminated by gasoline leaking from 

an underground storage tank. Sediments at the site, 

which is located in the New Jersey Coastal Plain, 

consist of horizontal layers of sand and clay. The two 

uppermost units of the ground-water-flow system, the 

perched-water zone and the unconfined regional 

aquifer, each contain a plume of gasoline-contaminated 

ground water. The areal extent and thickness of the 

perched-water zone varies with precipitation. The clay 

stringers that are responsible for the perching appear to 

be relatively impermeable in the northwestern part of 

the research site and leaky in the southeastern part. 

Monitoring well DEP-1, installed prior to USGS 

investigations, provides a conduit for water to flow 

down from the perched-water zone to the regional 

aquifer. The general direction of flow in the regional 

aquifer usually is to the east, but the direction varies 

from slightly north to east to east-southeast. Hydraulic 

gradients are less than 0.05, and estimates of average 

horizontal linear flow velocities range from 0.04 to 

0.36 ft/d. Elevated water levels in well DEP-1 indicate 

that directions and rates of ground-water flow near this 

well vary significantly from regional flow directions.

Gasoline as a separate-phase liquid has been 

detected in only two places: in the regional aquifer in 

well DEP-1 and in the perched-water zone near the 

location of the exhumed tank. Dissolved-hydrocarbon 

contamination in the perched-water zone consists of a 

stable elliptical plume about 80 ft long and 40 ft wide 

in the 11 to 14 ft depth interval. Ground-water flow 

rates are very low where contaminants currently reside 

in the perched-water zone. Hydrocarbon contamination 

in the regional aquifer was downgradient from well 

DEP-1. The location of the contaminant plume varied 

with changes in ground-water flow directions. The 

thickness of the plume in the regional aquifer usually is

15
less than 10 ft, and hydrocarbon-contaminant flow is 

predominantly horizontal. The concentration of total 

aromatic hydrocarbons decreased downgradient, in 

part as a result of dilution, dispersion, and 

biodegradation. Hydrocarbon degradation is affected 

by the rate of movement of organic and inorganic 

chemical species in the ground water. Natural 

attenuation mechanisms are significant throughout the 

site and account for many of the observed changes in 

water quality.

REFERENCES CITED

Bacdcckcr, M.J., Cozzarclli, I.M., and Phinncy, C.S., 1991, 

Distribution of organic and inorganic constituents in 

ground water at Galloway Township, New Jersey, in 

Mallard, G.E., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Montcrcy, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 287-293.

Bachr, A.L., and Fischcr, J.M., 1996, Overview of research 

on the transport, microbial degradation, and 

remediation of hydrocarbons at a subsurface gasolinespill site in Galloway Township, New Jersey in 

Morganwalp, D.W, and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting. 

Colorado Springs, Colorado, September 

20-24, 1993: U.S. Geological Survey Water Resources 

Investigations Report 94-4015.

Baker, R.J., Fischcr, J.M., Smith, N.P., Kochnlcin, S.A., 

and Bachr, A.L., 1991, Gas-chromatographic methods 

for determining the composition of the vapor phase in 

gasoline-contaminated soils, in Mallard, E.G., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Montcrcy, California, March 11-15, 

1991: U.S. Geological Survey Water Resources 

Investigations Report 91-4034, p. 280-286.

Cozzarclli, I.M., Bacdcckcr, M.J., and Hopple, J.A., 1991, 

Gcochcmical gradients in shallow ground water caused 

by the microbial degradation of hydrocarbons at 

Galloway Township, New Jersey, in Mallard, G.E., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

technical meeting, Montcrcy, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 243-249.

Cozzarclli, I.M., 1993, The biogcochcmical fate of organic 

acids in a shallow aquifer contaminated with gasoline: 

Charlottcsvillc, Va., University of Virginia, 

unpublished Ph.D. dissertation, 240 p.

Fischer, J.M., Baehr, A.L., and Smith, N.P., 1991, An

investigation of the transport and microbial degradation 

of hydrocarbons in the subsurface at a gasoline-spill 

site in Galloway Township, New Jersey, in Mallard, 

G.E., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Montcrcy, California, March 1I15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91 -4034, p. 243-249.

Gibs, J., Brown, G.A., Turner, K.S., MacLcod, C.L., 

Jclinski, J.C., and Kochnlcin, S.A., 1993, Effects of 

small-scale vertical variations in well-screen inflow 

rates and concentrations of organic compounds on the 

collection of representative ground-watcr-quality 

samples: Ground Water, March-April, 1993, v. 31, no. 

2, p. 201-208.

Hvorslcv, M.J., 1951, Time lag and soil permeability in 

ground-water observations: Vicksburg, Miss., 

Waterways Experiment Station, Corps of Engineers, 

U.S. Army, Bulletin No. 36, 50 p.

Owens, J.P., and Minard, J.P., 1979, Upper Ccnozoic 

sediments of the Lower Delaware Valley and the 

Northern Dclmarva Peninsula, New Jersey, 

Pennsylvania, Delaware, and Maryland: U'.S. 

Geological Survey Professional Paper 1067-D, 47 p.

16
Biogeochemical Fate of Organic Acid Metabolites

Produced During the Microbial Degradation of

Hydrocarbons in Ground Water

By Isabelle M. Cozzarelli 1 and Mary Jo Baedecker1

Abstract

A combined field and laboratory study of 

low-molecular-weight organic acids was 

undertaken to improve an understanding of the 

distribution of organic acids and the geochemical parameters that influence the prevalence of 

these compounds in ground water of a shallow 

aquifer contaminated with gasoline, in Galloway Township, New Jersey. The degradation 

of aromatic hydrocarbons from gasoline 

occurred in situ by oxidation-reduction reactions mediated by bacteria. The important 

reactions were aerobic degradation and reduction of nitrate, sulfate, and Fe(III). Temporal 

and spatial shifts in the biogeochemical reactions occurred in response to changes in hydrogeochemical conditions in the aquifer.

Low-molecular-weight aliphatic, aromatic, 

and alicyclic organic acids were associated 

with hydrocarbon degradation in oxygendepleted zones of the aquifer. Laboratory 

microcosm experiments demonstrated that the 

biogeochemical fate of specific organic acids 

observed in ground water varied with the structure of the acid and the availability of electron 

acceptors. Ben/oic and phenylacetic acid were 

degraded by the indigenous aquifer microorganisms when nitrate was supplied as an electron acceptor. Aromatic acids with two or more 

substituents on the benzene ring persisted 

under nitrate-reducing conditions. Organic 

acids produced during hydrocarbon degradation participate in processes such as metal complexation, sorption, and mineral dissolution; 

thus, understanding the biogeochemical fate of

U.S. Geological Survey, Rcston, Va.

organic acids is essential to predicting the 

geochemical evolution of shallow aquifers containing degradable organic compounds.

INTRODUCTION

Aromatic hydrocarbons are frequently 

reported in anoxic subsurface environments (for 

example, Reinhard and others, 1984; Barker and 

others, 1986; Baedecker and others, 1993; Lganhouse and others, 1993). In these environments, 

anaerobic degradation is an important attenuation 

process. The degradation of monoaromatic hydrocarbons has been demonstrated to occur in the 

absence of oxygen by a variety of microbially mediated reaction pathways, including nitrate reduction 

(Kuhn and others, 1988; Major and others, 1988; 

Hutchins and others, 1991), iron reduction (Lovley 

and Lonergan, 1990), sulfate reduction (Edwards 

and Grbic'-Galic', 1992; Haag and others, 1991), 

and methanogenesis (Wilson and others, 1986; 

Grbic'-Galic' and Vogel, 1987). As a result of 

these degradation processes, oxidized organic compounds, such as low-molecular-weight (LMW) 

organic acids, are formed. A variety of metabolic 

intermediates, including aromatic, aliphatic, and 

alicyclic organic acids, phenols, and aldehydes, 

have been identified in laboratory experiments during the anaerobic microhial oxidation of petroleum 

hydrocarbons (Grbic'-Galic' and Vogel, 1987; 

Kuhn and others, 1988). The few reports of LMW 

organic acids in shallow aquifers contaminated with 

petroleum products provide documentation that 

these compounds are associated with the in situ 

anaerobic degradation of aromatic hydrocarbons 

(Cozzarelli and others, 1990,1991; Wilson and others, 1990; and Godsy and others, 1992).

Organic acids are more geochemically reactive than the hydrocarbon precursors. Although 

there is little understanding of the importance of 

sorption of simple organic acids onto aquifer sediments, the sorption of these polar compounds is

17
known to occur onto oxide surfaces (for example, 

Kummert and Stumm, 1980), clay surfaces (Yariv and 

others, 1966), and organic coatings on mineral surfaces 

of marine sediments (Sansone and others, 1987). 

Organic acid anions also can form complexes with metals in solution (Sposito, 1989) and may thus facilitate 

the transport of dissolved metals in ground water.

Although organic acids may contribute significantly to the total organic carbon pool, little is known 

about the distribution and fate of LMW organic acids in 

ground-water environments. It is apparent that the 

transformation of aromatic hydrocarbons to more soluble, and potentially more reactive, intermediates may 

have significant implications to modeling the 

geochemical effects of organic solutes in ground water 

and the implementation of bioremediation technologies.

The research described in this paper is part of a 

larger ongoing study of the fate of LMW organic acids, 

produced during microbial degradation of aromatic 

hydrocarbons, in aquifers contaminated with petroleum products (Cozzarelli, 1993; Cozzarelli and others, 

1990, and 1994). The purpose of this paper is to document how the distribution of organic acids in a shallow 

Coastal Plain aquifer contaminated with gasoline is 

affected by the availability of electron acceptors in the 

ground water. Specifically, the prevalence and type of 

dissolved organic acids in the ground water are related 

to temporal changes in the concentrations of nitrate in 

the aquifer. The field observations are compared to laboratory microcosm experiments constructed with aquifer sediment in which nitrate was added as an electron 

acceptor.

FIELD SITE

The field site is in Galloway Township, New Jersey, 

10 mi west of Atlantic City. An underground storage 

tank on a farmer's property leaked, resulting in the 

accumulation of pockets of a separate gasoline phase in 

the saturated and unsaturated zones of this shallow 

aquifer (Fischer and others, 1991). The aquifer is in 

surficial sediments of the Bridgeton Formation, 

described as fluvial feldspathic deposits (Owens and 

Minard, 1979) of the Atlantic Coastal Plain. The upper 

20 ft of soil at the site is predominately fine- to 

medium- grained sand with iron-oxide coatings and 

discontinuous clay lenses and coarse-grained sand 

lenses throughout. Below a depth of 20 ft, coarsegrained sand predominates. At approximately 11 to 15 

ft below land surface, a clayey-sand layer contains I - to 

3-in.-thick, dense, clay lenses that have a low hydraulic 

conductivity (Fischer and others, these proceedings). 

Due to the presence of the clayey layer, a perched water 

table is present above the water table aquifer beginning

at about 10 ft below land surface. The water table 

fluctuated between approximately 14-17 ft below land 

surface during the period June 1989 - April 1991.

METHODS

Ground-Water Sampling. Water samples were 

collected seven times over a 2-year period between 

June 1989 - April 1991. For this study, ground-water 

samples were collected from stainless-steel probes 

installed at site VW9 as described previously (Cozzarelli and others, 1991; Fischer and others, 1991). Samples were collected from 10.7, 12.5, 15.5, 18 and 22 ft 

below land surface, in both the perched and regional 

water zones. The areal distribution of hydrocarbon 

contaminants and unstable constituents at this site have 

been described by Baedecker and others (1991).

Water samples were analyzed for geochemical 

constituents that are important reactants, intermediates, 

or end products involved in the degradation of the dissolved hydrocarbons. Redox-sensitive constituents, 

including dissolved organic carbon (DOC), bicarbonate (HCO^'), oxygen (O2 ), ammonia (NH^), hydrogen

*ji

sulfide (H2S), ferrous iron (Fe ), and methane (CH4 ), 

were analyzed as described in Baedecker and Cozzarelli 

(1992). Nitrate and sulfate were analyzed on filtered 

(0.4 urn) samples by ion-exchange chromatography. 

Water samples for hydrocarbon analyses were 

collected and analyzed using a liquid-liquid extraction 

technique as described by Phinney and Cozzarelli 

(these proceedings). Samples for analysis of organic 

acids were collected and analyzed by extraction with 

diethyl ether and quantified by gas chromatography 

with flame-ionization detection and gas chromatography/mass spectrometry (Cozzarelli, 1993).

Microcosm Experiments. To determine the biodegradability of selected aromatic organic acids found 

in ground water contaminated with gasoline hydrocarbons, microcosm experiments were conducted with 

nitrate added as electron acceptors. Sediment was collected for the microcosm experiments from the anoxic 

perched-water zone at a location 5 ft west of site VW9 

in April 1991 (fig. I). The cores were opened in a nitrogen (N2 )-fiIled glove bag at the field site and transferred to cleaned and sterilized glass jars. The jars 

were topped off with anoxic ground water collected at 

site VW9 at 12.5 ft below land surface (VW9-12.5) to 

minimize any headspace and kept on ice for transport 

to the laboratory. Twelve jars of sediment and an additional 2.5 L of water from VW9-I2.5 were collected 

for use in the construction of the microcosms.

18
WV21

USGS-8SP 

VW24*

WV22 USGS-7 ^S DEP-3

,WV13

USGS-2 S

_SITE I GALLOWAY 

L'OCATION^/'-' TOWNSHIP

DEP-1

VW20 USGS-4 Q Q USGS-3

.WV23

X VW9

WV6« .VW19 -VW10

VW11

."""DEM

.WV25

.VW12

\

VW16

,VW26

USGS-1

HOUSE

, DEP-4

EXPLANTION

LINE OF SECTION (section shown in Fischer 

and others, this volume.)

NESTED VAPOR PROBE

U.S. GEOLOGICAL SURVEY GROUNDWATER OBSERVATION WELL

NEW JERSEY DEPARTMENT OF 

ENVIRONMENTAL PROTECTION 

GROUND-WATER OBSERVATION WELL 

______/______\______________

MULTI-LEVEL GROUND-WATER SAMPLER 

NEUTRON PROBE ACCESS

THERMISTOR NEST 

SEDIMENT COLLECTION SITE

Figure 1. Map of the study area showing location of the surface gasoline phase and ground water sampling sites. X marks the 

location from which the sediment was collected for the microcosm studies. A description of the hydrogeology along cross 

section A-A is given by Fischer and others (1996).

The microcosms were constructed and solutions 

prepared in a N2 -filled glove bag in the laboratory. 

Approximately 40 g of sediment from the field and 40 

mL of organic-acid solution made with ground water 

from site VW9-12.5 were added to each 60mL BOD 

bottle. Five organic acids were added to the ground 

water for all the experiments to yield a concentration of 

approximately 0.05-0.20 mM of each compound. The 

acids added were benzoic acid (added as sodium benzoate), ortho-{o\uic acid, phenylacetic acid, 3,4- dimethylbenzoic acid, and 2,4,6-trimethylbenzoic acid. 

Nitrate was added to the organic acid solution in sufficient quantity (30 mM) such that the concentration 

would not limit degradation reactions. The pH of the 

organic acid solutions was adjusted (with I normal 

potassium hydroxide) to 6.0, the pH value measured at 

site VW9-12.5 in the field when ground-water samples 

were collected. The degradation intermediates and end 

products produced during the incubation period were 

monitored at intervals of 1,2, 4, 8, 16, and 32 days.

RESULTS AND DISCUSSION

Small-scale spatial and temporal changes in biogeochemical reactions at site VW9 resulted in significant vertical heterogeneity of chemical constituents in 

the aquifer. The concentrations of aromatic hydrocarbons, organic acids, and electron acceptors exhibited 

similar vertical distribution patterns throughout 1990 

(Cozzarelli and others, 1991). Oxygen and nitrate were 

depleted in the shallow (10.7 ft below land surface) and 

deeper part of the perched water (12.5 and 15.5 ft 

below land surface). Cozzarelli and others (1991) 

reported that in the deeper part of the perched water 

sulfate reduction and iron reduction resulted in a 

decrease in sulfate concentrations and an increase in

the concentrations of Fe in the ground water. The 

highest concentrations of hydrocarbons and organic 

acids were found at these depths.

Changes were observed, however, in the distribution of electron accepting species in April 1991 (fig. 2 

A,B). These changes result, in part, from changing 

hydrologic conditions at the site. Oxygen, nitrate, and

19
A

w

U

CO

Q

O

si

W 

Q

February 1990 

-n April 1991

0 2 4 6 8 10 12 14

CONCENTRATION OF NITRATE, IN MILLIGRAMS PER LITER

B

E

zH-<

e

<

oo 

Q

O

G3

ea

February 1990 

April 1991

0 5 10 15 20 25 30 35 40

CONCENTRATION OF SULFATE, IN MILLIGRAMS PER LITER

Figure 2. Concentrations of electron accepting species: A. NO3" and B. SO2"4, as a function of depth below land surface, in 

ground water at site VW9 in February 1990 and April 1991.

20
sulfate in the shallow perched water (10.7-ft depth) are 

supplied by vertical recharge from precipitation 

events, whereas, at greater depths, these electron 

acceptors are also supplied by mixing with ground 

water originating from recharge areas upgradient of 

the site. Low-hydraulic-conductivity layers that form 

the base of the perched-water zone may inhibit mixing, thereby limiting the supply of electron acceptors. 

The higher nitrate and sulfate concentrations in the 

upper perched water in 1991 (fig. 2) indicate that 

nitrate and sulfate reduction by anaerobic respiration 

were not limited by availability of electron acceptors. The increased availability of electron acceptors 

is probably a result of a combination of increased 

recharge in the spring of 1991 and a decrease in 

degradable hydrocarbons in the ground water (Cozzarelli, 1993).

The changes in concentration of specific classes 

and structures of organic acids over time in the ground 

water at site VW9 provide insight into the persistence 

of certain organic acid structures in the presence of different electron acceptors. In the presence of nitrate in 

the shallow perched water, aliphatic organic acids were

consumed at a rapid rate and did not accumulate in the 

ground water (Cozzarelli, 1993). This observation was 

in contrast to the observation of increased concentrations of aliphatic organic acids in the methanogenic 

zones of aquifers contaminated with petroleum products (Cozzarelli and others, 1994). Under the nitratereducing conditions of the shallow perched water, 

however, aromatic acids, as a class, were persistent.

The persistence of individual aromatic acids is 

different-largely dependent on their structures-in the 

perched water where nitrate reduction and denitrification were occurring (fig. 3). Because the metabolic 

intermediates of hydrocarbon degradation are both produced and consumed by microbiological processes, 

rates of degradation of the organic acids cannot be 

inferred from the changes observed in ground water 

over time. However, the relative persistence of the different compounds indicates that the simple aromatic 

acids, benzoic acid and phenylacetic acid, are either 

degraded more quickly when nitrate is available as an 

electron acceptor or that they were not produced as 

extracellular intermediates during hydrocarbon degradation under nitrate-reducing conditions. The organic

GALLOWAY SITE VW9, 10.7 FT BELOW LAND SURFACE

S

D

o

g

300

250

200

150

100

50

m BENZOIC ACID

E3 PHENYL ACETIC ACID

D o-TOLUIC

D 3,4 DIMETHYLBENZOIC ACID

0 2,4,6 TRIMETHYLBENZOIC ACID

T ERROR BARS

FEBRUARY 1990 APRIL 1991

Figure 3. The concentrations of individual aromatic organic acids: benzoic acid, phenylacetic acid, o/tho-toluic acid, 3,4-dimethylbenzoic acid, and 2,4,6-trimethylbenzoic acid in the ground water at site VW9 in February 1990 and April 1991 in the 

perched water at 10.7 feet below land surface.

21
acids with more substituents, ortho-toluic acid, 3,4- 

dimethylbenzoic acid, and 2,4,6- trimethylbenzoic 

acid, increase in concentration under the same conditions. Studies by other investigators have found that 

slight changes in the number or type of substituents on 

the aromatic ring causes compounds to be more or less 

resistent to biological degradation (see, for example, 

Horowitz and others, 1982; Berry and others, 1987).

The results from the microcosm experiments in 

which NO3 ~ was added as a potential electron acceptor

clearly establish that NO3 ~ reduction is coupled to the 

biological degradation of the simplest aromatic acids, 

benzoic acid and phenylacetic acid (fig. 4). Complete 

mineralization occurs, as evidenced by the loss of 

organic acids and production of bicarbonate in the 

microcosms over time. Although initiation of benzoic 

acid degradation appeared to occur first, the degradation of phenylacetic acid was observed before benzoic 

acid was depleted. Between days 16 and 32, after the 

benzoic acid was depleted, 0.093 mM of phenylacetic 

acid was degraded. For complete oxidation of that

amount of organic carbon, 0.74 mM of HCO3 ~ should 

have been produced. This amounts to a transfer of 36 

electrons. Over this time period, an increase of 0.50

mM of HCO3 ~ was actually measured, indicating that 

33 percent of the CO2 went into biomass, assuming 

there were no losses of aqueous CC>2 from the system. 

Alternatively, some carbon could be present as metabolic intermediates; however, none were detected during the organic acid analyses. If the oxidation of 

phenylacetic acid was coupled to nitrate reduction to

NH4+, 0.42 mM of NO3 ~ would need to be reduced to 

balance the electrons released. If, on the other hand, 

denitrification was occurring, 0.67 mM of NO3 ~ would

be reduced. The actual amount of NO3 " reduced was 

1.06 mM, indicating that the reduction of nitrate may 

have been coupled to the oxidation of other organic 

compounds in the microcosms as well.

In a study of the degradation of benzoic acid and 

phenylacetic acid by a sulfate reducer, Desulfosarcina 

strain DSU3, Sembiring and Winter (1989) observed 

that both substrates were utilized simultaneously when 

the concentrations of both compounds were less than 

3.5 mM. They also observed that benzoic acid 

degraded first when in the presence of phenylacetic 

acid an observation consistent with the results of the

microcosm experiment presented here in which NO3 " 

was added as an electron acceptor. However, in contrast to the current study, where no aliphatic acids were 

detected in the microcosms, Sembiring and Winter

o

1

s

S

£

H

2

z/

26

25

24

23

22

21 

7f)

i i i

Two data points at a given s

indicate that the microcosms 

* * in duplicate.

* 

* 

* .

« *

* 

i i i

impling time

were analyzed .. 1 sterile

\ -

0

0

-

viable

\

t -

10 15 20 25 30 35

2.5

2.0

0-5

0.0

viable

sterile 

\

0 5 10 15 20 25 30 35

0.25

0.20

0.15

0.10

0.05 -

sterile

viable

0 5 10 15 20 25 30 35

0.25

0.20

2 0.15

0.10

0.05

0.00

sterile

viable

15 20 25 30 

DAYS

35

Figure 4. The concentration of nitrate, bicarbonate, benzoic acid and 

phenylacetic acid as a function of time (in days), in the microcosm 

experiments in which nitrate was added as an electron acceptor.

22
(1989) found significant accumulation of acetic acid 

during the degradation of benzoic acid. This difference 

may indicate that the more reducing conditions, under 

which sulfate reduction occurs, result in the accumulation of acetate whereas nitrate-reducing conditions do 

not.

The lack of any degradation of the higher alkylated benzoic acids in the microcosms with nitrate 

added indicates that they were resistent to biodegradation under the conditions employed in the experiments. 

The observation, that the degradation of the organic 

acids by aquifer microorganisms is structure specific, is 

consistent with other reports of selective degradation of 

aromatic compounds in the literature (for example, 

Eganhouse and others, 1987 and 1993; Kuhn and others, 1989; Evans and others, 1991; Haag and others, 

1991). Several of the pathways proposed for anaerobic 

degradation of aromatic compounds require substituent 

removal before ring reduction and ring cleavage. The 

magnitude of the activation energy required to remove 

substituents depends upon chain length and position of 

the substituent. In particular, aromatic compounds 

with substitutions at the ortho position are more stable 

than aromatic compounds with substitutions at other 

positions.

SUMMARY

The fate of organic acids in a shallow aquifer contaminated with gasoline is controlled by a complex 

interplay between microbial degradation processes and 

geochemical and hydrologic changes in the aquifer. 

Organic acids were found to be associated with hydrocarbon degradation in oxygen-depleted zones of the 

aquifer. The organic acid pool changed in composition 

and concentration over time and in space as the availability of electron acceptors changed. The alkylbenzoic acids with two or more functional groups appeared 

to be the most persistent under a range of hydrogeochemical conditions. The availability of nitrate in 

particular, nitrate from infiltrating water-appeared to 

be an important geochemical parameter that resulted in 

low concentrations of simple aromatic acids in ground 

water. In contrast, alkylbenzoic acids with a greater 

number of functional groups were more persistent 

under nitrate-reducing conditions.

Laboratory microcosm experiments, designed to 

simulate field geochemical conditions, demonstrated 

that the biogeochemical fate of specific organic acids 

observed in the ground water varied depending on the 

structure of the acid and the availability of electron 

acceptors. The addition of nitrate as an electron acceptor enhanced the microbial degradation of benzoic acid 

and phenylacetic acid that were added to the experiments as substrates.

The results presented here, combined with our 

previous studies of organic acids in shallow aquifers 

(Cozzarelli and others, 1991 and 1994) indicate that the 

types and concentrations of LMW organic acids identified in ground water contaminated by petroleum hydrocarbons are related to the geochemical conditions 

under which the degradation of the organic compounds 

occurred.

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Owens, J.P. and Minard, J.P., 1979, Upper Ccnozoic 

sediments of the lower Delaware Valley and the 

Northern Dclmarva Peninsula, New Jersey, 

Pennsylvania, Delaware, and Maryland: U.S. 

Geological Survey Professional Paper 1067-D, 47 p.

Phinncy, C.S. and Cozzarclli, I.M., 1996, A comparison of 

two analytical techniques for identifying and measuring 

the concentrations of volatile hydrocarbons in ground 

water, in Morganwalp, D.W., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

24
Reinhard, M., Goodman, N.L., and Barker, J.F., 1984, 

Occurrence and distribution of organic chemicals in 

two landfill leachate plumes: Environmental Science 

and Technology, v. 18, p. 953-961.

Sansonc, F.J., Andrews, C.C., and Okamoto, M.Y., 1987, 

Adsorption of short-chain organic acids onto ncarshorc 

marine sediments: Geochimica ct Cosmochimica Acta, 

v. 51, p. 1889-1896.

Sembiring, T. and Winter, J., 1989, Anaerobic degradation 

of phenylacctic acid by mixed and pure cultures: Applied 

Microbiology and Biotechnology, v. 31, p. 84-88.

Sposito, G., 1989, The Chemistry of Soils: New York, 

Oxford University Press, 277 p.

Wilson, B.H., Smith, G.B., and Rccs, J.F., 1986,

Biotransformation of selected alkylbenzenes and 

halogcnated aliphatic hydrocarbons in methanogenic 

aquifer material A microcosm study: Environmental 

Science and Technology, v. 20, p. 997-1002.

Wilson, B.H., Wilson, J.T., Kampbcll, D.H., Blcdsoc, B.E. 

and Armstrong, J.M., 1990, Biotransformation of 

monoaromatic and chlorinated hydrocarbons at an 

aviation gasoline spill site: Geomicrobiology Journal, 

v. 8, p. 225-240.

Yariv, S., Russell, J.D. and Farmer, V.C., 1966, Infrared 

study of the adsorption of BA and nitrobenzene in 

montmorillonitc: Israel Journal of Chemistry, v. 4, p. 

201-213.

25
26
Method for Estimating the Rate of Microbial Degradation

of Hydrocarbons Associated with Bioventing at a 

Gasoline-Spill Site in Galloway Township, New Jersey

By Matthew A. Lahvis 1 , Arthur L. Baehr1 , and Jeffrey M. Fischer 1

Abstract

Bioventing is the enhancement of microbial 

degradation caused by the delivery of atmospheric 

oxygen to aerobically biodegradable contaminants 

in the subsurface through vapor-extraction remediation. Simulation of the redistribution of oxygen 

and carbon dioxide in the unsaturated zone following a period of vapor extraction provides a method 

for estimating the rate of microbial degradation of 

hydrocarbons. Rebound vapor-concentration data 

are used to calibrate a diffusive-transport model 

developed to quantify gas fluxes and the degradation rate. The resulting rates are indicative of the 

upper limit of aerobic degradation that can occur 

during or immediately after vapor extraction, 

when oxygen availability exceeds microbial utilization. Microbial degradation of is a significant 

hydrocarbon fate at a gasoline-spill site in Galloway Township, New Jersey.

INTRODUCTION

The distribution and fate of hydrocarbons in the 

subsurface depends on their tendency to dissolve in 

ground water, volatilize to the atmosphere, adsorb to 

sediment material, or undergo biochemical 

transformations. Energy-intensive methods to treat 

physically or remove the contaminant typically are 

hampered by the transport characteristics of the 

sediment and the physiochcmical properties of the 

hydrocarbon (its solubility, vapor pressure, and 

sorption potential). Restoration of a hydrocarboncontaminated aquifer may occur naturally, however,

U.S. Geological Survey, West Trenton, N.J.

through microbial degradation processes that 

molecularly transform hydrocarbons to harmless endproducts. This mechanism, as well as other natural 

attenuating mechanisms, such as volatilization to the 

atmosphere, commonly are the only viable means of 

remediation. Engineered attempts at remediation may 

he less effective in the field than in theory, especially at 

hydrogeologically complex contamination sites. 

Quantification of site-specific degradation rates may 

therefore he beneficial to regulators and environmental 

engineers challenged by the escalating costs of cleanup 

and faced with the consequences of partial remediation 

or no-action scenarios. In Wisconsin, for example, 

natural biodegradation must now be considered as a 

remediation option at hydrocarbon-contaminated sites 

(Wisconsin Department of Natural Resources, 1993).

Bioventing is the enhancement of microbial 

degradation caused by delivery of atmospheric oxygen 

to aerobically biodegradable contaminants in the 

subsurface through vapor-extraction remediation. 

Microbial degradation reactions at the ground-water 

table results in an accumulation of carbon dioxide 

((?(>>) and a reduction in oxygen (O2 ) in the gas phase 

of the overlying unsaturated zone. During vapor 

extractions at a gasoline-spill site at Galloway 

Township, New Jersey (tig. 1), near-atmospheric gasphase concentrations of CO2 and O2 are generated 

throughout the unsaturated zone. When vapor 

extraction is terminated, the subsequent redistribution, 

or rebound, of these gases in the unsaturated zone 

overlying the area of ground-water contamination 

provides a signature of microbial degradation as it 

occurred during vapor extraction. The degradation rate 

is quantified by calibrating a diffusive-transport model 

to rebound vapor-concentration data. As outlined by 

Baehr and others (1991), knowledge of the 

composition of unsaturated-zone gas coupled with 

transport modeling can be used to estimate degradation

27
.'b

r&

79' 74" 73°

40*

\

, GALLOWAY 

LOCATK>NA^TOwNSH|p

20 MiLFS 

r1 

0 20 KILOMETERS

\

\

0 10 20 30 40 50 FEET

10 15 METERS

.V

-

\

 Q e

\ I VW9 

Former as

tank

HOUSE

EXPLANATION

 Nested vapor probes

*» U.S Geological Survey ground-water

observation well 

S New Jersey Department of Environmental

Protection ground-water observation well

__________£______^____________

9 Multilevel ground-water s

C Neutron probe access

o Thermistor nest

8 Vapor-extraction well

VW9 Vapor probe number

Figure 1. Map of the research site at Galloway Township, New Jersey, showing location of Vapor Well 9 (VW9).
rates under unremediated conditions. This problem 

was addressed by Fischer and others (1992) and Lahvis 

(1993). This paper describes the application of the 

method of Baehr and others (1991) to obtain a sitespecific upper-bound estimate of the degradation rate 

when oxygen transport is not a limitation on aerobic 

degradation.

FIELD METHODS

Unsaturated-zone gas samples were collected 

from a network of vapor probes (methods discussed by 

Fischer and others, 1991) and their composition 

determined by using gas chromatography (Baker and 

others, 1991). Samples of perched water were 

collected from submerged vapor probes with a 

peristaltic pump and analyzed for dissolved 

hydrocarbons, dissolved oxygen, alkalinity, and pH. 

The distribution of organic and inorganic constituents 

in the ground water is described further by Baedecker 

and others (1991). Unsaturated-zone-sediment 

samples were collected with a split-spoon sampler and 

analyzed for total porosity and moisture content. Core 

samples were used to map the unsaturated-zone 

stratigraphy. The site geology is discussed further by 

Fischer and others (1991) and Fischer, Smith, and 

Baehr (1996). Some undisturbed sediment cores were 

used directly in laboratory column experiments to 

measure the effective diffusion coefficient (Fischer, 

Baker, Lahvis, and Baehr, 1996), a parameter required 

for vapor-transport modeling. A nest of thermistors 

and neutron probes was used to obtain additional 

depth-dependent temperature and moisture-content 

data, respectively, to help assess the diffusive-transport 

characteristics of the unsaturated-zone sediments.

Synoptic vapor- and ground-water sampling was 

conducted at the Galloway Township site to document 

conditions at the site prior to applying vapor extraction 

(Fischer and others, 1991). The distribution of CO2 

and O2 in the unsaturated zone is presented by Baehr 

and others (1991). High concentrations of dissolved 

hydrocarbons were present in the perched (10 feet 

below land surface), particularly near Vapor Well 9 

(VW9) (Baedecker and others, 1991), but were not 

detected in the overlying unsaturated zone except in a 

small region near the former location of the exhumed 

gasoline tank (Fischer, Smith, and Baehr, 1996). This 

observation indicates the presence of thin microbially 

active zone near the perched water table, where the rate

of hydrocarbon degradation exceeds the rate of 

hydrocarbon volatilization to the overlying unsaturated

zone.

MODEL APPLICATION

The diffusive-transport model developed to 

quantify microbial degradation in porous media 

implements the following one-dimensional analytical 

solution to the diffusion equation under the assumption 

of unsaturated-zone homogeneity:

dck "*IF = D> (1)

where the subscript k indicates that the equation 

applies to either CO2 or O2, Gk (g/cnv) is the gaseoushave concentration of the constituent, Dk (cnWs) is the 

effective diffusion coefficient, ak (unitless) is the storage coefficient, and L (cm) is the total length of the 

model domain. A homogeneous unsaturated zone 

implies that Dk and ak are constant, which in turn 

implies that the physical properties of the porous 

medium (total porosity, moisture content, and tortuosity) are constant. Because all degradation is simulated 

through use of a boundary condition a the perched 

water table, CO2 and O2 transport are uncoupled. 

Equation 1 incorporates the following assumptions: 

equilibrium partitioning of the mass between the gaseous ad aqueous phases, negligible adsorption, and no 

constituent sources or sinks.

This model was applied to simulate the 

redistribution of CO2 and O2 near VW9. Vaporconcentration data collected from vapor probes (VW9- 

6.0 and VW9-8.2 were used in the one-dimensional 

vapor-transport analysis. A conceptualization of the 

unsaturated-zone stratigraphy in the vicinity of VW9 is 

shown in figure 2. A high relative moisture content 

measured in the thin clay layer (unit #4), indicates that 

modeling the entire unsaturated zone as a uniform 

porous medium may be inappropriate. To account for 

this sharp discontinuity, the model domain was 

reduced to the three lowermost stratigraphic units 

(nos. 1-3) and the clay was modeled as a third-kind

29
Ground surface

20

40

60

80

100

120

140

160

180

200

220

240

260

280

,VW9-6.0

VW9T-7.7

\

VW9-8.2

DARK BROWN ORGANIC

CLAYEY SAND 

STRATIGRAPHIC UNFT *7

MEDIUM-BROWN

CLAYEY SAND

STRATIGRAPHIC UNIT #6

OFF-WHITE SAND 

STRATIGRAPHIC UNIT *S

WHITISH-GRAY CLAY 

STRATIGRAPHIC UNIT *4

OFF-WHITE SAND 

STRATIGRAPHIC UNIT «

ORANGISH-BROWN

CLAYEY SAND 

STRATIGRAPHIC UNIT *2

OFF-WHITE SAND 

STRATIGRAPHIC UNIT *1

Perched water table

Figure 2. Conceptualization of the unsaturated-zone stratigraphy in the vicinity of Vapor Well 9 (VW9) (Blackened 

rectangles represent vapor probes.

boundary, analogous to a leaky confining unit, as 

follows:

,atz = L (2)

where Dk ' (cm2/s) is the effective diffusion coefficient of the clay layer; b' (cm) is the clay-layer 

thickness; L (cm) is the length of the model domain; 

and Gk (g/cm3 ) is the constants concentration of 

the consVFtuent on the opposite site of the clay layer, 

typically designated as the atmospheric concentration.

The lower boundary corresponds to the top of 

the capillary zone, which is defined by a sharp gradient 

in the moisture content. An analytical solution was 

obtained for two boundary-condition cases: (I) a

constant flux, Jk((g/cm2)/s); and (2) a constant 

concentration, ( Gk (g/cnr), modeled as follows:

G, (z) =

dz

case (2) at z=0.

(3)

(4)

The constant-flux boundary condition implies 

constant substrate utilization, independent of transport 

in the unsaturated zone above, whereas the constantconcentration boundary condition implies timedependent degradation, directly influenced by 

unsaturated-zone transport. The boundary conditions 

mathematically span the range of microbial responses 

to unsaturated zone-transport, bracketing boundarycondition sensitivity.

The initial condition is specified as a linear 

piecewise interpolation through specified points as 

follows:

The solution to (1) subject to the boundary 

conditions (2), (3), or (4), and the initial condition (5) 

is obtained by applying the separation-of-variables 

solution technique as follows:

where v(z,t) is the transient component of the solution, 

and may be written as:

(7)

where:

) = cos /X~z for case l,or 

) = sn for case 2.

In general,

where Ak(z) is the initial condition comprised of m 

components, given by:

30
OZ

70

60 h

50

40

30

20

10

11 / 

. ¥

a Measured [CO2]: VW9-8.2

 Measured [CO2]: VW9-6.0

 Simulated [CO2]: VW9-8.2

 Simulated (CO2]: VW9-6.0

200 400 600 800 

TIME (HOURS)

1000 1200

Figure 3. Analytical model calibration to summer 1991 carbon dioxide vapor-concentration data with application of the 

constant carbon dioxide flux boundary condition.

and K(z) is the steady state solution defined as:

J,

case

case 2

For each case, the constants Cn are defined more specifically as:

m \M. B.

for case I, or

(8)

(8a)

(8b)

31
c- - y r

I n

Bi

r--cosL fk- 1 A, V "

for case 2,

where:

" Jk o M. = M + - for case l,or M = M, ki Dk ' ' Dk 

L + r b'

Dk

for case 2, and

B. = B°, - { \J,( -^ + -M ~| + G ) for Case 1, or B. = B°, - G. for case 2.

' L ^ k k' \ '"" ' ' "

The eigenvalues (kn ) in each case are determined numerically by the Newton-Raphson method and are given by:

Dk

I ~V I ~DkJ^n tan L Ik = for case 1, or tan L Ik = ^ ^ for case 2. V n M FT V » D '

~V

:3.32 were calculated for carbon dioxide and 

oxygen, respectively (Lahvis, 1993). Any agreement 

between the degradation rate calibrated by using CO2 

vapor-concentration data and the rate calibrated by 

using 02 vapor-concentration data supports the 

hypothesis of complete hydrocarbon mineralization.

Results for calibration to the CC)2 rebound data 

obtained during summer 1991 with application of the 

constant-flux boundary condition (case I) are 

illustrated in figure 3. A microbial degradation rate of 

Rhio=2.22E-IO ((g/cm2 )/s) was calibrated at the 

degradation boundary. Similar estimates were 

obtained from other model calibrations (Lahvis, 1993). 

This rate can be extrapolated to a sitewide estimate of 

pure hydrocarbon degradation (R^j0=16.4 L/yr, or 

4.33 gal/yr) by assuming a plume diameter of 1524 cm 

(50 ft) and an average hydrocarbon density of 

0.77 g/cm3 .

Calibrated estimates of the diffusion coefficient 

for the model domain (Dk ) compared favorably to 

laboratory-based estimates (Baker, 1993; Lahvis, 

1993). This agreement lends credibility to the estimate 

of flux, the other unknown parameter defined by model 

calibration.

In summary, the sitewide degradation-rate 

estimate, Rhio=16.4 L/yr, determined by means of 

model calibration to CO-, rebound data collected

RESULTS AND SUMMARY

For this analysis, the magnitude of the boundary 

condition (J k and Gk ) and the effective diffusion 

coefficients (Dk and Dk ')and are determined by means 

of model calibration by using a least-root-meansquares fit between measured and simulated rebound 

vapor-concentration data. Rebound vaporconcentration samples were collected over a 1,000- 

hour period after vapor extraction was terminated.

The relations among the rates of hydrocarbon 

degradation, carbon dioxide production, and oxygen 

consumption are defined under the assumption that 

aerobic degradation is the dominant microbialdegradation pathway and hydrocarbon mineralization 

is complete. (Baker and Baehr, 1996). The 

hydrocarbon-degradation rate in the perched water 

zone is then stoichiometrically related to the flux of 

CO-, and OT in the unsaturated zone as follows:

R, (9)

where R|,JO ((g/cm")/s) is the hydrocarbon-degradation 

rate, and rk (unitless) is the stoichiometric mass constant relating the mass of CO2 produced or mass of O2 

consumed to the mass of hydrocarbon degraded. Stoichiometric mass constants / = 3.19 and 2.28

32
during summer 1991, represents an upper bound of 

aerobic degradation that occurs during or immediately 

after vapor extraction when oxygen transport in the 

unsaturated zone is not a limitation. A similar rate 

estimate, Rbj0-15-8 L/yr (4.18 gal/yr), was obtained by 

calibrating a more comprehensive numerical transport 

model to the same data. The numerical model accounts 

for spatial variability in the diffusion and storage 

coefficients (Lahvis, 1993). The microbial degradation 

rate calibrated by using this approach indicates that 

microbial degradation of hydrocarbons is a significant 

process at the Galloway Township site, the rate of 

which exceeds the rate of volatilization into the 

atmosphere under unremediated conditions. This 

modeling approach provides regulators and 

environmental engineers with an ability to quantify the 

bioventing performance at other hydrocarboncontaminated sites.

REFERENCES CITED

Bacdcckcr, M.J., Cozzcrclli, I.M., and Phinncy, C.S., 1991, 

Distribution of organic and inorganic constituents in 

ground water at Galloway Township, New Jersey, in 

Mallard, G.E., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Montcrcy, California, March 1991: U.S. Geological 

Survey Water Resources Investigations Report 91- 

4034, p. 287-293.

Baehr, A.L., Fischcr, J.M., Lahvis, M.A., Baker, R.J., and 

Smith, N.P., 1991, Method for estimating rates of 

microbial degradation of hydrocarbons based on gas 

transport in the unsaturated zone at a gasoline-spill site 

in Galloway Township, New Jersey, in Mallard, G.E., 

and Aronson, D. A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

technical meeting, Montcrcy, California, March 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 250-255.

Baker, R.J., 1993, Hydrocarbon microbial degradation rate 

determination in columns of unsaturated porous media 

using vapor transport modeling: Philadelphia, Pa., 

Drcxcl University, Department of Civil Engineering, 

unpublished Ph.D. dissertation, 225 p.

Baker, R.J., and Bachr, A.L., 1996, Use of column studies 

and a reactive transport model to measure 

biodcgradation rates of hydrocarbons, in Morganwalp, 

D.W., and Aronson, D. A., cds., U.S. Geological Survey 

Toxic Substances Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, 

September 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Baker, R.J., Fischcr, J.M., Smith, N.P., Kochlcin, S.A., and 

Bachr, A.L., 1991, Gas chromatographic methods for 

determining the composition of the vapor phase in 

gasoline-contaminated soils, in Mallard G.E., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Montcrcy, California, March 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 280-286.

Fischcr, J.M., Bachr, A.L., and Smith, N.P., 1991, An

investigation of the transport and microbial degradation 

of hydrocarbons in the subsurface at a gasoline-spill 

site in Galloway Township, New Jersey, /'/; Mallard, 

G.E., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Montcrcy, California, March 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 243-249.

Fischcr, J.M., Bachr, A.L., and Smith, N.P., 1992, Field 

dctcrmination of degradation rates in shallow ground 

water by simulating the transport of oxygen and carbon 

dioxide in the unsaturated zone (poster): American 

Geophysical Union special session meeting, spring 

1992, Montreal, Quebec, Canada, unpublished.

Fischcr, J.M., Baker, R.J., Lahvis, M.A.. and Bachr, A.L., 

1996, Determination of vapor-phase diffusion 

coefficients for unsaturatcd-zonc sediments at a 

gasoline-spill site in Galloway Township, New Jersey, 

in Morganwalp, D.W., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting. 

Colorado Springs, Colorado, September 1993: U.S. 

Geological Survey Water-Resources Investigations 

Report 94-4015.

Fischcr, J.M., Smith, N.P., and Bachr, A.L., 1996, 

Hydrogcology, contaminant distribution, and 

biodcgradation processes at a gasoline-spill research 

site in Galloway Township, New Jersey, in 

Morganwalp, D.W., and Aronson, D.A., cds.. U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting. 

Colorado Springs, Colorado, September 1993: U.S. 

Geological Survey Water-Resources Investigations 

Report Report 94-4015.

33
Lahvis, M.A., 1993, Estimating rates of microbial Wisconsin Department of Natural Resources, 1993, Natural 

degradation of hydrocarbons at a gasoline-spill site biodcgradation as a remedial action option: Interim 

based on vapor diffusion modeling in the unsaturatcd guidance: Soil Release News, v. 3, no. 1, p. 1-17. 

zone: Philadelphia, Pa., Drcxel University, Department 

of Civil Engineering, unpublished Ph.D. dissertation, 

204 p.

34
Determination of Vapor-Phase Diffusion Coefficients for

Unsaturated-Zone Sediments at a Gasoline-Spill Site in

Galloway Township, New Jersey

By Jeffrey M. Fischer 1 , Ronald J. Baker1 , Matthew A. Lahvis 1 , and Arthur L. Baehr 1

Abstract

The rate of diffusion of gases through the 

unsaturated zone has been related to rates of 

microbial degradation of hydrocarbons at a gasoline-spill site in Galloway Township, New Jersey. 

The diffusion coefficient for the unsaturated 

porous medium is required to obtain rates of vapor 

movement from vapor-concentration measurements. Laboratory experiments were conducted on 

heterogeneous sediment cores to identify vaporphase concentration profiles and vapor fluxes. 

A one-dimensional vapor-diffusion model based 

on the Stefan-Maxwell equations was used to 

calculate diffusion coefficients from these experiments. Diffusion coefficients obtained by using 

this method were smaller than those predicted by 

using the Millington-Quirk equation, and were 

particularly significant for clayey sediments. Use 

of the theoretically calculated diffusion coefficients to calculate vapor flux and biodegradation 

rates at the study site would have led to overestimates of these rates when compared to those 

determined by using the experimental method.

INTRODUCTION

Studies of diffusive transport in the unsaturated 

zone of hydrocarbon vapors and inorganic-gas constituents that are indicators of biodegradation are being 

conducted at a gasoline-spill site in Galloway Township, N.J. (Baehr and others, 1991; Baehr and Fischer, 

1996; Baker and Baehr, 1996; Lahvis and others, 

1996). Diffusion of gases through the unsaturated zone 

is controlled by the vapor-constituent concentration 

gradient and the diffusion coefficient. Unsaturatedzone concentration profiles can be measured (Baker 

and others, 1991; Fischer and others, 1991), whereas

diffusion coefficients are determined by using some 

form of tracer experiment. The diffusion coefficient 

(D) of a vapor constituents in an unsaturated porous 

medium is commonly defined as follows:

D = Dh 9a T, (1) 

where

Dh is the diffusion constant of the vapor constituent in a 

bulk gaseous phase identical to that in the 

unsaturatcd-zone atmosphere, 

9a is the air-filled porosity, and 

T is the tortuosity of the vapor-filled pores.

Dh and 0a are measured independently of diffusion 

experiments. The tortuosity depends only on porespace configuration and is independent of vapor constituents. A widely used empirical approximation for 

tortuosity is the Millington-Quirk equation (Millington 

and Quirk, 1961):

(2)

where (J) is porosity.

U.S. Geological Survey, West Trenton, N.J.

This paper describes laboratory tracer experiments 

that were conducted on sediment cores from the Galloway Township research site. It describes and demonstrates the use of a one-dimensional model that was 

developed to estimate diffusion coefficients on the 

basis of the Stefan-Maxwell equations.

METHODS

Fickian diffusion is commonly assumed to be 

sufficient for calculating diffusive fluxes, or, if fluxes 

are known (as in a controlled experiment), for calculating tortuosity. Baehr and Bruell (1990) have shown that 

the Fickian assumption can lead to errors in estimated 

tortuosity values under certain conditions. This 

assumption can introduce substantial errors in subsequent calculations of biodegradation rates. A more rigorous approach for quantifying multicomponent vaporphase diffusive transport in an unsaturated porous 

medium is to use the Stefan-Maxwell equation (Baehr 

and Bruell, 1990).

35
Four cores were collected from the study-site location shown in figure 1. The sampling location is the site 

of concurrent in situ bioegradation studies (Lahvis and 

others, 1996) and sample collection for laboratory 

biodegradation experiments (Baker and Baehr, 1996). 

Cores were collected in 3.6-cm-i.d., 61 -cm-long aluminum tubes with a split-spoon sampler. Two of the four 

cores collected were used in laboratory vapor-diffusion 

experiments.

The column design used in the laboratory experiments is shown in figure 2. Cores were oriented vertically and a stainless-steel screen was attached to the 

bottom of the core liner to support the sediment. The 

tracer constituent is placed in a reservoir below the 

screen. For these experiments, pure liquid toluene was 

maintained in the reservoir. The liquid maintained a 

toluene-saturated airspace at the bottom of the core. 

If necessary, several centimeters of sediment were 

removed from the top of the column, and a cap was 

placed on the liner, creating an airspace. During the 

experiments, this space was purged continually with 

humidified gas (95 percent nitrogen and 5 percent 

oxygen, by volume). Purging with the sweep gas constantly removed vapor-phase toluene that had diffused 

through the column. Bracketing the column with toulene-saturated gas at the base and a sweep gas with a 

low toluene concentration at the top resulted in a vertical concentration profile in the sediment. Sampling 

ports were installed at about 9-cm intervals to determine this profile. Sampling ports consisted of 18-gauge 

stainless-steel syringe needles that were inserted 

through holes drilled in the core liner. The holes were 

sealed and the syringe needles were capped with 

syringe valves. Toluene concentrations were determined by using a gas chromatograph with a flameionization detector. Inorganic-gas concentrations (oxygen and carbon dioxide) were determined by using a 

gas chromatograph with a thermal-conductivity detector (Baker and others, 1991). Toluene flux at the top of 

the column was determined by multiplying the volumetric flow rate by hydrocarbon concentration in the 

sweep gas and dividing the result by the cross-sectional 

area of the column. When a steady vapor-phase composition, as indicated by an unchanging hydrocarbon profile, was reached in the column, vapor-concentration 

data were collected for use in model calibration. At the 

end of the experiments, the column was cut into 2-cm 

to 5-cm sections, and sediment lithology, porosity, and 

moisture content were determined.

A one-dimensional vapor-diffusion model was 

developed for analyzing laboratory column vaporconcentration and physical-property data to estimate 

tortuosity as a function of column depth. The model, an 

expansion of work by Baehr and Bruell (1990), incorporates the Stefan-Maxwell equations into a steadystate conservation-of-mass equation to define transport 

of each component of a multicomponent gas mixture in 

an unsaturated porous medium. The model has the 

ability to simulate one-dimensional vapor transport 

through a variably saturated porous medium that is heterogeneous with respect to porosity, moisture content, 

and the diffusion coefficient. Laboratory column 

vapor-concentration and physical-property data are 

simulated as a one-dimensional series of nodes. The 

model is calibrated by varying tortuosity values at 

nodes until the difference between observed and simulated constituent concentrations is minimized. Details 

of the model are presented in Baker (1993); additional 

uses of the model are presented by Baker (1993) and 

Baker and Baehr (1996).

The model requires data on vapor-phase constituents and porous-medium physical properties. Only 

steady-state conditions were considered in these simulations. Vapor-phase inputs include constituent concentrations and sample depths, as well as concentrations 

and fluxes of the constituents in the sweep gas. Porousmedium inputs include porosities and volumetric moisture contents. Initial estimates of tortuosity also are 

required; these were calculated by using Fick's law. 

Additional data on physical characteristics include 

atmospheric pressure, temperature, and sedimentcolumn length.

The ability of the model to accurately estimate tortuosity was tested by using data reported in Baehr and 

Bruell (1990) for a homogeneous dry sand column. The 

reported tortuosity for benzene diffusion through the 

column was 0.566. The simulated concentration profile 

matched the reported profile and the estimated tortuosity for 13 nodes averaged 0.576, with a standard deviation of 0.073. This result indicated that the model can 

be used to produce accurate estimates of tortuosity.

VAPOR-PHASE-DIFFUSION 

COEFFICIENTS

Core-2, the first core from the Galloway Township 

site to be analyzed, came from a depth of 0.6 to 1.2 m. 

It contained nonlayered, fairly uniform sediments composed of approximately 30 percent clay, 25 percent silt,

36
75* 74° 73*

39*

'SITE I GALLOWAY 

LOCATION/^TOWNSHIP

o 20 MILES 

HV 

0 20 KILOMETERS

X.

0 10 20 30 40 50 FEET

10 15 METERS

Core location

HOUSE

EXPLANATION

 Nested vapor probes

^ U.S. Geological Survey ground-water

observation well 

<S New Jersey Department of Environmental

Protection ground-water observation well

9 Multilevel ground-wate

C Neutron probe access

<> Thermistor nest

8 Vapor-extraction well

Figure 1. Galloway Township gasoline-spill study site showing location where cores for laboratory experiments were obtained.
3.8 centimeters 

inside diameter

-Headspa

 O-ring

Needle 

valve

Flow 

meter

Humidifier

Soil leve

Vapor s

Stainles

Media s

O-ring

-Liquid h

NOT TO SCALE

Figure 2. Column design for estimating tortuosity values of sediment cores. (Hydrocarbon reservoir and sweep-gas cap slip ov

sealed in place with 0-rings and lock nuts (not shown). Sediment level in core liner varied depending on amount of sample reco

ports are spaced about 9 cm apart. Number of ports varied with the amount of sediment recovery. Sweep gas was 95 percent ni

oxygen. Hydrocarbon used in experiments was toluene.)
and 45 percent coarse- to fine-grained sand. Other 

physical properties of the core are listed in table 1, in 

which estimated values of tortuosity from the simulations are compared to Millington-Quirk estimates. 

Simulated and actual vapor concentrations compare 

favorably, demonstrating a successful model calibration (fig. 3). Values for tortuosity derived from simulations were consistently greater than 1.0 (an infeasible 

value) for the upper 10 cm of the sediment column. 

No such problems were encountered in similar experiments in which repacked sediments were used (Baker, 

1993). Values greater than 1.0 probably indicate the

Table 1 . Physical properties of core subsections and tortuosity values calculated

by using the Millington-Quirk equation (theoretical) and those simulated with the

model

[g/cm3 , grams per cubic centimeter; V/V volume per volume; --, no data; >, greater than;

<, less than]

Sample

interval

(cm)

Bulk 

density

(g/cm3)

Total Air-filled

porosity porosity

(V/V) (V/V)

Tortuosity Tortuosity 

(theoretical) (simulated)

Core-2

0.0- 5.1 

5.1 - 10.2 

10.2- 15.2 

15.2-20.3 

20.3 - 25.4 

25.4 - 30.5 

30.5 - 35.0

.51 

.69

.79 

.76 

.83 

.77 

.71

0.407

.337

.297

.308

.280

.303

.328

0.311

.211

.156

.158

.177

.119

.134

Core 3

GEOLOGY

0 4 8 12 16 

TOLUENE CONCENTRATION. IN MILLIGRAMS PER LITER

0.394

.234

.149

.143

.225

.076

.085

0.073

.023

.025

.004

.017

0.0- 5.0

5.0- 10.0

10.0- 15.0

1 5.0 - 20.0 :

20.0 - 25.0

25.0-27.5

27.5 - 30.0

30.0 - 35.0

35.0 - 40.0

40.0 - 45.0

45.0-49.5

.44

.79

.87

'..03

.86

.99

.82

.75

.79

.62

.62

0.435

.296

.264

.202

.267

.217

.283

.312

.295

.361

.364

0.395

.186

.125

.080

.III

.001

.126

.241

.214

.308

.298

0.603

.225

.III

.068

.083

<.OOI

.100

.371

.314

.490

.447

>!.()

>I.O

>I.O

0.033

.039

.001

.029

.025

-

~

-

GEOLOGY

Fine sand

_ Clayey aand

Medium sand

Clayey sand 

Medium sand

0 40 80 120 160 

TOLUENE CONCENTRATION. IN MILLIGRAMS PER LITER

EXPLANATION

 Measured 

 Simulated

Figure 3. Measured and simulated toluene concentration with depth for homogeneous sediments (core-2) and heterogeneous 

layered sediments (core-3).

39
presence of a short-circuited flow path in the disturbed 

part of the upper core. Nondiffusive flux through the 

upper part of the volume does not affect diffusive flux 

through lower parts of the column.

Core-3, the second core from the Galloway Township site to be analyzed, came from a depth of 1.5 to 

2.1 m. It contained heterogeneous sediments with 

layers of sand, clay, and clayey sand (fig. 3). Physical 

properties of the layers are listed in table 1. Porosity 

and moisture-content values were assigned at layer 

interfaces to define layers in the model. Thus, values 

measured within a layer were entered at depths 1 cm 

above or below the interface to define the sediment 

discontinuity for the model. As for the previous core, 

estimated values of tortuosity were greater than 1.0 for 

the upper 15 cm of the column. No concentration gradient could be measured in the sand beneath the clay 

layers situated in the core at depths between 24.0 and 

37.0 cm because the diffusion coefficient of the underlying sand was so much greater than that of the clays, 

which were effectively a "transport bottleneck." Without a concentration gradient, it was not possible to 

calculate tortuosity in this layer; a separate laboratory 

experiment would need to be conducted on the sand 

alone to obtain this estimate. Estimates of tortuosity for 

the part of the core that was amenable to analysis with 

the model ranged from 0.001 to 0.033 (table 1), and 

simulated and actual vapor concentrations matched 

closely (fig. 3). The lower tortuosity value is for a thin 

clay lens at a core depth of 24 to 26.5 cm. Tortuosity 

values for the remaining layers were similar to those 

for core-2 because the physical properties of the 

sediment were similar.

Simulated tortuosity values for both cores generally were 2 to 20 times less than those predicted by 

using the Millington-Quirk equation (table 1) probably 

because the Millington-Quirk equation incorporates 

the assumption that the particles were spherical and the 

sediments contain considerable amounts of clay. Use of 

tortuosity values estimated from the Millington-Quirk 

equation to calculate vapor fluxes would have caused 

an overestimation of vapor fluxes at the site by an order 

of magnitude. Therefore, subsequent calculations of 

biodegradation rates at the site based on calculated 

vapor-transport values also would yield seriously 

overestimated values.

Although values of tortuosity calculated for cores 

in the laboratory are believed to be more representative 

of in situ conditions than theoretically obtained estimates, some uncertainty exists. Some sediment compaction probably occurred during sampling and while 

the cores were standing vertically in the lab. Further, 

moisture redistribution probably occurred as well. 

Compaction would cause tortuosity values to decrease 

and moisture redistribution would cause it to increase 

near the top of the core and decrease toward the bottom.

Although the model accurately simulated measured vapor concentrations, several concerns are being 

addressed. These include the sensitivity of tortuosity 

estimates to nodal spacing, and ways to reconcile differences in the vertical spacing of measured values 

(vapor concentration, moisture content, and porosity) 

and predicted values (vapor concentration and tortuosity). Also, to conserve computing time, an improvement in the algorithm used to minimize the difference 

between observed and simulated vapor concentrations 

may be made.

SUMMARY

Diffusion coefficients are required to quantify 

vapor fluxes through the unsaturated zone when the 

vapor-concentration distribution is known. Sediment 

cores from the unsaturated zone at the Galloway Township gasoline-spill research site were used to illustrate 

a method developed to obtain diffusion coefficients. 

The method involves the use of laboratory tracer experiments and the application of a one-dimensional model 

based on the Stefan-Maxwell equations. Physicalproperty data and vapor-concentration data from 

homogeneous and heterogeneous cores were used to 

successfully estimate tortuosity. Tortuosity in the upper 

part of the cores could not be estimated, however, 

possibly as a result of short-circuited flow through 

disturbed sediments. Tortuosity in the lower part of the 

layered core could not be simulated because concentration gradients could not be measured. Calculated tortuosity values ranged from 0.001 for a clay layer to 0.033 

for clayey sands. Simulated tortuosity values were 

about 2 to 20 times less than those predicted by using 

the Millington-Quirk equation. Use of tortuosity values 

estimated by using the Millington-Quirk equation 

would have resulted in the overestimation of vapor 

fluxes at the site and the overestimation of biodegradation rates based on estimated rates of vapor transport.

40
REFERENCES

Bachr, A.L., and Brucll, C.J., 1990, Application of the 

Stcfan-Maxwell equations to determine limitations of 

Fick's law when modeling organic vapor transport in 

sand columns: Water Resources Research, v. 26, no. 6, 

p. 1155-1163.

Bachr, A.L., Fischcr, J.M., 1996, Overview of research on 

the transport, microbial degradation, and remediation 

of hydrocarbons at a subsurface gasoline-spill site in 

Galloway Township, New Jersey, in Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 95-4015.

Bachr, A.L., Fischcr, J.M., Lahvis, M.A., Baker, R.J., and 

Smith, N.P., 1991, Method for estimating rates of 

microbial degradation of hydrocarbons based on gas 

transport in the unsaturatcd zone at a gasoline-spill site 

in Galloway Township, New Jersey. in Mallard, G.E., 

and Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Montcrcy, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 250-255.

Baker, R.J., 1993, Hydrocarbon microbial degradation rate 

determination in columns of unsaturatcd porous media 

using vapor transport modeling: Philadelphia, Pa., 

Drcxcl University, Environmental Studies Institute, 

unpublished Ph.D. dissertation, 237 p.

Baker, R.J., and Bachr, A.L., 1996, Use of column studies 

and a reactive transport model to measure biodcgradation rates of hydrocarbons, in Morganwalp, D.W., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 95-4015.

Baker, R.J., Fischcr, J.S., Smith, N.P., Kochnlein, S.A., and 

Bachr, A.L., 1991, Gas-chromatographic methods for 

determining the composition of the vapor phase in gasoline-contaminate soils, in Mallard, G.E., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Montcrcy, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 280-286.

Fischer, J.M., Bachr, A.L., and Smith, N.P., 1991, An investigation of the transport and microbial degradation of 

hydrocarbons in the subsurface at a gasoline-spill site 

in Galloway Township, New Jersey, in Mallard, G.E., 

and Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Montcrcy, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 243-249.

Lahvis, M.A., Bachr, A.L., and Fischcr. J.M., 1996, Method 

for estimating the rate of microbial degradation of 

hydrocarbons associated with biovcnting at a gasolinespill site in Galloway Township, New Jersey, in 

Mallard, G.E., and Aronson, D.A., cds., Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program-Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 95-4015.

Millington, R.J., and Quirk, J.P., 1961, Permeability of 

porous solids: Transactions of the Faraday Society, 

v. 57, p. 1200-1207.

41
42
Use of Column Studies and a Reactive 

Transport Model to Measure Biodegradation Rates

of Hydrocarbons

By Ronald J. Baker1 and Arthur L. Baehr 1

Abstract

A laboratory method for determining 

biodegradation rates of volatile hydrocarbons in 

unsaturaled porous media (subsurface sedimentary material) was developed. Glass columns were 

filled with samples of porous media from the 

unsaturated zone at a gasoline-contaminated site 

in Galloway Township, N.J. Vapor-phase hydrocarbons (benzene, toluene, or p-xylene) were 

added to the porous media at the bottom of the 

column and were transported vertically along a 

concentration gradient. Aerobic biodegradation of 

hydrocarbons is indicated by the production of 

carbon dioxide (CO2 ). The vapor-phase hydrocarbon and CO2 concentrations were monitored at 

several depths through sampling ports. Fluxes of 

hydrocarbons and CO2 leaving the column also 

were measured frequently. A mathematical model 

was used to calculate CO2 -production rates as a 

function of depth. These distributions and 

stoichiometric relations describing hydrocarbon 

biodegradation were used to calculate depthspecific hydrocarbon-degradation rates. Degradation rates generally increased with depth, and 

appeared to be coupled to soil-moisture content, 

which also increased with depth. Degradation 

rates also increased with the number of methyl 

groups on the aromatic ring, and rates were greater 

in unprocessed sediment than in sieved sand from 

which fine particles had been removed. This 

method could be used to determine optimum 

biodegradation conditions in the design of 

bioremediation strategies.

'U.S. Geological Survey, Trenton, N.J.

INTRODUCTION

An estimated 75,000 to 100,000 underground 

gasoline- and other fuel-storage tanks in the United 

States currently are leaking (Brown and others, 1985), 

causing widespread contamination of ground water by 

benzene, toluene, xylenes, ethylbenzene, and other 

toxic hydrocarbon components. The fact that hydrocarbons from petroleum products are degraded in the subsurface by indigenous microbes has been established 

(Atlas, 1984). Biodegradation rates must be quantified 

to assess the importance of this process as an attenuating mechanism under conditions associated with either 

in situ bioremediation or no-action scenarios.

Rates of biodegradation can be estimated by monitoring vapor-phase geochemical signatures of hydrocarbon biodegradation (for example, concentration 

gradients and fluxes of oxygen (O2 ), CO2 , and hydrocarbons). These gas constituents have been monitored 

extensively during ongoing research at the site of a 

crude-oil spill in Bemidji, Minn. (Baedecker and 

others, 1989), and at a gasoline-spill site in Galloway 

Township, N.J. (Baehr and Fischer, 1996).

One component of the research at the Galloway 

Township site is the determination of hydrocarbonbiodegradation rates in laboratory columns filled with 

samples of unsaturated porous media collected at the 

site. The columns are designed to maintain constant 

conditions over long periods of time (months), and to 

provide conditions that more closely approximate natural geochemical conditions than a closed-microcosm 

approach does. When the columns are aligned vertically, a stable moisture gradient forms that is representative of the moisture gradient typically observed in the 

unsaturated zone above the capillary zone. A stable 

gradient of vapor-phase hydrocarbon concentration 

also forms. This gradient is representative of conditions in the field, where contaminant concentrations 

typically decrease with distance from the contaminant 

source.

43
Aromatic hydrocarbons readily biodegrade under 

aerobic conditions (Atlas, 1984), resulting in consumption of O2 and production of CO2 and biomass. Molar 

ratios of O2 consumption, hydrocarbon consumption, 

and CO2 production can be derived stoichiometrically 

from relevant reactions. For example, representative 

reactions for toluene biodegradation are as follows:

 Mineralization of hydrocarbons to CO2 :

1 C7 H8 + 9 02 -> 7 C02 + 4 H 2 0 ( 1 )

Molar ratios: 1 .00 toluene : 9.00 O? : 7.00 CO2

 Conversion of hydrocarbons to biomass:

7C7 H 8 + 9H4 +9N03 - -> 9 C5 H 702N (2) 

+ 4 CO2 + 1 H 2O

Molar ratios: 1 .00 toluene : 1 .28 NO, : 0.57 CO2

Endogenous respiration of biomass to CO2 : 

C5 H7O2N + 5 O2 + H f ->- 5CO2 

+ NH4+ + 2 H 2O

(3)

Molar ratios: 1 .00 Biomass : 5.00 O: : 5.00 CO2 : 1 .00 NH/

Stoichiometric ratios allow for calculation of reaction 

rates for one chemical constituent in terms of another. 

This approach was used in this work, where hydrocarbon-utilization rates were not directly measurable, but 

were calculable from CO2 production rates.

This paper describes the development of the 

column apparatus used to ( 1 ) determine biodegradation 

rates, (2) measure biodegradation rates of hydrocarbons 

in columns of unsaturated porous media from the 

Galloway Township site, and (3) determine depthspecific hydrocarbon-biodegradation rates by use of 

a mathematical model.

EXPERIMENTAL METHOD AND 

MATERIALS

The experimental column design is shown in 

figure 1. Five 10-cm internal diameter glass columns 

each have six ports for vapor-phase sampling. A stainless-steel screen near the bottom of the column supports the porous media. The bottom reservoir provides 

a constant source of hydrocarbons, which evaporate 

and diffuse upward through the column. In the experiments reported here, pure liquid aromatic hydrocarbons were placed in the bottom reservoir. A gas-filled 

chamber above the porous media was purged continuously with a mixture of compressed gases referred to as 

the sweep-gas. The sweep-gas composition can be

varied to produce desired experimental conditions (for 

example, air for aerobic conditions and nitrogen for 

anaerobic experiments). A hydrocarbon-concentration 

gradient defined by high hydrocarbon concentration 

near the reservoir source and low concentration near 

the sweep gas was established. The composition of the 

gaseous phase of the porous media was determined by 

analyzing samples collected by syringe from the ports 

by using gas chromatography. Details of the methods 

are presented in Baker (1993).

Sediment samples were collected from the Galloway 

Township site with a hand auger. A white sand layer 

was sampled from the depth of the perched water table, 

about 3.5 m below land surface near the location vapor 

sampling well VW-9 (Fischer and others, 1996). This 

location is near the center of the hydrocarbon contaminant plume and also is a zone of active microbial degradation. Either the sand was directly placed in the 

glass columns (in the unprocessed-sediment experiment) or a fraction was prepared by sieving the sand 

and filling the columns with the predominant grain-size 

range of 0.25 to 1.0 mm (in the sieved-sand experiments). Sieved sand was prepared further by washing 

to remove fines and oven-drying at about 150 °C. The 

sieved sand was mixed with a small amount (2 percent) 

of unprocessed sediment from the area of vapor well 

VW-9 to inoculate it with microorganisms from the 

Galloway Township site. Additional inoculation was 

accomplished by flooding the porous media with 

hydrocarbon-contaminated ground water taken from 

VW-9, then draining by gravity.

Hater

NOT TO SCALE carbon

Figure 1. Porous-media-column design.

44
MATHEMATICAL MODELING

Mathematical modeling is required to determine 

the reaction rates for reactive vapor-phase chemical 

constituents (hydrocarbon, O2 , and CO2 ) as a function 

of column depth (Rk(Z)) and, therefore, as a function of 

gaseous-phase composition and moisture content, 

which vary with column depth. Reaction rates are 

determined by calibration of the model to experimental 

data. The model expands on the work of Baehr and 

Bruell (1990) to include reactive-transport quantification. The modeling approach is outlined here, and is 

presented in greater detail by Baker (1993).

If the time scale over which changes in reaction 

rates occur is long compared to the time scale defining 

the physical redistribution of gases, then it is appropriate 

to analyze the data from each synoptic sampling (sampling and analysis of the vapor-phase composition, 

requiring 2-3 days) with a steady-state model. Therefore, for each modeling application the reaction rates 

(Rk (Z)) can vary along the length of the column, and 

are fixed in time if a steady-state gas distribution is 

assumed. The long-term time dependence of the reaction rates can be defined by reapplying the steady-state 

model to the results of a subsequent synoptic sampling 

and redefining the reaction rates. This approach is 

defined as a quasi-steady-state approach-that is, the 

transient experimental behavior is approximated by a 

finite number of steady-state "snapshots."

The model is based on conservation-of-mass equations describing the transport and reaction of individual 

constituents of a multicomponent gas. For each constituent, the following equation applies:

k= 1,2,...., M (4)

where z is the distance in centimeters from the top of 

the porous media to a depth being considered in the column; Jk is the mass flux, in grams per square centimeter 

per second for the kth constituent; Rk(z) is the rate of 

production or consumption (reaction rate) of the klh 

constituent, in grams per cubic centimeter per second; 

and M is the number of constituents. Although the 

porous media may contain an aqueous phase, experimental conditions are such that the aqueous phase is 

stationary and is assumed to be in chemical equilibrium 

with the gaseous-phase. The fluxes, Jk , are defined by 

transport properties of the porous media, gaseous-phase 

concentration gradients, and boundary conditions

defined at the sweep-gas location. A key parameter 

used in the model is tortuosity (T), which relates bulk 

diffusion coefficients to the effective diffusion coefficients of gas-phase constituents in the porous media as 

follows

Dk = D>aT (5)

where Dk is the effective diffusion constant of k in the K u

porous media, D k is the bulk vapor-phase diffusion 

constant of k, 0a is the gas- filled porosity, and T is the 

tortuosity.

Although tortuosity can not be measured directly, it 

can be obtained by using a parameter-fitting technique 

included in the model in which concentration-gradient 

data and flux data for a gas-phase constituent are fitted 

to a depth-dependent distribution of tortuosity values. 

Any gas-phase constituent with a concentration profile 

that is not measurably affected by microbial degradation can be used in this calibration.

BIODEGRADATION RATES

Total rates of CO2 production and hydrocarbon 

biodegradation for each of the five experiments are 

shown in table 1. Sweep-gas CO2 concentration and 

sweep-gas flow rate were used to obtain the total CO2 

flux for each column which, at steady state, is equal to 

total CO2 production in the column. Hydrocarbondegradation rates were calculated from CO2-production rates by using stoichiometric relations, such as the 

one in equation 1.

For the three aromatic hydrocarbons studied, 

degradation rates increased as the number of methyl 

groups increased. Toluene degradation was more rapid 

in unprocessed sediment than in sieved sand, probably 

as the result of the presence of fine particles in the 

unprocessed sediment, which provided increased 

surface area and a more complete inorganic nutrient 

mixture for microbial activity.

Uniform hydrocarbon-degradation rates cannot be 

assumed over the length of a column because the 

columns are heterogeneous with respect to vertical 

moisture distribution and hydrocarbon concentration. 

Therefore, the degradation rates listed in table 1 are 

mass-averaged over the entire column, and depth-specific degradation rates (Rk(z)) are determined by application of the mathematical model.

45
Table 1. Total hydrocarbon biodegradation, experiments A-E 

[g, gram; ml, milliliter; mg/d, milligram per day

Experiment

A sieved

B sieved

C sieved

D sieved*

E unprocessed

Compound

Benzene

Toluene

p-Xylenc

Toluene

Toluene

Porous

mo/f io

dry mass 

(9)

6,653

6,681

6,699

6,663

7,124

Pore water

volume 

(ml)

163.4

217.7

229.2

309.7

263.9

C02

produced 

(mg/d)

3.7

5.3

16.6

0.0

27.7

Hydrocarbon

degraded 

(mg/d)

1.1

1.6

4.9

0.0

8.2

* sterile control

Tortuosity was determined as a function of column 

depth (T(Z)) through model calibration, and is plotted 

for Experiment A (sieved sand, benzene) in figure 2. 

Tortuosity near the lower boundary of the column was 

observed to be generally lower than tortuosity near the 

top of the column. This observation was anticipated 

because moisture content was higher at the bottoms of 

the columns than at the tops.

Depth-dependent CO2-production rates were 

determined by using the mathematical model; these 

rates were converted to hydrocarbon-degradation rates 

with the assumption of complete mineralization. 

Figure 3 shows benzene degradation as a function of 

column depth for Experiment A (sieved sand, benzene). 

Biodegradation rates were highest near the bottoms of 

the columns, where moisture levels were highest, yet

0.2 0.3 

TORTUOSITY

Figure 2. Tortuosity as a function of distance down the 

column, inoculated sieved sand with benzene.

-1 X NT10 0 1 X 10' 10 2 X 10' 10 3 X 10' 10 4 X 10' 10 

BENZENE DEGRADATION RATE, IN GRAMS PER CUBIC CENTIMETER PER SECOND

Figure 3. Degradation rate of benzene as a function of 

distance down the column for inoculated, sieved sand.

less than saturation. Biodegradation rates are expected 

to increase with moisture levels because the microbial 

community resides in the aqueous film adhering to the 

porous media, and because hydrocarbons are available 

for biodegradation only in dissolved form; therefore, 

the available mass of hydrocarbon substrate is proportional to the moisture content.

SUMMARY

A laboratory column experiment method in 

conjunction with a mathematical model was applied 

successfully to quantify hydrocarbon-biodegradation 

rates as a function of depth-dependent parameters 

(hydrocarbon concentration and moisture content) in 

unsaturated porous media. Under the experimental conditions presented, CC>2 flux was used as an indicator of 

hydrocarbon-degradation rates because the hydrocarbon and oxygen gaseous-phase concentration profiles

46
were not measurably affected by the microbial activity. 

This method could be used to determine optimum conditions for hydrocarbon biodegradation in the design of 

bioremediation strategies. Moisture, nutrient, and 

organic-compound substrate concentrations as well as 

temperature are variables that could be considered in 

the column-study design.

REFERENCES CITED

Atlas, R.M., 1984, Petroleum microbiology: New York, 

MacMillan Publishing Co., p. 99-106.

Bacdcckcr, M.J., Sicgcl, D.I., Bcnnctt, P., and Cozzarclli, 

I.M., 1989, Fate and effects of crude oil in a shallow 

aquifer I. The distribution of chemical species and 

gcochcmical facics, in Mallard, G.E., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program, Proceedings of the technical 

meeting. Phoenix, Ariz., September 26-30, 1988: U.S. 

Geological Survey Water-Resources Investigation 

Report 88-4220, p. 13-20.

Bachr, A.L., and Brucll, C.J., 1990, Applications of the 

Stcfan-Maxwcll equations to determine limitations of 

Pick's law when modeling organic vapor transport in 

sand columns: Water Resources Research, v. 26, no. 6, 

p. 1155-1163.

Bachr, A.L., and Fischcr, J.M., 1996, Overview of research 

on the transport, microbial degradation, and remediation of hydrocarbons at a subsurface gasoline-spill site 

in Galloway Township, New Jersey, in Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 95-4015.

Baker, R.J., 1993. Hydrocarbon microbial degradation rate 

determination in columns of unsaturatcd porous media 

using vapor transport modeling: Philadelphia, Pa.. 

Drcxcl University, unpublished Ph.D. dissertation, 

1993.

Brown, J.T., Noris, R.D., and Brubakcr, G.R., 1985. 

Pollution Engineering, v. 17, p. 25-28.

Fischcr, J.M., Baker, R.J., Lahvis, M.A., and Bachr, A.L., 

1996, Determination of vapor-phase diffusion coefficients for unsaturatcd-zonc sediments at a gasolinespill site in Galloway Township, New Jersey, in Morganwalp, D.W., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, Colorado 

Springs, Colo., September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

95-4015.

47
48
Method for Monitoring Biodegradation of Hydrocarbons

in Unsaturated Porous Media

By John G. Nolan1 , Ronald J. Baker 2 , and Arthur L. Baehr 2

Abstract

A method for monitoring reaction rates of 

reactants and products during biodegradation of 

hydrocarbons in unsaturated porous media was 

developed. Stainless-steel bioreactors were filled 

with porous media that had been collected from a 

gasoline-spill site and used in biodegradation 

column studies. The purpose of the bioreactor 

study was to test the hypothesis that the microbial 

population in the porous media taken from the 

columns was at steady-state with respect to 

biomass production and endogenous respiration. 

This hypothesis allows hydrocarbon-biodegradation rates to be calculated from carbon dioxide 

production rates or oxygen consumption rates if 

mineralization stoichiometry is assumed. Toluene 

and p-xylene were the hydrocarbons used.

INTRODUCTION

Gasoline spills cause widespread subsurfacecontamination problems in the United States. Remediation of gasoline-spill sites can be achieved by natural 

or enhanced biodegradation of gasoline hydrocarbons. 

Rates of microbial degradation of hydrocarbons are 

needed to estimate the rate of bioremediation. /// situ 

biodegradation rates can be inferred from results of 

microcosm experiments (Aamond and others, 1989) 

conducted with aqueous and solid samples collected 

from a contaminated site.

'Department of Civil Engineering, Drcxcl University. 

Philadelphia, Pa.

: U.S. Geological Survey, Trenton, N.J.

Typically, a microcosm experiment is conducted in 

a closed system. Because concentrations of biodegradable contaminants begin to decrease as soon as the 

microcosm is sealed, constant substrate concentrations 

and microbial populations are rarely achieved. In contrast, field conditions are open systems, in which quasisteady-state conditions with relatively stable microbial 

populations and hydrocarbon substrate concentrations 

would be expected. Therefore, kinetic information 

derived from microcosm experiments may not accurately reflect in situ biodegradation rates.

An alternative approach to estimating biodegradation rates in which conditions of hydrocarbon and oxygen (O2 ) concentrations and microbial populations are 

maintained in open columns of porous media is presented by Baker and Baehr (1996). In this approach, the 

hydrocarbon substrate is present in excess, and its utilization rate is not directly measurable. Therefore, stoichiometric relations describing hydrocarbon 

biodegradation are used to calculate hydrocarbon reaction rates from measured carbon dioxide (CO2 ) production rates or from O2 consumption rates. In order to 

verify the validity of such stoichiometric conversions, 

a means of determining complete mass balances of the 

principal bioreactive chemical constituents in known 

quantities of porous media is needed. Such a method 

was developed and verified in this study. The hypothesis was that mineralization (complete conversion to 

CO2 and water) is the dominant fate of the hydrocarbons in the columns, and that calculating hydrocarbon 

biodegradation from CO2 production or O2 consumption is appropriate. A stainless-steel bioreactor and a 

set of experiments were designed for testing this 

hypothesis. The objective of this study was to determine whether the microbial population in the porous 

media is at steady-state with respect to biomass production and endogenous respiration. This paper summarizes the experiments conducted to test this 

hypothesis. A detailed description is presented by 

Nolan (1993).

49
Experimental Method Table 1. Porous media used in bioreactors

Samples of porous media from the column experiments of Baker and Baehr (1996) were collected and 

and known amounts were placed into gas-tight stainless-steel bioreactors. The bioreactors were then spiked 

periodically (generally at 48-hour intervals) with a 

fixed amount of a hydrocarbon (either toluene or pxylene) and O2 . At the start of the experiment, known 

quantities of the specified hydrocarbon, O2 , and CO2 

were present in each bioreactor. Vapor samples 

extracted during the spiking process were analyzed by 

gas chromatography to quantify the mass of hydrocarbons, O2 , and CO2 present in the bioreactors at the 

time of sampling. Because all major bioreactive species were monitored, the reaction rates of the hydrocarbon, O2 , and CO2 could be determined at any time. 

Therefore, steady-state and mineralization assumptions could be tested.

The bioreactors consisted of !2-in.-long, l'/2 -in. 

I.D. Schedule 40 stainless-steel pipes with threaded 

stainless-steel endcaps. The endcaps were fitted with 

three vapor-sampling ports formed from three '/, 6-in. 

threaded holes cut into each bioreactor-two near the 

endcaps and one at the center. The port assembly 

included a threaded fitting sealed with a Viton O-ring 

('/ in.) and capped with a compression fitting (Vu, in.) 

and an 18-gauge stainless-steel syringe needle. The 

needle was capped with a syringe valve.

Unsaturated porous media for the experiments 

were taken from two columns in which unprocessed 

sediment from a gasoline-spill site in Galloway Township, N.J., had been exposed to aromatic hydrocarbons 

for a period exceeding 1 year (Baker and Baehr, 1996). 

One column experiment was conducted with toluene, 

the other with p-xylene. The porous media placed in 

each bioreactor were microbially acclimated to the 

hydrocarbon selected. Approximately 350 cm3 of sediment was packed into each bioreactor.

Materials (porous media) were taken from the 

toluene- and p-xylene-amended glass columns, and 

seven cylinder bioreactors were established and 

labeled as indicated in table 1. The sterile control bioreactors designated TC and XC were filled with a mixture 

of material from all three zones of the glass column, 

autoclaved for 2 hours at 120 °C, and dosed with silver 

nitrate (200-mg/L solution).

The bioreactors were spiked with CO2 -free air 

(about 79 percent nitrogen and 21 percent O2 by 

volume) containing known amounts of toluene or 

p-xylene. Every 48 hours a sample was collected from

Compound

Toluene

p-Xylene

Label

Tl

T2

T3

TC (control)

XI

X2

XC (control)

Location of material 

within glass column

Unsaturated /one

Water-table /one

Saturated /one

Mixture from all /ones

Unsaturated /one

Water-table /one

Mixture Irom all /ones

Water-filled 

porosity

0.0581

. 1 372

.2313

.1533

.1010

.111 4)

.1375

the bioreactor while an equal volume of hydrocarbon/ 

air mixture was injected. The sample was analyzed by 

gas chromatography as described in Baker (1993) to 

determine the hydrocarbon, O2 , and CO2 concentrations. A mass balance of these constituents was maintained throughout each experiment.

In order to maintain a mass balance, the quantities 

of hydrocarbons, O2 , and CO2 partitioned among the 

gaseous and aqueous phases in the bioreactors had to 

be considered. The solid phase was not considered 

because the sediment was not found to adsorb significant amounts of hydrocarbons, and would not be 

expected to retain O2 or CO2 . Aqueous-phase CO2 and 

O2 concentrations were calculated from vapor-phase 

concentrations if Henry's law is assumed. The cumulative numberofmillimoles of constituents consumed or 

produced in the reactor was monitored through time. 

Counts of total viable microorganisms were carried out 

at the beginning and end of each experiment.

BIODEGRADATION OF HYDROCARBONS

Hydrocarbon degradation and CO2 production in 

sterile control bioreactors TC and XC were negligible. 

Therefore, all hydrocarbon consumption and CO2 

production in the live reactors Tl through T3 and XI 

through X2 was attributed to microbial activity.

Table 2 shows cumulative millimoles of toluene 

degradation, CO2 production, and O2 consumption 

over time for experiments T1-T3. If mineralization is 

the dominant fate of the hydrocarbons, then the rates of 

change in the constituents are related according to the 

mineralization equation for toluene as follows:

1 C7 H8 + 9 O2 -> 7 CO2 + 4 H2O

For toluene mineralization, the relative reaction rates 

of reactants and products on a molar basis are 1:9:7 for 

toluene, O2 and CO2 , respectively.

50
Table 2. Results of experiments T1 through T3--Reactant and 

product reaction-rate data 

[Values are in cumulative millimoles]

Experiment

Tl

Tl

Tl

Tl

T2

T2

T2

T2

T3

T3

T3

T3

Time 

(days)

2.05

5.97

13.95

27.97

2.03

5.97

13.95

27.97

2.03

6.01

13.96

27.97

Toluene 

degraded

0.0295

.0763

.1452

.2712

.0184

.0542

.1216

.2385

.0178

.0539

.1253

.2437

Oxygen 

consumed

0.4334

1.0851

2.0599

3.6453

.1370

.3825

.6863

1.1832

.0024

.3730

.6708

1.2379

Carbon 

dioxide 

produced

0.2127

.5058

1 .6472

2.9848

.0890

.2723

.5894

1 .0460

.1347

.3414

.8236

1.2883

Normalized concentration data are shown in 

figures 1-3. The amount of each constituent degraded, 

produced, or consumed (from table 2) is divided by its 

corresponding stoichiometric coefficient from the 

mineralization equation. If the three lines showing 

rates of toluene degradation, O2 consumption, and CO2 

production as a function of time coincide, complete 

mineralization is indicated. If the slope of the hydrocarbon degradation line is less than those of the O2 consumption and CO2 production lines, then more O2 is 

being utilized and more CO2 is being produced than is 

accounted for by the degradation of the hydrocarbon. 

This observation indicates that there is net endogenous 

decay of microorganisms, and that the system is not in 

steady-state with respect to biomass. This condition 

was noted for Tl during the first 28 days of the experiment. When the slope of the hydrocarbon degradation

cc lu_ 

o

UJ

Q

X

o

CO

go

UJ

UJ

cc

CO 

UJ

cc

0.45

0.40

0.35

0.30

0.25

0.20

0.15

O-G TOLUENE

^-^ OXYGEN

B-S CARBON DIOXIDE

10 15 20 

TIME, IN DAYS

25

Figure 1. Stoichiometrically normalized cumulative toluene degradation, oxygen 

consumption, and carbon dioxide production as a function of time-results of 

experiment T1 (toluene, unsaturated zone).

51
CC

P

O

2

g

LU

O

g

P

COP

Q 

HI

HI 

CC

CO 

LU

i

o 

o

0.25

0.20

©-© TOLUENE

*-* OXYGEN

B-Q CARBON DIOXIDE

0.15

0.10

0.05

10 15 

TIME, IN DAYS

20 25 30

Figure 2. Stoichiometrically normalized cumulative toluene degradation, oxygen 

consumption, and carbon dioxide production as a function of time-results of 

experiment T2 (toluene, water-table zone).

line is greater than that of the O2 consumption and CO2 

production lines, more hydrocarbon is being consumed 

than can be accounted for by mineralization, which 

indicates utilization of toluene-carbon for biosynthesis 

at a rate exceeding that of endogenous decay. Again, 

the system is not in steady-state with respect to 

biomass. This condition was observed for T2 and T3 

during days 1-28.

Table 3 shows a comparison of measured cumulative hydrocarbon-degradation-rates to those calculated 

from measured cumulative CO2 production and those 

calculated from measured cumulative O2 consumption 

if the mineralization reaction is assumed. Two sampling times are reported for each experiment: halfway 

through the experiment and the last measurement 

made.

These data indicate that microbial populations 

were adjusting to the bioreactor environment during 

the experiment, either decreasing (Tl) or increasing 

(T2, T3, X2). Two questions need to be addressed for 

the application of the method: (1) is the bioreactor 

environment significantly different from the column 

environment, and could this difference lead to microbial-population changes, and (2) should mineralization 

stoichiometry be abandoned. The answer to the first 

question is that the saturated-unsaturated column environment was significantly different from that of the 

unsaturated bioreactors; hydrocarbon loading rates 

also were different. Therefore, adjustment of the 

microbial population would be expected. For the pxylene-amended bioreactors, counts of total viable 

microorganisms at the beginning and end of the exper52
DC

e

o

DC

I 

O

o

1

gQ 

LLJ

5

<

LU 

DC

to

LU

0.25

0.20

O-O TOLUENE 

OXYGEN 

CARBON DIOXIDE

0.15

0.10

^ 0.05

O 

O

10 15 

TIME, IN DAYS

20 25

Figure 3. Stoichiometrically normalized cumulative toluene degradation, oxygen 

consumption, and carbon dioxide production as a function of time-results of 

experiment T3 (toluene, saturated zone).

Table 3. Measured and Calculated Hydrocarbon Degradation Rates

Day

14 

14 

14

28 

28 

28

36

22

72 

46

Experiment

Tl 

T2 

T3

Tl 

T2 

T3

XI

X2

XI

X2

Millimoles of 

hydrocarbon 

consumed 

(measured)

0.2353 

.0842 

.1177

.4264 

.1494 

.1840

.3765 

.0442

.6497 

.1013

Calculated millimoles of hydrocarbon consumed

From CO 

production

0.1452 

.1216 

.1253

.2712 

.2385 

.2437

.3868 

.0900

.6916 

.1904

Percent 

error

+62.0 

-30.7 

-6.1

+57.2 

-37.3 

-24.5

-2.7 

-50.9

-6.0 

-46.8

FromO 

consumption

0.2289 

.0763 

.0745

.4050 

.1315 

.1375

.4265 

.0460

.6986

.1232

Percent 

error

+57.6 

-37.3 

-40.5

+49.3 

-44.9 

-43.6

+ 10.3 

-48.9

+ 1.0 

-35.3

53
iments are evidence of an adjustment in total viable 

count. The XI viable count remained relatively constant, decreasing slightly from 3.3 x 107 to 2.3 x 107 

viable cells per gram of sediment. This observation is 

consistent with the vapor-analysis reaction data, which 

closely follow the mineralization ratios of reactants and 

products. The X2 viable microbial count, however, 

increased from 3.3 x 103 to 4.9 x I0(l viable cells per 

gram of sediment, resulting in a loss of CO2 mass associated with biomass production. This phemonenon is 

consistent with the observed high rate of p-xylene utilization relative to the observed rate of CO2 production. 

The errors associated with calculating hydrocarbondegradation rates from CO2 production are significant; 

however, in the most extreme case, the rate of hydrocarbon degradation would be overestimated or underestimated by 50 percent on the basis of CO2 production. 

Therefore, the mineralization stoichiometry is valid.

CONCLUSIONS

A mass-balance approach to testing assumptions of 

biomass steady-state and hydrocarbon mineralization 

in porous media was developed. The stainless-steel 

bioreactors enabled long-term monitoring of bioreactive species in porous media while hydrocarbon 

and oxygen were added to the system periodically to 

approximate open-system conditions. The vaporanalysis-concentration data indicate that the mineralization hypothesis provides a reasonable stoichiometry 

for computing hydrocarbon-degradation rates from 

inorganic-gas-concentration data. Suggested improvements to the experimental method include extension of 

the period of the experiment in order to track the

transition to a mineralization-dominated process 

completely, while providing a more continuous (less 

incremental) way of introducing hydrocarbons and O2 

to the system. This could be achieved by increasing 

reactor size and sampling frequency and decreasing 

sample volume. Conditions can be maintained purposely to observe stoichiometry under conditions 

favoring biosynthesis (sparingly inoculated sediment) 

or decay (starvation conditions). Finally, radioactive 

tracers could be used to improve the accuracy of carbon 

mass-balance calculations.

REFERENCES

Aamand, J., Jorgcnscn, C., Arvin, E., and Jcnscn, B.K., 

1989, Microbial adaptation to degradation of hydrocarbons in polluted and unpolluted groundwatcr: Journal of Contaminant Hydrology, v. 4, no. 4, p. 299-312.

Baker, R.J, and Bachr, A.L., 1996, Use of column studies 

and a reactive transport model to measure biodegradation rates of hydrocarbons, //; Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 95-4015.

Baker, R.J., 1993, Hydrocarbon microbial degradation rate 

determination in columns of unsaturated porous media 

using vapor transport modeling: Philadelphia, Drcxel 

University, unpublished Ph.D., dissertation, 235 p.

Nolan, J.G.. 1993. Relationship between hydrocarbon biodegradation, oxygen consumption and carbon dioxide 

production in unsaturated porous media Philadephia, 

Drexel University, unpublished Master's thesis, 93 p.

54
Optimizing the Design of Vapor-Extraction Remediation 

Systems for Removal of Organic Contaminants From

the Unsaturated Zone

By Claire Welty1 , Craig J. Joss2, and Arthur L. Baehr3

Abstract

A new mathematical model has been 

developed that can be used in the design of vaporextraction and bioventing systems for removing 

organic contaminants from the unsaturated zone. 

The model has two major components: a threedimensional finite-difference model of air flow in 

unsaturated porous media (AIR3D), and an 

algorithm that couples the three-dimensional airflow model with mathematical programming 

(OPTAIR).

AIR3D is based on the flow of an ideal gas 

in porous media. The code consists of a 

preprocessor and postprocessor to be used with 

the U.S. Geological Survey ground-water flow 

simulator MODFLOW to solve the conservationof-mass equation for air flow in a discretized 

three-dimensional domain. OPTAIR couples 

AIR3D with an optimization algorithm, to allow a 

user to determine the optimal number and 

arrangement of well screens and rates of air 

withdrawal or injection in the domain to obtain 

the most efficient design.

INTRODUCTION

In situ vapor extraction has become a widely 

used technique for remediating unsaturated zones

'Drexel University, Department of Civil and Architectural 

Engineering, Philadelphia, Pa. 19104

2Applied Environmental Sciences, Ltd., P.O. Box 943, Bala 

Cynwyd, Pa. 19004

3 U.S. Geological Survey, West Trenton, N.J. 08628

contaminated with volatile chemicals. Figure 1 is a 

schematic of the process. Forcing air through the 

unsaturated zone also is recognized as a means of 

enhancing aerobic biodegradation of hydrocarbons. A 

method of predicting and quantifying both the flow 

paths of forced air in the unsaturated zone and the 

radius of influence resulting from air withdrawal or 

injection is needed for designing such vapor-extraction 

and bioventing systems. In addition, a method for 

determining the optimal number and arrangement of 

well screens and rates of air withdrawal and injection 

in the domain is desirable for optimizing cleanup 

efficiency. Two modeling tools have been developed 

for use in designing these systems for a given site: a 

three-dimensional numerical model of air flow 

through porous media, and an algorithm that couples 

the air-flow model with mathematical programming 

for evaluating the optimal combination of numbers, 

flow rates, and placement of air withdrawal, injection, 

and passive wells.

At the previous U.S. Geological Survey Toxic 

Substances Hydrology technical meeting, held in 

Monterey, California, in 1991, we presented a paper 

outlining our approach for modeling vapor-extraction 

systems in the unsaturated zone (Welty and others, 

199la). This paper describes the implementation of 

our approach and provides references for other reports 

resulting from this work.

DEVELOPMENT OF A THREEDIMENSIONAL AIR-FLOW MODEL

The three-dimensional equation for air flow is 

derived by combining the conservation-of-mass 

equation for air flow in unsaturated porous media and 

the Darcy equation for a compressible fluid. By 

assuming further that air density and pressure are

55
Release to atmosphere

Injected 

air

Treatment of 

withdrawn air

Vapor 

plume

Injected 

air

Water

Figure 1. Schematic diagram of the vapor-extraction process.

related by the ideal-gas law (Baehr and Hult, 1 99 1 ), 

the following field equation results:

(i)

where (f) = P2 , with P = air-phase pressure. The substitution (f) = P2 has been made to linearize the equation. 

The Cartesian coordinates have been aligned with the 

principal directions of anisotropy so that kxx, kyy, and 

k77 are the principal components of the air-phase permeability tensor. Air-phase permeability represents 

the permeability available to the gas phase in unsaturated porous media. It is not equivalent to intrinsic permeability as a result of the presence of liquid phases in 

the pore space.

Sa in eq. 1 is an air-storage term defined as

(2)

Sn = 9 K^"

where 0 is the air-filled porosity (assumed to be constant with respect to time) and fi is the dynamic viscosity of air. Equation 1 is linearized by assuming that 

<J> 1/2 in equation 2 is the (constant) prevailing atmospheric pressure. Equation 1 is derived by assuming 

that the Klinkenberg slip effect, the temperature gradient, and the elevation component of pneumatic head 

are negligible; the average molecular weight of air and 

air-filled porosity are constant; and the Darcy equation 

and the ideal-gas law apply. Baehr and Hult (1991)

table

evaluated the consequences of these assumptions and 

showed that the linear air-flow model given by eq. 1 is 

a good working model for vapor extraction and many 

other environmental applications.

The three-dimensional field equation for 

saturated ground-water flow is identical in 

mathematical form to eq. 1 when the following 

substitutions are made:

K; (i=x,y,z)

S.,

(3a)

(3b)

(3c)

where h is potentiometric head; Kxx , Kyy, and K77 are 

saturated hydraulic conductivities in the three principal directions of anisotropy; and Ss is specific storage. 

This mathematical equivalency can be used to advantage to solve eq. 1, because many well-documented, 

widely used computer codes have been published to 

solve the ground-water flow equation. We have taken 

the approach of using the ground-water-flow code 

MODFLOW (McDonald and Harbaugh, 1988) to 

solve eq. 1.

The air-flow model AIR3D consists of three 

components. First, a preprocessor (PREAIR) converts 

a pneumatic problem to a hydraulic-equivalent one.

56
Next, the hydraulic-equivalent problem is solved by 

using MODFLOW. For the last step, a postprocessor 

(POSTAIR) converts the hydraulic output to 

pneumatic units. PREAIR and POSTAIR are written 

in Fortran 77, and can be used on any computer with 

an installed Fortran 77 compiler. We have run these 

codes on PRIME 6350 and IBM 9121 mainframes, 

and on IBM and Macintosh personal computers. The 

three codes PREAIR, MODFLOW, and POSTAIR 

comprise the solution to eq. 1, which we have named 

AIR3D (Joss and Baehr, 1994).

In AIR3D, we model a well by specifying 

constant pressure along the well-screen length. (This 

approach also could be adapted for ground-water-flow 

modeling in place of specifying a uniformly 

distributed discharge over the length of a well screen.) 

This naturally simulates the variation in flow rate 

along the length of the well screen that occurs when a 

well passes through layers of varying air permeability. 

A disadvantage of this approach is that the volumetric 

and mass flow rates at the well corresponding to a 

specified pressure condition are not known until after a 

simulation is completed. If a particular volumetric or 

mass flow rate is desired, simulations must be 

performed iteratively to determine the well pressure 

that must be specified to yield the desired average flow 

rate. This procedure could be tedious for simulations 

involving multiple wells. The solution to this iterative 

problem is a default application of the optimization 

procedure discussed in the next section. In addition to 

the well-screen delineation, other boundary conditions 

can be specified to match a given application. For 

example, a macadam surface of limited area! extent 

can be part of the upper boundary, while the water 

table is fixed as a no-flow lower boundary.

Because vapor cleanup is proportional to 

calculated air-flow rates, modeling is an inexpensive 

alternative to field pilot testing. The accuracy of 

model predictions depends, however, on the 

availability of air-phase permeability data, which are 

site-specific. Joss and others (1992), and Baehr and 

Hult (1991) describe field techniques for evaluating 

unsaturated-zone air permeability. Baehr and Hult 

(1991) derived two-dimensional, axisymmetric 

analytical models for air flow to a single well in an 

unsaturated, homogeneous, anisotropic porous 

medium that can be used to solve the inverse air-flow 

problem (finding air-phase permeability from data 

obtained during a pneumatic test of the unsaturated 

zone). These analytic models have been used in the 

field application cited above and are implemented as

the Fortran 77 code AIR2D (J.C. Joss and A.L. Baehr, 

U.S. Geological Survey, written commun., 1993).

USE OF SIMULATION COUPLED WITH 

OPTIMIZATION OF THREE-DIMENSIONAL 

AIR FLOW

Although it is possible to find an efficient design 

for a vapor-extraction system by carrying out a 

number of simulations with various combinations of 

well-screen locations and air-injection/withdrawal 

rates, the most efficient design may not be found by 

using such a trial-and-error procedure. Many possible 

combinations of design parameters exist and 

identification of a best set of choices for test 

simulations may not be readily apparent for 

heterogeneous soils and complex site-boundary 

conditions. A more advanced level of design 

technology that builds on the numerical-simulation 

approach is formal optimization of the design 

variables, in which the best combination of number of 

wells, air-injection/withdrawal rates, and well 

locations is found by means of mathematicalprogramming (optimization) techniques commonly 

used in the field of operations research. A similar 

approach has been taken for two-dimensional aquifer 

management (Lefkoff and Gorelick, 1987). To 

optimize the design process, mathematicalprogramming algorithms can be coupled with AIR3D 

to determine the optimal set of design parameters for 

achieving stated venting objectives at a given site 

(Welty and others, 199la and 1991b). We have 

implemented this coupled model as the Fortran 77 

code OPTAIR (Welty and others, 1992). Typical 

objective functions that may be specified by a user 

include minimizing total costs, minimizing total-flow 

rates, or maximizing the flow of air over a subdomain.

OPTAIR consists of five modules, as illustrated 

by figure 2: PREAIR (from AIR3D), OPTAIR 1, 

MODFLOW, any standard optimization solver, and 

OPTAIR2. The physical conditions characterizing a 

site, including distribution of air-phase permeabilities, 

boundary conditions, and a chosen finite-difference 

grid, are specified as inputs to PREAIR. PREAIR is 

invoked to generate the hydraulic equivalent of the 

pneumatic problem written as MODFLOW input files. 

Trial well-screen locations, any physical or economic 

constraints, objective-function coefficients, and 

decision variables are specified as inputs to OPTAIR 1. 

For every trial well-screen location, OPTAIR 1 calls 

MODFLOW to evaluate the response of the physical 

system to the specified well-screen condition, and uses

57
Geologic information 

Boundary conditions 

Finite-difference grid

Preprocessor 

PREAIR

Trial well-screen

locations 

Constraints

Matrix generator 

OPTAIR1

MODFLOW

Optimization problem 

formulation

Optimization 

algorithm

Postprocessor 

OPTAIR2

Optimal flow rates 

(Hydraulic equivalent)

Optimal air-flow rates

and 

Well-screen locations

Optional application of

AIR3D 

{PREAIR, MODFLOW, POSTAIR)

Figure 2. Flow diagram showing the algorithm for OPTAIR.

the response-matrix method (Lee and Aronofsky, 

1958) to generate an optimization-problem 

formulation in standard mathematical-programming 

system (MPS) format. The current version of 

OPTAIR can generate any linear-programming 

problem; future versions will include mixedinteger/linear capabilities as outlined by Welty and 

others (1992), so that the fixed costs associated with 

well-screen installation can be included in specifying a 

cost-minimization problem.

Any optimization algorithm that reads MPSformatted input can be invoked to solve the linearprogramming problem as written by OPTAIR 1. We 

used LINDO (Schrage, 1991) on mainframes and 

personal computers. The output from the optimization 

algorithm specifies optimal flow rates at all trial wellscreen locations. An optimal flow rate of zero means 

that the trial well location should not be chosen as part 

of the final solution. The output from the optimization 

software is given in terms of the hydraulic-equivalent

problem, however, and must be converted to 

pneumatic units. This calculation is carried out by 

OPTAIR2. An independent check by the user is 

needed to determine whether the recommended flowpressure relations at stress points are physically 

obtainable that is, fall within a pump's capacity as 

defined by a rating curve.

INCORPORATION OF VAPORTRANSPORT FEATURES IN FUTURE 

RESEARCH

Joss (1993) developed a vapor-transport model 

to simulate chemical transport and degradation 

associated with vapor-extraction and bioventing 

systems. This model uses AIR3D to define the airflow field and allows estimation of transport of 

multiple constituents in the gas phase, including 

effects of partitioning into aqueous and solid phases,

58
chemical reactions, and biological degradation. As 

proposed by Welty and others (1991 a), the vaportransport model also can be coupled with optimization 

algorithms to include the effects of time-dependent, 

decreasing vapor-constituent concentrations on 

design of the optimal venting-well system. This 

would allow incorporation of the economics of off-gas 

treatment in designing an optimal system, because 

treatment costs depend on sizing of units which, in 

turn, depend on the time history of cleanup. In 

addition, the effects of a time-variable pumping 

schedule on the design of an optimal remediation 

system could be examined.

SUMMARY

This paper describes two numerical models we 

have developed for use in vapor-extraction-system 

design: AIR3D, a three-dimensional finite-difference 

model for simulating air flow in unsaturated porous 

media; and OPTAIR, a coupled simulationoptimization model that can be used to determine 

optimal pumping rates and well-screen locations in a 

three-dimensional unsaturated domain. AIR2D, based 

on an analytical solution, can be applied to data 

obtained during a pneumatic test of the unsaturated 

zone to determine in situ air-phase permeabilities. 

Extensions of this work to include chemical-transport 

considerations have been described. Making these 

codes available in the public domain will ensure 

maximum availability to geologists and engineers 

designing vapor-extraction and bioventing systems.

REFERENCES

Bachr, A.L., and Hult, M.F., 1991, Evaluation of

unsaturatcd-zonc air permeability through pneumatic 

tests: Water Resources Research, v. 27, no. 10, 

p. 2605-2617.

Joss, C.J., 1993, Multispccics vapor transport and reactions 

in unsaturated porous media under conditions of 

induced airflow: Philadelphia, Pa., Drcxcl University, 

Department of Civil and Architectural Engineering, 

unpublished Ph.D. dissertation, 112 p.

Joss, C.J., and Bachr, A.L., 1994, Documentation of

A1R3D, an adaptation of the ground-water-flow code 

MODFLOW to simulate three-dimensional airflow in

the unsaturated zone: U.S. Geological Survey Open 

File Report 94-533, 221 p.

Joss, C.J., Bachr, A.L., and Fischer, J.M., 1992, A field 

technique for determining unsaturatcd-zonc air 

permeability, in Proceedings of the symposium on soil 

venting, Houston, Texas, April 29-May 1, 1991: 

Washington, D.C., U.S. Environmental Protection 

Agency, Office of Research and Development, 

EPA/600/R-92/174, p. 149-170.

Lee, A.S., and Aronofsky, J.S., 1958, A linear 

programming model for scheduling crude oil 

production: Journal of Petroleum Technology, v. 10, 

no. 7, p. 51-54.

Lcfkoff, L.J. and Gorclick, S.M., 1987, AQMAN Linear 

and quadratic programming matrix generator using 

two-dimensional groundwatcr flow simulation for 

aquifer management modelling: U.S. Geological 

Survey Water-Resources Investigations Report 87- 

4061, 164 p.

McDonald, M.G., and Harbaugh, A.W., 1988, A modular 

three-dimensional finite-difference ground water flow 

model: Techniques of Water-Resources Investigations 

of the U.S. Geological Survey, book 6, chap. Al, 

586 p.

Schragc, Linus, 1991, User's manual for linear, integer, and 

quadratic programming with LINDO, Release 5.0: 

South San Francisco, Cal., The Scientific Press, 132 p.

Wclty, C., Bachr, A.L., Joss, C.J., and Dillow, J.J.A., 199la, 

Use of mathematical modeling coupled with a 

transport model to optimize the design of vapor 

extraction systems, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program-- Proceedings of the technical 

meeting, Montcrey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 276-279.

Wclty, C., Joss, C.J., and Bachr, A.L., 1992, A threedimensional air-flow model coupled with linear and 

mixcd-linear/intcgcr programming to optimize vapor 

extraction design: EOS, v. 73, no. 14, p. 124.

Wclty, C., Joss, C.J., Bachr, A.L., and Dillow, J.J.A., 1991b, 

Use of a three-dimensional air-flow model coupled 

with an optimization algorithm to aid in the design of 

soil venting systems, in Proceedings, American 

Petroleum Institute/National Water Well Association 

conference and exposition on petroleum hydrocarbons 

and organic chemicals in groundwatcr, Houston, 

Texas, November 20-22, 1991: Columbus, Ohio, 

Water Well Journal Publishing Company, p. 221 -232.

59
09
A Multispecies Transport Model to Simulate Vapor

Extraction and Bioventing Remediation of Unsaturated

Zones Contaminated with Gasoline

ByC.J. Joss 1 and A.L Baehr2

Abstract

Vapor extraction is an established method 

for removing spilled gasoline from the unsaturated 

zone. The technology involves the establishment 

of an unsaturated zone air-flow field by withdrawing and injecting air through dry well sand 

trenches. The success of a vapor-extraction installation depends fundamentally on the establishment 

of an air-flow field that intersects the contaminant 

and on the partitioning of the contaminant into the 

gaseous phase. Bioventing is a term applied to 

simultaneous venting and enhanced biodegradation of a contaminant. Aerobic biodegradation is 

stimulated by oxygen delivered to the unsaturated 

zone by air injection and withdrawal. A model to 

predict the transport and reaction of multiple 

chemical species in the gaseous phase under conditions of induced air flow has been developed to 

simulate the performance of vapor-extractionbased remediation. The model consists of a flow 

module and a reactive transport module. The flow 

module, named AIR3D, based on an adaptation of 

a ground-water flow simulator, defines the threedimensional air-velocity field within which the 

chemicals are transported and react. The transport 

module solves a coupled system of conservationof-mass equations, one for each chemical. Advection-domination transport is assumed, which 

allows the conservation-of-mass equations to be 

defined along streamlines. In this paper, model 

applications and assumptions are summarized.

'Department of Civil Engineering, Drexcl University, 

Philadelphia, Pa.

2U.S. Geological Survey, West Trenton, N.J.

INTRODUCTION

Hydrocarbon contamination of ground water 

resulting from gasoline spills is a serious problem 

throughout the world. Remedial technologies that take 

advantage of vapor transport include vapor extraction, 

bioventing, and air sparging. Vapor extraction 

removes hydrocarbons from the unsaturated zone by 

inducing the flow of air laden with volatile 

hydrocarbons to extraction wells and off-gas treatment 

systems (fig. 1). Bioventing generally refers to an 

auxiliary benefit achieved by vapor extraction, an 

increase in aerobic biodegradation of hydrocarbons in 

the unsaturated zone as a result of increasing oxygen 

recharge. Air-sparging technology effects remediation 

of contaminated ground water by the injection of air 

beneath the water table. Contaminated vapor then 

enters the unsaturated zone, where a vapor-extraction 

system can be used to collect and treat the gases.

A model to predict the advective transport and 

reaction of multiple chemical species in the gaseous 

phase under conditions of induced air flow has been 

developed to simulate the performance of vaporextraction based remediation. The model development 

is presented in detail by Joss (1993). The objective of 

this paper is to outline the model assumptions and 

capabilities.

MODELING METHOD

Air-flow paths can be complex as a result of the 

effects of sediment layering, moisture variation, 

multiple well locations, and boundary effects caused 

by surface structures and paving. Therefore, in 

general, a numerical simulation is required to provide a 

three-dimensional depiction of the induced air-flow 

field. The authors have developed a computer code, 

named AIR3D, for this purpose (Joss and Baehr

61
N

EXPLANATION

Line of equal pressure

Air flow

Valve

Well screen

Bentonite seal

Sand pack

Extraction

well Air-water 

separator

Pneumatic 

pump

Off-gas 

treatment unit

_ Secondary 

emmisions

NOT TO SCALE

Figure 1. Components of a vapor extraction system.

(1995). AIR3D is used to define the velocity of the 

gaseous phase without considering the transport of 

individual constituents (for example, N2, O2, and 

CO2 ). Therefore, in the modeling approach, air flow is 

uncoupled from transport. It is anticipated that AIR3D 

will be applicable for making basic vapor-extraction 

design decisions (for example, the number of wells and 

the depths of their screened intervals).

Predictions of hydrocarbon-specific recovery 

rates (for example, the time-dependent discharge of 

benzene to the atmosphere from a vapor-extraction 

system for remediating a gasoline-spill site) and 

bioventing performance require a transport model. The 

following system of equations (Joss, 1993) defines 

advection-dominated receive transport for any number 

of chemical species along a streamline in the induced 

flow field:

= 0,

k =

In this equation, k is the index for the kth 

constituent of the gaseous phase; Nt- is the number of 

gaseous-phase constituents that originate from and 

partition significantly into an immiscible phase (for 

example, the hydrocarbons of gasoline); and N is the 

total number of gaseous-phase constituents, where 

N>Nj would indicate that gases other than those 

originating from the immiscible phase (for example, 

O2 , CO2 and N2) are part of the simulation. The 

variable t is time, in seconds; £ denotes the length 

coordinate, in centimeters, along any streamline; and 

qC, in centimeters per second, is the magnitude of the 

gaseous phase specific discharge vector, q, in the 

streamline direction. Streamlines are defined 

numerically according to an uncoupled, steady-state 

prediction of the induced flow field, which can be 

generated by AIR3D. The divergence term, (V»q), also 

is evaluated numerically from the steady-state flow 

simulation and is non-zero because the gas phase is 

compressible. The variable Tk indicates either the 

gaseous-phase concentration of the k1 constituent, G^., 

in grams per cubic centimeter of the gaseous phase (if 

k>Nf) or the concentration of the kth constituent in the 

immiscible phase, 1^, in grams per cubic centimeter of 

the immiscible phase (if/;<#,). The definition of the 

storage coefficient ak, which is unitless, depends on the 

definition Tk, and is based on the multiphase

62
equilibrium model developed by Baehr and 

Corapcioglu (1987a, 1987b) and applied by Joss 

(1993). The reaction term, Rk, in grams per cubic 

centimeter of the porous media per second, allows for 

quantification of microbial reactions or any type of 

chemical reaction involving transformation from one 

constituent to another.

The model equations are solved numerically. 

The spatial derivatives dG^/dC, are evaluated by finitedifference approximations, given a descretization of 

the C, axis. A system of N ordinary differential 

equations, which defines the time-dependence of 

chemical concentrations at each node, results. The 

system of equations is solved by the modified-Euler 

algorithm. This solution technique does not require 

matrix inversions and, therefore, is very efficient. 

Because dispersion is neglected, the solution defined 

on any streamline is completely uncoupled from the 

solution on any other streamline. Therefore, the 

solution throughout the three-dimensional domain is 

refined by simply defining additional streamlines and 

repeating the algorithm. Furthermore, because 

equation (1) is a first-order partial differential equation, 

boundary conditions, in the form of specified 

concentrations, are required only for the upstream end 

of a streamline (either at land surface or at an injection 

well) located at ^=0. The numerical algorithm is 

presented in detail by Joss (1993).

SUMMARY

The environmental industry is faced with the 

task of removing volatile organic chemicals from the 

subsurface. Remediation techniques such as vapor 

extraction, bioventing, and air sparging are being used; 

therefore, a model capable of predicting chemical fate 

and transport in realistic induced air-flow fields is 

needed to facilitate design. The model outlined in this

paper was developed to satisfy the need. The model 

consists of a flow module and a reactive transport module. First, AIR3D is used to predict flow and pressure 

response in three-dimensional, heterogeneous, anisotropic, multiwell domains. Steady-state flow and pressure information from the flow simulator is transferred 

to the transport module through use of data files. Multispecies reactive transport is then modeled by applying 

the conservation-of-mass principle to each constituent 

of the gas phase. The assumption of advection-dominated transport results in a system of first-order partial 

differential equations defining the conservation-ofmass principle, which are solved along streamlines. 

The solution defined on any streamline is completely 

uncoupled form the solution on any other streamline. 

Therefore, the solution throughout the three-dimensional domain is refined simply by defining additional 

streamlines and repeating the basic algorithm. This 

algorithm is very efficient, allowing realistic threedimensional simulations to be performed on personal 

computers.

REFERENCES

Bachr, A.L., and Corapcioglu, M.Y., 1987a, A compositional multiphase model for ground water contamination by petroleum products, 1 Theoretical 

considerations: Water Resources Research, v. 23, 

no. l,p. 191-200.

Bachr, A.L., and Corapcioglu, M.Y., 1987b, A compositional multiphase model for ground water contamination by petroleum products, 2 Numerical solution: 

Water Resources Research, v. 23, no. 1. p. 201-213.

Joss, C.J., 1993, Multispecics vapor transport and reactions 

in unsaturated porous media under conditions of 

induced air flow: Philadelphia, Pa., Drcxcl University, 

Department of Civil Engineering, unpublished Ph.D. 

dissertation, 103 p.

63
64
A Comparison of Two Analytical Techniques For 

Identifying and Measuring the Concentrations of Volatile

Hydrocarbons in Ground Water

SyCurtis S. Phinney 1 and Isabella M. Cozzarelli 1

Abstract

Identification of monoaromatic hydrocarbons in ground water is commonly performed in 

circumstances where petrochemical contamination is suspected. Presence of certain of these 

species is considered indicative of petrochemical 

contamination. Most studies of ground-water 

contamination focus on the fate of benzene, 

ethylbenzene, toluene, and the xylenes (BTEX). 

However, a broad range of C6-Cj 0 aromatic 

hydrocarbons can be present in ground water 

contaminated with petroleum products. In this 

study, results of a comparison of two analytical 

techniques for the identifying and measuring the 

concentrations of C6-C| 0 monoaromatic hydrocarbons are presented. The techniques investigated are pentane/water microextraction 

followed by gas chromatography/flame ionization 

detection (GC-FID), and purge and trap (openloop) stripping of the sample followed by GC-Ion 

Trap detection (GC-ITD). Open-loop/GC-ITD 

allows the identification of 35 C6-C|Q aromatic 

hydrocarbons, compared to 21 identified by 

pentane/GC-FID. Compared to the pentane/GC 

method, higher sensitivity, higher selectivity, but 

lower precision were observed for samples 

analyzed by open-loop/GC-ITD.

Hydrocarbon concentrations measured by 

the two techniques were generally in good 

agreement, within 88.6-99.7 percent for total

U.S. Geological Survey, Reston, Va.

mono-aromatics. Concentrations of some trace 

constituents measured by pentane/GC-FID were 

higher under certain circumstances. For example, 

difficulty in resolving benzene from interferences 

by the pentane/GC-FID method sometimes 

resulted in considerably higher benzene and 

BTEX concentrations compared to open-loop/ 

GC-ITD. At a sample site where hydrocarbon 

concentrations were low, mean BTEX concentrations (n=10) were 661 |J,g/L (micrograms per liter) 

by pentane/GC-FID compared to 490 jj,g/L by 

open-loop/GC-ITD. At a high concentration sample site, mean BTEX concentrations (n=10) were 

5,077 jJ,g/L by pentane/GC-FID compared to 

5,089 ng/L by open-loop/GC-ITD.

INTRODUCTION

Techniques describing the detection and 

measurement of benzene, toluene, ethylbenzene, and 

the xylenes (BTEX) in ground water contaminated by 

petroleum products are well documented. These 

monoaromatic hydrocarbons are of interest because of 

the toxicity of the compounds and their relatively high 

solubilities and corresponding mobilities in groundwater systems (Wilson and others 1986; Barker and 

others, 1987). Recent studies have shown that, in 

addition to BTEX compounds, a broad range of other 

C6-C| 0 aromatic hydrocarbons are present in ground 

water contaminated with crude oil (Eganhouse and 

others, 1993), landfill leachate (Forst and others, 

1989), and gasoline (Baedecker and others, 1991). 

Information on these additional compounds is potentially useful in studying the fate of petroleum hydrocarbons in ground water. The demonstrated resistance

65
of certain monoaromatic hydrocarbons to degradative 

processes in the environment suggests their utility as 

indicators of petroleum contamination. These relatively stable constituents can be transported farther 

than the BTEX compounds and may become the most 

abundant compound class as they propagate through 

the ground water system (Eganhouse and others, 

1987). Improved knowledge of the fate and transport 

of a wider range of C6-C|o monoaromatic species 

could add considerably to the understanding of the 

physical and biological processes controlling the fate 

of petroleum hydrocarbons in shallow aquifers.

This study investigated the relative merits of 

two analytical methods that allow quantitative determination of C6-C|o aromatic hydrocarbons in ground 

water. The first is a pentane/water microextraction 

followed by gas chromatography with flameionization detection (pentane/GC-FID). The second 

method involves purge and trap stripping of the water 

sample with helium followed by gas chromatography 

with ion-trap mass-spectrometer detection (open-loop/ 

GC-ITD). The pentane/GC-FID method required less 

elaborate analytical instrumentation and provides 

reliable information for a wide range of low molecular 

weight aromatic hydrocarbons. The purge and trap 

technique coupled with mass spectrometric detection 

provided highly specific qualitative determinations 

and in the case of the ion trap mass spectrometer, also 

provides a lower quantitative method detection limit 

(MDL). However, this instrumentation is relatively 

expensive and requires a higher level of operator 

proficiency during the collection and subsequent 

reduction of sample data.

This paper presents a comparison of the specificity, sensitivity, and reproducibility of the two 

methods for high and low concentrations of monoaromatic hydrocarbons in ground-water samples from 

two well-characterized locations near a subsurface 

gasoline spill in Galloway Township, New Jersey 

(Fischer and others, 1991).

DESCRIPTION OF EXPERIMENT

Site description

The field site for this study is located in 

Galloway Township, N.J. This site has been under 

investigation by the U.S. Geological Survey and the 

New Jersey Department of Environmental Protection

since 1987. The site is approximately 10 mi. west of 

Atlantic City, N.J., and overlies the Cohansey Sand of 

the Kirkwood-Cohansey aquifer system (Zapecza, 

1989). Gasoline leaking from a single storage tank on 

a farm contaminated ground water at a shallow 

domestic well (Fischer and others, 1991). Two wellcharacterized sample locations were chosen to provide 

samples representing the range of contaminant concentrations encountered at the site. The first sample 

location, (Well VW-11, 22 feet deep, about 65 feet 

from the former gasoline tank), is relatively uncontaminated, whereas the second sample location (Well 

VW-9, 18 feet deep, about 25 feet from the former 

gasoline tank) had historically yielded water samples 

that contained high concentrations of monoaromatic 

hydrocarbons but no detectable free product.

Sample collection

Ground-water samples were collected from two 

stainless-steel nested monitoring probes, screened at 

depths of 18 feet and 22 feet, respectively. These 

probes have a 0.25 in. outside diameter (o.d.), and are 

slotted over a 6 in. interval; the probes were placed in 

a 4 in. augured hole. Water samples were withdrawn 

from the probes by fitting 1/4 in. inside diameter (i.d.) 

silicone tubing over the probe ends and withdrawing 

sample by use of a peristaltic pump. The sampling 

probes were flushed by allowing 10 to 12 L of ground 

water to pass through the sampling apparatus prior to 

sample collection. Samples for pentane/GC-FID 

analysis were collected in acid-washed 250 mL 

brown-glass bottles fitted with a Teflon/silicone 

septum cap. Samples for open-loop/GC-ITD were 

collected in commercially prepared 40 mL vials fitted 

with a Teflon/si licone septum cap.

Prior to being placed on ice for shipment to the 

lab, samples were preserved with mercuric chloride, 

which was added to produce a Hg concentration of 

40 mg/L. Deuterated or//?r;-xylene was added, as a 

recovery surrogate, to produce a concentration of 

0.20 mg/L in all open-loop/GC-ITD samples. The 

amount of sample submitted to the purge-and-trap 

device was adjusted with respect to the anticipated 

analyte concentration (that is, 20-400 uL). In the 

case of the pentane/GC-FID samples where the entire 

sample was extracted, water samples from wells were 

spiked with deuterated ortho-\y\ene to concentrations of 0.05 mg/L and 0.26 mg/L for the low-concentration and high-concentration wells, respectively.

66
Ten replicate samples were collected from each location 

for analysis by each technique. Collected sample volumes were approximately 250 mL for pentane/GC-FID 

and 40 mL for open-loop/GC-ITD analysis. Field 

blanks were prepared with deionized water in the fieldsite laboratory trailer and were spiked with deuterated 

ort/zo-xylene to a concentration of 0.20 mg/L.

ANALYTICAL TECHNIQUES

Pentane/GC-FID

In the laboratory, the 250 mL sample bottles 

were extracted with 2.0 mL of pesticide-grade pentane 

after first removing and discarding 2.0 mL of sample. 

The bottles were carefully resealed and inverted several times to disperse the solvent evenly throughout 

the sample. The bottles were then placed on an orbital 

shaker table and agitated for 1.5 hours. The pentane 

was allowed to rise to the top of the sample bottle by 

settling at 4 °C for 1 hour. The pentane fraction was 

then transferred to a 1.8 mL vial. Gas chromatographic analysis was performed on a high resolution 

capillary gas chromatograph equipped with a cool oncolumn injector and flame-ionization detector. The 

instrument was fitted with a 30 M x 0.32 mm DBWAX column, with a film thickness of 0.25 um. Chromatographic conditions were as follows: Oven: 20 °C, 

for 4 minutes, 4 °C per minute to 1 lOoC; 6 °C per 

minute to 175 °C; 8 °C per minute to 200 °C, 5 minute 

hold. Injector: 30 °C for 1 minute; 88 °C per minute to 

200 °C, held for 20 minutes. Detector: 240 °C, 

range=l, attenuation= 1. Helium carrier gas with flow 

controlled at 1.16 mL per minute, which corresponds 

to an average linear velocity of 21.4 cm/s.

Open-loop/GC-ITD

For open-loop/GC-ITD analysis, appropriate 

volumes of sample (20-400 fiL) were introduced into a 

5.0 mL syringe. The syringe was previously filled 

with an appropriate volume of interference-free 

dilution water, spiked with deuterated ethyl benzene 

(internal standard). This mixture was then injected 

into a liquid-sample concentrator (LSC). The transfer 

line from the LSC was fitted directly to the chromatographic column with a 1/16" stainless-steel union. 

Chromatographic separation of the sample components was performed on a 30 M x 0.25 mm DB-5

column, with a 0.25 um film thickness. The 5.0 mL 

water-sample mixtures were purged with He at 40 mL/ 

min. onto an adsorbent trap for 11 minutes. Water was 

selectively removed from the trap prior to desorbtion 

by ambient-temperature purging of the trap to atmosphere for a period of 5 minutes. Sample components 

sorbed onto the trap was desorbed onto the chromatographic column by heating the trap to 180 °C for 4 

minutes. The trap was then vented to atmosphere and 

baked at 225 °C for 7 minutes to condition the adsorbent for the next sample. The sample components 

were refocused during desorption by whole-column 

cryofocusing, that is, cooling the entire column oven 

to -40 °C using liquid nitrogen. The LSC transfer line 

and switching valve were held isothermally at 150 °C 

to enable efficient transfer of material up through the 

CIO monoaromatic hydrocarbons. Cryofocusing and 

subsequent gas-chromatographic separation of the desorbed components were performed under the following conditions: Oven: -40 °C, held for 4 minutes; 

50 °C per minute to 20 °C; 2.8 °C per minute to 88 °C; 

10 °C per minute to 200 °C, held for 5 minutes. 

Helium carrier gas, pressure controlled at 11.0 lb/in2 , 

which corresponds to an average linear velocity of 

30 cm/s at 200 °C column temperature. The column 

was interfaced directly to the ion trap detector (ITD), 

bypassing the open split interface, through the ITD 

transfer line and directly into the mass spectrometer 

ion source.

RESULTS

It was possible to identify 35 C6 -C]Q aromatic 

hydrocarbons by the open-loop/GC-ITD method 

compared to 21 components identified by the pentane/ 

GC-FID method. Higher specificity (attributable to 

elimination of coeluting interferences, for example), 

sensitivity (MDL's up to 455 times lower with GCITD than with GC-FID) and variability (GC-ITD 

coefficient of variation 2.6 times higher for total C6-C | () 

monoaromatics compared to GC-FID) were observed 

for samples analyzed by the open-loop/GC-ITD 

method compared to those analyzed by the pentane/ 

GC-FID method. Difficulty in resolving benzene from 

interferences, tentatively identified as cycloalkanes 

and branched alkenes, sometimes resulted in considerably higher benzene and BTEX concentrations by the 

pentane/GC-FID method compared to open-loop/GCITD. At the low-concentration sample site, mean 

BTEX concentrations (n= 10) were 661 |ag/L by

67
pentane/GC-FID compared to 490 Jug/L by open-loop/ 

GC-ITD. At the high-concentration sample site, mean 

BTEX concentrations (n=10) were 5,077 fag/L by 

pentane/GC-FID compared to 5,089 |Ug/L by openloop/GC-ITD. The measured concentrations of the 

total C6-C|Q monoaromatics were consistently in 

agreement between the two techniques at low and high 

concentrations (table 1). These measurements include 

14 compounds identified by ITD/MS that were 

obscured by interferences or coeluting analyte peaks 

by GC/FID detection. Using the pentane/GC-FID 

technique, the coeluting components were incorrectly 

identified, but their contribution to the overall FID 

response are reflected in the reported totals.

CONCLUSIONS

Open-loop stripping followed by mass-spectrometric detection of C6-C ]0 monoaromatic hydrocarbons yields results that are somewhat more variable 

than pentane/GC-FID, but the lower MDL's of openloop/GC-ITD are significant (table 2). The ability to 

identify and quantify individual components that may 

be poorly resolved chromatographically provides an 

opportunity to improve an understanding of the fate 

and transport of individual monoaromatic hydrocarbons in ground-water systems.

Pentane microextraction/GC-FID was generally 

found to yield highly reproducible data that was in 

good agreement with data from open-loop/GC-ITD, 

with the advantage of lower instrumentation costs and 

less complex operation. Pentane microextraction/GCFID has the disadvantages of lower overall sensitivity 

and no ability to distinguish coeluting analyte peaks 

and interferences from target compounds. These interferences give a positive bias on reported compound 

levels and were most pronounced on measurements of 

monoaromatic hydrocarbons at low concentration 

sample sites.

Table 1.-Concentrations of selected Q-Qo monoaromatic hydrocarbon in gasoline-contaminated ground water at Galloway 

Township, New Jersey

[mg/L, milligrams per liter; GC-FID, gas chromatography-flame ionization detection; ITD, ion trap detector; C.V., 

coefficient of variation]

Low Concentration Site

Pentane/GC-FID Open-Loop/ITD

High Concentration Site

Pentane/GC-FID Open-Loop/lTD

Compound mg/L CV. mg/L C.V. C.V. C.V.

Benzene

Toluene

Ethyl Benzene

O-Xylene

N-Propyl Benzene

1,2,4-Trimethyl Benzene

l-Ethyl-2-Methyl Benzene

1-Methyl Propyl Benzene

1,2,3,5-Tetramethyl Benzene

BTEX

Total C8 Monoaromatics

Total C9 Monoaromatics

Total CIO Monoaromatics

Total C6-C10 Monoaromatics

0.008

.008

.045

.254

.137

.615

.212

.009

.091

.661

.645

1.481

.854

2.996

70.898

3.201

4.914

2.851

3.478

3.646

3.182

3.560

4.066

3.480

3.379

3.634

3.511

3.435

<0.003

.008

.044

.210

.099

.494

.165

.008

.073

.490

.482

1.568

.596

2.654

N/A

4.772

6.761

4.609

5.273

4.594

5.968

5.804

5.059

5.195

5.207

4.821

4.748

4.816

0.438

1.466

.383

.713

.083

.585

.182

.004

.068

5.077

3.172

1.299

.497

6.974

3.006

2.901

3.102

3.133

2.999

3.149

2.797

3.104

3.112

2.805

3.016

3.181

3.030

2.841

0.336

1.730

.485

.770

.061

.586

.161

.004

.054

5.089

3.023

1.584

.324

6.997

9.712

7.536

8.335

7.317

7.770

8.832

7.999

7.027

7.593

7.346

7.127

8.130

7.427

7.475

68
Table 2. Method detection limits (MDL) for selected C6-Cio monoaromatic 

hydrocarbons

[FID, flame ionization detection; ug/L, micrograms per liter]

Compound

Benzene

Toluene

Ethyl Benzene

O-Xylenc

N-Propyl Benzene

1,2,4-Trimethyl Benzene

1,2,3,5-Tetramethyl Benzene

Pentane/FID 

JMDL 

(ug/L)

1.36

.739

.322

.254

.171

.178

.355

Open-Loop/lTD 

JMDL 

(ug/L)

0.003

.003

.002

.002

.003

.004

.004

1MDL = t(N-l, l-a = .99) X Sc

The ability to distinguish individual monoaromatic hydrocarbons in samples from ground-water 

monitoring wells improves the understanding of the 

fate and transport of hydrocarbons in ground-water 

systems. Use of the ion trap mass spectrometer as a 

chromatographic detector demonstrated distinct 

advantages in the measurement of trace or obscured 

components in these types of samples. These advantages include the ability to detect multiple components 

within a chromatographic peak and the ability to 

exclude interferences from the measured response of 

the target compounds. In addition the superior sensitivity of the ion-trap mass spectrometer allows the 

measurement of very low concentrations of monoaromatic hydrocarbons in water samples.

REFERENCES

Baedecker, M.J., Cozzarelli, I. and Phinney, C., 1991, 

Distribution of organic and inorganic constituents in 

ground water at Galloway Township, New Jersey, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. Geological 

Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, Monterey, Calif, 

March 11-15, 1991: U.S. Geological Survey Water 

Resources Investigations Report 91-4034, p. 287-293.

Barker, J., Patrick, G. and Major, D., 1987, Natural attenuation of aromatic hydrocarbons in a shallow sand aquifer: Ground Water Monitoring Review, v. 7, p. 64-71.

Eganhouse, R.P., Dorsey, T.F., and Phinney, C.S., 1987, Transport and fate of monoaromatic hydrocarbons in the subsurface, Bemidji, Minnesota, research site, in Franks, B.J.,

cd., U.S. Geological Survey Program on Toxic Waste 

Ground-Water Contamination Proceedings of the third 

technical meeting, Pensacola, Fla., March 23 27, 1987: 

U.S. Geological Survey Open-File Report 87-109. p. C-29.

Eganhouse, R.P., Dorsey, T.F., Phinney, C.S., and Wcstcott, 

A.M., 1993, Determination of C6-C| 0 aromatic hydrocarbons in water by purge-and-trap capillary gas chromatography: Journal of Chromatography, v. 628, p. 81-92.

Fischcr, J.M., Baehr, Arthur, and Smith, N.P., 1991, An 

investigation of the transport and microbial degradation of hydrocarbons in the subsurface at a gasolinespill site in Galloway Township, New Jersey, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. Geological Survey Toxic Substances Hydrology Program 

Proceedings of the technical meeting, Monterey, Calif, 

March 11-15, 1991: U.S. Geological Survey WaterResources Investigations Report 91-4034, p. 243-249.

Forst, C., Stieglitz, L., Roth, W. and Kuhnmunch, S., 1989, 

Quantitative analysis of volatile organic compounds in 

landfill leachatcs: International Journal of Environmental Analytical Chemistry, 37, p. 287-293.

Glaser, J., Focrst, D., McKee, G., Quave, S., and Buddc, W., 

1981, Trace analyses for wastewaters: Environmental 

Science and Technology, 15, p. 1426-1435.

Wilson, B.H., Smith, G., and Rccs, J., 1986, Biotransformations of selected alkylbenzencs and halogcnated aliphatic hydrocarbons in mcthanogcnic aquifer 

material A microcosm study: Environmental 

Science and Technology, v. 20, p. 997-1002.

Zapecza, O.S., 1989, Hydrologic framework of the New 

Jersey Costal Plain. Regional Aquifer-System Analysis - Northern Atlantic Coastal Plain: U.S. 

Geological Survey Professional Paper 1404-B, 49 p.

69
70
Overview of Research on Use of Hydrologic, Geophysical,

and Geochemical Methods to Characterize Flow and

Chemical Transport in Fractured Rock at the

Mirror Lake Site, New Hampshire

By Alien M. Shapiro1 and Paul A. Hsieh2

Abstract

Research efforts at the Mirror Lake ToxicSubstances Hydrology Program research site in 

Grafton County, New Hampshire have focused on 

the development of equipment, field testing 

procedures, and interdisciplinary interpretive 

approaches of characterizing ground-water flow 

and chemical transport in fractured rock over 

distances that range from tens of meters to kilometers. This paper summarizes the results of current 

research in bedrock hydrogeology in local- and 

regional-scale investigations at the Mirror Lake 

site. Local-scale investigations are conducted over 

distances of tens of meters and focus on identification of (1) fractures and fracture properties on 

exposed surfaces, (2) fractures in the subsurface 

using borehole and surface geophysics, and 

(3) hydraulic and transport properties of fractures 

by means of hydrologic testing. In regional-scale 

investigations, controlled hydrologic testing 

cannot be conducted to identify hydraulic and 

transport properties of the bedrock. Regional-scale 

investigations have focused on (1) methods of 

collecting hydrologic and geochemical information in heterogeneous bedrock environments, and 

(2) interdisciplinary methods of synthesizing these 

data by ground-water-flow and chemical-transport 

modeling to infer hydraulic and transport properties of the bedrock.

U.S. Geological Survey, Rcston, Va.

U.S. Geological Survey, Mcnlo Park, Calif.

INTRODUCTION

The U.S. Geological Survey (USGS) Mirror Lake 

Toxic-Substances Hydrology Program research site is 

located in Grafton County, at the southern extent of the 

White Mountains in central New Hampshire (fig. 1). 

The Mirror Lake site was selected for investigations in 

fractured rock as a part of a nationwide program to 

describe the fate of contaminants in subsurface environments. Although there are no known contaminants 

in ground water at the Mirror Lake site, the equipment, 

field techniques and interdisciplinary interpretive 

methods that are being developed in this study area are 

directly transferable to other fractured-rock sites where 

ground-water has been contaminated. Research at the 

Mirror Lake site is focused on characterizing groundwater flow and chemical transport in fractured rock 

over distances of tens of meters to kilometers. The 

purpose of this paper is to review the current research 

being conducted at the Mirror Lake site. An earlier 

summary of bedrock research activities at the Mirror 

Lake site is in Shapiro and Hsieh (1991).

SITE DESCRIPTION

Mirror Lake is located at the lower end of the 

Hubbard Brook valley in the southern part of the White 

Mountains of central New Hampshire (fig. 1). Hubbard 

Brook, which drains the valley, flows into the Pemigewasset River. The topography of the area generally is 

steep along mountain sides and flat along the Pemigewasset River. Although ground-water flow extends 

beyond the surface-water drainage basin associated 

with Mirror Lake, regional field investigations have 

concentrated on the Mirror Lake drainage basin. The 

Mirror Lake drainage basin occupies 85 ha. Land-surface elevations range from 213m above sea level at the

71
71°42' 71°41'30"

43°57'

43°56'30"

Well site 

Well field

R1

New 

Hampshire

Map/area

1000 2000 FEET

300 600 METERS

Figure 1. Location of the Mirror Lake watershed in New Hampshire and the location of bedrock wells in the vicinity of 

Mirror Lake.

lake surface to 481 m at the top of the drainage divide. 

Three perennial streams flow into Mirror Lake, and 

Mirror Lake drains through an outlet stream into Hubbard Brook.

Glacial drift overlies the bedrock in much of the 

Mirror Lake area. The drift thickness ranges from 0 to 

55 m and consists mainly of silty and sandy till, with 

numerous cobbles and boulders. Kame terraces, consisting of fine- to coarse-grained sandy ice-contact 

deposits, are found at several levels on the mountain 

sides. In the area between Mirror Lake and Hubbard 

Brook, the silty sand and gravel form what is believed 

to be a delta-type deposit laid down when Hubbard 

Brook adjusted its grade to the Pemigewasset River 

during deglaciation. Mirror Lake is believed to be a 

kettle lake overlying a bedrock saddle, where bedrock

valleys descend to the north and south. The ridge north 

of the lake is a moraine overlying the northern bedrock 

valley, where drift thickness is as much as 55 m.

The bedrock in the Mirror Lake area is a Silurian 

sillimanite-grade schist that has been extensively 

intruded by late Devonian granite (C.C. Barton, U.S. 

Geological Survey, written commun., 1993). The schist 

and granite are cut by dikes of pegmatite of unknown 

age, possibly a residual differentiate of the granitic 

intrusions. All three rocks are cut by lamprophyre 

dikes, whose age is believed to be middle Jurassic and 

early Cretaceous. At a road cut along Interstate Highway 93, directly east of Mirror Lake, the exposed bedrock consists of a complex suite of rock types. The 

granitic intrusions are present as dikes, irregular pods, 

and anastomosing fingers that range in width from

72
centimeters to tens of meters. Pegmatite typically is in 

the form of dikes, centimeters to meters wide, that 

cross-cut both the schist and the granite. The lamprophyre dikes, having widths of centimeters to meters, 

are less common than the other three rock types in the 

Mirror Lake area.

All components of the hydrologic system in the 

Mirror Lake drainage basin are being monitored. Flow 

of all streams leading to Mirror Lake are measured with 

flumes having continuous recorders; precipitation is 

measured at two locations in the drainage basin; climatological information for energy budget and evaporation studies have been collected since 1983; water 

levels in bedrock wells, drift piezometers, and watertable wells are continuously recorded (Winter, 1984, 

1985). Bedrock wells have been drilled in two clusters, 

referred to as the FSE and CO well fields (fig. 1). Bedrock wells also have been areally distributed in the 

vicinity of Mirror Lake. In early 1993, there were 13 

bedrock wells in the FSE well field, 4 bedrock wells in 

the CO well field and 12 bedrock wells areally distributed in the Mirror Lake area. Depths of bedrock wells 

range from 60 to 230 m. At most bedrock-well sites, 

piezometers are installed at one or more locations in the 

saturated drift above the bedrock surface. Water-table 

wells, whose screens intersect the water table, have 

been installed throughout the watershed and at most 

bedrock-well sites.

Ground-water flow in the Mirror Lake area is 

typical of mountain-valley lerranes of the New 

England uplands. Water from precipitation and snowmelt infiltrates into the subsurface, flows through the 

glacial drift and fractured bedrock, and discharges into 

streams, lakes and rivers. Estimates of recharge to 

ground water vary from 0.2 m/yr (meters per year) to 

0.45 m/yr (T. C. Winter, U.S. Geological Survey, oral 

commun., 1993). The major features controlling 

ground-water flow in the Mirror Lake area are (I) the 

steep slope of the mountain sides, (2) the low hydraulic 

conductivity of the drift and the bedrock, and (3) the 

presence of surface-water bodies such as streams, 

Mirror Lake and the Pemigewasset River. In most of 

the Mirror Lake watershed, the water table is close to 

land surface, typically less than 10 m below land surface. The only exception is the moraine ridge north of 

Mirror Lake where depth to water table exceeds 20 m. 

The presence of a shallow water table suggests that 

local topography affects the local pattern of groundwater flow.

Hydraulic conductivities of the glacial drift and 

bedrock in the Mirror Lake area have been estimated 

from results of hydraulic tests. Based on slug tests in 

piezometers, the hydraulic conductivity of the till is 

estimated to range between 10~() m/s (meters per second) and 10"5 m/s (Wilson, 1991). Hydraulic conductivity of the sands in the kame terraces is approximately 

10 times higher than the till. The results of single-hole, 

straddle-packer tests using 5-m long test intervals indicate that the bedrock hydraulic conductivity is highly 

heterogeneous, varying over many orders of magnitude 

(Hsieh and Shapiro, 1996). The average hydraulic 

conductivity over the upper 100 m of the bedrock is 

3x 10" 7 m/s. Single-hole tests, however, sample the rock 

in a small region around the bedrock well and do not 

necessarily represent the hydraulic conductivity controlling regional flow through the bedrock.

CURRENT RESEARCH

Current research at the Mirror Lake site is focused 

on local- and regional-scale investigations of groundwater flow and chemical transport in fractured rock. 

The local-scale investigations are conducted over 

distances of tens of meters and focus on identifying 

(1) fractures and fracture properties on exposed surfaces, (2) fractures in the subsurface using borehole 

and surface geophysics, and (3) hydraulic and transport 

properties of fractures through hydrologic testing 

(Hsieh and others, 1993). When applied individually, 

these methods generally yield limited information on 

fracture characteristics. However, when applied 

together, these methods become powerful tools that 

enable an interdisciplinary approach to the study of 

ground-water flow and chemical transport in fractured 

rock. In regional-scale investigations, controlled 

hydrologic testing cannot be conducted to identify 

hydraulic and transport properties of the bedrock. 

Hydraulic and transport properties must be inferred 

from hydrologic, geologic and geochemical information. Regional-scale investigations at the Mirror Lake 

site have focused on (1) methods of collecting hydrologic and geochemical information in heterogeneous 

bedrock terranes, and (2) the synthesis of this information by means of ground-water-flow and chemicaltransport modeling.

73
Local-Scale Investigations

The principle problem encountered in characterizing ground-water flow and chemical transport in 

fractured rock is the physical identification and mathematical characterization of the heterogeneous hydraulic properties of the rock. Fractures are the conduits for 

fluid movement in the subsurface; therefore, identification of their location, interconnectivity and hydraulic 

properties is paramount in the characterization of 

ground-water flow and chemical transport over distances of tens of meters. Hence, a major research effort 

at the Mirror Lake site is the investigation of geometrical properties of fractures. Understanding the geometric properties of fractures provides a basis for making 

inferences and predictions of flow and transport in the 

subsurface. Barton (1996) conducted an extensive 

investigation of fracture properties and rock types on 

four vertical highway road cuts and one horizontal 

glacial pavement. Rock types were delineated, and 

fractures with trace lengths greater than 1 in were 

mapped for orientation, surface roughness, aperture, 

trace length, connectivity and signs of mineralization. 

From this investigation, the granite was shown to be 

more heavily fractured and to have shorter and more 

planar fractures than schist, whereas fractures in schist 

have greater fracture roughness than those in the granite. In addition, the connectivity of fractures in the 

granite and schist was poor in comparison to studies of 

fracture connectivity in other areas of the country. This 

investigation indicates that fluid moves through highly 

tortuous paths in the bedrock.

Fracture mapping is conducted only on exposed 

surfaces. Research on characterizing fractures in the 

subsurface is being conducted by means of surfaceand borehole-geophysical surveys. Geophysical surveys generally yield an inferred distribution of a specific rock-mass property from which the presence of 

fractures must be inferred. Anomalies in the rock-mass 

properties attributed to fractures, however, also could 

be the result of variations in lithology; therefore, the 

correlation of several geophysical techniques as well as 

geologic information and hydrologic testing must be 

used to characterize the geometry of fractures in the 

subsurface.

Lieblich and others (1991) and I laeni and others 

(1993, 1996) discuss results of surface-geophysical 

surveys using azimuthal-seismic refraction and azimuthal-DC resistivity to detect the orientation of 

steeply dipping fractures. These methods infer the 

strike of a dominant, saturated, steeply dipping set of

fractures from the azimuthal variation in the seismic 

velocity and the apparent resistivity of the rock, respectively. Azimuthal seismic-refraction data collected 

near the CO well field indicate that the principle strike 

of fractures is approximately N. 22° E.--an orientation 

similar to that of the principle orientation of fractures 

mapped on the road cut east of Mirror Lake. The principle 

strike of fractures on the road cut is approximately 

N. 30° E. Results from azimuthal-DC resistivity using 

a square array were conducted at the same site and 

yielded a similar orientation for the principle strike of 

a steeply dipping set of fractures. An azimuthal-DC 

resistivity survey using a linear array also was conducted at the same site. This survey indicated a principle 

fracture strike of N. 352° E. The linear array requires a 

large current-electrode spacing that may have emphasized environmental interferences such as a buried telephone cable. Also, voltage-electrode spacing should be 

large relative to the fracture spacing a condition that 

may not have been satisfied at this site. The interpretation of these results indicates that multiple geophysical 

methods must be used to obtain an overall interpretation of fracture orientation (Lieblich and others, 1991),

The results of surface-exposure mapping indicated 

the importance of identifying distributions of subsurface rock types and fractures in characterizing groundwater flow. Techniques for identifying fractures and 

rock types in bedrock wells for hydrologic investigations are discussed by Johnson (1996) and Paillet 

(1996). Johnson (1996) describes the use of a boreholecolor television camera to identify rock types, fractures 

and other conditions in the borehole, including mineralization in and near fracture surfaces. The borehole 

camera reduces the need to core bedrock wells to identify the distribution of rock type. Paillet (1996) discusses a systematic approach to the characterization of 

fractures in boreholes using a combination of standard 

geophysical logging tools that identify the general 

structure and lithology of the rock, and advanced borehole geophysical techniques that identify fractures and 

their potential for conducting fluid. The acoustic televiewer log shows the location and orientation of fractures intersecting boreholes. A sensitive heat-pulse 

flowmeter measures the fluid velocity in the borehole. 

When used under ambient hydraulic conditions in the 

borehole, permeable fractures having different hydraulic heads in the borehole can be identified. When used 

in conjuction with pumping in the borehole, the location of the most permeable fractures can be identified.

74
Results from the heat-pulse flowmeter can also be 

used to infer fractures forming hydraulic connections 

between adjacent boreholes (Paillet, 1996). By pumping one bedrock well and measuring the fluid velocity 

in an adjacent bedrock well, the location of fractures 

responding to the hydraulic stress can be identified. 

Paillet and others (1992) interpreted the transient 

velocity response to identify hydraulic properties of 

fractures and propose conceptual models of fracture 

connectivity between boreholes

The acoustic televiewer log and borehole television 

camera can identify fractures at the face of the borehole 

wall including their strike and dip. The distance these 

fractures extend into the rock and their interconnectivity with other fractures cannot be identified by these 

borehole methods. A major research effort at the Mirror Lake site has been the evaluation of geophysical 

methods that image fractures away from boreholes. 

Three methods of characterizing fractures in the rock 

mass have been applied at the FSE and CO well fields 

at the Mirror Lake site: single-hole, directional electromagnetic (EM) imagining; EM tomography; and seismic tomography.

Haeni and others (1993) discuss results of using a 

single-hole, directional EM tool that measures reflections of electromagnetic wave energy radiated from the 

transmitter in the tool. Water-filled fractures in the 

vicinity of a borehole, and fractures intersecting the 

borehole can be inferred from the reflected signal. The 

results of using this investigative tool in the CO well 

field showed that the same fracture zone could be 

detected from different boreholes, and the lateral extent 

and orientation of fractures could be inferred.

Seismic and EM tomography also infer the location of fractures in the rock mass away from borehole 

walls. Unlike the single-hole directional tool, where the 

transmitter and receiver are located in the same borehole, EM tomography is conducted by radiating 

electromagnetic energy from one borehole and monitoring its reception at an adjacent borehole (Wright and 

others, 1996a,b). The transmitted energy is delayed by 

the propagation distance and by the different dielectric 

properties of the water and rock matrix. By collecting 

data of sufficient density at a variety of angles (between 

the source and receiver) over the length of the boreholes, the spatial distribution of dielectric properties in 

the rock volume between the boreholes can be inferred 

by an inversion algorithm. Wright and others (1996a) 

also discuss the results of seismic-tomography surveys 

conducted between pairs of bedrock wells in the FSE

well field. Seismic tomography is based on principles 

similar to those of EM tomography; however, instead 

of using electromagnetic energy, an acoustic signal is 

generated in one borehole, and the first-arrival of the 

pressure wave is monitored in an adjacent borehole. 

A spatial distribution of the seismic velocity in the 

volume of rock between the two wells can be inferred 

using an inversion algorithm; areas of low-seismic 

velocity imply zones of fracturing, and zones of highseismic velocity imply intact rock matrix.

Wright and others (1996b) compare the seismic 

and EM tomograms with each other and with fractures 

identified from acoustic televiewer logs. Results of the 

seismic- and EM-tomography surveys identified the 

same general zones of fracturing in the rock mass in 

wells separated by 10 m in the FSE well field. However, 

the orientation of the fractures intersecting the borehole 

(as identified from the acoustic televiewer log) was not 

the same as that of the fracture zones inferred from the 

seismic and EM tomograms. Results of hydrologic 

testing confirm the orientation of the fracture zones as 

identified by the tomograms, implying that the orientation of permeable fractures intersecting boreholes is 

not necessarily the orientation of permeable zones in 

the rock mass.

Seismic and electromagnetic signals respond to 

both fractures and lithology; therefore, dielectric 

permittivity and seismic velocity may not correlate 

with hydraulic properties of the rock. Thus, tomography must be conducted in conjunction with hydrologic 

testing to identify permeable structures in the bedrock. 

The use of hydrologic testing to characterize fracture 

locations and their hydraulic and transport properties is 

another major research effort being conducted at the 

Mirror Lake site. Hydrologic testing, consisting of 

hydraulic and tracer testing, is the most direct method 

of investigating flow and transport properties in the 

subsurface. Hydrologic tests generally involve artificially inducing a perturbation (a change in fluid pressure or chemical composition) in the subsurface and 

measuring the resulting response. The flow metering 

during pumping, as discussed by Paillet and others 

(1992) and Paillet (1996), are examples of hydraulic 

testing. Hsieh (1994), Hsieh and Shapiro (1996) and 

Shapiro (1996) discuss hydraulic and tracer tests conducted in the FSE well field to quantify hydraulic and 

transport properties of the bedrock.

Hsieh (1996) and Hsieh and Shapiro (1996) discuss hydraulic tests conducted by installing inflatable 

packers in all bedrock wells in the FSE well field to

75
eliminate the artificial vertical permeability induced by 

open boreholes. An hydraulically isolated interval in 

one bedrock well was pumped continuously, and fluid 

pressure responses were monitored in hydraulically 

isolated intervals in other bedrock wells. The results of 

these tests indicated the bedrock does not respond as an 

equivalent homogeneous porous medium. Instead, 

there are several clusters of highly permeable fractures 

of limited areal extent. These highly permeable fracture 

clusters are only connected through less transmissive 

fractures.

Shapiro (1996) discusses radially converging 

tracer tests conducted in one of the highly permeable 

fracture clusters identified in the FSE well field. New 

equipment and field techniques of conducting tracer 

tests in fractured rock were developed for these tracer 

tests (Shapiro and Hsieh, 1996), the results of which 

showed that there is significant variability in rock 

properties that can affect chemical transport, even 

within a highly permeable fracture zone.

Other interpretive methods of identifying hydraulic 

and transport properties of the rock were developed to 

take advantage of environmental tracers. Goode and 

others (1993) used the natural variability in the radon 

content of water sampled during an aquifer test to 

estimate the effective porosity of the bedrock and the 

vertical leakage from the overburden. The temporal 

variability in radon concentration of water measured at 

the pumped well was attributed to the mixing of water 

sources; water in the glacial drift was assumed to have 

a different radon concentration than water in the bedrock.

Regional-Scale Investigations

On the scale of kilometers, hydrologic tests in 

fractured rocks become ineffective because, over such 

distances, changes in pressure caused by fluid injection 

or pumping are generally too small to measure, and 

transport times become too long for tracer tests to be 

practical. In order to investigate phenomena affecting 

flow and transport on the regional scale, the response of 

the ground-water system to natural perturbations and 

long-term human disturbances must be monitored. 

Monitoring hydrologic and gcochemical phenomena in 

bedrock terranes requires the design of equipment and 

techniques that account for the heterogeneous hydraulic 

properties of bedrock.

Hsieh and others (1996) discuss methods of measuring hydraulic heads in multiple intervals in a single 

bedrock well. The monitoring equipment converts an

open borehole to the equivalent of a multilevel piezometer. The equipment is removable, thereby allowing 

access to the bedrock well for hydrologic and geophysical testing when desired. The long-term installation of this equipment in bedrock wells also prevents 

the mixing of waters of differing chemistry because 

inflatable packers are used to hydraulically isolate 

hydraulically discrete intervals in the borehole.

Shapiro and others (1996) briefly discuss methods 

of collecting water samples for geochemical and isotopic analysis from bedrock wells. Because of the large 

variability of the hydraulic conductivity of fractures, 

water samples for geochemical and isotopic analyses 

cannot be collected by pumping an open borehole. 

A sample from an open borehole would yield water 

mixed from all permeable fractures intersecting the 

borehole. Instead, water samples are collected after 

isolating a single fracture or fracture zone in the 

borehole using inflatable packers.

A major research effort at the Mirror Lake site is 

the interpretation of geologic, hydrologic, and geochemical information to infer regional properties of 

ground-water flow and chemical transport in the bedrock. Rosenberry and Winter (1993) used 10 years of 

streamflow, lake-discharge, and climatological data to 

estimate the components of bedrock recharge and discharge in the water budget of Mirror Lake. Water from 

bedrock is estimated to account for up to 4 percent of 

all inflows to the lake and up to 1 percent of all outflows 

from the lake. These estimates are approximately 10 

percent of ground-water inflows to and outflows from 

the lake; however, they are subject to great uncertainty 

because they are small relative to the major components of the water budget for the lake. These estimates 

of bedrock fluxes to and from the lake are being used to 

calibrate ground-water-flow models in the Mirror Lake 

drainage basin and its vicinity.

Harte and Winter (1996) used a numerical, crosssectional-flow model of the drift and bedrock to investigate processes affecting bedrock recharge from the 

overlying glacial drift. Results of model simulations 

were compared with the distribution of the hydraulic 

head along an hypothesized path of ground-water flow 

running through the northwestern catchment area of the 

Mirror Lake drainage basin. The locations of bedrock 

recharge in the Mirror Lake area are attributed to lateral 

trends in the bedrock horizontal hydraulic conductivity, 

which is hypothesized to decrease with increasing 

elevations in the drainage basin to reproduce measured 

head gradients between the drift and bedrock.

76
Shapiro (1993) conducted a numerical simulation 

ofareal (two-dimensional) regional ground-water flow 

in a heterogeneous bedrock terrane, where the heterogeneity was assumed to be the result of a random distribution of constant transmissivity blocks. The random 

distribution of transmissivity resulted in the poor connectivity of highly transmissive zones in the bedrock. 

The conceptual model was based on results of hydraulic testing in the FSE well field that illustrated the poor 

hydraulic connection between highly permeable fracture zones (Hsieh, 1996; Hsieh and Shapiro, 1996). 

The numerical simulations showed that hydraulic 

heads increased with increased elevation of upgradient 

locations in a manner similar to that observed in the 

Mirror Lake drainage basin. Thus, it is hypothesized 

that the poor connectivity of highly permeable zones in 

the bedrock could also play a role in the regional distribution of hydraulic heads.

A major research effort at the Mirror Lake site has 

been the analysis of ground-water chemistry and environmental isotopes to identify transport properties of 

the bedrock and processes affecting chemical transport 

in bedrock terranes over distances of kilometers. As 

water flows in the subsurface, its chemical composition 

evolves as the water reacts with the rock along the flow 

path. Understanding the chemical changes in ground 

water can help develop conceptual models of regional 

ground-water flow and chemical transport. In addition, 

environmental isotopes, such as tritium, and chlorofluorocarbons (CFC's) in ground water can be used to 

estimate the residence times of shallow ground-water.

Water samples were collected from discrete 

intervals in bedrock wells and analyzed for a variety of 

dissolved solids, dissolved gases, and a variety of stable and radioactive isotopes. Busenberg and Plummer 

(1996) measured the concentrations of CFC's in 

ground-water samples. By using historical records of 

CFC concentrations in the atmosphere, the date of 

ground-water recharge can be estimated. Drenkard and 

others (1996) estimated ground-water ages on the basis 

of the ratio of the parent-daughter isotopes tritium and 

helium. Many ground-water samples contained anomalously high helium concentrations; however, plausible 

corrections could be made for other sources of helium, 

and ground-water ages were estimated. The ages estimated from the tritium-helium and CFC analyses were 

similar. However, these ages do not correspond well 

with the presumed date of the maximum atmospheric 

tritium concentration. Both the tritium-helium and 

CFC ages appear to be approximately 6 years younger

than ground-water ages estimated by using the maximum tritium concentration as the date of the maximum 

atmospheric tritium input. This difference is still being 

investigated and may be the result of ground-water 

flow through the unsaturated zone where the air phase 

can allow further equilibration with atmospheric conditions, or it may be a result of the mixing of waters of 

various ages, either naturally or through the collection 

of water samples.

Ground-water ages indicate the residence time of 

the water since it was recharged; however, the distance 

the water has traveled must also be identified to estimate the regional flow velocity. In heterogeneous bedrock terranes, the distance the water has traveled is not 

readily identifiable. Shapiro and others (1996) proposed a simple conceptual model to use changes in the 

concentration of the bicarbonate ion as a second measure of the ground-water-residence time in the bedrock, 

and as a means of estimating the average ground-water 

velocity in the bedrock. The model considers transport 

through a single fracture and diffusion of bicarbonate 

ions into the fracture from the rock matrix. When used 

in conjunction with the ground-water ages estimated 

from CFC or tritium-helium concentrations, an average 

ground-water velocity can be estimated to reproduce 

the measured residence time and bicarbonate concentration at sampling locations.

The basis of the conceptual model proposed by 

Shapiro and others (1996) stems from discussions by 

Wood and others (1996) regarding the significance of 

diffusion in defining the ground-water chemistry. 

Wood and others (1996) noted that residence times of 

less than 50 years are not sufficient for dissolution of 

minerals in the granite and schist along fracture faces 

to produce the concentration of dissolved solids in the 

water samples collected in the bedrock. Instead, it is 

hypothesized that water in the void space of the unfractured granite and schist is in chemical equilibrium with 

the associated minerals in the granite and schist. In particular, small amounts of calcite (1 percent by weight) 

have been found in samples of the granite. Thus, the 

equilibrium bicarbonate concentration in the rock 

matrix is significantly higher than the concentration in 

the water recharged through the overburden and into 

the bedrock. The concentration gradient between water 

in the rock matrix and water in the fractures results in 

the mass flux of bicarbonate ions into fractures.

Wood and others (1996) conducted laboratory 

experiments to identify the effective diffusion coefficient and porosity of the intact rock matrix to verify the

77
role of diffusion from the rock matrix. Porosity of granite samples was approximately 1.5 percent, and the 

effective diffusion coefficient of l37Cs in granite 

(including the effects of retardation, tortuosity and 

porosity) was approximately 6 X 10' n m2/s. In addition, 

iron-staining in the rock matrix adjacent to fractures 

has been noted on outcrops and cores, and this is a 

further indicator of the effects of diffusion in the 

ground-water chemistry (Barton, 1996; Wood and 

others, 1996). Preliminary calculations of precipitation 

rates for ferric hydroxide using the laboratoryestimated porosity and diffusion coefficients indicate 

that ferric hydroxide bands can form over tens of years 

as the result of the diffusion of dissolved oxygen into 

the rock matrix. These timeframes are consistent with 

laboratory diffusion experiments and the variability of 

bicarbonate concentration in the ground water in the 

bedrock, which is assumed to be the result of diffusion 

from the rock matrix.

SUMMARY

Research at the Mirror Lake Toxic-Substances 

Hydrology Research site has focused on characterizing 

the physical properties that can define ground-water 

flow and chemical transport in fractured rock over 

distances that range from meters to kilometers. The 

characterization of ground-water flow and chemical 

transport over tens of meters has focused on identifying 

the geometrical and hydraulic properties of fractures. 

Mapping of surface exposures, use of surface and borehole geophysical techniques, and hydrologic testing 

have been used to identify properties of fractures that 

influence ground-water flow and chemical transport. 

Over distances associated with regional ground-water 

flow, processes affecting ground-water movement and 

chemical transport are inferred through the synthesis of 

hydrologic and geochemical data by means of groundwater and chemical-transport modeling.

REFERENCES

Barton, C.C., 1996, Characterizing bed rock fractures in 

outcrop for ground-water hydrology studies An 

example from Mirror Lake, Grafton County, New 

Hampshire, in Morganwalp, D.W., and Aronson, D.A., 

cds., U.S. Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting. 

Colorado Springs, Colo., September 20-24, 1993: U.S.

Geological Survey Water-Resources Investigations 

Report 94-4015.

Buscnbcrg, E., and Plummcr, L.N., 1996, Concentrations of 

chlorofluorocarbons and other gases in ground waters 

at Mirror Lake, New Hampshire, in Morganwalp, D.W., 

and Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Drcnkard, S., Torgcrsen, T, Wcppernig, R., Parley, K., 

Schlosscr, P., Michcl, R., L., Shapiro, A.M., and Wood, 

W.W., 1996, Helium isotope analysis and tritiumhelium age dating in the Mirror Lake basin, Grafton 

County, New Hampshire, in Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Goodc, D.J., Hsich, PA., Shapiro, A.M., Wood, W.W., and 

Kraemcr, T.F., 1993, Concentration history during 

pumping from a leaky aquifer with stratified initial concentration, in Shcn, H.W., ed., Hydraulic Engineering, 

Proceedings of 1993 National Conference, July 25-30, 

1993, San Francisco: New York, American Society of 

Civil Engineers.

Haeni, P.P., Lane, J.W., Licblich, D.A., 1993, Use of surface-geophysical and borehole-radar methods to detect 

fractures in crystalline rocks, Mirror Lake area, Grafton 

County, New Hampshire, in Hydrogcology of Hard 

Rocks, 24th Congress of the International Association 

of Hydrogcologists, June 28 - July 2, 1993, Oslo, 

Norway, p. 577-587.

Hacni, P.P., Lane, J.W., Licblich, D.L., and Barton, C.C., 

1996, Using surface geophysical and bedrock outcrop 

data to identify fractures near Mirror Lake, New Hampshire, in Morganwalp, D.W., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. 

Geological Survey Water-Resources Investigations 

Report 94-4015.

Hartc, P.T., and Winter, T.C., 1996, Factors affecting 

recharge to crystalline rock in the Mirror Lake are, 

Grafton County, New Hampshire, in Morganwalp, 

D.W., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program-Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

78
Hsieh, P. A., 1996, An overview of field investigations of 

fluid flow in fractured crystalline rocks on the scale of 

hundreds of meters, in Stevcns, P.R., and Nicholson, 

T.J., eds.. Joint U.S. Geological Survey, Nuclear Regulatory Commission Workshop on Research Related to 

Low-Level Radioactive Waste Disposal, May 4-6, 

1993, National Center, Rcston, Virginia, Proceedings: 

U.S. Geological Survey Water- Resources Investigations Report 95-4015.

Hsich, P.A., Perkins, R.L., and Rosenberry, D.O., 1996, 

Field instrumentation for multi-level monitoring of 

hydraulic head in fractured bedrock at the Mirror Lake 

site, Grafton County, New Hampshire, in Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Hsich, P.A., and Shapiro, A.M., 1996, Hydraulic characteristics of fractured bedrock underlying the FSE well 

field at the Mirror lake site, Grafton County, New 

Hampshire, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting. 

Colorado Springs, Colo., September 20-24, 1993: U.S. 

Geological Survey Water-Resources Investigations 

Report 94-4015.

Hsich, P.A., Shapiro, A.M., Barton, C.C., Hacni, P.P., 

Johnson, C.D., Martin, W.W., Paillct, F.L., Winter, 

T.C., Wright, D.L., 1993, Methods of characterizing 

fluid movement and chemical transport in fractured 

rock, //; Cheney, J.T., and Hepburn, J.C., Field Trip 

Guidebook for the Northeastern United States; 1993 

Boston GSA, Contribution No. 19, Department of 

Geology and Geography, University of Massachusetts, 

Amhcrst, Mass., p. Rl-30.

Johnson, C.D., 1996, Use of a color borehole video camera 

to identify lithologics, fractures and borehole conditions in bedrock wells in the Mirror Lake area, Grafton 

County, New Hampshire, //; Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Lieblich, D.A., Lane, J.W., Haeni, P.P., 1991, Results of 

integrated surface-geophysical studies for shallow subsurface fracture detection at three new Hampshire sites, 

in Expanded Abstracts with Biographies, 61st Annual 

International Meeting, November 10-14, 1991: Society 

of Exploration Gcophysicists, Houston, Texas, p. 553- 

556.

Paillet, F.L., 1996, Using well logs to prepare the way for 

packer strings and tracer tests- Lessons from the 

Mirror Lake study, in Morganwaip, D.W., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the technical 

meeting, Colorado Springs, Colo., September 20-24, 

1993: U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Paillct, F.L., Novakowski, K., and Lapccvic, P., 1992, Analysis of transient flows in boreholes during pumping in 

fractured formations, in Society of Professional Well 

Log Analysts Annual Logging Symposium, 33rd, Oklahoma City, Oklahoma, p. S1 -21.

Rosenberry, D.O., and Winter, T.C., 1993, The significance 

of fracture flow to the water balance of a lake in fractured crystalline rock terrain, in Hydrogcology of Hard 

Rocks, 24th Congress of the International Association 

of Hydrogcologists, June 28 - July 2, 1993, Oslo, 

Norway, p. 967-977.

Shapiro, A.M., 1993, The influence of heterogeneity in 

regional hydraulic properties of crystalline rock, in 

Hydrogcology of Hard Rocks, 24th Congress of the 

International Association of Hydrogeologists, June 28- 

July 2, 1993, Oslo, Norway, p. 125-136.

Shapiro, A.M., 1996, Estimating effective porosity in fractured crystalline rock using controlled-traccr tests, //; 

Stevcn, P.R., and Nicholson, T.J., cds., Joint U.S. Geological Survey, Nuclear Regulatory Commission Workshop on Research Related to Low-Level Radioactive 

Waste Disposal, May 4-6, 1993, National Center, 

Reston, Virginia, Proceedings: U.S. Geological Survey 

Water- Resources Investigations Report 95-4015.

Shapiro, A.M., and Hsieh, P.A., 1991, Research in fractured 

rock hydrogcology Characterizing fluid movement 

and chemical transport in fractured rock at the Mirror 

Lake drainage basin. New Hampshire, in Mallard, G.E., 

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Montercy, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91 -4034, p. 155-161.

Shapiro, A.M., and Hsich, P.A., 1996, A new method of performing controlled injection of traced fluid in fractured 

crystalline rock, in Morganwalp, D.W., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the technical 

meeting, Colorado Springs, Colo., September 20-24, 

1993: U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Shapiro, A.M., Wood, W.W., Busenbcrg, E., Drcnkard, S., 

Plummcr, L.N., Torgcrsen, T., and Schlosscr, P., 1996, 

A conceptual model for estimating regional groundwater velocity in bedrock of the Mirror Lake area.

79
Grafton County, New Hampshire, in Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Wilson, A., 1991, Distribution of hydraulic conductivity in 

the glacial drift at Hubbard Brook Experimental Forest, 

West Thornton, New Hampshire, unpublished Senior 

Thesis, Department of Earth Sciences, Dartmouth 

College, Hanover. New Hampshire, 56 p.

Winter, T.C., 1984, Gcohydrologic setting of Mirror Lake, 

West Thornton, New Hampshire: U.S. Geological Survey Water-Resources Investigations Report 84-4266, 

61 p.

Winter, T.C., 1985, Physiographic setting and geologic origin of Mirror Lake, in Likens, G.E., cd.. An Ecosystem 

Approach to Aquatic Ecology: New York, SpringerVcrlag, p. 40-53.

Wood, W.W., Shapiro, A.M., and Hsich, P.A., 1996, Observational, experimental and inferred evidence of solute 

diffusion in granite rocks: Examples from the Mirror 

Eakc watershed, Grafton County, New Hampshire, in 

Morganwalp, D.W., and Aronson, D.A., cds., U.S. Geological Survey Toxic Substances Hydrology ProgramProceedings of the technical meeting, Colorado 

Springs, Colo., September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

94-4015.

Wright, D.E., Olhocft, G.R., and Grovcr, T.P., I996a, Velocity, amplitude and dispersion EM tomography in fractured rock, 1993, in Morganwalp, D.W., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the technical 

meeting, Colorado Springs, Colo., September 20-24, 

1993: U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Wright, D.E., Olhocft, G.R., Hsich, P.A., Majer, E.E., 

Paillct, F.E., and Eanc, J.W., 1996b, Electromagnetic 

and seismic tomography compared to borehole acoustic 

televiewer and flowmctcr logs for subsurface fracture 

mapping at the Mirror Eakc site. New Hampshire, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. Geological Survey Toxic Substances Hydrology ProgramProceedings of the technical meeting, C'olorado 

Springs, Colo., September 20-24. 1993: U.S. Geological Survey Water-Resources Investigations Report 94- 

4015.

80
Characterizing Bedrock Fractures in Outcrop for Studies

of Ground-Water Hydrology: An Example from Mirror

Lake, Grafton County, New Hampshire

By Christopher C. Barton 1

Abstract

The fractured-rock-hydrology study site at 

Mirror Lake, New Hampshire provides an 

opportunity to use the pavement method to 

characterize bedrock fractures at four highway 

roadcuts and one glacial pavement adjacent to the 

bedrock-hydrology test wells at Mirror Lake. The 

method entails the mapping of the fracture-trace 

network down to fracture-trace lengths of 1 meter 

on a detailed geologic base map. These maps, in 

turn, form the basis for quantifying the scaling, 

interconnectedness, spatial distribution, and tracelength distribution for the fracture network. The 

characteristics of each mapped fracture include 

orientation, roughness, mineralization, aperture, 

and tectonic features, such as slickensides. The 

combined fracture data for the five pavement 

maps made to date indicate the following: 

Fracture orientation is highly variable with a 

preferred strike azimuth of approximately 30°E., 

preferred fracture dips of 7°NW., 50°SE., and 

82°SE. Fracture roughness coefficients (RC) 

range from 0 to 18 RC values, with a mode at 5 to 

6. Fracture aperture ranges from a lower cutoff of 

0.005 mm to a maximum of 20.6 mm, and the 

frequency distribution of fractures follows a 

power-law function with a scaling exponent of 

-1.5. Fracture-trace length ranges from a lower 

cutoff of 1 m to a maximum of 24.6 m, and the 

frequency distribution follows a power-law 

function with a scaling exponent of-2.4. Fracture 

connectivity within the network is low compared

U.S. Geological Survey, Denver, CO

to that of other sites around the country. Fracture 

mineralization includes iron-oxide coatings that 

appear have been biologically precipitated by 

iron-fixing bacteria in the ground water and 

Liesegang bands that extend up to about 1 meter 

into the rock matrix. A paleohydrologic map of 

the iron-stained fracture traces reveals that ground 

water flowed in only part of the available network 

of the open and interconnected fractures.

INTRODUCTION

Fracture networks exist in almost all rocks at or 

near the surface of the Earth. Where the fractures are 

open and interconnected, a portion of the network can 

conduct groundwater flow, and especially so over the 

short time scales associated with the extraction or 

injection of fluids in a bedrock well. Despite the 

importance of individual conductive fractures within 

the network, fractures are generally poorly 

characterized in most ground-water studies, because of 

the inaccessibility of buried fractures to direct 

observation and to the inability of surface and crosshole 

geophysical methods to image individual fractures and 

fracture networks to the same degree of detail as they 

can be mapped on surface exposures. Borehole logging 

and core sampling together permit measurement of 

fracture orientation, roughness, mineralization, 

aperture, and tectonic features. However, these 

characteristics may not be representative, because only 

a very small part of the fracture is recovered or exposed 

in the borehole wall or recovered in core, and statistical 

corrections are of limited value. It is impossible to 

determine fracture-trace length, spatial distribution of 

fracture traces and intersections, fracture connectivity, 

fracture abutting relations, or the size, shape, and spatial

81
distribution of blocks bounded by fractures on the basis 

of borehole data. The primary limitation of borehole 

logs and cores is that they represent a one-dimensional 

sampling of a three-dimensional fracture network. The 

use of lineament analysis to attempt to locate individual 

fractures, fracture networks, and zones of high 

permeability is scientifically unsound because 

lineaments generally fail to coincide with fractures on 

the ground (Throckmorton, 1987) and because 

lineament analyses are highly nonreproducible because 

of user bias and the position of the lighting source in 

areal photographs (Wise and others, 1985). A recent 

study of the relation of lineaments and lineament 

intersections to the production of water in crystaline 

rocks shows no statistically significant correlation 

between lineaments and water production (Maybee and 

others, 1990). Studies of fractures from aerial 

photographs have proven to have extremely limited use 

in this region because of the extensive glacial cover. Of 

course, even with complete bedrock exposure, areal 

photographs provide only strike (not dip) and spatial 

patterns and densities. Preliminary studies at this site 

indicate that linear features on areal photographs do not 

correspond to fractures on the ground.

The purpose of this paper is to describe a 

method developed for directly sampling and 

measuring characteristics of individual fractures and 

fracture networks that are relevant to fluid flow in 

fractured rock, and that form a basis for constructing 

volumetric synthetic fracture networks that are 

conditioned to the individual and statistical properties 

of the fractures. The method, called the pavement 

method, is an integrated method applicable to sites 

where there are natural or excavated subplanar 

exposures of bedrock (Barton and Hsieh, 1989).

The U.S. Geological Survey study of flow and 

transport in fractured crystaline rock at Mirror Lake, 

New Hampshire (fig. 1) has provided an opportunity 

to apply the pavement method to quantify the 

individual properties of fractures and fracture 

networks in a complex geologic setting. The fracture 

characteristics reported in this study are the 

benchmarks against which results of borehole, surface, 

and cross-hole geophysical methods can be compared.

PAVEMENT METHOD

The basis of the method is the complete and 

detailed mapping of the bedrock geology and of 

fracture traces over a wide range of length scales on

large subplanar exposures of bedrock (fig.3). The 

resulting map permits characterization of properties of 

the fracture network including the spatial scaling of 

the fractures, the connectivity of the network, tracelength distribution, and fracture density; none of these 

properties can be quantified by geophysical or by 

standard methods for characterizing fractures in 

outcrop (Kulander and others, 1979; McClay, 1987). 

Where fracture staining and (or) mineralization is 

present, it is possible to construct paleohydrologic 

maps of the conductive fractures as described below. 

In addition to its use in constructing fracture and 

paleohydrologic maps, the pavement method also 

allows for the quantitative measurement of individual 

fracture characteristics, including orientation, surface 

roughness, aperture, staining and (or) mineralizaion, 

and any other characteristics that might be needed to 

determine the temporal and spatial origin and 

propagation of the fracture.

The pavement method has several inherent 

weaknesses: (1) It is only a two-dimensional sample of 

a three-dimensional fracture network, and (2) the 

apertures of individual fractures measured in outcrops 

are affected by unloading. The objective of a pavement 

study is to quantify, in surface exposures, the 

characteristics of individual fractures and fracture 

networks that are analogous to those being 

hydrologically sampled in the subsurface. Those who 

use this method must be careful not to include fractures 

that are caused by exposure, weathering, blasting, and 

stress release caused by excavation. At the Mirror Lake 

site, the fractures formed in the subsurface are readily 

identified because of their mineral coatings and 

percipitates from paleo ground-water flow.

MIRROR LAKE STUDY SITE

Location

The Mirror Lake hydrologic study site is located 

adjacent to the southwestern corner of the White 

Mountain physiographic province in central New 

Hampshire (fig. 1). The study site includes most of the 

Mirror Lake watershed. The study site includes very 

few exposures of bedrock (<2% exposure) because of 

a blanket of glacial deposits. The few natural 

exposures that do exist are small and covered with 

lichen and moss, making rock type, contacts, and 

fractures difficult to identify and measure. The

82
74°41'30" 74°41

EXPLANATION

Topographic contour of altitude of 

land surface, contour interval is 

100 meters. Datum is sea level

Wells drilled for fractured 

bedrock hydrology study

Figure 1. Map showing location of the Mirror Lake fractured 

exposing fractures and bedrock geology along Interstate 93,

thickness of glacial deposits prohibits the clearing of 

bedrock exposures, such as those excavated in other 

study areas (Barton and Hsieh, 1989). However, the 

interstate 93 roadcut provides spectacular exposure of 

the same bedrock present where the hydrology test 

wells have been drilled immediately to the west 

(fig. I). The roadcut through a small hill of exposed 

bedrock has produced five exposures (four subvertical 

faces exposed by blasting, and one subhorizontal face 

that was glacially cleaned and polished) that were used 

for purposes of mapping and studying the fractures 

and the geology (fig. 1). Recent quadrangle bedrock 

mapping by J.B. Lyons and R.H. Moench 

(unpublished) shows that the bedrock types of the 

highway roadcut underly the hydrology study site and, 

therefore, it is reasonable to assume that the fracture 

characteristics measured in the roadcut are statistically 

identical to fractures accessible only by drilling in the 

hydrology study site.

l-rock-hydrology site, ground- water wells, and the roadcuts 

Grafton County, New Hampshire.

Geologic Setting

As shown on the preliminary geologic map of 

New Hampshire (Lyons and others, unpublished), the 

bedrock underlying the Mirror Lake study site lies in 

the Merrimack synclinorium (Central Maine Terrane). 

Bedrock at the site includes pelites of the upper and 

lower members of the early Silurian Rangeley 

Formation (438-428 MY (million years before 

present)). These pelites were multiply folded and 

metamorphosed during the Acadian orogeny (midDevonian) and attained the lower sillimamte grade 

(approximately 600°C and 4 kb, equivalent to burial to 

approximately 15km). A strong metamorphic foliation 

in the Rangeley Formation strikes N.25°- 45°E. and 

dips 60°- 67°NW. and 40°- 80°SE. Although not shown 

on the preliminary geologic map of New Hampshire, 

the Rangeley is locally extensively intruded by dikes 

and by anastomosing fingers and pods of late Devonian 

Concord granite (370-365 MY). The Rangeley and

83
687 POINTS

EXPLANATION

-12- Line of equal fracture density. Contour 

interval is 1.0 percent

111

O

z 111

QC 

D 

O 

O 

O

LL. 

O

O

B

90

80

70

60

50

40

30 

20

1Q

n

_

-

-

-

-

-

_

-

418 points

-

-

-

-

-

-

 ,

I i i

0-2 3-4 5-6 7-8 9-1011-1213-1415-1617-1819-20

ROUGHNESS COEFFICIENT (RC)

ou

Z 70

LLJ

o 60

DC 

LUQ- 50

Z

~> 40

Z 30

HI

0 20

HI

£ 10

ft

 i i i i i i i i i i i i i i i i i i i i i i i i i 

.

-

\-

Frequency (percent) = 20.4 (aperture)" 1 -53

r 2 = n.89

444 points

-

_

-

^L

P^L^^ 

I I FI i i i i i i i i i i i i i i i i i i i i 

ou

70

60

50

40

30

20

10

0

 i i i i i i i i i i i i i i i i i i i i i i i i i

Frequency (percent) = 224.6 (aperture)'2 - 41

r2=0.93

I 643 points

\

- \

. \

i i \

~ \

I l^hr- I I ' ' I I I I I I I I I I I I 

5 10 15 20 

APERTURE, IN MILLIMETERS

25

D

5 10 15 20 

TRACE LENGTH, IN METERS

25

EXPLANATION

Granite, 1-93, this study 

Schist, 1-93, this study 

Pegmatite, 1-93, this study

Volcanic tuff, Yucca

Mountain, NV 

Quartz diorite, Cedar

City.UT

Limestone, Lannon, WI 

Sandstone, Golden, CO

BLIND

10 15 20 25 

TRACE LENGTH, IN METERS

ABUTTING CROSSING

Figure 2. Quantitative characteristics of fractures exposed on the I-93 Roadcut; A. Contoured lower-hemisphere projection 

of poles to fractures; 8. Histogram of roughness coefficient (RC) frequency for fractures; C. Histogram of aperture frequency 

for fractures; D. Histogram of trace-length frequency for fractures; E. Plot of trace length versus aperture; F. Ternary diagram 

of percentages of fracture intersection and terminations for fractures. Data from sites mapped at Yucca Mountain, Nevada 

and elsewhere shown for comparison.

84
Concord are cut by dikes and interfolial intrusions of 

pegmatite of unknown age that possibly are a residual 

differentiate of the Concord granite. All rocks are cut 

by lamprophyre dikes whose age is not specifically 

known but is likely middle Jurassic to early Cretaceous 

(190-95 MY) on the basis of the ages of similar dikes 

elsewhere in the region (McHone, 1984). As shown by 

Richmond and Fullerton (1991), the flanks of hills are 

blanketed by ground moraine of boulder till at the 

higher elevations and by outwash sand and gravel in the 

stream valleys. Locally, these deposits are as much as 

10 m thick and were laid down during late Wisconsin 

glaciation which peaked in this region 22,000 to 12,000 

years ago. The glaciation produced a limited number of 

bedrock outcrops by the plucking action of the ice on 

the lee sides of hills, the clearing of ridges, and the 

polishing of glacial pavements on some hill tops. The 

ice retreated from this region 15,000 - 12,000 years ago.

METHODOLOGIES FOR 

CHARACTERIZING FRACTURES

The pavement method of mapping and 

characterizing fractures and fracture networks used in 

this study is described in Barton and Larsen (1985), 

and Barton and Hsieh (1989). Each of the measured 

characteristics is described below, except for scaling 

and density, which have not yet been measured. The 

pavement method permits an accurate and complete 

sampling of the individual and network characteristics 

of fractures. Pavements created by blasting and 

removal of bedrock commonly exhibit fractures 

induced by blasting, stress release caused by 

excavation, and weathering. Such fractures show no 

staining and are excluded in this study.

FRACTURE CHARACTERISTICS

Orientation

Orientation has historically been considered the 

single most important characteristic of fractures. 

Where the orientation of a number of fractures is 

similar, they are said to form a set. When the poles to 

fracture planes are plotted on an equal-area lowerhemisphere projection and their concentrations 

contoured, as shown on figure 2A, the fracture 

orientations are observed to be scattered but exhibit a 

preferred strike azimuth of approximately 30 °E. with 

preferred dips of 7°NW., 50°SE., and 82°SE.

Fracture-Surface Roughness

Fracture-surface roughness is an important 

characteristic because it controls the aperture variation 

and, thereby, the channeling of flow between the 

fracture walls. Fracture-surface roughness is 

measured where an unweathered surface can 

accommodate a 15-cm-long contour gauge. The 

roughness profiles are converted to a roughness 

coefficient by visual comparison to a standard set of 

profiles of known RC, which range (in integer values) 

from 0 to 20 (Barton and Choubey, 1977). RC ranges 

from 0 to 18 at this site, and has a mode at 5 to 6 

(fig. 2B).

Fracture Aperture

Aperture is a major characteristic controlling the 

characteristics of flow in individual fractures. The 

functional dependence of flow rate on aperture in 

rough fractures is a topic of much study, as is the 

functional relation between hydraulic aperture and 

mechanical aperture. Mechanical apertures are 

measured on surface exposures, however, the removal 

of adjacent and overlying rock and exposure to surface 

weathering has undoubtedly affected the mechanical 

apertures. In contrast, apertures measured in 

boreholes are affected by mechanical abrasion of the 

drill bit and by stress redistribution and concentration 

around the borehole walls. It is not possible, therefore, 

to measure mechanical apertures that have not been so 

affected, either at the surface or in boreholes. 

Nevertheless, an imperfect measure of aperture is 

preferable to no measure. Scaling analysis of aperture 

frequency for apertures fixed at depth by 

mineralization (Barton and Hsieh, 1989) suggests that 

unloading affects the exact values, but not the form, of 

the aperture distribution.

A representative aperture is determined by 

visual inspection at places where weathering has been 

minimal and mineralization absent. Aperture is 

measured with an automotive feeler gage or a finely 

divided ruler. Aperture ranges from a lower cutoff of 

0.005 mm to a maximum of 20.6 mm, and the 

frequency distribution follows a power law with a 

scaling exponent of-1.5 (fig. 2C).

85
Fracture-Trace Length Fracture Mineralization

Fracture-trace lengths are measured from 

pavement maps or directly on the pavement surface. 

Figure 2D includes the exposed lengths of fractures 

whose traces extend beyond the edges of the 

pavement. Truncation affects the exact values but not 

the overall form of the distribution. Fracture-trace 

length ranges from a lower cut off of 1 m to a 

maximum of 24.6 m, and the frequency distribution 

follows a power law with a scaling exponent of-2.4 

(fig. 2D). There is no discernable correlation between 

fracture-trace length and aperture (fig. 2E). Such a 

correlation might be found for isolated nonintersecting 

fractures, but one should not be expected for networks 

of interconnected fractures.

Fracture Connectivity

The fluid-flow-related properties of a fracture 

network are affected by the degree to which the 

fractures are interconnected. Fractures that are not 

highly interconnected can contribute only to 

storativity. Connectivity can be represented by the 

ratios of the three types of fracture termination or 

interaction: (1) fractures that terminate in the rock 

matrix (blind endings), (2) fractures that cross other 

fractures, and (3) fractures that abut other fractures. 

The percentage of blind endings, crossings, and 

abutments mapped at this site are shown on figure 

2F. The network of fractures at this site is poorly 

interconnected relative to other networks mapped 

throughout the country, and, thus, the ground-water 

flow at this site is expected to be highly channelized 

and tortuous between wells.

Minerals deposited on fracture faces can be 

useful for determining the paleohydrology of a 

fracture network. All of the natural fractures (as 

opposed to those that were created by blasting) have 

coatings of iron oxide and (or) manganese oxide. Iron 

oxide has also penetrated as much as 1 m into the rock 

matrix adjacent to fracture traces, in places as a 

continuous band and in other places as a rhythmic 

series of Liesegang bands. Surface-attaching ironfixing bacteria detected in water pumped from well 

FSE-6 could have been responsible for the deposition 

of iron oxide on the fracture surfaces, indicating that 

the iron oxide records the paths of paleo ground-water 

flow through the fracture network. The bacterial 

protein secretions may also have plugged the fracture 

permeability.

PALEOHYDROLOGIC MAP OF FRACTURE 

FLOW

A map of fracture traces for a part of the roadcut 

on the eastern side of the southbound lane of 1-93 is 

shown in figure 3. The parts of the fracture traces 

shown as solid lines are coated with iron oxide that 

precipitated from ground water before these rocks 

were exposed by excavation. Note that ground water 

did not flow in all of the open fracture traces. The 

backbone is a term used in physics to denote the flow 

path across a percolation cluster. The term is applied 

here to denote that portion of the fracture network that 

sustained flow. In fractures where flow did occur, the 

direction of flow was approximately perpendicular to 

the plane of the map. Note that there is no percolation 

across the map.

EXPLANATION

Boundary

Fracture trace, iron oxide staining

Fracture trace, no iron oxide staining

Figure 3. Map of fracture traces exposed in 1-93 roadcut. Portion of fracture traces coated with iron oxide precipitated from 

ground water are solid.

86
SUMMARY

Pavement maps provide the basis for 

quantifying the scaling, interconnectedness, spatial 

distribution, and trace-length distribution of fractures 

and the fracture network. The characteristics of each 

mapped fracture at the study site were measured, 

including, orientation, roughness, mineralization, 

aperture, and slickensides. The fracture data for the 

five maps made to date indicate the following: 

Fracture orientation is highly variable, the preferred 

strike azimuth is approximately 30°E., and preferred 

dips are 7°NW., 50°SE., and 82°SE. Fracture 

roughness coefficients range from 0 to 18, with a mode 

at 5 to 6. Fracture aperture varies from a lower cutoff 

of 0.005 mm to a maximum of 20.6 mm, the frequency 

distribution follows a power-law function with a 

scaling exponent of -1.5. Fracture-trace length ranges 

from a lower cutoff of 1 m to a maximum of 24.6 m, 

and the frequency distribution follows a power-law 

function with a scaling exponent of -2.4. Fracture 

connectivity within the network is low compared to 

that at other sites across the United States. Fracture 

mineralization includes iron oxide coatings that could 

have been biologically precipitated by iron-fixing 

bacteria in the ground water and Liesegang bands that 

extend as much as a meter into the rock matrix. A 

paleohydrologic map of part of the fracture network 

reveals that ground water does not flow in all the 

available open interconnected fractures. The fracture 

characteristics reported here are benchmarks against 

which the results of borehole, surface, and cross-hole 

geophysical studies of fractures can be compared.

REFERENCES

Barton, C.C., and Larsen, Eric, 1985, Fractal geometry of 

two-dimensional fracture networks at Yucca Mountain, 

southwest, Nevada, in Stephansson, Ove, ed.. 

Fundamentals of rock joints: Lulca, Sweden, Centck 

Publishers, p.77-84.

Barton, C.C., and Hsieh, P.A., 1989, Physical and

hydrologic-flow properties of fractures: American 

Geophysical Union, Field Guide Book T385, 28th 

International Geological Congress, 36 p.

Barton, N.R., and Choubey, Vishnu, 1977, The shear 

strength of rock joints in theory and practice: Rock 

Mechanics, v. 10, p. 1-54.

Kulander, B.R., Barton, C.C., and Dean, S.L., 1979, 

Applications of fractography to core and outcrop 

investigations: U.S. Department of Energy METC/SP79/3, 174 p.

Lyons, J.B., and Moench, R.H., unpublished bedrock 

geologic map of the Woodstock, New Hampshire 

quadrangle, 1:24,000 scale .

Lyons, J.B., Bothner, W.A., Moench, R.H., and Thompson, 

J.B., Jr., editors, 1989, review copy of unpublished 

Bedrock Geologic Map of New Hampshire, 1:250,000 

scale.

Maybee, S.B., Hardcastle, K.C., and Wise, D.U., 1990, 

Correlation of lineaments and bedrock fracture fabric: 

implications for regional fracturcd-bedrock aquifer 

studies, preliminary results from Georgetown, Maine, 

in Proceedings of the Focus Conference on Eastern 

Regional Ground Water Issues: National Water Well 

Association, p. 283-297.

McClay, K.R., 1987, The mapping of geologic structures, 

Halstcd Press, New York, 161 p.

McHonc, J.C., 1984, Mcsozoic igneous rocks of northern 

New England and adjacent Quebec: Summary, 

description map, and bibliography of data sources: 

Geological Society of America, Map and Chart Series, 

MC-49.

Richmond, G.M., and Fullerton, D.S., cds.,1991.

Quaternary Geological Map of the Boston 4° x 6° 

Quadrangle, United States and Canada: U.S. 

Geological Survey Miscellaneous Investigations Scries 

Map I-1420 (NK-19), 1:1,000,000 scale.

Throckmorton, C.K., 1987, Photogcologic study of smallscale linear features near a potential nuclear-waste 

repository site at Yucca Mountain, Southern Nyc 

County, Nevada: U.S. Geological Survey Open-File 

Report 87-409, 54 p.

Wise, D.U., Funicicllo, R., Parotto, M., and Salvini, F., 

1985, Topographic lineament swarms clues to their 

origin from domain analysis of Italy: Geological 

Society of America Bulletin, v. 96, p. 952-967.

87
88
Use of a Borehole Color Video Camera to Identify

Lithologies, Fractures, and Borehole Conditions in

Bedrock Wells in the Mirror Lake Area, Grafton County,

New Hampshire

ByCarole D. Johnson1

Abstract

A submersible color camera was used to 

describe bedrock lithologies and fractures in boreholes at the U.S. Geological Survey fractured-rock 

research site near Mirror Lake, Grafton County, 

New Hampshire. From June through August 1992, 

video surveys were completed in 29 bedrock wells 

that ranged in depth from 60 to 230 meters. Use of 

the submersible camera was prompted by a need to 

verify and provide additional descriptions of rock 

types identified in the wells. In two of the wells 

from which bedrock core was collected, video 

images together with drill cuttings were used to 

determine lithologies. These lithologies corresponded to lithologies determined directly from 

bedrock core samples collected from two wells. 

For wells from which core was not obtained, video 

images were used to improve the interpretations of 

the rock types that were based only on initial logs 

made at the time of drilling and later detailed 

examinations of drill cuttings. In addition, the 

images were used to inspect the conditions of the 

borehole walls for angularity, stability, or blockage.

INTRODUCTION

The Mirror Lake area is the site of a multidiscipline research effort to characterize fluid movement 

and chemical transport in fractured crystalline rock 

over a range of scales from meters to kilometers

'U.S. Geological Survey, Bow, N.H.

(Shapiro and Hsieh, 1991). The bedrock domain in the 

Mirror Lake area is typical of New Hampshire and of 

most crystalline rock terranes associated with orogenic 

belts and, therefore, is structurally complex. The bedrock is predominantly sillimanite-grade pelitic schists 

and gneisses that have been intruded by anatectic granites, pegmatites, and basalts (Lyons and others, 1986). 

Fractures that have developed in response to tectonic 

stresses (Hardcastle, 1989) and sheeting fractures 

(Trainer and others, 1987) serve as conduits for fluid 

movement and chemical migration. The bedrock is 

overlain by 10 to 50 m of glacial deposits that cover 

approximately 97 percent of the bedrock surface. 

Therefore, the characterization of the rock types relies 

on subsurface exploratory drilling and other techniques. Twenty-nine bedrock wells, ranging in depth 

from 60 to 230 m, were drilled between 1979 and 1992 

by use of the air percussion method. From June through 

August 1992, a color video camera was used to survey 

the wells and improve the interpretation of the subsurface. This paper describes the field procedure developed for identifying rock types, fractures and borehole 

conditions in the bedrock wells by use of a color video 

camera in the Mirror Lake area of New Hampshire.

DESCRIPTION OF EQUIPMENT

The equipment consists of a downhole camera and 

light source, an electrically powered winch, control 

unit, and video tape recorder (VTR) and monitor. The 

color video camera and light source are encased in a 

9.2-cm-diameter housing, which is suspended from 

double steel-wrapped coaxial cable. On the land surface, a powered winch is used to raise and lower the 

camera in the borehole. A camera control unit encodes

89
the depth of the camera; processes the analog image of 

the borehole; and provides downhole control of the 

light intensity, focus, and iris (or lens aperture), all of 

which enhance the video image. The processed image, 

including a superimposed digital depth readout, is sent 

to a VTR and a high-resolution monitor. This camera 

has two light attachments that allow for two different 

perspectives of the well: one looking down the well and 

the other looking at the side of the well. In order to view 

and adequately record the image, the VTR should have 

fast and slow forward and reverse speeds, freeze frame, 

four heads, and a real-time counter.

CAMERA-SURVEY TECHNIQUES

Two video surveys were completed in each of the 

29 bedrock wells. For the first survey, the camera was 

set up to look down the borehole. From this perspective 

(fig. 1 a), the borehole wall closest to the camera lens 

appears at the edge of the video image, and the bottom 

(the more distant views) of the well appears at the center. In the actual video image, however, the bottom of 

the well is blocked by the light attachment, which is 

supported by two rods. A planar feature (such as a fracture) that intersects the borehole at 90 degrees appears

as a circle, whereas a planar feature that intersects the 

borehole at less than 90 degrees appears as an ellipse. 

During the first survey, the occurrence of foliation, 

fractures, zones of borehole enlargement, and major 

changes in rock type and texture, were described. In 

addition, places to be closely viewed during the second 

survey were noted.

In the second survey, a light attachment that has a 

360-degree rotational mirror tilted at 45 degrees to the 

camera lens to permit a side view of the borehole wall 

was used. A wide-angle lens directed downward 

toward the mirror allows a simultaneous view of the 

borehole wall adjacent to the mirror as well as a view 

down the well at the borehole below the mirror. The 

result is a composite image of the two views (fig. Ib). 

The mirror view appears at the center of the image. 

Surrounding the mirror view is the wide-angle view 

that looks beyond the edges of the mirror, and down the 

well, much like the perspective shown in figure la. 

This expanded view is useful when searching for 

individual fractures identified in the first survey or 

when tracing a fracture that is connected to or near 

other fractures. During the second survey, the rocks 

were viewed in greater detail than in the first survey, 

and more precise measurements of the locations and 

dips of major fractures were made.

BOREHOLE VIDEO IMAGES

(a) Downhole view (D) Side-looking view

breakout zone

top of fracture

vertical limb 

of fracture

Figure 1 . Example of borehole video images: (a) Downhole view of the well at a depth of 57 meters, showing a fracture in granite 

intersecting the well bore; (b) Side-looking mirror view superimposed on the downhole view at a depth of 55 meters.

90
IDENTIFICATION OF ROCK TYPES, 

FRACTURES, AND BOREHOLE 

CONDITIONS

Rock fragments and chips diverted up the hole by 

the force of the air and water from the drill rig were 

collected and described over 1.5-m-long intervals. 

Although mineralogy, texture, grain size, and color of 

the cuttings could be adequately described, identification of rock types was difficult. Rock chips measuring 

1 to 2 mm in length are not always large enough to 

exhibit features of the source rock, such as foliation or 

banding. When rock chips from a short (1- to 1.5-m 

long) drilling interval include chips of granite and 

schist, it is impossible to determine from drill cuttings 

whether the well penetrated a schist xenolith in a granite, a granite dike in a schist host, or a contact between 

two rock types. Video surveys were completed in the 

29 bedrock wells to positively identify the rock types. 

In two of the wells from which core was collected, 

lithologies compiled after examination of the drilling 

chips and borehole video surveys corresponded to the 

lithologies determined directly from 100 m of bedrock 

core. Thus, the video images provide a direct verification 

of fractures and the contacts between rock types. In 

addition, video images can be used to describe fracture 

zones that are missing or completely rubbled in the 

bedrock core. For wells from which core was not collected, video images were used to improve the interpretation of the rock types based only on drilling logs and 

drill cuttings. The video images were used to describe 

texture, grain-size, color, presence of fractures, foliation, 

folds, and faults, as well as the condition of the borehole. From the video images, the borehole wall could 

be identified as circular, extremely enlarged, or angular 

and jagged. The data from the video logs were used for 

assessing the stability and integrity of the borehole and 

for planning other tests, including the placement of 

packers for hydraulic or solute-transport tests.

ROCK TYPES

Four major rock types (schist, granite, pegmatite, 

and basalt) have been identified in the bedrock outcrops, drill cuttings and video images in the Mirror 

Lake area. Geologic names were not assigned to the 

rocks encountered in the boreholes, because of the 

localized nature of the sampling. However, in the following list, the schist and gneiss probably correspond 

to the metamorphic rocks of the Rangely Formation of

the Silurian age (Lyons and others, 1986). The granite 

is most likely anatectic, two-mica granite of the 

regional Concord Intrusive Suite of the Devonian age 

(Armstrong and Boudette, 1984; Lyons and others, 

1986). These granites, derived from the melting of 

preexisting rocks, were injected into the overlying host 

rocks and commonly are found in sheetlike, tabular 

bodies. The basalt probably corresponds to the lamprophyric intrusions of Mid-Jurassic through Early Cretaceous age (McHone, 1984) and are found cross-cutting 

all other rock units.

Schists and gneisses. Schists and gneisses are usually 

easy to identify by the foliated biotite, muscovite, 

and sillimanite. Some schist sections of the wells 

exhibit isoclinal folding, augens, banding, and (or) 

large felsic sections. Metasedimentary rocks typically are black, brown, green, yellow, or white. 

The coarse-grained biotites and muscovites reflect 

the light and appear bright and shiny, whereas 

fine-grained schists appear dark. In general, the 

schists and gneisses are much darker than the 

granites. The mirror attachment permitted a 

detailed view of some structural and stratigraphic 

features in the schists, such as the isoclinal fold in 

schistose bedding cored with a quartz layer, shown 

in figure 2.

BOREHOLE VIDEO IMAGE

of an isoclinal 

fold in schist

Figure 2. The nose of an isoclinal fold in schist seen using the 

side-looking mirror at a depth of 200 meters. Bedding layers can 

be traced around the nose of the fold and down the limbs of the 

fold.

91
Granite. I\\Q granite is typically equigranular,

medium-to-coarse grained, leucocratic, creamy 

white to gray with an occasional greenish tinge. 

Iron staining along fracture faces and within the 

rock matrix can be seen clearly in some of the 

video images. These rocks are usually medium 

grained and have a sugary texture. They occasionally exhibit weak foliation of biotites and muscovites, and sometimes contain strongly foliated 

biotite schlieren, which are oriented parallel to 

major structural features in the host rock (Armstrong and Boudette, 1984). These linear features 

are easily visible in the video image and can usually be differentiated from fractures. Local variations in granite composition, such as increased 

biotite or quartz content, appear as variations in 

shading along the borehole wall.

Pegmatites. Pegmatites usually are the lightest and 

most reflective rocks in the boreholes. The contacts between the pegmatite and their host rocks 

were viewed with the mirror attachment. Some 

contacts were sharp and straight, whereas others 

were graded or wavy. Large individual crystals of 

muscovite, biotite, feldspar and quartz were 

observed and their lengths measured.

Basalt. The darkest and least common rocks in the 

wells are the basalts. These rocks reflect little light 

because of their extremely fine-grained matrix and 

mafic-mineral content. However, felsic inclusions, 

including feldspar phenocrysts and cavity fillings, 

are highly reflective and discernible in the video 

image. Basalts crosscut both the granite and the 

schist, are present at a variety of depths and in a 

variety of widths, and range from subhorizontal to 

subvertical. Chill margins and vesicles were 

viewed with the mirror attachment.

FRACTURES

Fractures appeared in the video image as dark lines 

and were differentiated from other linear features by a 

change in relief or the continuity and smoothness of the 

wall. When the light source was below the fracture, the 

fracture was in shadow. In contrast, the borehole wall 

appeared smooth and continuous at other dark linear 

features, such as narrow basalt dikes and biotite 

schlieren. The borehole was significantly enlarged at 

some fractures. Other fractures had breakout zones 

only along part of the fracture, as shown in figure la.

During the first survey, fractures were described 

as being either single fractures or as fracture zones 

(multiple fractures that occur over a short distance). In 

addition, the intersection of the fracture along the entire 

borehole or only part of the borehole was noted. In the 

second survey, the fractures were further described as 

being horizontal, moderately dipping, or steeply 

dipping. Fracture apertures were only qualitatively 

described as being very tight, narrow, or wide, or a 

fracture zone with breakout. No attempt was made to 

measure the fracture width. Fractures were also 

described for the occurrence of mineralization or mineral coatings, including iron (with some migration into 

the rock matrix), grayish white clay, white calcite, and 

light white to brownish yellow quartz. Black manganese coatings were impossible to differentiate from 

dark minerals.

BOREHOLE CONDITIONS

In addition to providing rock type and fracture 

information, the video images were used to determine 

borehole conditions, including roughness and angularity of borehole walls, fractures with broken-out zones 

or borehole enlargement, alteration of the rock mass, 

scoring or rifling caused by drilling, blockage of 

the drill hole, and possible zones of falling rocks. 

The borehole shown in figure 3 is no longer circular,

BOREHOLE VIDEO IMAGE

angular,

eniarged and jagged 

fracture zone

Figure 3. Downhole view of an enlarged and jagged breakout 

zone at a fracture zone at a depth of 165 meters, whereupon the 

well caved in.

92
but is now angular, jagged, and enlarged at this fracture 

zone. This well was originally drilled 30 m deeper than 

shown, but it caved in at this fracture zone. Enlarged 

and angular sections of the borehole that were identified in the video surveys were avoided when placing 

packers for hydraulic testing, tracer tests, and longterm water-level monitoring. Information on the condition of the borehole can be used to explain the results 

of other borehole tests.

SUMMARY AND CONCLUSIONS

Video cameras are effective tools for identifying 

rock types, fractures, foliation, folds and enlarged 

zones in the boreholes. Interpretations of rock types 

using video logs and drill cuttings agreed with interpretations of the core. In contrast, rock types could not 

always be determined from analysis of drill cuttings. 

The video images provided information about the distribution of rock types and the extent of the fractures 

that was not available from previously collected data, 

including data obtained by conventional drilling and 

borehole geophysical methods. Based on these merits, 

video cameras are a practical method for obtaining 

borehole data. The interpretation of the rock types from 

the video surveys will be used to calibrate borehole 

geophysical signals.

REFERENCES

Armstrong, F.C., and Boudette, E.L., 1984, Two-mica granites; Part A, Their occurrence and petrography: U.S. 

Geological Survey, Open-File Report 84-0173, 36 p.

Hardcastle, K.C., 1989, Analysis of faults and fracture

fillings along a cross section of the New England Appalachians: Clues to the later brittle deformation of an 

orogenic belt [published thesis]: Amhcrst, University 

of Massachusetts, 242 p.

Lyons, J.B., Bothner, W.A., Moench, R.H. and Thompson, 

J.B., Jr. eds.,1986. Interim geologic map of New 

Hampshire: Concord, N.H. New Hampshire Department of Resources and Economic Development, OpenFile Report 86-1, 1 sheet, scale 1:250,000.

McHone, J.G., 1984, Mesozoic igneous rocks of northern 

New England and adjacent Quebec: Summary, description map and bibliography of data sources: Geological 

Society of America, Map and Chart Series, MC-49.

Shapiro. A.M., and Hsieh, P.A., 1991, Research in fractured 

rock hydrogeology: Characterizing fluid movement and 

chemical transport in fractured rock at the Mirror Lake 

drainage basin, New Hampshire, in Mallard, G.E., and 

Aronson, D.A., cds.,Toxic Substances Hydrology Program Proceedings of technical meeting, Montcrcy, 

Calif. March 11-15, 1991: U.S.Geological Survey 

Water-Resources Investigations Report 91 -4034, 

p. 155-161.

Trainer, F.W., 1987, Hydrogeology of the plutonic and 

metamorphic rocks, in Back, W., Rosenshcin, J.S., and 

Seaber, P.R., eds., Hydrogeology: Boulder, Colo., Geological Survey of America, The Geology of North 

America, v. O-2, p. 367-380.

93
94
Fracture Detection in Crystalline Rocks, Mirror Lake 

Area, Grafton County, New Hampshire

By P.P. Haeni 1 , John W. Lane, Jr. 1 , C.C. Barton2, and David A. Lieblich 1

Abstract

Six surface-geophysical methods were used 

to detect saturated fractures in the upper 60 m 

(meters) of bedrock at the U.S. Geological survey's fractured-rock research site in the Mirror 

Lake area, Grafton County, New Hampshire. 

Crystalline bedrock, consisting of foliated schists 

intruded by granite, pegmatite, and gabbro, underlies 3 to 10m of glacial drift throughout the study 

area. Surface-geophysical methods included 

azimuthal seismic refraction, azimuthal Schlumberger direct-current- (DC) resistivity, squarearray DC-resistivity, inductive-terrain conductivity, very-low-frequency (VLF) terrain resistivity, 

and ground-penetrating radar (GPR).

Azimuthal seismic-refraction and DCresistivity methods measured directionally dependent physical properties of the crystalline rock at 

the Camp Osceola well field and in a ballfield 

75 m southeast of Mirror Lake. The interpretation 

of the seismic-refraction P-wave data is that the 

primary fracture strike is 022.5° (degrees) with a 

secondary strike at 127°. The orientation of the 

anisotropy and probably the fractures, as determined from a quantitative interpretation of the 

original P-wave data, is 037°. The interpretation 

of the DC-resistivity data is that the primary fracture strike is 030° with a secondary strike at 150°. 

Inductive-terrain conductivity and VLF-terrain 

resistivity data show very small anomalies that 

have been interpreted as possible fractures or 

fracture zones having a strike of 045°. Processed

'U.S. Geological Survey, WRD, Hartford, CT. 

U.S. Geological Survey, Geologic Division, Denver, CO.

GPR data were used to determine the depth to 

bedrock and to locate numerous subhorizontal 

reflectors, which are interpreted to be fractures, 

fracture zones, and (or) foliation in the bedrock.

The strike interpreted from the surfacegeophysical data correlates with the strike determined from bedrock outcrops, which are located 

150 m from the Camp Osceola well field along 

Interstate 93. The outcrop data indicate a fracture-strike frequency maximum at 030° and a 

secondary maximum at 000°. Fracture dips range 

from 20° to 90°.

INTRODUCTION

The U.S. Geological Survey established a 

fractured-rock research site in the Mirror Lake area, 

Grafton County, New Hampshire (fig. 1) for the purpose of conducting multidisciplinary research on the 

flow of contaminants in fractured bedrock. As part of 

the study, surface-geophysical surveys to detect bedrock fractures and detailed mapping of bedrock outcrops have been conducted. This paper presents the 

results of surface-geophysical surveys and the mapping of fractures in bedrock outcrops in the Mirror 

Lake area.

Surface-geophysical methods have been used 

previously in a number of studies to detect saturated 

fractures in bedrock (Haeni and others, 1993; Lewis 

and Haeni, 1987). In most of these studies, a single 

geophysical method was used to define the predominant strike of steeply dipping fractures. The methods 

most commonly used were seismic refraction or 

direct-current- (DC) resistivity arrays rotated about 

a fixed centerpoint in order to measure azimuthal 

changes in the physical properties of the rock.

95
71°42' Orf

U.S. GEOLOGICAL SURVEY WATER RESOURCES DIVISION

MIRROR LAKE BASE MAP - STATE PLANE PROJECTION ZONE 4676

NATIONAL GEODETIC VERTICAL DATUM OF 1929

2GOI FEET

7 1°4 1' 23*

IOOO

I ] I I I

50O METERS

43°56 I 28"-

43°56'22" I 

Figure 1. Location of U.S. Geological Survey fractured-rock research 

site and geophysical-survey sites, Mirror Lake area, Grafton County, 

New Hampshire. (From Haeni and others, 1993, fig. 1.)

Center of azlmuthal. 

DC resistivity and 

refraction aurveya.

CAMP OSCEOLA WELL FIELD

96

500 FEET

i IT r i i

0 IOOMETFRS
More recent studies have used multiple geophysical 

methods (Lieblich and others, 1991).

For this study, surface-geophysical surveys were 

conducted at two sites the Camp Osceola well field 

and at a field (hereafter referred to as the ballfield) 

located 75 m southeast of Mirror Lake, both of which 

are located in the Mirror Lake area, Grafton County, 

New Hampshire (fig. 1). At both sites, crystalline bedrock, consisting of foliated schists intruded by granite, 

pegmatite, and gabbro, underlies 3 to 10m of glacial 

drift. Analysis of fracture data from outcrops on Interstate Route 93, located 150 m from the Camp Osceola 

well field, indicates a fracture-strike frequency maximum at 030° and a secondary maximum at 000° (fig. 

2). Fracture dips range from 20° to 90°.

SURFACE-GEOPHYSICAL METHODS 

AND RESULTS OF SURVEYS

Azimuthal Seismic-Refraction Method

The seismic-refraction method uses an impulsive source to generate elastic waves that propagate 

through the subsurface. The critically refracted 

(head) wave, either P or S, is detected at geophones on 

the surface when the seismic velocity increases at 

depth. The effects of single fractures or small, isolated 

fracture zones are unlikely to be seen in P-wave 

refraction surveys. Azimuthal refraction methods may 

be able to detect an anisotropic distribution of fractures throughout the rock. For a single, saturated, 

steeply dipping fracture zone, a seismic-velocity maximum occurs in the direction of the fracture strike, and 

a seismic-velocity minimum occurs at 90° to the fracture strike. Velocity and (or) thickness variations in 

the material overlying the refractor can also contribute 

to azimuthal velocity variations. A P-wave refraction 

experiment that has successfully detected fractures in 

crystalline rocks is described in Park and Simmons 

(1982).

Azimuthal seismic-refraction data using Pwaves were collected at the ballfield in 1990 (fig. 1) by 

rotating a refraction line in 022.5° -increments about a 

common centerpoint. For each line, a 5.7-kg (kilogram) sledge hammer served as the energy source at 

each shot location. Twelve vertical-displacement 

7-Hz (hertz) geophones were spaced 3.1 m apart, and 

there were five shot points per seismic line. The near 

end-shot was 3.0 m from each end of the geophone

line and the far end-shot was 35.1 m from each end of 

the geophone line. Additionally, a center shot between 

geophones 6 and 7 was used. A signal-enhancement 

seismograph was used to record the data. Elevations 

were recorded for each shot location and each geophone position to within 0.03 m.

Unlike some other geophysical methods used in 

this study, analysis and interpretation of seismicrefraction field data must be performed to obtain 

P-wave velocity for a given direction. For this study, 

velocity analysis was done using the Hobson-Overton 

method, developed by Scott and others (1972). The 

interpreted azimuthal plot of the seismic-velocity data 

(fig. 2) indicates a primary fracture and (or) foliation 

strike at 022.5° and possibly a secondary strike at 127° 

(Lieblich and others, 1991).

Additional seismic-refraction surveys using 

multicomponent shear and P-waves were conducted in 

1992 as part of an effort to identify the magnitude of 

the true anisotropy. Preliminary results from the interpretation of a 9-component walkaway-noise survey 

and a radial survey indicate that the bedrock is anisotropic. A quantitative interpretation of the original 

P-wave data indicates that the principal direction of 

the azimuthal anisotropy, and probably the strike of 

the fractures, is 037°.

Azimuthal Schlumberger DC-Resistivity 

Method

The DC-resistivity method uses either a directcurrent source or an alternating-current source with a 

frequency below 10 Hz. The current is transmitted 

into the ground through two electrodes, and the 

receiver measures the potential difference (voltage) 

at two other electrodes. The apparent resistivity of the 

subsurface can then be calculated for a specific electrode spacing and geometry. The DC-resistivity 

method measures the apparent resistivity of a volume 

of rock, which is usually large compared to the scale 

of and the spacing between individual fracture zones. 

For a single, saturated, steeply dipping fracture zone, 

an apparent resistivity maximum occurs in the direction of the fracture strike and an apparent resistivity 

minimum occurs at 90° to the fracture strike.

An azimuthal DC-resistivity survey using the 

Schlumberger array was conducted at the Camp 

Osceola well field in 1991. Data were collected every 

45°, and the AB/2 spacing was expanded from 3 m to 

30 m. Qualitative interpretation of these data indicate

97
70 .

60 _

50 

O 40 

LJ

Z> 

O

CK 3° ' 

L_

20 

10 

n = 534

0 10 20 30 40 50 60 70 80 90 100 110 120 130 140 150 160 170 180

STRIKE AZIMUTH (DEGREES)

EXPLANATION

Line of equal seismic velocity 

in kilometers per second

Interpreted primary fracture strike 

Possible secondary fracture strike

Data points showing Interpreted seismic 

velocity through bedrock at Indicated 

azimuthal directions with straight line 

Interpolations between points.

014.5 Compass azimuth, In degress

Figure 2. Fracture strike frequency at Route 93 outcrop (upper) and seismic P-wave velocity (lower) plotted against azimuth, 

Mirror Lake area, Grafton County, New Hampshire. (From Haeni and others, 1993, fig. 2.)

98
a fracture zone and (or) foliation strike at 045°. 

Azimuthal DC-resistivity surveys using the Schlumberger array also were conducted at the ballfield in 

1991 and 1992. The soundings in 1991 were collected 

every 22.5° using a DC-resistivity system, and the 

AB/2 spacings were expanded from 3 m to 40 m.

The soundings in 1992 were collected every 15° with 

an ABEM automatic sounding system, and the AB/2 

spacings were expanded from 3 to 40 m. The results 

of the 1991 survey are shown in figure 3. The fracture strike interpreted from both data sets is 352°.

EXPLANATION

Line of equal apparent resistivity, 

In ohm meters

Interpreted primary fracture strike

Data points showing apparent resistivity 

measured at MN/2 - 2 meters through 

bedrock at Indicated azlmuthal directions 

with straight line Interpolations between points.

Data points showing apparent resistivity 

measured at MN/2 - 4 meters through 

bedrock at Indicated azlmuthal directions 

with straight line Interpolations between points.

000 Compass azimuth. In degress

1000

EXPLANATION

Line of equal apparent resistivity. 

In ohm meters

Interpreted primary fracture strike 

Possible secondary fracture strike

 Data points showing apparent resistivity

at the Indicated azlmuthal directions with 

straight line Interpolation between points.

000 Compass azimuth, In degress

Figure 3. Schlumberger (upper) and square array (lower) apparent resistivity plotted against azimuth at the ballfield, Mirror 

Lake area, Grafton County, New Hampshire.

99
Square-Array DC-Resistivity Method

The square array was originally developed to 

take the place of Wenner or Schlumberger arrays when 

a dipping subsurface, bedding, or foliation was 

present. Techniques for analyzing the directional 

resistivity information provided by the square-array 

method were developed by Habberjam (1972, 1975). 

The method has not been in widespread use, but two 

studies that used the square array to detect fractures 

are described by Darboux-Afouda and Louis (1989) 

and Sehli( 1990).

A DC-resistivity survey using the square array 

is conducted similarly to a traditional linear array 

(Habberjam and Watkins, 1967). The measurement 

location is at the centerpoint of the square and the 

array size (A) is the length of the side of the square. 

For depth sounding, the array is expanded symmetrically about the centerpoint, in increments of A+/2 

(Habberjam and Watkins, 1967). Major advantages of 

the square-array method are that the array has approximately twice the sensitivity to anisotropy as Schlumberger arrays, and the array occupies less surface area 

for a given depth of penetration (Darboux-Afouda and 

Louis, 1989). For a single zone of oriented, saturated, 

steeply dipping fractures, the apparent resistivity minimum measured by a square array is oriented in the 

same direction as the fracture strike.

Square-array data were interpreted in two different ways to determine the fracture strike graphically 

(using square-array data collected every 15°) or 

analytically (using crossed square-array data collected 

every 45°). The strike was interpreted graphically by 

plotting the apparent resistivity for azimuthal square 

arrays against the azimuth of the measurement; the 

fracture strike is perpendicular to the direction of maximum resistivity. Graphical display and interpretation 

of the individual square-array data were performed on 

a rosette diagram using methods of Taylor and Fleming (1988). The strike was interpreted analytically by 

using crossed square-array data (two square arrays 

separated by an angle of 45°). Methods for analyzing 

crossed square-array data are described by Habberjam 

(1975).

A square-array DC-resistivity survey was conducted in the ballfield in 1992 that consisted of six 

square-array soundings separated by a rotational angle 

of 15° about the array centerpoint. The A-spacings of 

the arrays were expanded from 5 to 50 m in logarithmic intervals for each sounding. The data from the six 

square arrays show a significant change in apparent

resistivity for different azimuthal orientations for all 

A-spacings. The azimuthal apparent resistivity values 

for the 50-m A-spacing (fig. 3) have a primary maximum resistivity at 120° and a secondary maximum 

resistivity at 060°. A graphical interpretation is that a 

primary fracture zone is present with a strike of 030° 

and a secondary fracture zone has a strike of 150°. 

From the data collected with the six square arrays, 

there are three independent crossed square-array data 

sets available for further analysis. Analytical interpretation of these data indicates a fracture strike at 

024° and an apparent anisotropy value (the ratio of the 

maximum to the minimum resistivity) of 1.33.

Close inspection of the data, field site, and survey design can explain the discrepancy between the 

Schlumberger and square-array data, and also illustrate possible weaknesses of using azimuthal Schlumberger array surveys for fracture detection: (1) A large 

current-electrode separation (AB/2) is required by the 

Schlumberger array. This caused a buried telephone 

line to be included within the array that was not 

included within the smaller square array. (2) A large 

voltage-electrode separation (MN/2), as compared to 

the fracture spacing, is required by the Schlumberger 

array to detect fractures. The largest MN/2 separation 

used in this study was 4 m, which may not have been 

large enough to meet the theoretical requirements. 

(3) Lithologic inhomogeneities present near the voltage electrodes may have affected the Schlumberger 

data . In view of these considerations, the squarearray results are interpreted to be more representative 

of the conditions at the ballfield than the Schlumberger 

array results.

Inductive Terrain-Conductivity Method

Horizontal and vertical dipole inductive terrainconductivity data were collected in the ballfield along 

three lines oriented east-west and five lines oriented 

north-south. A station spacing of 0.5 m and a coil 

separation of 20 m was used to collect the data. Field 

measurements could be read to an accuracy of 

0.2 mS/m. Where unstable readings were obtained, 

multiple readings were made and, if necessary, stations 

were occupied at later times.

The data were interpreted manually by looking 

for an anomaly typically associated with vertical conductors. The anomalies present on the lines, at this 

site, were extremely small (less than 0.4 mS/m) and 

could only be identified when multiple parallel lines

100
were analyzed. One possible fracture or fracture zone 

with a strike of 045° was interpreted from the field 

data.

Very-Low-Frequency Terrain Resistivity 

Method

VLF data were collected in the ballfield along 

three lines oriented east-west. Measurements were 

taken every 1 m using the Cutler, Maine and Annapolis, Maryland VLF stations. In addition, a portable 

transmitter with a 1 km single strand of wire was used 

to obtain a third data set.

The data sets from the remote VLF stations 

were virtually identical. The signal obtained with the 

portable transmitter was stronger than that of the 

remote transmitters, and this signal was used for the 

interpretation. VLF anomalies were very small (less 

than 20 percent of the horizontal primary field), and 

the only way that fractures could be identified was by 

comparing parallel VLF data lines. Three fractures or 

fracture zones were interpreted from the VLF data, all 

with a strike of 045°.

Ground-Penetrating Radar Method

The ground-penetrating radar (GPR) method 

uses a control unit and transmitter and receiver antennas, which may be housed in the same unit, to generate 

and detect reflected electromagnetic (EM) waves. The 

advantage of GPR for fracture detection is that the 

conductivity and dielectric permittivity are substantially different in a fluid-saturated fracture or fracture 

zone as compared to those properties in the host rock. 

In previous investigations at other sites, Ulriksen 

(1982) obtained reflections from individual fractures 

or fracture zones in granodiorite, and Imse and Levine 

(1985) obtained strong scattering responses from 

steeply dipping fractures in carbonate rocks in northern New York.

A radar unit with a bistatic antenna configuration that operated at a center frequency of 80 MHz 

(megahertz) was used in this study. The antennas were 

towed about 30 m behind a vehicle moving at approximately 4.8 km/hr. A scan rate of 25.6 scans per second 

and a time scale of 900 ns was used. The gain was 

adjusted in the field and the signal was digitally 

recorded to allow for further processing. The records 

were marked for every 9.1 m of ground covered. 

Seven profiles, three east-west and four north-south,

were run in the ballfield. A reflector, interpreted as the 

top of bedrock, is present at the top of most of the 

radar records. Numerous subhorizontal reflectors 

within the bedrock are interpreted as fractures or foliation planes, but no vertical fractures were identified.

SUMMARY

Six surface-geophysical methods were used to 

detect saturated fractures in crystalline bedrock at the 

U.S. Geological Survey's fractured-rock research site 

in the Mirror Lake area, Grafton County, New Hampshire. Geophysical surveys were run during 1990-92 

at the Camp Osceola well field and at a ballfield 

located 75 m southeast of Mirror Lake.

Interpreted azimuthal seismic-refraction P-wave 

data from the ballfield show a primary fracture (and) 

or foliation with a strike of 022.5° and a secondary 

fracture and (or) foliation pattern with a strike of 127°. 

The orientation of the anisotropy and probably the 

fractures, as determined from a quantitative interpretation of the original P-wave data, is 037°. The azimuthal Schlumberger DC-resistivity method produced 

inconsistent results and was shown to be less reliable 

at detecting fracture strike than the square-array 

method. The interpreted azimuthal Schlumberger data 

from the Camp Osceola well field showed a fracture 

zone with a strike of 045°, whereas the interpreted data 

from the ballfield showed a direction of 352°. Interpreted square-array measurements at the ballfield 

showed a primary fracture with a strike of 030° and a 

secondary fracture direction of 150°. The two electromagnetic methods, VLF terrain conductivity and 

inductive terrain conductivity, had very small anomalies that could only be tentatively identified as being 

caused by fractures (oriented 045°) when multiple 

lines were correlated with each other. Processed GPR 

data showed numerous subhorizontal reflectors that 

were interpreted as fractures or foliations in the rock. 

In addition, the depth to the top of bedrock was determined from the GPR data.

The strike interpreted from the surfacegeophysical data correlates with the strike determined 

from bedrock outcrops, which are located 150 m from 

the Camp Osceola well field along Interstate 93. The 

outcrop data indicates a fracture-strike frequency 

maximum at 030° and a secondary maximum at 000°. 

Fracture dips range from 20° to 90°.

101
NOTICE

Although the information in this document has 

been funded wholly or in part by the U.S. Environmental Protection Agency under interagency agreement DW14935104 to the U.S. Geological Survey, it 

does not necessarily reflect the views of the U.S. Environmental Protection Agency, and no official endorsement should be inferred.

REFERENCES

Darboux-Afouda, R., and Louis, P., 1989, Contribution dcs 

incsurcs dc Tanisotropic clcctriquc la recherche dcs 

aquifres dc fracture en milieu cristallin au Benin: Geophysical Prospecting, v. 37, p. 91-105.

Habbcrjam, G.M., 1972, The effects of anisotropy on square 

array resistivity measurements: Geophysical Prospecting, v. 20, p. 249-266.

Habbcrjam, G.M., 1975, Apparent resistivity, anisotropy 

and strike measurements: Geophysical Prospecting, 

v. 23, p. 211-247.

Habbcrjam, G.M.. and Watkins, G.E., 1967, The use of a 

square configuration in resistivity prospecting: Geophysical Prospecting, v. 15, p. 221-235.

Hacni, P.P., Lane, J.W., Jr., and Licblich, D.A., 1993, Use of 

surface-geophysical and borehole-radar methods to 

detect fractures in crystalline rocks. Mirror Lake Area, 

Grafton County, New Hampshire, in Banks, Sheila, 

and Banks, David, cds., Hydrogcology of Hard Rocks. 

Mcmoircs of the XXI Vth Congress, Oslo, Norway, 28 

June to 2 July 1993: Oslo, International Association 

of Hydrologists, p. 577-587.

Imsc, J.P., and Lcvine, E.N., 1985, Conventional and state 

of the art geophysical techniques for fracture detection: 

National Water Well Association Annual Eastern 

Regional Groundwatcr Conference, 2nd , Portland, 

Maine, 1985, Proceedings, p. 18-36.

Lewis, M.R., and Haeni, P.P., 1987, The use of surface 

geophysical techniques to detect fractures in bedrock an annotated bibliography: U.S. Geological 

Survey Circular 987, 14 p.

Licblich, D.A., Lane, J.W., Jr., and Haeni, P.P., 1991,

Results of integrated surface-geophysical studies for 

shallow subsurface fracture detection at three New 

Hampshire sites, in Expanded Abstracts with Biographies, SEG 61st Annual International Meeting, Houston, Texas, November 10-14, 1991: Houston, Texas, 

Society of Exploration Gcophysicists, p. 553-556.

Park, S., and Simmons, G., 1982, Crack induced velocity 

anisotropy in the White Mountains, New Hampshire: 

Journal of Geophysical Research, v. 87, no. 84, 

p. 2977-2983.

Scott, J.H., Tibbetts, B.L., and Burdick, R.G., 1972, Computer analysis of seismic refraction data: U.S. Bureau 

of Mines Report of Investigations 7595, 95 p.

Schli, A.S., 1990, Contribution of electrical prospecting to 

the geophysical study of discontinuous media, in International symposium on applications of geophysics to 

water prospecting in arid and semi-arid areas, Aug. 22- 

24, 1990, Luzcrn, Switzerland, unpaginatcd.

Taylor, R.W., and Fleming, A.H., 1988, Characterizing 

jointed systems by azimuthal resistivity surveys: 

Ground Water, v. 26, no. 4, p. 464-474.

Ulrikscn. C.P.F., 1982, Application of impulse radar to civil 

engineering: Hudson, N.H., Geophysical Survey Systems, Inc., 179 p.

102
Use of Well Logs to Prepare the Way for Packer 

Strings and Tracer Tests: Lessons from the

Mirror Lake Study

By Frederick L. Paillet1

Abstract

Although well logs can never replace conventional aquifer-testing techniques, a carefully 

designed suite of borehole measurements can identify the distribution of fractures along well bores. 

This information can be used to improve the effectiveness of hydraulic testing and water sampling. 

After nearly a decade of logging research at the 

Mirror Lake site, we have developed a combination 

of conventional and new techniques that have 

proven especially effective in preparing the way for 

subsequent installation of packer strings and tracertest analysis. The approach consists of four steps: 

(1) characterizing the general structure and lithology of the rock mass by interpreting conventional 

well logs, (2) identifying and characterizing individual fractures by obtaining acoustic televiewer 

and other borehole image logs, (3) identifying the 

few fractures intersecting the well bore that produce water during pumping by running highresolution flow logs in pumped and observation 

boreholes, and (4) inferring the pattern of fracture 

connectivity in the rock mass adjacent to boreholes 

by analyzing transient flow in boreholes at carefully selected locations between producing fractures. The application of this analysis allows 

geochemists and hydrologists to identify those 

locations where packers can be set to prevent communication between conductive fractures and to 

design efficient procedures for otherwise timeconsuming aquifer and tracer tests.

U.S. Geological Survey, Denver, CO

INTRODUCTION

Fractured-bedrock formations are some of the 

most heterogeneous aquifers investigated in studies of 

toxic-substance dispersal in ground water. Characterization of a fractured formation requires extensive 

series of water-sample analyses, straddle-packer injection tests, and tracer studies to infer the complicated 

three-dimensional distribution of transmissivity within 

the rock mass. The problem confronting the hydrologist is especially complex where boreholes encounter 

numerous fractures, all of which potentially can communicate with each other in an open borehole. Preliminary study is needed to identify the specific depths in 

boreholes where a limited string of packers can be 

used to isolate the primary flow zones.

The U.S. Geological Survey has been conducting a long term study to develop new analytical techniques for applying borehole geophysical logging to 

the characterization of fractured-bedrock aquifers at 

the Mirror Lake site. Since this study began in 1984, 

logs have been run in 17 boreholes at two borehole 

arrays and in 12 isolated boreholes scattered over the 

study area (fig. 1). This report reviews the suite of 

conventional and recently developed experimental 

logs that has been developed for the specific purpose 

of preparing the way for subsequent packer installation and tracer tests. Well logging is important in this 

regard because it provides quick and efficient characterization of the fracture distribution along the well 

bore, identification of the very few hydraulically 

important fractures that are connected to large-scale 

flow paths, and information concerning the possible 

connections between those flow paths in the adjacent 

rock mass.

103
74° 41' 0" 74° 41' 30"

CONTOUR INTERVAL IS 50 METERS 

\ DATUM IS SEA LEVEL 

BOREHOLE ARRAYS BASE IS FROM U.S. GEOLOGICAL SURVEY

\ WOODSTOCK, NEW HAMPSHIRE 1:24,000 - 1980 

BEDROCK BOREHOLES

\

*FSE-3

Figure 1 . Location of isolated boreholes and the Forest Service East (FSE) and Camp Osceola (CO) borehole arrays at the 

Mirror Lake, New Hampshire fractured rock study site.

104
EQUIPMENT

Conventional borehole geophysics (well logging) is useful in defining the general lithology of rock 

masses penetrated by boreholes, but is not very useful 

in identifying individual fractures. Caliper logs indicate those intervals where the borehole is enlarged by 

fracturing. Natural gamma logs indicate where fractures could be associated with lithologic contacts or 

where radioisotopes could be deposited as partial fracture fillings. Epithermal-neutron, electrical resistivity, 

and acoustic logs indicate where rock properties have 

been altered adjacent to fractures. All of these logs 

sample rock properties over volumes too large to be 

affected by fracture permeability directly, but each of 

these logs provides useful indications of the direct 

(mechanical weakening) and indirect (alteration) 

effects of fractures on the bulk properties of rocks 

(Keys, 1979; Paillet, 1993a).

The acoustic borehole televiewer (BHTV) log is 

a device that produces a photograph-like image of the 

borehole wall by recording the intensity of ultrasonic 

reflections from the borehole wall (Zemanek and 

others, 1970; Keys, 1979). The images produced by 

the BHTV system can be used to identify the strike, 

dip, and relative size of fractures that intersect the 

borehole; however, the BHTV system characterizes 

fractures only in the immediate vicinity of the borehole wall where the appearance of the fracture has 

been affected by drilling.

The heat-pulse flowmeter (HPFM) is a highresolution flowmeter designed to measure vertical 

flows in well bores over the range of 0.04 to 10 L/min 

(Hess, 1986). Flows are measured by recording the 

time interval between the electrical generation of a 

small parcel of heated water in the measurement section of the tool and the arrival of that pulse at a thermistor located 2 cm above or below the heat source. 

Profiles of the vertical flow in boreholes are generated 

by measuring flow at discrete depth points where the 

tool is held stationary and a downhole bladder is 

inflated to fill the annulus (Hess and Paillet, 1990). 

The flowmeter is used to measure flow produced by 

the influence of naturally occurring differences in 

hydraulic head between fractures intersected by the 

borehole, and the changes in that flow field produced 

by pumping from one of the boreholes in the well 

field.

FRACTURE CHARACTERIZATION AND 

FLOW ANALYSIS

The analysis of fractured rock aquifers developed over the course of the Mirror Lake study consists 

of four steps: (1) identification of the general structure 

and lithology of the host rock formation, (2) identification and characterization of the fractures intersecting 

the borehole, (3) identification of fractures conducting 

flow during cross-borehole aquifer tests, and (4) 

development of specific hydraulic models for flow 

paths between boreholes based on changes in the flow 

field over time during cross-borehole flow tests.

Geophysical well logs can be used to determine 

the general structure of a rock mass, and BHTV image 

logs can be used to identify the distribution of fractures along the well bore. These applications of logs 

are illustrated schematically in figure 2A and 2B, and 

are described in detail by Paillet (199la), Paillet and 

Kapucu (1989), and Keys (1979). The results are 

especially effective when logs indicating alteration, 

such as the resistivity log, are compared to image logs, 

such as BHTV, that show how the patterns of alteration are related to specific sets of fractures. As indicated in the figure, these results can be useful in 

inferring the locations of unaltered and unfractured 

intervals that could be used to isolate fracture zones 

from each other.

Numerous studies have shown that only a small 

subpopulation of fractures conduct flow during aquifer 

tests, and that fracture connectivity is more important 

in controlling the distribution of flow than local fracture aperture (Paillet, 1991b; Long and others, 1982). 

High-resolution flow measurements made with the 

HPFM during borehole pumping, or in observation 

boreholes when an adjacent borehole is pumped, are 

very effective in identifying those few fractures conducting flow during aquifer tests or cross-borehole 

flow tests (fig. 2C). Even when the depths of fractures 

conducting flow into or out of pumped and observation boreholes are known, it is still not known how 

those fractures are connected to allow flow through the 

surrounding fractured rock. Some of these connections 

can be inferred by correlating fractures and fracture 

zones between boreholes or by projecting inflow and 

outflow points identified during aquifer tests across the 

borehole array. Such projections are described in 

detail by Paillet and others (1987) and Paillet (1991 a). 

This technique has been expanded to address the transient evolution of the measured flowfield during the 

course of cross-borehole flow tests. For example,

105
B OBSERVATION WELL LOGS

CALIPER, 

IN CM

RESISTIVITY, 

IN OHM-METERS TELEVIEWER 

PUMPED WELL OBSERVATION WELL 12 15 18 0 1000 2000 3000

OBSERVATION WELL 

FLOW, IN LITERS PER MfNUTE

-0.2 -0.1 0.0 0.1 0.2

DOWNFLOW

EARLY^

AMBIENT 

FLOW

UPFLOW

a

LATE

PUMPED 

FLOW

FLOW TRANSIENTS,

0.15

0.0

-0.15

z

5 tr 0.15

Ul 

Q.

6 0.0

-0.15

u! 0.15

0.0

-0.15

I

T

UPFLOW

DOWNFLOW

0 100 200 300

TIME, IN MINUTES 

PUMP ON PUMP OFF

Figure 2. Schematic illustration of fracture characterization: (A) two fracture zones intersected by the pumped and 

observation boreholes; (B) caliper, televiewer, and electrical resistivity logs for the observation borehole; (C) vertical-flow 

profiles obtained during cross-borehole flow tests used to define locations where water enters and exits boreholes under 

ambient conditions, shortly after the pumping begins (early), and after pumping for more than 1 hour (late); and (D) transient 

flows measured at specific depths in the observation borehole during repeated aquifer tests.

106
previous results have shown that the vertical profile of 

flow varied with time during cross-borehole flow tests 

(fig. 2C). Flow transients can be measured by continuously recording the variation in flow measured when 

the experiment is repeated with the flowmeter stationed at specific depths (a, b, and c in fig. 2C). These 

transient flows conform to certain type curves (Paillet, 

1993b) such as those illustrated in figure 2D.

TRANSIENT-FLOW TYPE CURVES AND 

ANALYSIS

The type-curve analytical technique developed 

from the Mirror Lake study is based on the recognization of characteristic flow responses for measurements 

made in casing above all fractures that intersect the 

well bore and measurements made between pairs of 

fractures communicating with each other along the 

open borehole during the course of the experiment 

(fig. 3). Transient-flow curves of type a (fig. 3) are 

obtained when measurements are made in casing. 

These transients are characterized by a maximum flow 

(Qmax) anc* a time interval required to achieve that 

maximum (AT).

Interaction between the two sets of conductive 

fractures is determined by the shape of the type curves 

for flow along the well bore at measurement station b 

(type curves bj, b2 and b3 in fig. 3; Paillet and others, 

1992). If the two sets of fractures do not communicate 

with each other in the rock mass between the two 

boreholes, the flow transient will build up to a quasisteady value as in type curve bj. However, if there are 

permeable connections between the two fractures in 

the intervening rock mass, the quasi-steady flow will 

not be achieved and a flow transient similar to type 

curve b2 will be measured. In the extreme situation 

where the two fracture zones are short-circuited in the 

intervening rock mass, type curve b will be exactly 

proportional to type curve a (type curve b3 in fig. 3).

Two field examples of these type curves are 

given in figure 4. In figure 4A, the flow in the open 

hole between boreholes matches type curve b|, indicating that the two fractures are completely isolated in 

the surrounding rock mass. However, the flow in the 

open borehole is similar to type curve b3 , in figure 4B 

indicating that the two fractures are short-circuited 

near the well bore. The strange shape to the transient 

measured in casing in figure 4B represents a pump 

failure during that phase of the aquifer test and not a 

departure from the expected type curve.

PUMPED WELL OBSERVATION WELL

PUMP ON PUMP OFF

TIME, IN MINUTES

FLOW MEASURED AT STATION A FLOW MEASURED AT STATION 

IN OBSERVATION WELL IN OBSERVATION WELL

Figure 3. Schemcatic illustration of (A) transient flow in ar 

observation borehole when the pumped borehole 

intersects a single major fracture zone, and (B) the 

observation borehole provides a connection between the 

major zone and a deeper fracture zone; Qmaxb2> ^3 denote 

type curves obtained at depth stations a and b under 

various conditions.

Although these results seem clear-cut and useful, one note of caution is warranted. The analogy 

between flow-transient type-curve analysis and standard hydraulic tests is clear, but the transient measurements are made in the presence of open boreholes. 

The well-bore storage and communication along the 

open boreholes in the well field affect these measurements just as much as they affect conventional aquifer 

tests. The flow-transient analysis could be improved 

by installing packers at appropriate locations in all of 

the open boreholes; this technique will be explored in 

future studies. Even without such measures, we find 

the flow-transient technique useful in making a

107
PUMPING RECOVERY

l.U 

UJ1-

Z

2 0.5

CE

UJ

jg 0.0.

UJ

H

Z

- -0.5 (

0

u. 

.1 n

PUMPING IN FSE-4 ,

FLOW IN FSE-2

t

i

 ,

^OCCCD O O OO (

)

,

5

)

?

L

IBfrzooo o o .

I

"

0 20 40 60 80 100 12: 140 160 180 200 

TIME, IN MINUTES

PUMPING .RECOVERY

u.o

UJ h-

^ 0.4

i

PUMPING IN FSE-8

. FLOWINFSE-11

oc 0.2 

UJ ( )

g 0.0 I

uj k

t U^M.-

2 -0.2

0-0.4

u. 

.n K

**> nn J

^^-O O O O 0 CEC

III!

-

ft) -

) ~ ° CO

i

f

>

)

0 20 40 60 80 100 120 140 160 180 200 

TIME, IN MINUTES

EXPLANATION

O FLOW IN CASINGABOVE BOTH FRACTURES

 FLOW IN BOREHOLEBETWEEN THE PAIR OF FRACTURES

Figure 4. Examples of transient-flow response when the 

pumped well intersects a single major fracture zone, and the 

observation borehole connects the primary fracture zone with 

a deeper fracture zone: (a) example where the secondary 

zone is hydraulically isolated from the primary fracture zone; 

(B) exampoe where the secondary zone is hydraulically 

connected to the primary fracture zone.

preliminary evaluation of fracture connectivity within a 

few days after boreholes have been drilled. The speed 

with which these measurements can be made indicates 

that the results will be useful in the subsequent design of 

conventional but much more time-consuming hydraulic 

tests such as those described by Hsieh and others 

(1987), and Shapiro and Hsieh (1991).

SUMMARY

A systematic approach to fracture characterization using borehole geophysics has been developed over nearly a decade of geophysical logging in 

open bedrock boreholes at the Mirror Lake study 

site. One of the most important applications of 

these logging techniques is conducting a preliminary 

investigation of fracture hydrology so that fracture 

zones can be isolated with packers soon after boreholes are drilled. The information obtained during 

this preliminary investigation is also helpful in planning efficient and effective aquifer tests and tracer 

studies.

The fracture-characterization approach 

developed during the Mirror Lake study consists of 

four steps: (I) using conventional geophysical measurements such as caliper, temperature, gamma, neutron, resistivity, and acoustic logs, to characterize 

the general structure and lithology of the rock mass 

along the well bore; (2) using borehole image logs, 

such as the televiewer to indicate the characteristics 

of individual fractures intersecting the borehole; 

(3) using high-resolution flow measurements to 

identify fractures that conduct flow during aquifer 

tests; and (4) using data obtained during the transient 

evolution of the flow field during aquifer tests to 

infer the boreholes. These techniques have proven 

effective in quickly and efficiently identifying 

depths where a few packers can be used to isolate 

the major flow zones within the aquifer. The transient-flow analysis can be used to infer whether fractures conducting flow into or out of observation 

boreholes during aquifer tests are completely isolated from each other in the surrounding rock, or 

whether there is a hydraulically conductive fracture 

connection between these two fractures. Although 

this suite of geophysical measurements will probably never replace conventional aquifer tests, the 

application of these techniques in the Mirror Lake 

study indicate that an effective borehole logging 

program began immediately after boreholes are 

drilled can greatly improve the effectiveness of 

subsequent hydraulic tests and tracer studies.

REFERENCES

Hess, A.E., 1986, Identifying hydraulically conductive fractures with low-velocity borehole flowmeter: Canadian 

Geotechnical Journal, v. 23, no. 1, p. 69-78.

Hess, A.E. and Paillet, F.L., 1990, Applications of the

thermal-pulse flowmeter in the hydraulic characterization of fractured rocks, in American Society for Testing and Materials Special Technical Publication 1101, 

p. 99-112.

108
Hsieh, P.A., Neuman, S.P., Stiles, G.K., and Simpson, E.S., 

1987, Field determination of the three-dimensional 

hydraulic conductivity tensor of anisotropic 

media 2, Methodology and application to fractured rocks: Water Resources Research, v. 21, 

no. 11, p. 1667-1676.

Keys, W.S., 1979, Borehole geophysics in igneous and 

metamorphic rocks, in Society of Professional Well 

Log Analysts 20th Annual Logging Symposium, 

Tulsa, Okla., 1979, Transactions: Houston, Society of 

Professional Well Log Analyst, p. OO1-OO26.

Long, J.C.S., Remer, J.S., Wilson, C.R., and Witherspoon, 

P.A., 1982, Porous media equivalents for networks for 

discontinuous fractures: Water Resources Research, 

v. T8, p. 645-658.

Paillet, F.L., 199la, The 1st core hole at Mirror Lake. New 

Hampshire comparing geophysical logs to core and 

cross-hole flow logging, in Mallard, G.E. and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey Calif, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 162-171.

Paillet, F.L., 1991b, Use of geophysical well logs in evaluating crystalline rocks for siting of radioactive-waste 

repositories: The Log Analyst, v. 32, no. 2, p. 85-107.

Paillet, F.L., 1993a, Application of borehole geophysics in 

the characterization of flow in fractured rocks, in 

Trends in Geophysical Research: p. 127-153.

Paillet, F.L., 1993b, Using borehole geophysics and crossborehole flow testing to define hydraulic connections 

between fracture zones in bedrock aquifers: Journal of 

Applied Geophysics: v. 30, p. 261-279.

Paillet, F.L., Hess, A.E., Cheng, C.H. and Hardin, E.L., 

1987, Characterization of fracture permeability with 

high-resolution vertical flow measurements during 

borehole pumping: Ground Water, v. 25, no. 1, p. 28- 

40.

Paillet, F.L., and Kapucu, K., 1989, Characterization of 

fracture permeability and fracture flow modeling at 

Mirror Lake, New Hampshire: U.S. Geological 

Survey Water-Resources Investigations Report 

89-4058, 49 p.

Paillet, F.L., Novakowski, Kent, and Lapcevic, Pat, 1992, 

Analysis of transient flows in boreholes during pumping in fractured formations, in Society of Professional 

Well Log Analysts 33rd Annual Logging Symposium 

Transactions, Oklahoma City, Okla., 1992, Transactions: Houston, Society of Professional Well Log 

Analysts, p. SI-21.

Shapiro, A.M. and Hsieh, P.A., 1991, Research in fracturcdrock hydrogcology: Characterizing fluid movement 

and chemical transport in fractured rock at the Mirror 

Lake drainage basin. New Hampshire, in Mallard, G.E. 

and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings 

of the technical meeting, Monterey Calif, March 11- 

15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91 -4034, p. 155-161.

Zemanek, Joseph, Glenn, E.E., Norton, L.J., and Caldwcll, 

R.L., 1970, Formation evaluation by inspection with 

the borehole televiewer: Geophysics, v. 35, p. 254- 

269.

109
110
Velocity, Attenuation, and Dispersion Electromagnetic 

Tomography in Fractured Rock

By David L. Weight1 , Gary R. Olhoeft 1 , andThomas P. Grover1

Abstract

Hole-to-hole electromagnetic tomography 

data have been acquired from 15 pairs of wells at 

the Mirror Lake site, Grafton County, New Hampshire, using a high-speed borehole radar system 

designed and built by the U.S. Geological Survey. 

Although tomography has good theoretical and 

mathematical underpinnings, higher resolution, 

better data-acquisition methods, better tomographic processing and better visualization of the 

processed data than are currently available are 

needed if tomography is to achieve full potential 

as a geophysical tool for hydrologists. A 3-dimensional display of two velocity tomograms from 

data collected at the FSE well field at the Mirror 

Lake site illustrates the potential for mapping 

fracture zones in three dimensions. In addition, 

we show examples of velocity, attenuation, and 

dispersion tomograms made from a single data set 

in the well pair FSE4-FSE1. Comparisons 

between these three tomograms and with projected fractures from acoustic televiewer logs 

indicate that velocity, attenuation, and dispersion 

tomograms, though having some correlations with 

each other and with fractures, differ in shape. 

Because tomograms respond to lithology as well 

as to fractures, it might be necessary to use more 

than one type of tomogram to help discriminate 

between the response to fractures and the 

response to lithology in tomograms, unless the 

lithology is uniform.

'U.S. Geological Survey. Denver, Colo.

INTRODUCTION

A radar/tomography system incorporating a 

high-speed digital-data-acquisition system has been 

designed and built by the U.S. Geological Survey 

(USGS) (Wright and others, 1989; Wright and others, 

1990). This system has provided the digitally recorded 

data from which hole-to-hole tomograms for fracture 

mapping at the Mirror Lake site, Grafton County, New 

Hampshire (fig. 1) have been produced. The geohydrologic setting of this area is given in Winter (1984). 

The rock from wells FSE4 and FSE1 is largely granite, 

pegmatite, and schist (C.D. Johnson, U.S. Geological 

Survey unpublished data in Bow, N.H. office). The 

tomograms were produced with software written at the 

USGS (Olhoeft, 1988). The objective of this paper is 

to illustrate similarities and differences between velocity, attenuation, and dispersion tomograms and point 

out the need to examine all three types of tomograms 

for correlation with fracture and lithologic information 

to help hydrologists in their research, the objective of 

which is to develop improved 3-dimensional (3-D) 

hydraulic flow models for fractured rock.

EQUIPMENT AND METHODS

The data from which we are producing electromagnetic tomograms were obtained with a hole-tohole short-pulse system. A short-pulse transmitter, for 

which the peak in the amplitude spectrum is at 60 

megahertz (MHz), radiates energy through the volume 

of rock to be studied. In the rock at Mirror Lake the 

wavelength at 60 MHz is about 2 in. The waveform is 

retarded in time, in response to increases in dielectric 

permittivity; attenuated in amplitude, in response to 

increases in electrical conductivity; and broadened in 

time in response to frequency dependent processes 

such as scattering. We record data while the transmitter

111
EXPLANATION

250 Topographic contour 

10m interval 

Datum is sea level

FSE 10

Well and identifier

43° 56'35'

43° 56'30'

25i

50

i

100 METERS

100 200 300 FEET

Figure 1. Location of the FSE well field at the Mirror Lake site, Grafton County, New Hampshire.

112
and receiver are moving, typically at 12 cm/s 

(25 ft/min). At this speed, we average 512 waveforms 

and record at 17-cm intervals. The transmitter and 

receiver move in synchronism in their respective boreholes at offsets of 0, +5 m, -5 m, +10 m, and -10m. 

Other offsets or methods of data acquisition are possible with the system. The full waveform is displayed on 

a color monitor in real time and recorded on a highcapacity magneto-optical disk drive.

DATA PROCESSING

After the data are acquired, the waveforms are 

picked automatically for first arrival time, amplitude, 

and first-cycle width. The picked data must be supplemented by accurate survey and borehole deviation 

data to produce quality tomograms. To determine the 

quality and consistency of the entire data set, the first 

arrival times are used with the measured surface positions and borehole depth/deviation data to perform a 

consistency test (Olhoeft, 1993). The raypath distance 

divided by the first arrival time is the average velocity 

of electromagnetic propagation between the boreholes. 

The square of the ratio of the speed of light in vacuum 

to this velocity is the relative dielectric permittivity. 

For each offset between transmitter and receiver, the 

apparent relative permittivity is computed using first 

arrival time and raypath distance for the perturbed 

receiver-tool position. A set of curves for five offsets 

is generated. These curves are the average for the 

entire hole-pair data set. If the data set is consistent 

and there are errors neither in instrument timing nor in 

knowledge of the tool positions, the five curves should 

intersect where the probe deviation equals zero and at 

the average relative permittivity of the medium. In 

practice, some error occurs, so there is a spread in the 

intersections of the five curves. The spread is an indication of the quality of the overall data set (fig. 2). 

For the FSE4-FSE1 data set the spread is small (all 

curves cross inside one unit square) indicating that the 

data are consistent, but there is a cumulative positioning indicated error of about 2 m. The cumulative positioning error is the sum of all surveying errors, 

borehole-deviation measurement errors, and toolpositioning errors. The picked data are then processed using a filtered back projection algorithm 

(Olhoeft, 1988) to produce the following tomograms. 

The tomograms can be displayed on a workstation or 

personal computer in 3-D, and rotated, tilted, and 

zoomed in real time as an interpretive aid.

TOMOGRAMS AND FRACTURES

As noted previously, a radiated electromagnetic 

pulse is modified by the electrical properties of rock 

and water as the pulse propagates through fractured 

and/or inhomogeneous rock. Three processes are 

easily distinguished and can be measured from 

recorded data. The first process is first-arrival-time 

retardation. First arrival time is proportional to the 

distance between the transmitter and receiver and 

inversely proportional to average velocity. Velocity is 

inversely proportional to the square root of the relative 

dielectric permittivity, defined above. Because unfractured crystalline rocks, such as granites, commonly 

have relative dielectric permittivites in the range of 

perhaps 5 to 7, whereas water has a relative dielectric 

permittivity of about 81, it is expected that fracture 

zones containing significant amounts of water will 

manifest lower electromagnetic wave velocities than 

adjacent unfractured rock. Therefore, a velocity tomogram might be expected to image the location and the 

relative amount of water in fracture zones. The greater 

the proportion of water in a given volume of rock, the 

slower the velocity. Thus, in a velocity tomogram, we 

expect that local velocity minima might indicate the 

presence of water-filled fractures. Velocity tomograms 

are presented in figures 3 and 4a and discussed later in 

this paper.

A second process is attenuation. As a pulse 

propagates through rock, it is diminished in amplitude 

both by geometrical "spreading loss" and by ohmic 

losses in the rock that convert electromagnetic energy 

to heat. Electrical resistivity, or its reciprocal, conductivity, controls ohmic loss. The higher the resistivity, 

the lower the loss. Unfractured crystalline rock can 

have resistivities of thousands of ohm-meters, whereas 

water, unless it is very fresh, will commonly be in the 

tens to low hundreds of ohm-meters. Thus, it might be 

expected that an attenuation tomogram would image 

the presence of fractures by areas of high attenuation.- 

The terms "high" and "low" here are relative to the 

particular rock being imaged. The calculated numerical values that produce a tomogram are not as important as the variations of those values that produce the 

visual image.

The third process is dispersion, manifested by 

broadening of the radiated pulse shape as the pulse 

propagates through a dispersive medium. It is common 

to define a dispersive medium to be one in which the 

phase velocity is a function of frequency. In practice, 

however, any process that removes energy from a

113
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CUMULATIVE TOOL POSITION ERRORS, IN METERS

Figure 2. Relation between relative dielectric permittivity and cumulative transmitter-receiver tool position errors for transmitter-receiver offsets of 0, +5 m, -5 m, +10 m, -10 m for well pair FSE4-FSE1. The spread in the intersections of these five 

curves is a measure of the consistency of the data set. The average relative dielectric permittivity is 6.55 with error bars of +/- 

0.24. The average cumulative tool position error is -2.04 m.

propagating wave in a frequency-dependent manner 

can produce pulse broadening. Possible examples, in 

addition to frequency-dependent phase velocity, 

include frequency-dependent attenuation, scattering 

and diffraction from embedded inhomogeneities, and 

frequency-dependent reflection and transmission as a 

wave traverses a planar fracture geometry. The firstcycle width of the received pulse is used here as a 

measure of dispersion. Other parameters may also be 

used to characterize dispersion. Dispersion generated 

by any given fracture or fracture zone depends on 

many factors including fracture size and geometry, 

wavelength, and angle of incidence. Forward and 

inverse modeling of fracture-generated dispersion is 

difficult and there is much room for study in this area.

In this paper, we merely assert, based on experimental 

observation, that dispersion, as well as velocity and 

attenuation, can usefully correlate with fractures.

A complicating factor for the interpretation of 

tomograms is that rocks themselves are rarely spatially 

homogeneous even within a single rock type such as 

granite, and if the rock encountered by a pair of wells is 

not all of the same type, as it is not in wells FSE1 and 

FSE4, the resulting heterogeneity in electrical properties that will be imaged by a tomogram may be indicative of lithology, fracturing, or both. An additional 

complication is that the presence of water-filled fractures does not always translate into high transmissivity, 

as has been pointed out by Paillet (Paillet and others, 

1987; Paillet and Kapucu, 1989; and Paillet, 1991), who

114
EXPLANATION

LOCATION OF ZONES 

WITH FRACTURES OF 

HIGH HYDRAULIC 

TRANSMISSIVITY

100

Figure 3. A pair of velocity tomograms in well pairs FSE1-FSE2 and FSE2-FSE4. The arrows indicate a zone that is known 

to have fractures with high hydraulic transmissivity. The large cube is 100 meters on a side and the grid is at 10 meter 

intervals. The gray scale goes from dark (low velocity) to light (high velocity).

has noted that large fracture apertures as measured 

with the acoustic televiewer do not always have high 

hydraulic transmissivity. Despite these complications, 

however, tomograms imaging rock electrical properties between wells at Mirror Lake do correlate with 

zones containing fractures of high hydraulic transmissivity.

Figure 3 shows a pair of velocity tomograms 

between well pairs FSE1-FSE2, and FSE2-FSE4 that

successfully image zones of fractures known to have 

high hydraulic transmissivity. These well pairs are at 

right angles to each other, so the view is into the tomograms as if looking into the pages of a half-opened 

book. The large cube is 100 m on a side, and the grid 

intervals are 10 m. The velocity gray-scale runs from 

dark (slow) to light (fast). The terms "slow" and "fast" 

are relative to the global average velocity in the tomogram. The actual values are not very important to the

115
EXPLANATION 

Possibly permeable 

Probably permeable 

Water Producing

HIGH

LOW

i

UJ

m

CD

1

LL 

O

a.

50

60

70

80

90

40 -i

50

60 -

70

80

90

30

40

50

70

E4 (a) FSE1 FSE4~ (b) FSE1 FSE4(c) FSE1

Figure 4. Three tomograms produced from the same data set between wells FSE4 (left) and FSE1 (right): (a) a velocity 

tomogram, (b) an attenuation tomogram, and (c) a dispersion tomogram. For this case it appears that (a) correlates best with 

the fractures that have high hydraulic transmissivity (marked with arrows). In each tomogram the gray scale is from dark (low) 

to light (high).

116
question of whether the tomograms visually image 

fracture zones. The arrows indicate the location of a 

zone that contains hydraulically permeable fractures 

(Paillet, 1991).

Figure 4 shows three tomograms produced from 

the same data set taken between wells FSE4 (left side 

of each panel) and FSE1 (right side). Figure 4a is a 

velocity tomogram formed from first arrival time 

picks, figure 4b is an attenuation tomogram formed 

from amplitude measurements, and figure 4c is a dispersion tomogram formed from pulse-width measurements. Superimposed on the tomograms are 

projections of fractures determined from acoustic televiewer logs of holes FSE4 and FSE1 (Paillet, 1991; 

Wrightand others, 1996). The fractures seen in well 

FSE1 were projected in the plane defined by the FSE4- 

1 pair halfway to well FSE4 and the fractures in well 

FSE4 were projected halfway to well FSE1. The overlays are inexact because the tomograms include the 

effects of well deviation, and are therefore wider at the 

bottom that at the top, whereas the fracture projections 

neglect well deviation. Approximately 0.5 meter zones 

along each well have been excluded from each tomogram because calculated values immediately adjacent 

to the wells are not considered reliable. The gray scale 

runs from low to high in each tomogram. The actual 

calculated values for velocity, attenuation, or dispersion are not particularly important for the purpose of 

identifying visual correlations with fractures and thus 

have been omitted. According to Paillet, the projection 

angles are subject to errors of a few degrees, and it 

appears that relatively few fractures project very far in 

the FSE well field. Thus, the indicated projections are 

not definitive, but comparisons of projected fractures 

to the tomograms is of considerable interest. When 

well FSE4 is pumped and a flowmeter is used to find 

fractures that carry significant flow, it is determined 

that there are two fractures in well FSE1 and two in 

well FSE4 that have much higher hydraulic transmissivity than any other fractures. The projections of 

these fractures are denoted by the heaviest lines. The 

two fractures intersecting well FSE4 that have the 

highest transmissivity are in the 39 to 47-m deep zone. 

Of these, 80 percent of the water flows through the 

upper fracture. Consider the zone from 39 to 47 m 

deep in the tomograms. In this zone, the velocity 

tomogram (fig. 4a) seems to be responding to the presence of the fractures. Presumably there is more water 

in this zone than in adjacent zones. The attenuation 

tomogram (fig. 4b) shows an increase in attenuation in

this zone compared to the attenuation in the area 

below, but the local attenuation maximum is several 

meters above this zone. The dispersion tomogram has 

only subtle features in the region of the high-transmissivity fracture projections, but strong sharp features in 

other regions may correlate with other fractures.

CONCLUSIONS

Electromagnetic tomograms map changes in 

electrical properties. The presence of water-filled fractures will alter the local electrical properties, but so 

will lithologic differences. The tomograms presented 

here show correlation with some projected fractures 

but not others. In addition, the presence of water-filled 

fractures does not, in itself, produce high hydraulic 

transmissivity. Consequently, we believe that crossborehole tomography will be a valuable tool for characterizing and mapping fractures, but must be used in 

conjunction with other geophysical measurements to 

achieve its full potential. We also conclude that more 

than one type of tomographic processing might be 

needed if we are to improve the ability to separate the 

effects of lithology from those caused by the presence 

of water-filled fractures.

ACKNOWLEDGMENT

We acknowledge F.L. Paillet for providing fracture projections from wells FSE1 and FSE4 and for 

providing digital data for fractures mapped with the 

acoustic televiewer in all of the FSE wells.

REFERENCES

Olhocft, G.R., 19X8, Interpretation of holc-to-holc radar 

measurements, in Third Technical Symposium on 

Tunnel Detection Proceedings, Jan., 12-15, 19X8: 

Golden, Colorado School of Mines, p. 616-629.

Olhocft, G.R., 1993, Velocity, attenuation, dispersion and 

diffraction holc-to-holc radar processing, in Fourth 

Tunnel Detection Symposium on Subsurface Exploration Technology Proceedings, April 26-29, 1993: 

Golden, Colorado School of Mines, p. 307-320.

Paillet, F.L., Hess, A.E., Chcng, C.H., and Hardin, E.L., 

1987, Characterization of fracture permeability with 

high-resolution vertical flow measurements during 

borehole pumping: Ground Water, v. 25, no. 1, p. 28- 

40.

Paillet, F.L., and Kapucu, Kernel, 1989, Characterization of 

fracture permeability and fracture flow modeling at

117
Mirror Lake, New Hampshire: U.S. Geological Survey 

Water-Resources Investigations Report, 89-4058, 49 p.

Paillct, F.L., 1991, Comparing geophysical logs to core 

and cross-hole flow logging in the Mirror Lake drainage basin, in Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program-- Proceedings of the technical meeting, 

Montcrcy, Calif., March 11-15, 1991: U.S. Geological Survey Water-Resources Investigations 

Report 91^034, p. 162-171.

Winter, T.C., 1984, Gcohydrologic setting of Mirror Lake, 

West Thornton, New Hampshire: U.S. Geological 

Survey Water Resources Investigation Report 84-4266, 

61 p.

Wright, D.L., Bradley, J.A., and Hodge, S.M., 1989, Use of 

a new high-speed digital data acquisition system in airborne ice-sounding: Institute of Electrical and Electronics Engineers Transactions on Gcoscicncc and 

Remote Sensing, v. GE-27, p. 561-567.

Wright. D.L., Hodge, S.M., Bradley, J.A., Grovcr, T.P., and 

Jacobcl, R.W., 1990, A digital low-frequency, surfaceprofiling ice radar system: Journal of Glaciology v. 36, 

p. II2-I2I.

Wright, D.L., Olhocft, G.R.. Hsich, P.A., Majcr, E.L., Paillct. F.L., and Lane, J.W. Jr., 1996, Electromagnetic and 

seismic tomography compared to borehole acoustic 

televiewer and flowmctcr logs for subsurface fracture 

mapping at the Mirror Lake site. New Hampshire, in 

Morganwalp, D.W., and Aronson, D.A., cds., 1995, 

U.S. Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the Technical Meeting, Colorado, September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

118
Electromagnetic and Seismic Tomography Compared to 

Borehole Acoustic Televiewer and Flowmeter Logs for 

Subsurface Fracture Mapping at the Mirror Lake Site,

New Hampshire

By David L. Wright1 , Gary R. Olhoeft1 , Paul A. Hsieh2, Ernest L Majer3 , 

Frederick L Paillet1 , ancfJohn W. Lane Jr.4

Abstract

Among the techniques used in research at 

the Mirror Lake site, Grafton County, New 

Hampshire, are electromagnetic (EM) and seismic 

tomography, borehole radar, borehole acoustic 

televiewer, and borehole flowmeter. Of these 

techniques, tomography and radar can probe 

several tens of meters between or around 

boreholes at that site with resolutions in the order 

of 1 meter, whereas televiewer data provides great 

detail at the borehole wall but little penetration. 

Flowmeter data, along with hydraulic tests and 

tracer tests provide information on hydraulic 

connectivity, but hydraulic paths between the 

wells can not be inferred from these data alone. 

We find from side-by-side comparison of 

electromagnetic and seismic tomograms of rock 

properties between wells FSE1 and FSE4 that 

both types of tomograms show the presence of 

fractures of high hydraulic transmissivity. We 

present velocity tomograms for two EM systems 

and one seismic system, attenuation tomograms 

for the two EM systems, and projections of 

fractures derived from acoustic televiewer logs. 

The pulsed transmitters used in the EM 

tomography at Mirror Lake produced 

wavelengths in the granite of about 2 m. The

U.S. Geological Survey. Denver, Colo. 

2U.S. Geological Survey, Menlo Park, Calif.

Lawrence Berkeley National Laboratory, Berkeley, Calif. 

4,U.S. Geological Survey, Hartford, Conn.

seismic system produced wavelengths of about 

1 m. Seismic and EM tomography can detect the 

presence of fractures whose aperture is much 

smaller than a wavelength but can not resolve 

fractures whose spacing is much less than the 

wavelength. Resolution the ability to 

distinguish two nearby objects from one 

another is determined by a number of factors, 

including spatial data density, but resolution 

usually can not be better than about half the 

wavelength used for probing regardless of 

whether the method is EM or seismic. It is 

possible to achieve some resolution improvement 

in low-attenuation environments, but a needed 

step to achieve maximum benefit from 

tomography is correlation of the tomograms with 

other hydrologic and geophysical information. 

Tomography is an art that is not fully mature but 

can be expected to improve in the future.

INTRODUCTION

Data have been recorded between 15 pairs of 

wells in the FSE field (fig. 1) with the 

radar/tomography borehole system developed by the 

U.S. Geological Survey (USGS) (Wright and others, 

this proceedings), and 18 pairs have been examined 

with a hole-to-hole seismic tomography system 

developed by Lawrence Berkeley National 

Laboratory. In addition, one pair of wells was studied 

tomographically using equipment leased from ABEM, 

a Swedish company. Because it was of some interest to 

compare results from three systems, all three were run

119
FSE 10

EXPLANATION

 250 Topographic contour 

10 m interval 

Datum is sea level 

well and Identifier

43° 56'35'

43° 56'30' \

25 50

i

100 METERS

100 200 300 FEET

Figure 1. Location of tne FSE well field at the Mirror Lake site in central New Hampshire. The geohydrologic setting of this 

area is given in Winter, 1984.

in the well pair FSE4-FSE1. This afforded a rare 

opportunity to see results in the same pair of wells that 

were obtained by different systems and processed by 

three independent groups by means of three different 

sets of software. Results of the borehole acoustic televiewer, flowmeter, and packer tests, which indicated 

the presence of the high-transmissivity fractures, were 

compared to the tomograms. The tomographic studies

at Mirror Lake are intended to aid hydrologists in the 

development of adequate 3-dimensional fracture-flow 

models, suggest other hydrologic applications of 

tomography, indicate limitations for applications of 

tomography in its present state in fractured crystalline 

rock, and identify directions in which the state-of-theart could be enhanced for toxic-substances hydrologic 

studies.

120
well 1

TRANSM I TIER

STATION 

LOCAT I ONS

Well 2

RECE I VER

STATI ON

LOCATIONS

Figure 2. Generic tomography schematic for electromagnetic or seismic tomography. The number, spacing, and relative 

locations of the transmitter and receiver stations may vary.

BASICS OF TOMOGRAPHY

Tomography between boreholes is conceptually 

and mathematically similar to what is done when a 

medical computer-aided-tomography (CAT) scan is 

made, but the distances covered in borehole 

applications are much greater, the angular coverage is 

incomplete, and the resolution is much lower. The 

method of collecting data from boreholes is to record a 

number of measurements with a transmitter and 

receiver or string of receivers at various locations so 

that a number of raypaths pass through any particular 

volume of rock (fig. 2). For any given raypath, an 

average number for velocity, amplitude, or some other 

quantity is measured. If a water-filled fracture zone is 

present, it would be expected, for example, that the 

velocity would be lower for both electromagnetic 

(EM) and seismic waves than it would be in 

unfractured rock. Mathematical algorithms are applied 

to calculate an array of values for the mapped quantity 

at particular locations in the rock. The region between 

two boreholes is gridded, and calculated values are 

assigned to each area element, commonly referred to 

as a pixel. Frequently a smoothing algorithm is 

applied to visually enhance the display. The display

can be presented by using a color scale, a gray scale as 

in figures 3 and 4, contouring, or some combination of 

methods. Many algorithms are used for producing 

tomograms. Each has strengths and weaknesses. Much 

of the literature on tomography is dominated or 

motivated by medical applications (Herman, 1980). 

Applications of tomography to geophysics has grown 

rapidly, however, and there is now a substantial and 

growing literature applicable to geophysics (Dines and 

Lytle, 1979; Natterer, 1986; Kak and Slaney, 1988). 

There is still much art involved in producing 

tomograms. A "good" tomogram will be well 

constrained by the data from which it is derived and 

will visually reveal features of interest to the 

hydrologist, geologist, or geophysicist. Things that 

might be of great interest to the geologist, such as 

lithology, may be of less interest to the hydrologist 

who wants to know where the fractures are between 

wells and which fractures are hydraulically connected, 

and to quantify the hydraulic conductivity. A good 

tomogram from the hard-rock fracture-flow 

hydrologist's point of view will need to help answer 

those questions.

121
VELOCITY, IN METERS 

PER MICROSECOND

VELOCITY, IN 

KILOMETERS 

PER SECOND

100 FSE1

FSE1

EXPLANATION

The

location of fractures 

that have high 

transmissivity

FSE4 (a) FSE1

5.5

5.4

117 118 119 120 121 122 123 124

HORIZONTAL DISTANCE FROM FSE4, 

IN METERS 0 5 10 15

Figure 3. Electromagnetic velocity tomogram produced with the ABEM system (a), electromagnetic velocity tomogram 

produced with USGS system (b), seismic velocity tomogram produced with the Lawrence Berkeley National Laboratory 

system (c).

122
HIGH

LOW

ATTENUATION, 

IN DECIBELS PER 

KILOMETER 

(UNCAUBRATED)

HORIZONTAL DISTANCE FROM FSE4, 

IN METERS

0 5 10 15

HIGH

EXPLANATION 

Possibly permeable 

Probably permeable 

Water Producing

LOW

40 -

100J

(b) FSE1 100

FSE1 FSE4 FSE1

Figure 4. Electromagnetic attenuation tomogram produced with the ABEM system (a), electromagnetic attenuation 

tomogram produced with the U.S. Geological Survey system (b), fracture projections produced from the acoustic televiewer 

logs (c). The fractures with high hydraulic transmissivity are indicated with heavy lines.

123
EQUIPMENT AND MEASUREMENT 

METHODS

The ABEM radar system transmitter used in this 

study radiates a short pulse whose amplitude spectrum 

peaks at about 60 MHz. The transmitter was in well 

FSE1 and the receiver was in well FSE4. Data were 

recorded with the transmitter and receiver stationary at 

3-m intervals.

In the case of the USGS radar, the 60-MHz 

transmitter (not identical to the ABEM transmitter) 

was in well FSE4 and the receiver was in well FSE1. 

Theoretically, the same system should produce 

identical tomograms, no matter which hole the 

transmitter was in, provided all other factors are equal. 

The method of data recording and processing did 

differ from those employed using the Swedish system, 

however. With the USGS system, the boreholes are 

logged with the probes moving at about 12 cm/s 

(25 ft/min) at constant depth offsets from one another. 

An exceptionally high-speed data acquisition system 

designed and built by the USGS (Wright and others, 

1989) allowed averaging 512 waveforms for signal-tonoise improvement while recording the averaged data 

every 17 cm. Because half the data were used in 

producing the tomograms, the effective data interval 

was about 35 cm. Five different offsets were used. The 

maximum offset used was +/-10 m, so the steepest 

angle was about +/-380 .

The seismic equipment was developed and 

fabricated by scientists at the Lawrence Berkeley 

National Laboratory (LBL). The equipment consists of 

a seismic source, a receiver string, and a datarecording system. For the 1991 survey from which the 

displayed tomogram is derived, the seismic source 

consisted of a high-voltage, high-fidelity amplifier 

which delivers electrical pulses of 5 kV amplitude and 

20 ms width through a coaxial cable to a downhole 

piezoelectric source. In this pulse mode, the generated 

seismic signal has a dominant frequency content in the 

5 to 8 kHz range. For a seismic velocity of 5,000 m/s, 

the wavelength is approximately 1 m. Both the source 

and receiver stations were at 1 -m intervals. The datarecording system consists of a computer with 16-bit 

analog-to-digital converter.

All three systems record the full waveform, so 

that any refinements in picking and processing could 

be applied to the original data in the future. All three 

systems also provide real-time computer-screen 

displays of the recorded waveforms so that the system 

operator can instantly detect gross system malfunction 

and have some assurance of the integrity of the 

recorded data without requiring playback or analysis.

One experimental factor should be mentioned that has 

been a problem for the radar systems, however. If the 

time base drifts, it will induce errors in the data. It is 

known that both the USGS and the ABEM systems 

sometimes have exhibited drift. For the 1992 field 

season, a new experimental procedure was used with 

the USGS system to allow detection, measurement, 

and correction for time-base drift, should it occur. The 

USGS data processing also includes a unique 

consistency test of the data quality (Olhoeft, 1988).

TOMOGRAMS AND FRACTURE 

PROJECTIONS AT THE FSE4-FSE1 WELL 

PAIR

Figure 3 is a side-by-side display of three 

velocity tomograms between well FSE4 (left side) and 

well FSEl (right side). The gray scale renditions 

printed here are inferior in detail; color displays are 

commonly superior. Figure 3a shows a velocity 

tomogram produced from data acquired using the 

ABEM system. The tomogram was produced using 

proprietary ABEM software. The algorithm used by 

ABEM is of the conjugate gradient type. All of the 

possible raypaths were used in producing the ABEM 

tomogram. Because some of the rays that are produced 

by connecting every measurement station in well 

FSE4 to every station in well FSEl are nearly vertical, 

the angular coverage is very good for this tomogram. 

Figure 3b shows a velocity tomogram produced from 

data recorded with the USGS system and processed 

with software written by the USGS (Olhoeft, 1988). 

The tomogram was produced using a variational 

filtered back-projection algorithm. Figure 3c was 

produced from data acquired with the LBL borehole 

seismic system. The manually picked first arrival time 

data were processed using an algebraic reconstruction 

technique (Peterson and others, 1985).

An attenuation tomogram produced with the 

ABEM system and software is shown in figure 4a. An 

attenuation tomogram produced with the USGS 

system and software is shown in figure 4b. The 

attenuation scales in figures 4a and 4b are not 

calibrated, but the relative values are sufficient for 

purposes of visual comparison. Figure 4c is produced 

by projecting fractures recorded with the acoustic 

televiewer from wells FSE4 and FSEl (Paillet and 

others, 1987, Paillet and Kepucu, 1989, Paillet, 1991, 

Paillet, 1996). The fractures recorded in well FSE4 are 

projected in the plane of the wells from FSE4 halfway 

to FSEl and similarly for the fractures recorded in 

well FSEl. The projections are subject to errors of a 

few degrees, and the projections and other evidence

124
indicate that few, if any, fractures project in a straight 

line from one well to the other and that the flow 

between the wells does not follow a single welldefined path. There are probably hydraulic 

connections between fractures; therefore, if one 

overlays the fracture projections on the tomograms 

with the idea of finding correlations with the 

tomograms, one should bear in mind that the evidence 

indicates that relatively few fractures can be projected 

very far in the FSE well field.

COMPARISONS OF THE TOMOGRAMS 

AND FRACTURE PROJECTIONS

All the tomograms begin at about 20 m depth, 

slightly below the bottom of the steel casing in both 

wells. When well FSE4 is pumped and the hydraulic 

transmissivities of various fractures are measured by 

the flowmeter, a small number of fractures dominate 

hydraulic flow. The high-transmissivity fractures are 

shown in figure 4c as heavy lines. The arrows along 

the sides of each tomogram indicate the location of 

those high-transmissivity fractures except for the one 

at a depth of about 18 m in well FSE 1. The slope of the 

arrows roughly indicates the dip of the waterproducing fractures recorded with the acoustic 

televiewer and confirmed to be a water-producer by 

the flowmeter. Most (80 percent) of the water that 

enters well FSE4 when it is pumped flows through the 

fracture indicated by the upper arrow.

At well FSE4, the velocity lows in figures 3a 

and 3b seem to reflect a dense nest of fractures just 

above the upper of the two arrows (see fig. 4c). The 

seismic tomogram of figure 3c also show a response. 

At well FSE1, figure 3a shows a velocity low about 2 

m above the indicated fracture location. Figure 3b 

shows a larger zone of relatively low velocity. In none 

of the tomograms can the flow paths between wells be 

defined with precision, though hints may be seen. It is 

possible that substantial flow even leaves the plane of 

the wells, a situation that cannot be determined by 2- 

dimensional tomography. Directional borehole radar, a 

capability of the ABEM system, might perhaps reveal 

whether such flow occurs. It is evident that the 

tomogram in figure 3a has been smoothed to a greater 

degree than the tomogram of Figure 3b. Some detail 

might have been lost in the process, although some 

might argue that the USGS data have not been 

smoothed enough. Near-horizontal features appear to 

be more sharply defined in figure 3b as a result of the 

eight times greater spatial density of data. On the other

hand, figure 3a indicates some steep contours in the 

50- to 70-m depth range, and the 80- to 100-m range. 

Such features are not evident in figure 3b. The limited 

angular coverage of the USGS data might be expected 

to produce tomograms that would distort or miss just 

such long, near-vertical features. A conclusion is that 

both high spatial data density and wide angular 

coverage are desirable. There are some differences 

among figures 3a, 3b, and 3c in the hydraulically 

important 39- to 47-m deep zone. Two minima in the 

velocity in figure 3a occur adjacent to the wells. In 

figure 3b, on the other hand, a minimum velocity is 

attained near the middle of the zone. Figure 3c shows 

velocity lows both along the boreholes and in the 

interior. If figures 3b and 3c are accurate in this 

respect, they could indicate probable flow paths. 

Finally, it should be noted that the appearance of 

visual shapes can sometimes be considerably altered 

by rescaling the color or gray-scale representation of 

the data without altering the underlying data in any 

way.

The gross structure of figures 4a and 4b is 

similar. Both show a closed area of relatively low 

attenuation at about the 30-m depth, rapidly grading to 

relatively high attenuation higher in the section. The 

USGS plot does not extend to a depth of 20 m, but the 

ABEM tomogram indicates relatively high attenuation 

in the upper right-hand corner that correlates with the 

projection of the water-producing fracture intersecting 

well FSE1 at a depth of about 18 m. Both tomograms 

also show a zone of relatively high attenuation at 

about the 40-m depth. The two patterns are remarkably 

similar in shape but appear to be offset from one 

another by about 3 m . This local attenuation high is 

several meters above the projection of the waterproducing fracture in well FSE1. The ABEM 

tomogram has a distinctive boot-shaped area of 

moderate attenuation, with the toe of the boot pointing 

down and to the right and ending at about 70 m depth. 

There is a narrow strip adjacent to well FSE1 from 

about 47 m to the toe of the boot where the attenuation 

is low. In figure 4b, a zone of low attenuation is also 

seen, but the precise shape is somewhat different. 

Another zone of low attenuation is seen in the depth 

interval from about 75 to about 90 m in figure 4b with 

a corresponding area shown in figure 4a. The contours 

in the lower left-hand corner of figure 4a oriented 

down and to the right appear to correlate with fracture 

projections in the corresponding part of figure 4c, 

although, in other areas, correspondence to fracture 

projections is not clear.

125
CONCLUSIONS

The EM and seismic tomograms do correlate, 

though imperfectly, with each other and with some 

known fractures. Interpretations of tomograms for 

fracture mapping is made more difficult because 

tomograms map lithology as well as fractures. 

Detailed lithologic logs of wells FSE1 and FSE4 are 

not yet available, but it is known that at various depths 

in well FSE1, the rock is schist or granite, and, in well 

FSE4, the rock is pegmatite, granite, and schist. 

Granites, schists, and pegmatites can have overlapping 

ranges of electrical and elastic properties that govern 

the propagation of electromagnetic and seismic waves. 

It appears that, of the two EM tomograms, the ABEM 

tomograms are superior for detecting steeply dipping 

features, and the USGS tomograms are superior for 

detecting nearly horizontal features but would benefit 

from steeper angular coverage, which would not be 

difficult to obtain, although it would require some 

changes to the tomography-processing algorithm.

Higher resolution is desirable. In the rock 

observed in wells FSE1 and FSE4, the attenuation is 

low enough that it would be possible to use 

frequencies higher than 60 MHz (perhaps 120 MHz) 

for EM tomography. The trade-off is that, in using 

higher frequencies, the attenuation will increase; thus, 

in many cases, especially in widely spaced wells, there 

will not be sufficient signal to work with. This is true 

for either EM or seismic systems. Other possibilities 

include examination of the data to see whether 

diffraction tomography might be possible (Olhoeft, 

1988) and emphasizing what happens to the highfrequency end of the amplitude spectrum (dispersion 

tomography) in reprocessing existing data (Wright and 

others, this proceedings).

There is much work yet to be done. Tomography 

has tended to be done in isolation with little or no 

reference to other borehole geophysics or lithology 

information. It is time to begin the task of data 

integration in the pursuit of improving our understanding 

of ground-water flow through fractured rock.

REFERENCES

Dines, K.A., and Lytlc, R.J., 1979, Computerized

geophysical tomography: Institute of Electrical and 

Electronics Engineers Proceedings, v. 67, no., 7, 

p. 1065-1073.

Hacni, P.P., 1993, Use of surface geophysics and borehole 

radar methods to detect fractures in crystalline rocks at 

the Mirror Lake area, Grafton County, New 

Hampshire, in Hydrogcology of Hard Rocks: 

International Association of Hydrogcologists 

Congress, 24th, Oslo, 1993 [Proceedings], p. 577-587.

Kak, A.C., and Slancy, Malcolm, 1988, Principles of 

computerized tomographic imaging: New York, The 

IEEE Press, 329 p.

Nattcrcr, Frank, 1986, The mathematics of computerized 

tomography: Chichcstcr, John Wilcy and Sons Ltd., 

222 p.

Olhoeft, G.R., 1988, Interpretation of holc-to-holc radar 

measurements: in Third Technical Symposium on 

Tunnel Detection Proceedings, Jan., 12-15, 1988, 

Colorado School of Mines, Golden, CO, p. 616-629.

Paillct, F.L., Hess, A.E., Chcng, C.H., and Hardin, E.L., 

1987, Characterization of fracture permeability with 

high-resolution vertical flow measurements during 

borehole pumping: Ground Water, v. 25, no. 1, p. 28- 

40.

Paillct, F.L., and Kapucu, Kcmal, 1989, Characterization of 

fracture permeability and fracture flow modeling at 

Mirror Lake, New Hampshire: U.S. Geological Survey 

Water-Resources Investigations Report, 89-4058, 49 p.

Paillet, F.L., 1991, The I st core hole at Mirror Lake, New 

Hampshire comparing geophysical logs to core and 

cross-borehole flow logging: U.S. Geological Survey 

Toxic Substance Hydrology Conference, Montercy, 

Calif., Proceedings, March 11-15, 1991, WaterResources Investigation Report 91 -4034, p. 162-171.

Paillet, F.L., 1996, Using well logs to prepare the way for 

packer strings and tracer tests - lessons from the Mirror 

Lake study in Morganwalp, D.W., and Aronson, D. A., 

cds., 1995, U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Pctcrson, J.E., Paulsson, B.N.P., and McEvilly, TV., 1985, 

Applications of algebraic reconstruction techniques to 

crosshole seismic data: Geophysics, v. 50, no. 10, 

p. 1566-1580.

Winter, T.C., 1984, Gcohydrologic setting of Mirror Lake, 

West Thornton, New Hampshire: U.S. Geological 

Survey Water-Resources Investigation Report 84- 

4266,61 p.

Wright, D.L., Olhoeft, G.R., and Grovcr, T.P., 1996,

Velocity, amplitude, and dispersion EM tomography in 

fractured rock, in Morganwalp, D.W., and Aronson, 

D.A., cds., 1995, U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Wright, D.L., Bradley, J.A., and Hodge, S.M., 1989, Use of 

a new high-speed digital data acquisition system in 

airborne ice-sounding: Institute of Electrical and 

Electronics Engineers Transactions on Geoscience and 

Remote Sensing, v., GE-27, p. 561-567.

126
Hydraulic Characteristics of Fractured Bedrock

Underlying the FSE Well Field at the Mirror Lake Site,

Grafton County, New Hampshire

By Paul A. Hsieh 1 and Alien M. Shapiro2

Abstract

At the Mirror Lake site in Grafton County, 

New Hampshire, 13 wells are drilled in a 120-m 

(meter) by 80-m area, known as the FSE well field, 

the characterize the hydraulic and transport properties of the underlying fractured bedrock. Borehole geophysical logs and downhole video camera 

images show that each well intersects 20 to 60 

fractures within the upper 60 m of bedrock. However, single-well hydraulic tests show that only 

one to three fractures in each well are highly transmissive. Multiple-well hydraulic tests suggest that 

the few highly transmissive fractures connect with 

one another locally to form fracture clusters. 

These highly transmissive fracture clusters, in 

turn, connected to one another by fractures that are 

comparatively lower in transmissivity by several 

orders of magnitude. In such a fractured-bedrock 

setting, the response to a multiple-well hydraulic 

test is entirely different from the response in a 

homogeneous aquifer. These finding s suggest that 

it is necessary to identify and characterize explicitly the highly transmissive fracture clusters in 

order to analyze multiple-well hydraulic tests at 

the FSE well field.

INTRODUCTION

Development and testing of field methods to 

determine the hydraulic and transport properties of 

fractured rocks are major objectives of a field study at 

the U.S. Geological Survey Toxic Substances 

Hydrology Program research site in the Mirror Lake

'U.S. Geological Survey, Menlo Park, Calif. 

2U.S. Geological Survey, Reston, Va.

area, Grafton County, New Hampshire. Within the 

Mirror Lake area, the FSE well field, which occupies a 

120-m 80-m area (fig. 1), serves as an experimental site 

for detailed characterization of fractures in the 

underlying bedrock. This paper presents some 

preliminary results of single-well and multiple-well 

hydraulic tests for fracture characterization.

FSE2 FSE1

FSE7

FSE5

FSE8

FSE10

FSE6 FSE11

FSE13

0 10 20 METERS

Figure 1. Map showing arrangement of wells in the FSE 

well field.

127
BEDROCK AND FRACTURES AT THE FSE 

WELL FIELD

Overlain by approximately 20 m of glacial drift, 

the bedrock at the FSE well field consists of schist that 

is extensively intruded by granite and pegmatite. At a 

nearby road cut on highway 1-93, the exposed bedrock 

shows a complex distribution of rock types (CC. 

Barton, U.S. Geological Survey, written communs., 

1993). The granitic intrusions have the form of dikes, 

irregular pods an anastomosing fingers, ranging from 

centimeters to tens of meters thick. The pegmatite 

intrusions have the form of dikes, centimeters to meters 

thick, cross cutting both the schist and the granite. 

Although basaltic dikes have been observed in the 

Mirror Lake area, they have not been encountered in 

wells drilled at the FSE well field.

Thirteen wells are drilled into bedrock at the 

FSE well field. Well construction begins with mudrotary drilling through the glacial drift at the top 3 m of 

bedrock. A 15-cm inside-diameter (i.d.) steel casing is 

cemented in the drill hole. After the cement has 

hardened, drilling continues by air-percussion to create 

a 15-cm-diameter open hole to the desired depth in the 

bedrock. The only exception to this procedure was the 

construction of well FSE5, which was first cored as a 

9-cm-diameter hole and then reamed to a diameter of 

15 cm. Eight of the 13 wells in the FSE well field are 

75 to 85 m deep. The shallowest well is 61 m deep; the 

deepest well is 229 m deep.

Rock types and fractures encountered in the 

wells are determined by examining drill cutting, 

viewing the borehole walls with a downhole video 

camera (C.D. Johnson, U.S. Geological survey, written 

communs., 1993), and detecting fractures with a 

borehole televiewer (Paillet, 1985; Paillet and Kapucu, 

1989). The drill cuttings and video images show that 

the FSE wells penetrate variable thicknesses of schist, 

granite, and pegmatite. However, there is little to no 

apparent correlation between rock types in neighboring 

wells. This lack of correlation is consistent with the 

complex patterns of rock-type distribution on the 1-93 

road cut.

of bedrock. Transmissivities of these fractures are 

determined by single-well hydraulic tests. These tests 

employ two inflatable packers to isolate a section of the 

well. Water is either injected or withdrawn from the 

test interval, and the flow rate and pressure change are 

recorded. By assuming steady, horizontal, radial flow, 

the transmissivity of the test interval is calculated by 

the Theim equation (as discussed by Bear, 1979). 

Because the intact rock is much less transmissive than 

the fractures, the computed transmissivity is attributed 

entirely to the fractures that intersect the test interval. 

If only one fracture intersects the test interval, the 

calculated transmissivity is the combined 

transmissivity of the fractures. Because of practical 

considerations, single-well hydraulic tests in the FJSE 

wells are conducted with 4.5-m- long test intervals. 

This means that fractures spaced less than 4.5 m apart 

are tested together as a group. Infractors sections of 

wells are not tested.

The distribution of fractures and their 

transmissivities in well FSE 11, which is 85-m deep and 

is cased over the upper 25 m, are shown in figure 2.

(b)

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TRANSW1SSIVITY, IN 

METERS SQUARED PER SECOND

SINGLE-WELL HYDRAULIC TESTS

Borehole-televiewer logs show that each FSE 

well intersects 20 to 60 fractures within the upper 60 m

Figure 2. (a) Vertical distribution of transmissivity determined from single-well hydraulic tests in well FSE11. Bar 

thickness indicates length of test interval, (b) Borehole televiewer log of well FSE11.

128
The borehole-televiewer log shows than the well 

intersects approximately 25 fractures below the casing. 

Because most of these fractures are tested in groups, it 

is not possible to determine the transmissivity of each 

fracture. Nevertheless, the test results shown in figure 

2 clearly indicates that only a few fractures have a high

c "}

transmissivity (greater than 10 m /s), whereas most 

of the fractures have a much lower transmissivity (less 

than 2x10~8 m2/s). Highly transmissive fractures are 

present at depths of 46 m and 80 to 85 m. By 

measuring borehole flow while pumping the open well, 

Paillet and others (1987) conclude that highly 

transmissive fractures are present at the same locations. 

With the exception of well FSE7, which does not 

intersect any highly transmissive fractures, characteristics of fractures at all the wells in the FSE well field 

are similar to those at FSE11; that is, 20 to 60 fractures 

are observed in each well within the upper 60 m of 

bedrock. Of the fractures, only a few are highly 

transmissive. These observations suggest that the 

highly transmissive fractures could play an important 

role in controlling fluid flow in the FSE well field.

transmissive fracture, and monitoring drawdown in all 

the other highly transmissive fractures. The hydraulic 

connection between these fractures are inferred from 

the response to pumping. For example, if two highly 

transmissive fractures are connected by other highly 

transmissive fractures in the bedrock, the drawdowns 

in the two fractures should be nearly identical. In 

contrast, if two highly transmissive fractures are not 

connected by other highly transmissive fractures in the 

bedrock, the drawdowns in the two fractures could be 

significantly different.

A hypothesized distribution of fractures in the 

vertical section between wells FSE1 and FSE6 is 

shown in figure 3. Results of multiple-well hydraulic 

tests suggest that the highly transmissive fractures 

intersected locally to form four fracture clusters, 

denoted by I to IV in figure 3. The fracture clusters are, 

in turn, connected to one another by fractures of 

comparatively lower transmissivity. Preliminary 

analysis of test data indicates that these fractures are 

several orders of magnitude less transmissive than the 

fractures that form clusters I to IV.

MULTIPLE-WELL HYDRAULIC TESTS

The hydraulic connections between the highly 

transmissive fractures intersected by the FSE wells are 

investigated by multiple-well hydraulic tests. To set up 

a test, the highly transmissive fractures in each well are 

isolated from one another by installing packers in 

unfractured parts of the borehole. The packers divide 

each well into separate intervals, each interval 

continuing either (a) a single highly transmissive 

fracture, or (b) several highly transmissive fractures 

that are close to one another (that is, a few meters 

apart). In the latter case, the closely spaced fractures 

are assumed to be part of a well-connected fracture 

cluster, so there is no attempt to isolate these fractures 

from one another. Although each interval also contains 

fractures of comparatively low transmissivity, it is the 

highly transmissive fracture (or fracture cluster) that 

dominates the hydraulic head in the interval. 

Therefore, the hydraulic response observed in each 

interval represents the hydraulic response of the highly 

transmissive fracture in that interval.

During a multiple-well hydraulic test, water is 

pumped at a constant rate from one interval, and the 

hydraulic-head responses field setup described above, 

the this amounts to pumping from a highly

FSEG FSE9 FSE5 FSE4 FSE1

T

Glacial Drill

Highly Iransmissive fracture cluster 

Network ol lower Iransmissivily Iraclures

Figure 3. Diagram showing hypothesized distribution of 

fractures in vertical section between wells FSE1 and FSE6. 

Four clusters of highly transmissive fractures occur within a 

network of less transmissive fractures.

Drawdowns during one of the multiple-well 

hydraulic tests are shown in figure 4. During this test, 

conducted in August 1992, water from fracture cluster 

HI is pumped from well FSE6 at a steady rate of 

10 L/min. Within fracture cluster III, the drawdown at 

well FSE6 (pumped interval) is somewhat greater than

129
the drawdown at well FSE9, because drawdown must 

be greater tat the pumped interval for converging flow 

to enter. Within fracture clusters I, II, IV, drawdown 

appears to be internally uniform. For example, at 

fracture cluster II, identical drawdowns are observed at 

wells FSE5, FSE4, and FSE1, located 57.9 m, 88.4 m, 

and 101.3 m respectively from well FSE6. This 

uniformity of drawdowns at different distances from 

the pumped well is direct evidence that the fractures in 

the cluster are in good hydraulic connection. The fact 

that the hydraulic heads in fracture clusters I, II, and IV 

respond to pumping from fracture cluster III, but the 

responses are quite different from one another, 

suggests that the four fracture clusters are connected to 

each other by fractures of comparatively lower 

transmissivity.

10

0.1

0.01

0.001

t> A I r.iclure Cluslor I

I I i.icluio Cluster II

» I liiuiuu Cluster III

> * ri.-n.uifo Cluster IV

.1-_____L_ 

0.1 11 10 100 1000 10000 

TIME SINCE START OF PUMPING. IN MINUTES

Figure 4. Drawdown during multiple-well hydraulic test 

conducted in August 1992.

The drawdown responses described above is 

entirely different from the drawdown response in a 

homogeneous aquifer, in which the drawdown 

decreases progressively with increasing distance from 

the pumped well. At the FSE well field, the highly 

transmissive fracture clusters exert a strong control on

the drawdown response during a multiple-well 

hydraulic test. This finding suggests that it is necessary 

to identify and characterize the highly transmissive 

fracture clusters explicitly in order to analyze the test 

response. Quantitative analysis of the test data is the 

focus of the ongoing work.

SUMMARY AND CONCLUSIONS

Results of single-well hydraulic tests at the FSE 

well field show that, of the 20 to 60 fractures 

intersected by each of the 13 wells, only one to three 

are highly transmissive. Results of multiple-well 

hydraulic tests suggest that these highly transmissive 

fractures are interconnected locally to form fracture 

clusters. These fracture clusters are, in turn, connected 

to each other by a network of fractures several orders 

of magnitude lower in transmissivity. In such a setting, 

the drawdown response to a multiple-well hydraulic 

test is entirely different from the drawdown response in 

a homogeneous aquifer. These findings suggest that it 

is necessary to identify and characterize the highly 

transmissive fracture clusters explicitly in order to 

analyze multiple-well hydraulic tests at the FSE well 

field.

REFERENCES

Bear, Jacob, 1979, Hydraulics of Groundwatcr: New York, 

McGraw-Hill, 567 p.

Paillet, F.L., 1985, Geophysical well log data for study of 

water flow in fractures in bedrock near Mirror Lake, 

West Thornton, New Hampshire: U.S. Geological 

Survey Open-File Report 85-3340, 27 p.

Paillet, F.L., and Kapucu, Kcmal, 1989, Fracture

characterization and fracture-permeability estimates 

from geophysical logs in the Mirror Lake watershed, 

New Hampshire: U.S. Geological Survey WaterResources Investigations Report 89-4058, 49 p.

130
A New Method of Performing

Controlled Injection of Traced Fluid in

Fractured Crystalline Rock

By Alien M. Shapiro1 and Paul A. Hsieh2

Abstract

A new method is presented for the 

controlled injection of traced fluids into 

permeable intervals in bedrock wells. Prior to the 

tracer injection, an inflatable packer is used to seal 

fractures in the injection interval, allowing 

downhole mixing of the tracer with borehole 

water. After a uniform mixture is achieved, the 

packer is deflated to inject the traced fluid into the 

fractures. The injection apparatus requires smooth 

borehole walls and no more than a few closely 

spaced permeable fractures in the injection 

interval. In this paper, operation of the equipment 

is presented for the instantaneous injection of a 

known volume of traced fluid into the formation. 

The equipment can also accommodate other 

injection procedures, such as the continuous 

injection of traced fluids into the formation.

INTRODUCTION

Controlled-tracer tests are conducted in 

subsurface terranes to identify formation properties 

such as effective porosity and dispersivity. These 

tracer tests are conducted under either hydraulically 

stressed or ambient-flow conditions by introducing a 

known mass of a tracer into the formation and 

monitoring its spatial distribution at several times or 

its temporal variation at one or more locations.

Regardless of the type of hydraulic conditions 

used to conduct the tracer test, the identification of

! U.S. Geological Survey, Reston, Virginia 

U.S. Geological Survey, Menlo Park, California

formation properties requires a knowledge of the flow 

regime and a knowledge of the temporal or spatial 

distribution of the tracer concentration after it has been 

introduced into the formation. It is also crucial to have 

an explicit knowledge of the temporal variation of the 

tracer concentration as it is introduced into the 

formation. If the method of introducing the traced fluid 

into the formation is not well known, then the 

interpretation of the tracer test becomes ambiguous.

In this paper, a newly designed injection 

apparatus is presented for the controlled injection of 

traced fluids into permeable sections of bedrock wells 

in crystalline rock. The operation of this equipment 

permits the injection of a known volume and 

concentration of a traced fluid into the formation over 

a specific period of time. In particular, the application 

of this equipment to the instantaneous injection of a 

traced fluid into fractured rock is discussed.

EFFECT OF VARIABLE HYDRAULIC 

PROPERTIES ON THE INJECTION OF 

TRACED FLUIDS

The bedrock of the Mirror Lake watershed in 

the Hubbard Brook Experimental Forest in Grafton 

County, New Hampshire (fig. 1) has been a site for the 

development of field techniques and interpretive 

methods of characterizing fluid movement and 

chemical transport in fractured rock (Shapiro and 

Hsieh, 1991). The bedrock in the Mirror Lake 

watershed is a sillimanite-grade schist with extensive 

intrusions of granite and pegmatite, and lesser 

amounts of basalt. The variability in hydraulic 

conductivity, fracturing and rock type along the length 

of borehole H-l are shown in figure 2. The hydraulic 

conductivity was measured using single-hole, fluid131
71°42' 71°41'30"

43°57'

43°56'30"

Well site 

Well field

R1

New 

Hampshire

Map/area

1000I 2000 FEET 

_I

300 600 METERS

Figure 1. Location of Mirror Lake watershed in New Hampshire.

injection test over 5-m-long intervals along the length 

of the borehole. Only intervals containing fractures 

were tested; the hydraulic conductivity of the 

unfractured rock matrix was assumed to be below the 

detection limit of the injection-test apparatus, which is 

approximately 10~ 10 m/s. The rock type was identified 

using a borehole television camera (C. D. Johnson, 

U.S. Geological Survey, written commun., 1992) and 

the location of fractures was determined from an 

acoustic televiewer log (F. L. Paillet, U.S. Geological 

Survey, written commun., 1992).

The fracturing and hydraulic conductivity 

shown in figure 2 are typical of boreholes in 

crystalline rock. The fractures are not uniformly 

spaced, and the hydraulic conductivity varies over

4 orders of magnitude with abrupt changes over 

relatively short distances. Not all of the fractured 

sections of the borehole have measurable hydraulic 

conductivity. In some instances, the injection rates 

were below the detection limit of the testing apparatus, 

and thus, a measurable hydraulic conductivity could 

not be assigned to the interval.

Because of the extreme variability in the 

hydraulic conductivity, the controlled injection of a 

traced fluid into the formation cannot be conducted in 

an open borehole. The volume and concentration of 

the traced fluid entering the formation at each of the 

permeable sections of the borehole can not be 

explicitly identified. Also, it would be difficult to 

quantify the concentration of the tracer residing in the

132
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LOGARITHM OF

HYDRAULIC CONDUCTIVITY,

IN METERS PER SECOND

Figure 2. Hydraulic conductivity, acoustic televiewer and lithologic logs for borehole H-1 in the Mirror Lake watershed, 

Grafton County, New Hampshire.

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borehole and then diffusing into the formation. 

Consequently, it is necessary to design a controlled 

environment for injecting traced fluid into bedrock 

formations for the purposes of conducting controlledtracer tests.

A straddle-packer assembly is commonly used 

to isolate a permeable section of a bedrock well for 

hydraulic testing or the injection of traced fluids. A 

straddle-packer assembly consists of two packers, 

each with a flexible bladder that can be enlarged by

133
mechanical or pneumatic means to seal against the 

borehole wall. The interval between two packers is 

hydraulically isolated from the remainder of the 

borehole. A traced fluid can be injected into the 

straddled interval, and then the straddled interval can 

be flushed with tracer-free water. This procedure, 

however, does not guarantee that all of the traced fluid 

will be forced into the formation. Because of density 

differences between the ambient fluid and the traced 

fluid, or incomplete flushing in the straddled interval, 

some of the tracer can reside in the straddled section of 

the borehole and diffuse into the formation over time. 

This will affect the character of the temporal or spatial 

distribution of the tracer concentration in the 

formation and could result in erroneous interpretations 

of formation properties. Reducing the length of the 

straddled interval such that only the fracture is 

hydraulically isolated would eliminate this problem. 

However, it is difficult to reduce the distance between 

packers to less than 0.5 m, which could still cause 

problems in identifying the explicit temporal variation 

of the tracer concentration as it is introduced into the 

formation.

NEW TRACER INJECTION METHOD

To provide for the controlled injection of traced 

fluids into permeable sections of bedrock wells in 

crystalline rock, a new tracer injection system has 

been designed to seal off the permeable fractures 

intersecting the borehole between the straddle packers. 

The equipment is schematically depicted in Figure 3a. 

The equipment consists of 3 packers (labeled A, B, 

and C), and two valves controlling the "injection" and 

a "return" tubing.

The packers labeled A and C are inflated prior to 

the start of the hydraulic test to hydraulically isolate a 

section of the borehole where there is a permeable 

fracture (fig. 3b). Packer B is positioned so that it will 

block off the permeable fracture or fractures in the 

interval straddled by the top (A) and bottom (C) 

packers when it is inflated. Thus, relatively smooth 

borehole walls and only one or two closely spaced 

permeable fractures in the injection interval are 

required. These conditions are most likely to be 

encountered in crystalline rock.

Information on fracture location and hydraulic 

conductivity must be obtained prior to conducting the 

tracer test. For example, the location of fractures can 

be identified using a combination of borehole

geophysical logging tools such as the acoustic 

televiewer and caliper logs (Paillet, 1991) and a 

borehole television camera. The permeable sections of 

the borehole can be identified by using a flow meter 

while pumping water from the open borehole to 

identify locations where fluid enters the borehole from 

the formation (Hess, 1986). More exact measurements 

of hydraulic conductivity can be made using injection 

tests, similar to those described in the previous section 

(the results of which are shown in fig. 2).

Prior to starting the tracer test, the middle 

packer is inflated to block off the permeable fracture in 

the straddled interval (fig. 3c). A tracer solution is then 

prepared at land surface. The concentration of this 

tracer solution should be sufficient to achieve the 

desired injection concentration when it is mixed with 

the volume of tracer-free water residing between 

packers A and C. The injection and return valves are 

opened, and the tracer solution is then pumped through 

the "injection" tubing, and it is circulated between the 

packers. The mixed water returns to the surface 

through the "return" tubing and to the fluid reservoir 

containing the tracer solution (fig. 3d). With the 

middle packer inflated, none of the tracer is introduced 

into the formation. The tracer solution is circulated 

until it is well mixed in the reservoir at land surface 

and in the volume between packers A and B, and B 

and C. Once this is achieved, the injection and return 

valves are closed, and the volume and concentration of 

the water in the reservoir at land surface is measured.

The injection of the traced fluid is started by 

deflating the middle packer and opening the injection 

valve. The tracer is pumped down the injection tube 

and is forced into the formation (fig. 3e). The injection 

of traced fluid into the formation is terminated by 

inflating the middle packer to block the permeable 

fractures in the straddled interval (fig. 3f). The volume 

of water remaining in the fluid reservoir at land 

surface is then measured to determine the volume of 

the traced fluid injected into the formation.

The tracer solution between packer A and B, 

and B and C is then flushed with tracer-free water. 

This is done to ensure that the tracer is not introduced 

into the formation accidentally during the test or once 

the test is completed. The flushing of the tracer is 

conducted by opening the injection and return valves 

and using tracer-free water to flush the straddled 

interval through the "injection" tubing. The water 

exiting from the "return" tubing is monitored until the 

background concentration is achieved.

134
Injection 

Tubing

Injection 

Valve

Return 

Tubing

Injection 

Tubing

\

Return 

Valve

Packer A

Fracture 

Packer B

T .

Injection 

Valve

Packer C

NOT TO SCALE

Return 

Tubing

\

Return 

Valve

Packer A

r

Fracture 

Packer B

Packer C

Injection 

Tubing

r .

Injection 

Valve

(C)

Return 

Tubing

\

Return 

Valve

Packer A

from

fluid Injection 

reservoir Valve

(open)

Fracture 

Packer B

(D)

Packer C

NOT TO SCALE

^

1?* 

sllr

|

A

s

ti *? Bi

$

Return 

Tubing

-q- ^

\ t0 Return fluid 

Valve reservoir

(open) 

Packer A

^^^ Fracture 

Packer B

* Traced Fluid 

Packer C

Figure 3. (a) New straddle packer assembly used for the controlled injection of traced fluids, (b) Packers A and C inflated to 

isolate injection interval in bedrock well, (c) Packer B inflated to seal permeable fracture in the injection interval, (d) Injection 

and return valves opened and traced fluid is circulated to mix in volume between packers A and B, and B and C. (e) Packer B 

deflated and injection valve opened to inject traced fluid into permeable fracture, (f) Packer B inflated to terminate injection; 

injection and return valves opened and tracer-free water circulated to flush traced fluid between packers A and B, and B and C.

135
Injection 

Tubing

Return 

Tubing

from

fluid Injection 

reservoir Valve 

(open)

\

Return

Valve

(closed)

Packer A

Injection 

Jubing

tracer-free /X 

water Injection 

Valve 

(open)

Fracture (F)

Packer B

Traced fluid Packer CTraced fluid

NOT TO SCALE

Return 

Tubing

to

Return waste 

Valve 

(open)

Packer A

dtfr Fracture 

Packer B

Packer C

Figure 3. (a) New straddle packer assembly used for the controlled injection of traced fluids, (b) Packers A and C inflated to 

isolate injection interval in bedrock well, (c) Packer B inflated to seal permeable fracture in the injection interval, (d) Injection 

and return valves opened and traced fluid is circulated to mix in volume between packers A and B, and B and C. (e) Packer 

B deflated and injection valve opened to inject traced fluid into permeable fracture, (f) Packer B inflated to terminate 

injection; injection and return valves opened and tracer-free water circulated to flush traced fluid between packers A 

and B, and B and C Continued.

SUMMARY walls and no more than a few closely spaced

permeable fractures.

In this paper, operation of the new injection 

apparatus is presented for the instantaneous injection 

of a known volume of traced fluid into the formation.

The interpretation of controlled tracer 

experiments in subsurface terranes requires a 

knowledge of the temporal variation in the 

concentration of the traced fluid as it is injected into 

the formation. Because of abrupt spatial changes in 

hydraulic properties in fractured rock, it is difficult to 

perform the controlled injection of traced fluid in open 

sections of bedrock wells. Density differences 

between the ambient fluid and the traced fluid, and the 

lack of adequate circulation of the traced fluid in the 

borehole usually results in a residual concentration in 

the borehole that diffuses into the formation.

A new method for the injection of traced fluid 

into permeable intervals in bedrock wells uses a string 

of three inflatable packers. This newly designed 

injection apparatus permits the controlled injection of 

traced fluids into fractures within a hydraulically 

isolated section of a borehole. Prior to the tracer 

injection, the middle packer is used to seal off the 

fractures in the injection interval, allowing downhole 

mixing of the tracer with borehole fluid. After a 

uniform mixture is achieved, the middle packer is 

deflated to allow injection of the traced fluid into the 

formation. The equipment requires smooth borehole

The equipment can also accommodate other injection 

procedures, such as the continuous injection of traced 

fluids into the formation.

REFERENCES

Hess, A., 1986, Identifying hydraulically conductive 

fractures with a slow-velocity borehole flowmctcr: 

Canadian Geotechnical Journal, v.23, no. 1, p. 69-78.

Paillct, F.L., 1991, Use of geophysical well logs in

evaluating crystallinerocks for siting of radioactivewaste repositories: The Log Analyst,v. 32, no. 2, 

p. 85-107.

Shapiro, A.M., and Hsieh, P.A., 1991, Research in

fracturcd-rock hydrogeology: Characterizing fluid 

movement and chemical transport in fractured rock at 

the Mirror Lake drainage basin, New Hampshire, in 

Proceedings of the U.S. Geological Survey Toxic 

Substances Hydrology Program Technical Meeting, 

G.E. Mallard and D.A. Aronson, eds., Monterey, 

March 11-15, 1991, U.S. Geological Survey WaterResources Investigations Report 91 -4034, p. 155-161.

136
Field Instrumentation for Multilevel Monitoring of

Hydraulic Head in Fractured Bedrock at the Mirror Lake

Site, Grafton County, New Hampshire

By Paul A. Hsieh 1 , Richard L. Perkins2, and Donald O. Rosenberry3

Abstract

An adjustable and removable instrument, 

consisting of commercially available components, 

was developed for monitoring hydraulic heads at 

several depth intervals in bedrock wells. The 

intervals are isolated from each other by packers. 

Each interval is connected to an open pipe, so that 

the water level in the pipe equals the hydraulic 

head in the interval. The water levels are monitored by a potentiometer-float system and automatically recorded by a datalogger. The water 

levels are also measured manually by a waterlevel probe to check periodically the automatic 

measurements. The hydraulic-head monitoring 

instrument can be removed from the well to 

accommodate activities such as geophysical logging, water sample collection, or hydraulic testing. This instrumentation has been installed at 12 

bedrock wells at the Mirror Lake site.

INTRODUCTION

Accurate measurement of hydraulic head is 

essential for monitoring fluid flow in the subsurface. In 

particular, a measurement of the vertical distribution 

of hydraulic head is needed to determine the vertical 

component of fluid flow. In shallow alluvial material, 

the vertical distribution of hydraulic head can be 

measured by installing a cluster of piezometers 

screened at different depths. In most bedrock 

environments, the cost of drilling and casing multiple

*U.S. Geological Survey, Menlo Park, CA

2U.S. Geological Survey, Bow, NH

3 U.S. Geological Survey, Lake wood, CO

wells to different depths is prohibitively high, so that 

the hydraulic heads at different depths must be 

measured in a single well. However, an open well can 

perturb the natural flow field by connecting highly 

permeable fracture zones that are unconnected in the 

absence of the well. Therefore, packers are generally 

needed to divide the well into separate intervals and to 

isolate the highly permeable fracture zones from one 

another. To complicate matters, the well could be 

needed for other purposes, such as collecting water 

samples, geophysical logging, and testing new 

instruments. Accordingly, any instrument installed in 

the well must be removable. This paper describes a 

removable instrument developed for multilevel 

monitoring of hydraulic heads in bedrock wells at the 

U. S. Geological Survey Toxic Substances Hydrology 

research site in the Mirror Lake area, Grafton County, 

New Hampshire.

DESIGN CRITERIA

To accommodate the many uses of a bedrock 

well, the hydraulic-head monitoring instrument is 

designed according to the following criteria: (1) The 

instrument had to divide the well into several intervals 

so that the hydraulic head in each interval could be 

measured separately. (2) Each interval had to be 

connected to an open pipe, so that the hydraulic head 

could be measured either manually with a water-level 

probe, or automatically with a potentiometer-float 

device or an electronic pressure transducer. The 

manual measurements would provide checks on the 

automatic measurements. (3) The instrument had to be 

removable so that the well can be used for purposes 

such as geophysical logging, water quality sampling, 

or hydraulic testing. Upon completion of work, the 

instrument would be reinstalled in the well. (4) The

137
instrument had to be adjustable so that the monitored 

interval could be modified should the need arise. (5) 

The instrument had to be made of components that are 

commercially available to minimize the need for 

specialized fabrication.

DESCRIPTION OF INSTRUMENT

Figure 1 is a schematic diagram showing the 

instrument set up for multilevel monitoring of 

hydraulic head in a bedrock well. The well is cased 

through the overburden and is left as an open hole 

below the casing. Two packers are suspended on 3- 

cm-inside-diameter (i.d.) steel pipes. When inflated 

with air, the rubber bladder of the packer swells and 

presses against the borehole wall, creating a pressuretight seal. The two packers separate the well into three 

intervals: A, B, and C. The hydraulic head in each 

interval is a composite of the hydraulic heads in the 

transmissive fractures that intersect the interval. In 

many cases, the interval contains a single, highly 

transmissive fracture or fracture zone that dominates 

the hydraulic head in the interval.

5-centimeter insidediameter steel pipe

5-centimeter insidediameter PVC pipe

15-centimeter inside- 

"" diameter well casing

5-centimeter insidediameter PVC pipe

Interval A

Interval B

3-centimeter insidediameter steel pipe

Interval C

NOT TO SCALE

Each interval is connected to a 5-cm i.d. pipe so 

that the water level in the pipe equals the hydraulic 

head in the interval. The 5-cm i.d. is needed to 

accommodate a potentiometer-float system for 

automatic water-level measurement. Hydraulic head in 

the bottom (C) interval is measured in the 5-cm i.d. 

steel pipe that is connected to the 3-cm i.d. steel pipe 

on which the packers are suspended. The hydraulic 

head in the middle (B) interval is measured in a 5-cm 

i.d. polyvinyl-chloride (PVC) pipe that is connected to 

the interval by means of a nylon tubing. The hydraulic 

head in the upper (A) interval is the water level in the 

well casing. This water level is also measured through 

a 5-cm i.d. PVC pipe to avoid tangling the float with 

tubing in the well. Not shown in figure 1 are the nylon 

tubing used to inflate the packers.

A potentiometer-float system (Rosenberry, 

1990a) is used to measure water levels in the 5-cm i.d. 

pipes (fig. 2). The wire on which the float and 

counterweight are suspended is wrapped around the 

wheel of a potentiometer. As the water level changes, 

the float moves, turning the wheel and changing the 

electrical resistance of the potentiometer. This change 

in resistance is converted back to water level and 

recorded by a datalogger. Under controlled laboratory

Wheel

Platform

Counterweight

Water level in pipe

NOT TO SCALE

Wire

5-centimeter insidediameter pipe

Float

Figure 1. Instrument for multilevel monitoring of hydraulic 

head in a bedrock well.

Figure 2. Potentiometer-float method for measuring water 

level in the 5-centimeter-inside-diameter pipe.

138
conditions, this potentiometer-float system can detect 

a water level change of 0.1 cm. However, during longterm operation in the field, the wire may gradually slip 

against the potentiometer wheel, causing a larger error. 

A study by Rosenberry (1990b) shows that, during 

long-term operation, water levels are measured by a 

potentiometer-float system with an accuracy of ±1 cm.

In principle, the well can be divided into more 

than three intervals by installing more than two 

packers. In practice, the number of intervals is limited 

by the number of pipes that can fit inside the well 

casing. Because a minimum inside diameter of 5 cm is 

needed to accommodate the float, only three pipes can 

fit inside a 15-cm i.d. well casing. Reducing the pipe 

diameter allows additional intervals to be monitored, 

but requires the use of a smaller float, which lessens 

the sensitivity of the recorder to water-level changes. 

Accurate measurement of water-level change in a 

smaller diameter pipe may require the use of 

alternative water-level sensors such as electronic 

pressure transducers.

INSTALLATION OF INSTRUMENT

Before installing the hydraulic-head monitoring 

instrument, some preliminary information is needed to 

select appropriate locations for packer placement. In 

general, packers are placed between the highly 

transmissive fractures to isolate them from each other. 

These highly transmissive fractures can be identified 

by hydraulic testing with straddle packers, or by 

measurement of flow in borehole during pumping 

(Paillet and others, 1987). A packer should be placed 

where the borehole wall is smooth and unfractured to 

achieve a pressure-tight seal. A caliper log is 

extremely helpful for identifying locations where the 

borehole wall is rough or broken; such sections in the 

borehole should be avoided. Borehole-image logs, 

such as those made by borehole televiewers and video 

cameras, provide detailed information on fracture 

locations.

Once the packer placements are selected, the 

instrument can be installed in the well. In general, a 

well-servicing vehicle (work-over rig) or a tripodwinch assembly is needed to lower the packers and 

pipes into the well. The depth range of each interval 

can be easily adjusted by cutting the 3-cm i.d. steel 

pipes to the desired length. Installing the three 5-cmdiameter i.d. pipes requires special care, because all 

three pipes are lowered simultaneously into the well.

The PVC pipes are affixed to the steel pipe by steel 

bands.

After the packers are lowered into position and 

are inflated, a platform is attached to the well head. 

The potentiometers are then mounted on the platform. 

The well-head installation, including the datalogger, is 

covered by an aluminum box, which provides shelter 

from rain and snow.

MONITORING HYDRAULIC HEADS IN THE 

MIRROR LAKE AREA

Monitoring hydraulic heads in fractured 

bedrock is part of an interdisciplinary study on fluid 

flow and chemical transport in fractured rocks at the 

U. S. Geological Survey Toxic Substances Hydrology 

research site in the Mirror Lake area in central New 

Hampshire (fig. 3). As discussed by Shapiro and Hsieh 

(1991), the objectives of this study are to develop 

monitoring and testing methods for characterizing 

ground-water flow and solute transport in bedrock, 

and to establish a site for long-term study. Although 

the Mirror Lake area is not contaminated, the 

techniques and understanding developed in this study 

are directly transferable to other fractured rock sites 

where ground water has been contaminated with 

chemical wastes.

As of early 1993, multilevel hydraulic-head 

monitoring instruments have been installed in 12 

bedrock wells. Hydraulic heads are recorded 

automatically once every hours. Each well is inspected 

once a week to check that the instrument is working 

properly. Several packers tend to leak air and must be 

repressurized from time to time. During each visit, 

water levels are measured manually for comparison 

with the automatic measurements. As more experience 

is gained from this operation, the frequency of visit 

will decrease to approximately one visit per month.

Figure 4 illustrates the hydraulic heads recorded 

at well site FS3 during 1991. During this period, 

hydraulic heads were manually measured once a week. 

Interval A in the bedrock well extends from the 

bottom of casing (depth of 11.6 m) to a depth of 25.5 

m below land surface. Interval B extends from a depth 

of 26.2 m to 59.2 m. Interval C extends from a depth 

of 59.9 m to the bottom of the well (depth of 196.6 m). 

In addition to the heads in the bedrock well, figure 4 

also shows the heads in three nearby piezometers 

installed at depths of 4.3 m, 6.7 m, and 8.8 m in the 

glacial drift overlying the bedrock. The decrease of 

hydraulic head with depth suggests a downward 

component of flow from the glacial drift into the 

bedrock at this well site.

139
43°57' -

43°56'30' -

Figure 3. Location of the FS3 well site in the Mirror Lake area in New Hampshire.

SUMMARY

An adjustable and removable instrument has 

been developed to monitor the hydraulic heads in 

packer-isolated intervals in a bedrock well. The 

hydraulic head in each interval can be monitored by 

measuring the water level in an open pipe connected to 

the interval. Because the pipes are open, automatic 

water-level measurements can be made using a 

potentiometer-float system. Periodic manual 

measurements using a water-level probe provide a 

check of the automatic measurements. The instrument

276

uj 274 

273

O 271 

5 270 

269 

268 ^

Z 267

9 266

^ 265

264

£

PIEZOMETER DEPTH 4.3. 6.7. 8.8 METERS

INTERVAL

INTERVAL C

1/1/91 3/2/91 5/2/91 7/2/91 9/1/91 11/1/91 1/1/92 

DATE

Figure 4. Hydraulic heads in intervals A, B, and C in 

bedrock well FS3 and in three nearby piezometers installed 

in glacial drift at depths of 4.3,6.7, and 8.8 meters below 

land surface.

can be removed from the well to accommodate 

activities such as geophysical logging, water-sample 

collection, and hydraulic testing. This instrumentation 

has been installed at 12 bedrock wells at the Mirror 

Lake site.

REFERENCES

Paillet, F.L., Hess, A.E., Cheng, C.H., and Hardin, E.L., 

1987, Characterization of fracture permeability with 

high-resolution vertical flow measurements during 

borehole pumping: Ground Water, v. 25, no. 1, p. 28-40.

Roscnberry, D.O., 1990a, Inexpensive groundwater

monitoring methods for determining hydrologic budgets 

of lakes and wetlands: in National Conference on 

Enhancing the States' Lake and Wetland Management 

Program, 1989, Proceedings, Chicago, Illinois, 

sponsored by the U. S. Environmental Protection 

Agency, North American Lake Management Society, 

and Northeastern Illinois Planning Commission, p. 123- 

131.

Rosenberry, D.O., 1990b, Effect of sensor error on

interpretation of long-term water-level data: Ground 

Water, v. 28, no. 6, p. 927-936.

Shapiro, A.M., and Hsieh, P.A., 1991, Research in fractured 

rock hydrogeology: Characterizing fluid movement 

and chemical transport in fractured rock at the Mirror 

Lake drainage basin. New Hampshire, in Mallard, 

G.E., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Monterey, California, March 11- 

15, 1991: U. S. Geological Survey Water-Resources 

Investigations Report 91 -4034, p. 155-161.

140
Factors Affecting Recharge to Crystalline Rock in the 

Mirror Lake Area, Grafton County, New Hampshire

By Philip T. Harte1 and Thomas C. Winter2

Abstract

The effects of local and regional-flow 

processes on recharge to crystalline rock are 

investigated by use of numerical, crosssectional models of ground-water flow, 

comparison of vertical-head gradients between 

lower piezometers in drift and upper bedrock 

wells, and chemical mass-balance models of 

net differences in water chemistry. This paper 

describes factors that control recharge to 

crystalline rock in the Mirror Lake area. Four 

principal factors that affect bedrock-recharge 

patterns are (1) relief of land and bedrock 

surface above ground-water discharge areas, 

(2) lateral trends in bulk-rock horizontal 

hydraulic conductivity, (3) local topographic 

features, and (4) drift stratigraphy. Factors 1 

and 2 control the regional distribution of 

bedrock recharge, whereas the local 

distribution of bedrock recharge is controlled 

by factors 3 and 4.

INTRODUCTION

The spatial distribution of bedrock recharge 

from overlying unconsolidated deposits is an 

important factor in delineating areas susceptible to the 

advective transport of contaminants. Crystalline-rock 

aquifers in many areas of the country that have been 

glaciated have few surface exposures. The low 

hydraulic conductivity of these aquifers can cause the 

water table to be shallow and to he present primarily in

'U.S. Geological Survey, Bow, NH 

U.S. Geological Survey, Lakcwood, CO

the overlying unconsolidated drift in humid 

environments. Few studies have evaluated the 

distribution of bedrock recharge from overlying drift, 

although Keeler and others (1988) have indicated that 

unconsolidated deposits can control rates of bedrock 

recharge.

This paper describes factors that affect bedrock 

recharge from overlying glacial drift. Two numerical, 

steady-state cross-sectional models of ground-water 

flow are used to evaluate recharge processes: (1) A 

model of a generalized hillside-valley system 

representative of New England and (2) A model of the 

Mirror Lake area. Results of the modeling 

experiments are compared to the distribution of 

vertical head between glacial drift and bedrock at 14 

locations, and to the vertical distribution of groundwater chemistry at 2 locations (FS3 and FS1) in the 

Mirror Lake area (fig. 1).

FACTORS AFFECTING BEDROCK 

RECHARGE FOR HYPOTHETICAL 

SYSTEMS

The model of the hypothetical setting is 

patterned after the generalized characteristics of the 

physiography and hydrogeology encountered in the 

Mirror Lake area and elsewhere in the glaciated 

northeast United States (fig. 2). The geologic setting 

consists of fractured crystalline rock overlain by a 

largely continuous mantle of glacial till, which in 

turn is overlain in places by stratified drift. A detailed 

discussion of the setting, finite-difference model 

(McDonald and Harbaugh, 1988), model 

assumptions and boundaries, and sensitivity tests is 

given in Harte (1992). The bulk hydraulic properties 

of the fractured crystalline rock are viewed as an 

equivalent porous medium over the dimensions of 

the investigation.

141
0.4

0.3

0.2

0.1

-0.1

... 1 ... I ... 1 ... 1 ...,.., , ,..,...,.,

FS3-A VHEAD BEDROCK WELL 

1 AND OPEN ZONE

i

1 | 1

T~

a.

H

S?

^

CJ

o

0

I

1

^ UJ

7- W 

H li.

a<

H 1

< ! 

T* 1 

li. <

t

s

< 1

1 1

J : 

 > :

. ;

gj 166 172 176 180 184 188 192 196 200 204 208 212 216 220 224 228 232 236 240 244 248 252 256 260 264 

MIDPOINT OF BEDROCK OPEN INTERVAL. IN METERS ABOVE SEA LEVEL

(a)

71*42' 7 1*41' 30*

43' 

57'

43* 

56' 

30*

i_

 BEDROCK OUTCROP 

WELL AND NUMBER

 SHALLOW ROCK 

FS3C WELL AND NUMBER

SHALLOW ROCK 

BOREHOLE AND 

NUMBER

100 200 300 METERS

CONTOUR INTERVAL 10 FEET (3.048m) 

NATIONAL GEODETIC VERTICAL 

DATUM OF 1929 _____ I____

Base from U.S. Forest Service 

Hubbard Brook Watershed. 1956

(b)

Figure 1. Vertical-head gradients between lower drift and upper bedrock from bedrock well nests (a), and location of well 

nests and shallow rock boreholes in the Mirror Lake area, Grafton County, New Hampshire (b) (modified from U.S. Forest 

Service, written commun., 1956).

142
HILLTOP 

254

25.4

2.54

0.254

0.0254

0.00254

254

25.4

2.54

0.254

0.0254

0.00254

HILLSIDE

HORIZONTAL INCREASE

VALLEY

AME TERRACE 

WITH TILL

CONDITIONS SIMULATED 

Horizontal Increase 

Horizontal Decrease 

Katne Terrace with Till

 DOWNWARD FLOW

 UPWARD FLOW

100

10

0.1

0.01

CONDITIONS SIMULATED . 

Base 

Two Bedrock Knobs

 DOWNWARD FLOW

 UPWARD FLOW

0.001

100

10

1.0 

0.8 

0.6 

0.' 

0.2 

0

Hilltop Hillside Rivr »«»»ey

Till Katne terrace

..\..... Stratlflsr: drift

0.1

0.01

0.001

0.6 

0.4 

0.2 

0

25

O O

0 < z*22*2

0 0.10 0.20 0.30 0.40 0.50 0.60 0.70 0.80 0.90 

PROPORTION OF HORIZONTAL DISTANCE FROM 

HILLTOP TO CENTER OF RIVER VALLEY

1.0

Figure 2. Diagrams of hypothetical hydrogeologic section and simulated distribution of bedrock recharge and discharge from 

overburden.

Conditions tested include hydraulic 

conductivity variations of the drift and bedrock, 

modification of physiographic characteristics such as 

slope, basin-aspect ratio, and small topographic 

irregularities on the uniformly sloping terrain. 

Distribution of vertical-flow rates (bedrock recharge 

and discharge) from the overlying drift are shown in 

figure 2 for selected conditions. Results presented 

include simulations of layered units consisting of 

homogenous and isotropic material (glacial drift and 

bedrock), homogeneous and isotropic drift and 

heterogeneous crystalline rock, and small topographic 

features superimposed onto the uniform sloping

hillside. Results from the layered homogenous and 

isotropic units serve as the reference "base" case. For 

the "base" case, each hydrogeologic unit has a 

uniform horizontal hydraulic conductivity in the 

following descending relation: stratified 

drift>till>bedrock.

Results indicate that besides relief of the land and 

bedrock surface and the distribution of ground-water 

discharge areas (sinks), which are the primary controls 

on topographically driven flow systems, the 

predominant controls on distribution patterns of bedrock 

recharge are lateral trends in bulk-rock hydraulic 

conductivity (horizontal heterogeneity) and topography.

143
The curves shown in figure 2 illustrate the following 

features: (1) Maximum recharge and discharge occur at 

changes in water-table slope coincident with changes in 

land-surface slope, (2) relatively small topographic 

features can cause localized recharge and discharge 

patterns in a regional recharge or discharge zone, and (3) 

relatively small horizontal trends in bulk-rock hydraulic 

conductivity can result in large variations in recharge 

and discharge zones.

Lateral trends in bulk-rock horizontal hydraulic 

conductivity (horizontal heterogeneity) significantly 

affect the size of bedrock recharge areas. The bulkrock hydraulic conductivity can vary as a result of 

spatial-distribution patterns in fracture density and 

connectivity. Variation in bulk-rock hydraulic 

conductivity of one order in magnitude results in a 50- 

percent variation in the size of the bedrock recharge 

area from that produced under base conditions. The 

two tests shown in figure 2 include increasing and 

decreasing the hydraulic-conductivity distribution 

with decreasing altitude. The change in hydraulic 

conductivity of one order of magnitude is similar to 

the inferred hydraulic-conductivity distribution 

implied by topographic variations in well yields 

(LeGrand, 1967, Sever, 1964, and Knopman and 

Hollyday, 1993). The ripple effect in the curves is 

caused by a step-like approximation to a linear trend in 

hydraulic conductivity. Decreasing hydraulic 

conductivity with decreasing altitude results in a small 

regional recharge area and a large regional discharge 

area. Conversely, increasing hydraulic conductivity 

with decreasing altitude results in a large regional 

recharge area and a small regional discharge area.

Small topographic features cause localized 

bedrock recharge in regional recharge and discharge 

zones because they cause inflections in the water-table 

slope. Although this concept is not new and has been 

reported by others (Toth, 1962, 1963; and Freeze and 

Witherspoon, 1966-68), the results indicate a high 

degree of head sensitivity to topography in low 

hydraulic-conductivity aquifers because the water table 

is generally shallow and is very similar to land surface. 

Bedrock is recharged under downward inflections in 

the water table corresponding to downward inflections 

in land surface. Bedrock discharges under upward 

inflections in the water table corresponding to upward 

inflections in land surface. If local topographic features 

are associated with variations in drift stratigraphy, 

these features can exert greater control over bedrock 

recharge patterns (Harte, 1992).

FACTORS AFFECTING BEDROCK 

RECHARGE IN THE MIRROR LAKE AREA

Results of a numerical model of the crosssectional, ground-water flow in the Mirror Lake area 

(fig. 3) show the importance of relief and local 

ground-water sinks on regional flow in bedrock. The 

results presented are the best-fit approximation 

(closest match) between simulated and observed 

heads and show two important features of this low 

hydraulic-conductivity aquifer. The first feature is the 

steep hydraulic gradient on the hillside and the 

shallow depth to the water table beneath the hilltop. 

The second is the strong control of local ground-water 

sinks (Mirror Lake) relative to regional ground-water 

sinks (Pemigewasset River). The flow system 

associated with Mirror Lake extends to nearly the full 

depth of the simulated ground-water system. The lake 

also receives a small amount of ground-water 

discharge from the eastern side, the side toward the 

regional ground-water sink.

The distribution of vertical-head gradients 

between lower drift and upper bedrock indicates that 

bedrock receives recharge across much of the upland 

adjacent to Mirror Lake (table 1 and fig. 1). Verticalhead gradients show a strong correlation with landsurface altitude (coefficient of determination is 0.75) 

and bedrock-surface altitude (coefficient of 

determination is 0.73).

Further comparison of vertical-head gradients 

between lower drift and upper bedrock indicates that 

discharge from bedrock is restricted to areas where the 

bedrock water-level-sample midpoint is less than 

213 m above mean sea level (table I), which is the 

lake level for Mirror Lake. The sample midpoint is the 

average vertical distance of the open bedrock interval 

and is the reference point for the bedrock water level. 

The strong relation between altitude of bedrock 

midpoint and bedrock recharge and discharge 

indicates that the flow system is significanlty affected 

by relief relative to local ground-water sinks.

The predominance of bedrock recharge 

determined from vertical-head gradients along the 

lower third of the hillside indicates the possible 

occurrence of lateral trends in bulk-rock hydraulic 

conductivity. An increase in horizontal hydraulic 

conductivity with a decrease in altitude, such that low 

hydraulic-conductivity rocks underlie hilltops and 

high hydraulic-conductivity rocks underlie valleys, 

would produce the observed distribution pattern of 

bedrock recharge. Otherwise, the lower hillside would 

be a bedrock discharge area.

144
AREA SH5WN 

IN FIGURE 1

r J *

0 200 400 800 METERS 

I "-r-1 S

BEDROCK OUTCROP 

.__ DRAINAGE DIVIDE 

Q NESTED WELLS

CONTOUR INTERVAL 61 METERS 

NATIONAL GEODETIC VERTICAL 

DATUM OF 1929

Base from U.S. Geological Survey 

Woodatock, N.H., 1:24,000. 1980

A

METERS 

700 1

DIRECTION OF FLOW

LINE OF EQUAL HEAD, IN METERS

200 

SEA LEVELALTITUDE OF MEASURED 

WATER LEVEL. IN METERS

1000 2000 3000 FEET 

(b)

Figure 3. Location of line of (a) regional section and the Mirror Lake watershed, and (b) geohydrologic section showing lines 

of equal head and generalized pathlines from best-fit model

145
Table 1. Vertical-head gradients in the Mirror Lake area, Grafton 

[m, meters; m/m, meters per meters; minus sign indicates upward

County, New Hampshire 

flow; , no data]

Well cluster

FS3

FS3C

FSl

FS2

Rl

TRl

TR2

Tl

FSE

CO

Kl

K2

K3

Well description

Upper drift

Lower drift

Upper bedrock

Upper drift

Lower drift

Upper bedrock

Upper drift

Lower drift

Upper bedrock

Lower drift

Upper bedrock

Upper drift

Lower drift

Upper bedrock

Upper drift

Lower drift

Upper bedrock

Drift

Upper bedrock

Drift

Upper bedrock

Upper drift

Lower drift

Upper bedrock'

Upper drift

Lower drift

Upper bedrock'

Upper drift

Lower drift

Upper bedrock

Upper bedrock

Upper drift

Lower drift

Bedrock

Upper drift

Lower drift

Bedrock

Upper drift

Lower drift

Bedrock

Well number

FS3-14

FS3-29

FS3-A

FS3C-14

FS3C-29

FS3C

FSl-WT

FSl -35

FS I -A

FS2-22

FS2-A

RI-WT

R I -55

RI-A

TRl-WT

TRI-L32

TRl -A

TR2-23

TR2-A

TI-8

Tl-A

FSE6-WT

FSE6-59

FSE6-A

FSE2-WT

FSE2-43

FSE2-A

COI-WT

CO1-I8

COI-A

CO2-A

KI-WT

K I -39

Kl

K2-WT

K2-4I

K2

K3-WT

K3-6I

K3

Altitude of 

measurement 

point 

(m above sea 

level)

275.09

275.02

274.90

274.04

274.18

274.69

262.58

262.68

262.33

255.8S

255.93

257.29

257.31

257.05

249.23

249.34

249.41

233.38

232.95

229.62

229. 1 5

242.99

243.03

242.90

240.98

241.01

241.07

217.78

217.91

217.97

2I6.SS

215.37

2I5.3S

215.41

215.76

215.62

215.19

208. 1 6

207.83

207. 1 6

Altitude of 

average water 

level (m above 

sea level)

271.74

270.53

26X.80

269.59

269.51

268.54

259.83

260.70

256.36

247.77

247.34

25 1 .52

251.72

25 1 .06

233.24

231.76

230.24

232.14

232.26

228.05

224.95

238.51

237.49

235.79

235.80

234.20

234.07

217.7S

217.91

216.23

215.51

212.84

211.91

2 1 2.67

210.37

210.46

211.31

206.09

204.65

206.44

Midpoint of 

open interval 

(m above sea 

level)

270.91

265.42

252.83

269.38

264.87

261.03

257.86

250.55

230.61

248.47

234. 1 8

246. 1 0

239.51

23 1 . 1 9

230.09

208.58

192.42

226.10

20S.67

227.50

210.28

23S.I1

224.52

214.45

234.67

226.92

208.29

214.15

211.75

191.41

1 96.66

2 1 1 .99

205.59

1 86. 1 1

20X.6I

202.51

1 S2.09

201.75

1 89.53

169.07

Average 

vertical-head 

gradient 

(m/m)

--

0.22056

. 1 3656

.01149

.2523S

-.11917

.21774

.03006

--

-.02963

.07839

--

.06856

.09434

--

-.00734

0.18000

-

.07480

. 1 S088

_

.20709

.01358

--

.003X0

-.01874

.02242

-

.14524

-.03897

--

-.01552

-.04145

-

.11798

-.08723

I'1st i niii ted water level

The distribution of piezometer nests, which 

are located predominantly on hillside terraces, can 

contribute to the observed distribution of bedrock 

recharge. The stratigraphy of the drift in most 

terraces, including FS3 and FS3C, FSl, and FSE 

(fig. 1) consists of an upper, relatively permeable 

sandy layer (kame terrace deposit) overlying a

lower, less permeable till layer. These conditions 

can produce local areas of bedrock recharge within 

a regional discharge zone because the water table is 

commonly elevated in the terrace as a result of 

infiltration of precipitation recharge through the 

sand that exceeds vertical flow through the 

underlying till (Harte, 1992).

146
FS3 Sand

DRAFT WELL SCREEN 

;S1-35^ AND WELL NUMBER

FS1-A BEDROCK OPEN HOLE 

MIDPOINT AND WELL 

NUMBER

V DIRECTION OF GROUND- 

^* WATER FLOW

FS3-

Not to scale

(a)

l-35£

DRAFT WELL SCREEN 

AND WELL NUMBER

FS1-A BEDROCK OPEN HOLE 

MIDPOINT AND WELL 

NUMBER

\ DIRECTION OF GROUND- 

^ V WATER FLOW

FS1 Sand

Not to scale

(b)

Figure 4. Schematic diagrams showing (a) bedrock 

recharge distribution produced in a regional-flow system 

caused by trends in horizontal heterogeneity, and (b) 

bedrock recharge distribution produced in a local-flow 

system caused by variations in topography.

Local topography can also affect bedrock 

recharge and discharge. Upward and downward 

inflections of the land and corresponding inflections in 

the water table can produce discharge and recharge. 

Bedrock recharge occurs at piezometer nests TRl, Tl, 

FS3C, and FS2 (fig. I) near a downward inflection in 

the slope of the land. Bedrock discharge occurs at 

piezometer nest TR2 near an upward inflection in the 

slope of the land (fig. I). However, local topographic 

features cannot fully explain the observed distribution 

of bedrock recharge and discharge. For example, 

bedrock recharge occurs at piezometer nest Rl 

although it is near an upward inflection in slope of the 

land (fig. I). However, the local topography produces 

divergence of vertical flow in the lower drift (table I).

Upward flow occurs from lower to upper drift and 

downward flow occurs from the lower drift to upper 

bedrock.

BEDROCK RECHARGE PATTERNS 

BASED ON WATER CHEMISTRY

Model experiments indicates two possible 

bed rock-recharge patterns resulting from either 

regional or local control. Solute and isotopic 

chemistry of the water are used here to define which of 

these proccses is most probable.

Two possible vertical-flow patterns resulting 

from regional and local controls on bedrock recharge 

are shown for part of a hillside between two adjacent 

terraces (fig. 4). Regional control on bedrock recharge, 

caused by lateral trends in bulk-rock hydraulic 

conductivity, results in bedrock recharge over the entire 

area (fig. 4a). Local control on bedrock recharge, 

caused by local topography or drift stratigraphy, 

results in the formation of local-flow cells between 

adjacent terraces (fig. 4b). Vertical-head gradients 

between lower drift and upper bedrock indicate bedrock recharge occurs under each terrace (table I). Linear interpolation of heads between terraces indicates 

that bedrock recharge also might occur between terraces. However, the effects of either flow pattern on 

water chemistry at the FSI terrace should result in distinct chemical signatures, which would help verify the 

primary bedrock-recharge process.

Mass-balance and fractionation water-chemistry 

processes are modeled to assess the evolution of water 

chemistry as a result of various flowpaths. The 

interactive computer code "NETPATH" (Plummcr and 

others, 1991) is used to interpret net geochemical 

mass-balance reactions between initial and final water 

along a possible hydrologic flowpath. The principle 

behind mass-balance modeling of the evolution of 

water chemistry is to reconstitute final water chemistry 

through source/sink reactions with initial water 

chemistry.

The conceptual chemical model used here is 

based on water-chemistry reactions in silicate terrains 

(Garrels, 1967; Bricker and others; 1968). These 

reactions indicate that the dominant effect on water 

chemistry in silicate terrain is reaction of carbondioxide enriched water, from contact with the 

atmosphere (open system), with plagioclase feldspars 

and biotite to form clay minerals. The initial and final 

concentrations of elements involved in these reactions,

147
sodium, silica, calcium, potassium, and carbon, are 

used to constrain model reactions between proposed 

evolutionary waters. An additional constraint is 

imposed on the system by modeling carbon-13 

concentrations with calcite and carbon dioxide inputs 

from the atmosphere (open system) and(or) organic 

decay. Calcite has been observed in petrographic rock 

analysis of crystalline rock from the Mirror Lake area 

(W.W. Wood, U.S. Geological Survey, oral commun., 

1993). The ground-water-chemistry data used in the 

analysis includes major cations and anions and the 

carbon-13 isotope (W.W. Wood, U.S. Geological 

Survey, written commun., 1992) from lower drift and 

upper bedrock wells at FS3 and FS1 (table 2).

The evolutionary paths of water chemistry 

evaluated are shown in table 3. The initial and final 

water chemistry is tested by use of the conceptual 

chemical model to determine the feasibility of 

flowpaths. The feasibility of flowpath existence is 

determined by checking whether final waters can be 

reconstituted from the specified reactions discussed in 

the conceptual chemical model and the initial water 

input. The flowpaths evaluated include vertical flow 

between lower drift and upper bedrock at each terrace 

and lateral flow between terraces.

Net differences in water chemistry between lower 

drift and upper bedrock under the FS3 and FSI terraces 

satisfy the constraints of the specified conceptual 

chemical model and vertical-head gradient, and indicate 

that bedrock recharge occurs under each terrace. 

Satisfactory reaction results are obtained that produce 

clay minerals from dissolution of feldpars and biotite. 

Upper bedrock cores collected from beneath each 

terrace reveal the presence of clay minerals. The 

modeled chemical reactions also indicate calcite 

dissolution under the FS3 and FS1 terraces and an 

increase in the heavier carbon isotope, carbon-13, 

which agrees with carbon-13 data. The thermodynamic 

potential of these waters supports calcite dissolution.

Net differences in water chemistry between 

terraces fails to support the presence of a local-flow 

cell connecting terrace FS3 with FS1. Final waters in 

the lower drift at FS1 could not be reconstituted from a 

mixing input of initial drift and bedrock waters from 

FS3. If a local-flow cell connected the two terraces, 

bedrock discharge would occur on the uphill side of 

the FS1 terrace and waters at FS1 lower drift should 

reflect a component of bedrock-derived waters, if the 

relative amount of flow in the till and bedrock are 

similar.

Table 2. Chemistry of selected lower drift and upper 

bedrock ground water from the Mirror Lake area, Grafton 

County, New Hampshire

[°C\ degrees Celsius, mg/L, milligrams per liter; (.ig/L, micrograms per 

liter; %>, pans per thousand; ft, difference from a standard; VPDB, ViennaPecDee Belemile; <, less than; --, no data

Chemical 

constituent

PH

Eh (volts)

Temperature (°C)

Alkalinity. (mg/L 

as HCO3 ) "

Calcium as C'a

(mg/L)

Magnesium as Mg 

(mg/L)

Sodium as Na

(mg/L)

Potassium as K

(mg/L)

Iron as Fc ((.ig/L)

Strontium as Sr

Chloride as Cl

(mg/L)

Sulfatc as SO4 

(mg/L)

Fluoridc as F

(mg/L)

Bromium as Br

(mg/L)

Nitrate as NO-, 

(mg/L)

Phosphate, ortho 

dissolved as

P04 (mg/L)

Silicate as SiO2 

(mg/L)

Carbon - 1 3 as <*n/^ °/

O L /OO

relative to

VPDB

FS3-22 

(Lower 

drift)

6.9

.227

10.0

26.8

4.7

1.2

3.6

1.6

5,300

30

.5

14

.12

<.OI

<.0t

<.OI

22

-24.40

FS3-A 

(Upper 

bedrock)

7.1

-.025

8.5

61

11

3.9

4.6

3.7

1,000

93

1.2

8.6

.20

.06

-

29

-12.65

FS1-35 

(Lower 

drift)

7.7

.43

8.5

39

9. 1

1. 6

3.6

1. 7

1 3

37

.6

6.6

.28

<.()!

.07

<.()!

21

-19.80

FS1-A 

(Upper 

bedrock)

7.2

.154

9.5

39

5.9

1.8

4.4

1.6

3,900

48

.5

12

.33

<.()!

<.()!

<.()!

21

-17.10

The results of the water-chemistry analysis of 

lower drift and upper bedrock waters underlying the 

FS3 and FS1 terraces indicate that upper bedrock 

waters are primarily derived from vertical flow 

through lower drift, although some lateral flow in 

bedrock between terraces might occur. The analysis of 

water chemistry suggests bedrock recharge occurs 

over the entire area between terraces. This 

interpretation is consistent with available head data.

148
Table 3. Summary of flowpaths evaluated by use of chemical-reaction modeling to test the presence of a local-flow cell 

between FS3 and FS1 terrace, in the Mirror Lake area, Grafton County, New Hampshire

Initial 

waters

Final 

waters Results

FS3-22(Lower drift) 

FSI-35 (Lower drift) 

FS3-22( Lower drift)

FS3-A (Upper bedrock)

FS3-22(Lower drift) 

FS3-A (Upper bedrock)

FS I-35 (Lower drift) 

FS3-A (Upper bedrock)

FS3-22(Lower drift) 

FS3-A(Upper bedrock)

FS3-A (Upper bedrock) 

FS I-A (Upper bedrock) 

FS I-35 (Lower drift)

FS I-A (Upper bedrock) 

FS I-A (Upper bedrock)

FS I-A (Upper bedrock) 

FSI-35 (Lower drift)

Chemistry indicates evolution of waters; therefore, docs not 

disprove tlowpath. System closed to CO2

Chemistry indicates evolution of waters; therefore, docs not 

disprove flowpath. System closed to CC>2

Chemistry indicates evolutionary waters; therefore, docs not 

disprove flowpath. Works for open system to atmospheric 

input of CO2 or fermentation reaction.

Chemistry indicates noncvolutionary waters; therefore, 

flowpath is unlikely.

Chemistry indicates evolutionary waters from FS3-22 but not 

from FS3-A; therefore, strong influence of lateral flow in 

bedrock is unlikely. Need open system of CO2 or 

fermentation reaction for evolution of FS3-22 waters to 

FSI-A waters.

Chemistry indicates evolutionary waters primarily from FSI35 (86 percent) and small amount from FS3-A (14 percent); 

therefore, vertical flowpath is not disproved with some 

lateral flow. Works for closed system.

Chemistry indicates non-evolutionary waters; therefore, 

flowpath is unlikely.

CONCLUSIONS

The predominant controls on distribution of 

recharge to crystalline rock in a mountainous New 

England setting are (I) relief of land and bedrock 

surface above ground-water sinks, (2) lateral trends in 

horizontal hydraulic conductivity in bedrock, (3) local 

topographic features, and (4) drift stratigraphy. 

Regional distribution of bedrock recharge is controlled 

by factors I and 2. Local distribution of bedrock 

recharge is controlled by factors 3 and 4.

The distribution of bedrock recharge in the 

Mirror Lake area is controlled primarily by relief of 

land and bedrock surface above ground-water sinks 

and lateral trends in horizontal hydraulic conductivity. 

The predominant distribution of vertical-head 

gradients indicate that bedrock recharge occurs 

throughout the lower parts of the hillside. The large 

area of bedrock recharge covering the low parts of the 

hillside probably results from lateral trends in bulkrock horizontal hydraulic conductivity where the least 

permeable rocks underlie hilltops and the most 

permeable rocks underlie low-lying areas. Bedrock 

recharge appears to be less affected by local 

topographic features and drift stratigraphy than by 

lateral trends in bulk-rock horizontal hydraulic 

conductivity.

REFERENCES CITED

Brickcr, O.P., Godfrey. A.E., and Cleaves, E.T., 1968, 

Mineral-water interaction during the chemical 

weathering of silicates: American Chemical Society, 

Advances in Chemistry, scries no. 73, p. 128-142.

Freeze. R. A., and Withcrspoon, P.A., 1966, Theoretical 

analysis of regional groundwatcr flow: 1. Analytical 

and numerical solutions to the mathematical model: 

Water Resources Research, v. 2, p. 641-656.

___ 1967, Theoretical analysis of regional groundwatcr 

flow: 2. Effect of water-table configuration and 

subsurface permeability variation: Water Resources 

Research, v. 3, p. 623-624.

____ 1968, Theoretical analysis of regional groundwatcr

flow: 3. Quantitative interpretations: Water Resources 

Research, v. 4, p. 581-590.

Garrels, R.M., 1967, Genesis of some ground waters from 

igneous rock, Researches in geochemistry 2: N.Y., 

John Wilcy and Sons, p. 405-420.

Hartc, P.T., 1992, Regional ground-water flow in crystalline 

bedrock and interaction with glacial drift in the New 

England Uplands: University of New Hampshire, 

Durham, N.H., published masters thesis, 147 p.

Kcclcr, C.K., Van Dcr Kamp, G., and Cherry, J.A., 1988, 

Hydrogcology of two Saskctchewan tills, I) Fractures, 

bulk permeability, and spatial variability of downward 

flow: Journal of Hydrology, v. 101, p. 97-121.

Knopman, D.S., and Hollyday, E.F., 1993, Variation in 

specific capacity in fractured rocks, Pennsylvania: 

Ground Water, v. 31, no. 1, p. 135-145.

149
LcGrand, H.E., 1967, Ground water of the piedmont and 

blue ridge provinces in the southeastern states: U.S. 

Geological Survey, Water-Resources Circular no. 538, 

lip.

McDonald, M.G., and Harbaugh, A.W., 1988, A modular 

three-dimensional finite-difference ground-water flow 

model: U.S. Geological Survey Techniques of WaterResources Investigations, chap. Al, book 6.

Plummcr, L.N., Prcstcmon, E.C., and Parkhurst, D.L., 1991, 

An interactive code (NETPATH) for modeling NET 

gcochcmical reactions along a flow PATH: U.S. 

(ieological Survey Water-Resources Investigations 

Report 91-4078, 227 p.

Sever, C.W., 1964, Geology and ground-water resources of 

crystalline rocks, Dawson County, Georgia: U.S. 

Geological Survey Information Circular no., 30, 5 p.

Stcwart, G., 1968, Drilled water wells in New Hampshire: 

New Hampshire Department of Resources and 

Economic Development Mineral Resources Survey. 

Part XX, 58 p.

Toth, J., 1962, A theory of ground water motion in small 

drainage basins in Central Alberta, Canada: Journal of 

Geophysical Research, v. 67, no, 11, p. 4375-4387.

___ 1963, A theoretical analysis of ground-water flow in

small drainage basins in Central Alberta, Canada: 

Journal of Geophysical Research, v. 68, no, 11, p.

4795-4812.

150
Concentrations of Chlorofluorocarbons and Other Gases 

in Ground Water at Mirror Lake, New Hampshire

ByEurybiades Busenberg andL. Niel Plummer1

Abstract

The concentrations of Chlorofluorocarbons 

(CFC's) were measured in samples of ground 

water from 51 different depth intervals in wells 

completed in the low-permeability glacial till and 

fractured crystalline rocks at Mirror Lake. The 

CFC concentrations were used to model waterrecharge ages. The ground-water ages generally 

increased with depth from land surface and were 

consistent with the reconstructed tritium ( 3H) concentrations. A 6°C water-recharge temperature 

was obtained from the nitrogen (N2) to argon (Ar) 

ratios from the wells completed in the glacial till. 

The dissolved N2, Ar, and helium concentrations 

generally increased with depth from land surface. 

Excess air in ground water ranges from about 0 to 

19 cubic centimeters per liter in some deep wells. 

In addition to the excess air, some wells contain 

significant amounts of excess N2 . Enrichment in 

the concentrations of sulfate, calcium, and magnesium were present in some ground waters with 

CFC-recharge ages of 1963 to 1973.

INTRODUCTION

Flow of ground water through fractured rocks is 

difficult to characterize because of the complexity of 

the fracture network, the extreme variability in 

permeability and hydraulic properties of the bedrock, 

and difficulty of predicting the direction of fluid flow. 

Fractured-rock systems have the potential to transport 

fluids rapidly over large distances in a few highly 

permeable fractures. The age of the ground water can 

be a valuable tool in determining flow velocities, 

determining the hydraulic conductivities, and 

evaluating the direction of flow in such systems.

'U.S. Geological Survey, Mail Stop 432, Reston, VA

The Miror Lake watershed area is approximately 

85 ha and is located at the eastern end of the Hubbard 

Brook Experimental Forest in Grafton County in 

central New Hampshire. A map of the area and the 

location of wells are given in Shapiro and others 

(1996). The precipitation and stream-water chemistry 

has been extensively studied for more than 20 years 

(Likens and others, 1977; 1985; Driscoll and others, 

1989). The hydrogeologic setting has been described 

by Winters (1984). At Mirror Lake, the ground water is 

present in two distinct units the overburden and the 

crystalline bedrock. The overburden consists largely of 

low-permeability glacial tills that store most of the 

ground water. The crystalline rocks have a network of 

fractures through which regional flow is thought to 

occur. Water samples collected from wells completed 

in the overburden and from discrete intervals in wells 

completed in crystalline rocks were used to determine 

CFC concentrations and water ages. Dichlorodifluoromethane (CFC-12), trichlorofluoromethane 

(CFC-11), and trichlorotrifluoroethane (CFC-113) are 

the most useful CFC's (Busenberg and Plummer, 1992; 

Busenberg and others, 1993) for dating recent 

recharge. Plummer and others (1993) compare and 

contrast the utility of the CFC's with that of tritium/ 

helium ( 3 H/3 He) (Tolstikhin and Kamensky, 1969), 

and krypton-85 ( x5Kr) (Rosanski and Florkowski, 

1979) for dating young ground water.

This paper presents a summary of the results of 

CFC-dating of ground water at Mirror Lake for water 

samples collected in 1991 and 1992. A total of more 

than 290 ground-water samples were analyzed for 

concentrations of CFC-11 and CFC-12. In addition to 

concentrations of CFC-11 and CFC-12, concentrations 

of CFC-113 in the 1992 samples were also determined. 

In addition, concentrations of dissolved nitrogen (N2 ), 

argon (Ar), carbon dioxide (CO2 ), and methane (CH4) 

in the ground waters were measured.

151
CONCENTRATIONS OF DISSOLVED 

GASES IN GROUND WATERS

Concentrations of N 2 , Ar, CO2 , and CH4 were 

measured in 26 ground-water samples collected at 

Mirror Lake. Concentrations of N2 and Ar generally 

increased with depth as shown in figure 1A and table I. 

The partial pressure of N 2 in many ground waters 

exceeded 1 atmosphere, and many deep waters were 

observed to degas when pumped to the surface; this 

degassing may have affected the results slightly. The 

composition of sample 10 (table 1) was substantially 

modified by an air leak, and the concentrations of N 2 

and Ar in sample 45 (table I) are uncertain because of 

analytical problems.

The N2 and Ar concentrations are shown in 

figure IB. The water-recharge temperature of the 

sample on the diagram can be obtained by following a 

diagonal line (parallel to the temperature lines) from 

the sample to the recharge-water-temperature axis. The 

excess-air content of the sample can be obtained by 

following a diagonal line (parallel to the excess-air 

lines) from the sample to the excess-air axis. Possible 

trends in the evolution of dissolved gases in a ground 

water are shown in the lower right-hand corner of 

figure I B. Presence of radiogenic Ar will shift the 

sample composition (tilled circle) vertically as shown. 

Bacterial reduction of the solute nitrate to N2 gas will 

shift the dissolved-gas composition horizontally. 

Addition of excess air will shift the composition 

diagonally. The slope of the line is determined by the 

N2 to Ar ratio in air. In addition to the rccharge-water 

temperature, these diagrams provide valuable 

information on the recharge mechanism of the ground 

waters (Mazor and Verhagen, 1983, 1984).

Ground waters from wells completed in the 

overburden (table 2) plot near the water in equilibrium 

with ail' (WE!A) or zero-excess-air line and give a gasrecharge temperature of about 6°C. Most ground water 

from the crystalline fractured rocks deviate from the 

WEA line, and the deviation is generally greater for 

waters obtained from the deeper packed intervals. The 

waters contain far more N 2 than could be derived from 

the reduction of nitrates because the nitrate 

concentrations in precipitation (Likens and others, 

1977), and in the shallow ground water are very low 

(W.W. Wood and A.M. Shapiro, U.S. Geological 

Survey, written commun., 1993). The high Ar 

concentrations in some waters could not have been 

derived from the radioactive decay of potassium-40 in

the tens of years since recharge. The ground waters 

could have been contaminated with compressed air 

because modern air was used to power the pumps; 

however, mass-balance calculations on CFC's indicate 

that the excess air was not modern air. The excess N2 

and Ar can not be entirely attributed to excess air 

because the ratios of these two gases in many ground 

waters are not those found in air. A possible source of 

excess N? may have been leakage form N 2 -filled 

packers. Even though no pressure drops were observed 

in the packers, this source of N2 cannot be totally 

dismissed at this time. In addition, high concentrations 

of He were observed in deep ground waters (Drenkard 

and others, 1996), and this He could not have been 

introduced by the sampling.

Similar high N2 and Ar concentrations were 

observed in ground waters from fractured rocks from 

Yosemite National Park in California (J.W. Bochers, 

U.S. Geological Survey, written commun., 1992), and 

high N2 and possible He concentrations have been 

reported in waters from boreholes in crystalline rocks 

in Sweden (Hermansson and others, 1991). It is 

conceivable that the dissolved-gas compositions may 

have been modified by the addition of gases released 

from the crystalline rock. N2 , Ar, and He arc trapped in 

minerals and fluid inclusions of igneous and 

metamorphic origin (Roedder, 1972, p. J.I32-.U62 and 

table 2). Deep circulating ground waters may be 

discharging at Mirror Lake. Mixing of these waters 

with modern recharge could be producing the 

anomalous dissolved-gas compositions. Additional 

dissolved-gas samples were collected in 1993, and the 

isotopic composition of the N2 indicates that the 

ground-water samples were not contaminated with 

compressed air or with NT from the NT-filled packers 

(J.K. Bohlke, U.S. Geological Survey, oral commun., 

1993). The 1993 results have not been fully evaluated.

THE CHLOROFLUOROCARBONDATING METHOD

CFC's are stable synthetic organic compounds 

that were first manufactured in the I930"s and are 

entirely of anthropogenic origin (Lovelock, 1971). 

Detectable concentrations of CFC-12, CFC-11, CFC1 13 are present in post-1940, 1945, and 1973 waters, 

respectively. A detailed description of the method and 

the special procedures that are required to collect and 

analyze water samples containing CFC's are described 

in detail elsewhere (Busenberg and Plummer, 1992;

152
45

40

i

!35

 ^7* (J

3d25 

g^

SS20

15.

Nitrogen concentration in 

equilibrium with air at 6°C

100 200 300 400 500 

DEPTH BELOW LAND SURFACE, IN FEET

1.2

cc

1.0

si

0.4

10 20 30 

NITROGEN CONCENTRATION, IN

40 50 

MILLIGRAMS PER LITER

Figure 1 A. Concentrations of nitrogen as a function of 

depth, at Mirror Lake, New Hampshire. The wells completed 

in the glacial tills are of less than 65 feet deep. The composition of sample 10 was modified by an air leak into the sampler. Symbols are well identifiers and entries above symbols 

are sample numbers, listed in table 2.

Figure 1B. Dissolved nitrogen and argon concentrations in 

ground water collected at Mirror Lake. (Concentrations were 

normalized to a total pressure of 1 atmosphere.)

Buscnberg and others, 1993). The concentrations of 

CFC's in water vary as a function of the atmospheric 

partial pressures are calculated from the measured CFC 

concentrations in ground water from the ground-water 

recharge temperature by means of Henry's Law; the 

partial pressures were then compared to those depicted 

in the reconstructed atmospheric growth curves of 

Busenberg and others (1993) and Elkins and others 

(1993).

Uncertainties in CFC-modeled ages result from 

(1) differences in concentrations of CFC's in the 

unsaturated-zone air at the water table from the 

northern-hemisphere continental troposphere (such

differences occur in semiarid regions with thick 

unsaturated zones; Weeks and others, 1982); (2) 

elevated atmospheric CFC concentrations in industrial 

or large urban areas (Hester and others, 1974); (3) 

introduction of CFC's from the pump, tubing, or other 

sampling equipment; (4) contamination of ground 

water form sewage or other sources (Busenberg and 

Plummer, 1992; Dunkle and others, 1993); (5) removal 

of CFC's by sorption onto lignite or other organic 

matter in the aquifer (Ciccioli and others, 1980; Russell 

and Thompson, 1983); or (6) removal of CFC's by 

biogenic degradation in highly anaerobic environments 

(CFC-11 degrades more rapidly than CFC-12; Lovley 

and Woodward, 1992).

Recharge dates of water containing CFC's were 

calculated using a recharge temperature of 6°C; results 

of the calculations are given in table 2. The most reliable 

dates were obtained from use of CFC-12 concentrations. 

About 35 percent of waters (table 2) were contaminated 

with CFC-11 and CFC-113, and their measured 

concentrations exceeded the concentrations that would 

be in equilibrium with the 1992 atmosphere. The 

principle source of contamination appears to have been 

the plastic water-discharge tubing connecting the pump 

to the CFC-sampling apparatus, and the rubber packers. 

Rubber and plastic tubing sorbs significant amounts of 

CFC-11 and CFC-113, which are slowly released into 

the ground water during sampling (Reynolds and others, 

1990).

GROUND-WATER AGES AT MIRROR LAKE

The volume-weighted concentrations of 3 H in 

the recharge waters were calculated from the 

reconstructed 3 H concentrations (R.L. Michel, U.S. 

Geological Survey, written commun., 1993) and the 

Mirror Lake precipitation data (U.S. Forest Service, 

written commun., 1993), and then were decayed lo the 

year 1992 (fig. 2A). The model-CFC-12 concentrations 

in water at the recharge temperature of 6°C are also 

shown in figure 2A. The CFC-12 recharge dates of 

ground waters as a function of their 3 H content are 

given in table 1 and figure 2B. CFC-12 recharge dates 

refer to the time since the recharge water was isolated 

from the unsaturated-zone atmosphere, and are not 

necessarily the dates of precipitation. The dashed 

horizontal line shows the estimated 3H concentrations 

in the recharge calculated by the model for Mirror 

Lake; these concentrations are the average of the 3 H 

concentrations of two snow samples collected in 1991

153
Table 1. Concentration of gases and excess air and nitrogen in ground 

water from Mirror Lake, New Hampshire.

[Concentrations ofgases are in milligrams per liter, and excess gases in cubic ccntimeteis pci 

liter. NT, nitrogen; Ar, argon; C()->, carhon dioxide, CII 4 , methane; It. leel]

Sample Well name and 

No. depth interval

1

2

7

8

9

10

11

12

13

16

17

18

19

27

28

33

34

39

43

44

45

46

47

48

49

50

CO-1 104-1 14 ft

CO-1 420-430 ft

FS1-17 ft

FS1-25 ft

FSl-35 ft

FS1 121-136 ft

FS1 309-324 ft

FS1 348-363 ft

FS2 134-149.3 ft

FS3 C-24 ft

FS3 C-29 ft

FS3 60-75.1 ft

FS3 294-309.1 ft

HI 107-122 ft

HI 178-193 ft

Rl-35 ft

Rl-55 ft

Tl 57-72 ft

TR1 170-190 ft

TR1 271-299.9 ft

TR1 298.6-327.4 ft

TR1 390-397.8 ft

TR1 4S2-489.S ft

TR2 151-170.1 ft

TR2 270-289. 1ft

TR2 426-445.1 ft

Average 

depth (ft)

109.0

425.0

17.0

25.0

35.0

128.5

316.5

355.5

141.7

24.0

29.0

67.6

301.6

114.5

1S5.5

35.0

55.0

64.5

180.0

285.5

3130

393.9

4S5.9

160.6

279.6

435.6

Measured concentrations

N2

21.81

27.37

19.82

20.05

22.98

35.63

23.18

24.78

24.99

21.59

18.57

24.16

27.04

22.64

22.82

19.54

22.31

22.62

24.10

26.17

28.50

27.60

38.03

25.14

34.00

42.19

Ar

0.9364

.9231

.7527

.7579

.8257

1.0923

.8219

.8258

.8942

.7604

.7869

.8625

.7830

.8262

.7515

.7561

.8271

.7896

.7534

.8042

.7879

.8100

.8524

.7975

.9453

1.0355

02

0.0

.1

.2

2.7

3.2

3.2

.4

.2

.0

7.1

.0

.0

.0

.1

.1

10.1

.2

.2

.1

.0

.0

.0

.0

.0

.0

.0

C02

2.4

2.7

3.2

0.0

.0

9.4

22.5

85.6

17.2

25.8

18.6

6.0

5.2

2.1

2.9

74

.0

10.2

.9

1.4

2.3

1.4

1.4

1.9

5.6

6.5

CH4

0.07

.00

.28

.00

.00

.00

.00

.01

.24

.18

.09

.00

.00

.00

.06

.Ot)

.00

.00

.01

.09

.00

.02

.00

.00

.00

.00

Excess gases

Air

0.00

6.77

.15

.35

2.85

15.00

3.45

6.45

3.48

3.35

.00

3.46

2.10

2.13

.00

.00

1.44

3.88

.00

2.95

3.15

3.75

6.80

3.00

13.20

19.00

N2

0.00

.00

.00

.00

.00

.00

.00

.00

.00

.00

.00

.00

5.70

.00

3.20

.00

.00

.00

4.50

3.80

6.50

4.70

12.80

2.85

2.95

6.10

(W.W. Wood and R.L. Miclicl, U.S. Geological 

Survey, oral commun., 1993). The measured H 

eoneentrations in ground water from shallow 

overburden wells are significantly lower than the 

estimated 3 II eoneentrations in modern recharge. 

Several reasons may explain the discrepancy between 

the observed 3 H and estimated 3 H concentrations in the 

recharge water. The estimated 3 H concentrations in 

recharge may be too high if the recharge occurs in the 

summer months when ^H concentrations in 

precipitation are low compared to the concentrations 

during winter and spring, if the recharge water takes 

several years to reach the water table, or if the water 

sampled in the overburden wells consisted of a mixture 

of recent recharge with pre-1940's water discharged 

from the fractured bedrock. It is unlikely that 

significant recharge occurs during the summer months 

because of the high evapotranspiration (Likens and 

others, 1977). Recharge may take several years to 

reach the water table because the glacial till has low

permeability at some locations; however, the 

overburden is highly heterogeneous and rapid recharge 

may occur at other locations.

Note that 35 TU was the highest measured *\\ 

concentration in the ground waters; this concentration 

is one order of magnitude lower that the H 

concentration modeled at the 1963-bomb peak (figure 

2A). Also note that the 3 H peak in figure 2B does not 

correspond to the 1963-bomb 3 H peak. Both of these 

results suggest that the samples are mixtures of waters 

of different ages. The observed mixing or dispersion of 

these waters may be natural or may represent an artifact 

of sampling. The concentrations of CFC's and H are 

the volume-weighted average of the mixtures. The 

apparent age of the mixed waters depends on the 

proportions as well as the concentrations of the 

components in the mixture. The input function of the 

tracers (figure 2A) will significantly affect the volumeweighted concentration of the mixture. The C'FC-12 

apparent age will be slightly younger than the volume154
Table 2. Calculated chlorofluorocarbon water-recharge dates, tritium concentrations, 

pumping rates, field observations, and dissolved gases in ground water from Mirror Lake, New 

Hampshire

[CFC-12, chlorofhiorocarbon-12; CFC-1 I, chlorofluoroc;irhon-l I; CFC-1 13, clilorofluorocarhon-l 13; TU. (riliuni 

units; L/inin, liters per niiiuite; <, older than; cone., eoneentration; A, leakage from above; U, leakage from below; 

I), water degassing at surface; NT, nitrogen; Ar, argon; Com., contaminated; , not applicable or not sampled; PP, 

problems purging well; Cu, samples collected with copper tubing; pi I. pi I of the water; (), overburden well]

Oldest Oldest 

Sample Well name and CFC-12- CFC-11- 

No. depth interval based date based date

1 CO-1 104- 114 ft

2 CO- 1420-430 ft

3 CO-3 26-40 ft

4 CO-3 78- 102 ft

5 CO-3 140-155 ft

6 CO-4 92-98.8 ft

7 FS-1 17ft

8 FS-1 25 ft

9 FS-1 35 ft

10 FS-1 121-136 ft

11 FS-1 309-324 ft

12 FS-1 348-363 ft

13 FS-2 134-149.3 ft

14 FS-2 236-25 1.3 ft

15 FS-2 364-399.3 ft

16 FS-3C-24 23-24 ft

17 FS-3C-29 28-29 ft

18 FS-3 60-75.1 ft

19 FS-3 294-309.1 ft

20 FS-3 365-380. 1ft

21 FS-3 530-545. 1ft

22 FSE-32 30-32 ft

23 FSE-4341-43ft

24 FSE-4 438-445 ft

25 FSE-5 108-1 15 ft

26 FSE-693.4-122ft

- H-l Open Hole

27 H-l 107-122 ft

28 H-l 17S-193 ft

29 IS- 156-58.8 ft

30 IS- 1364-379.8 ft

31 IS- 1400-406.8 ft

32 IS- 1466.5-473.3 ft

33 R- 135 ft

34 R- 155 ft

35 R-l 65-80.1 ft

36 R-l 110-125. 1ft

37 R-l 147-162.1 ft

38 R-l 480-495. 1ft

39 T- 157- 72 ft

40 T-l 155-157 ft

41 T-l 220-237 ft

42 TR-163ft

43 TR-1 170- 190 ft

44 TR-1 271-299.8 ft 

45 TR-1 298.6-327.4 ft

46 TR-1 390-397. 8 ft

47 TR-1 482-489.8 ft

4S TR-2 151-1 70. 1ft

49 TR-2 270-289.1 ft

50 TR-2 426-445.1 ft

1967.0

1955.5

Modern

19695

1966.5

19725

Modern

19S5.0

19755

19675

< 1940.0

< 1940.0

1974.0

1966.5

1966.0

1992.6

1989.5

1960.5

1959.5

1955.0

19570

1987.0

1978.0

< 1940.0

1955.0

19535

19735

19685

19670

19695

19705

1971.0

19600

1991.7

1989.5

19730

1958.5

1952.5

1965.0

19710

1985.5

1974.5

1982.5

1966.5

1962.5 

19620

1958.5

1953.5

1948.5

1949.0

19520

19640

1957.0

1964.0

1960.5

19475

19705

Cont.

Cont.

Cont.

Cont.

19655

19635

19875

Cont.

19680

Cont.

1992.5

19890

Cont.

Cont.

Cont.

Cont.

Cont.

1973.0

1963.5

19505

19740

19630

19650

19635

19710

19740

19680

Cont.

19465

Cont.

Cont.

Cont.

19680

19530

1954.0

1958.5

1984.0

1966.5

Cont. 

Cont.

1978.0

19675

< 19450

1960.5

1958.5

Oldest Tritium 

CFC-113- cone, 

based date TU

<1973

<1973

Cont.

1992

1981

1979

...

...

1977

1975

1975

1984

Modern

1989

Cont.

Cont.

Cont.

1975

Cont.

Cont.

1981

1975

Cont.

1982

Cont.

1977

Cont.

Cont.

1976

<1973

<1973

1985

Cont.

<1973

<1973

...

1987

Cont. 

Cont.

Cont.

Cont.

1982

1984

Cont.

15.60

4.80

13.80

18.00

10.30

17.20

11.80

15.60

31.20

20.20

.03

.05

.07

.04

.02

12.90

12.80

11.10

2.84

2.10

2.48

14.30

21.90

.29

16.60

14.10

27.70

25.70

26.50

1.53

12.20

17.00

13.40

6.57

1.96

.52

21.80

12.90

1.58

12.30

21.60

9.60 

5.53

1.17

1.00

.76

.18

-0.13

Pumping 

rate 

L/min.

1.80

3.30

1.60

5.00

5.10

3.60

.08

.08

.10

5.50

.SO

1.20

.28

.45

.40

.05

5.40

1.20

5.40

4.60

...

...

.24

...

1.50

120

4.40

3.80

4.90

4.60

Variable

.25

Dry

4.10

5.40

4.70

1.00

4.50

4.70

4.20

.50

.69

3.30 

1.00

.80

.80

6.10

4.40

4.50

Field 

observations

_

_

...

...

A

0

0, PP, Cu, pHll

0, PP, Cu, pH 8.5

...

D

D

D

0, Around casing

0

A

_

...

0

0

...

"

...

_

_

...

...

0, Cu

0, PP, Cu, PH 9.7

A.B

A

_

...

...

0, Cu

...

A,B, D 

A,B, D

D

D

A

...

...

Excess 

dissolved 

gases

Ar

Air

Normal

Normal

Air

Air leak

Air

Air

Air

...

...

Air

Ar

Air

N2

_

...

...

Ar

N2

...

...

Normal

Ar

...

...

Air

...

...

...

N2

N2 

Air, No

Air, N2

Air, No

Air, No

Air, No

Air, No

weighted average age of the sample; however, any 

mixture that contains a small amount of bomb-peak 

water will have high 3 H concentrations. The mixing of 

waters may partly account for the apparent shift of the 

 *H peak in figure 2B.

Figure 2B shows that samples 13, 14, 15,38, and 

41 (open circles) have essentially no " H and must be 

older than their apparent CFC-12 ages. Indeed, samples 

40 and 41 have significantly older CFC-11 and CFC113 ages. Samples 13 and 14 have virtually no 3 H and 

were almost certainly contaminated with CFC-11 and, 

possibly, with CFC-12 during sampling.

CALCIUM, MAGNESIUM, AND SULFATE 

CONCENTRATIONS

The concentrations of calcium (Ca), magnesium 

(Mg), and sulfate (804) in the ground waters are shown 

as a function of the CFC-12 date of recharge in figure 

3A. Also shown is the U.S. Environmental Protections 

Agency's (1988) estimates of sulphur dioxide (SO2 ) 

emissions for the northeastern United States, as well as 

the SO2 emissions normalized to the amount of 

precipitation at Mirror Lake. The normalized SO2 

concentrations were obtained by solving the equation 

v ), where E and P are the annual SO2

155
400

300

100

O 30

i=t/>

o:

O

o 

o

oi= 

05;

ID Z

20

 Model tritium concentration In 

recharge in tritium units (TU)

o Modal chlorofluorocorbon 1 2 

concentrotion in 

picogroms per X 

kilogrom of -- -- ' *

1963 Tritium' 

Bomb Peak I

- 1992 Tritium

- Concentration

'?» s0'V ' V ' , 381W 1?

A

B

YEAR OR DATE OF RECHARGE (CFC-12)

Figure 2A. Modeled concentrations of tritium and chlorofluorocarbon-12 in recharge waters at Mirror Lake, New Hampshire. Tritium concentrations were corrected for radioactive 

decay to the year 1992.

Figure 2B. Relation of tritium concentration to year of 

recharge based on chlorofluorocarbon-12 (CFC-12) concentrations in ground water. Open circles are chlorofluorocarbon12 model dates that are inconsistent with the tritium of chlorofluorocarbon-11 and chlorofluorocarbon-113 concentrations. Symbols are well identifiers and entries above symbols 

are sample numbers, listed in table 2.

emissions and precipitation, respectively. The 

subscripts i and max represent the / year and 

maximum amount observed, respectively. The data in 

figure 3 indicate that there is correlation among the Ca, 

Mg, and SO4 concentrations in ground water, 

atmospheric SO2 emissions, and the amount of 

precipitation. The high concentrations of SO4 are 

present in ground water with CFC-dates of 1963 to 

1973. CFC-model recharge dates are consistent with 

high emissions of SO2 , and lower-than-normal rainfall. 

The results of Driscoll and others (1989) imply that the 

SO4 and cation concentrations in precipitation peaked 

at approximately this time at Mirror Lake. No chemical 

data for precipitation at Mirror Lake are available prior 

to 1963.

CONCLUSIONS

Results from this study show that CFC's can be 

used to data ground waters from fractured rocks. 

Tritium, chemical, and hydrological information were 

useful in evaluating the CFC dates.

The CFC and 3 H results indicate that 

considerable mixing of waters can occur as a result of 

sampling and also naturally in fractured-rock aquifers, 

resulting in a shifting of CFC-based ages to a slightly 

younger age than suggested by ^H concentrations. 

CFC-12 recharge dates refer to the time since the 

recharge water was isolated from the unsaturated-zone 

atmosphere and are not necessarily the age of the 

precipitation. The precipitation may take several years 

to reach the water table because of the low permeability 

of the glacial till overburden at some locations.

An average recharge temperature of 6"C was 

determined from the dissolved gas concentrations in 

ground water from wells in overburden. Anomalous 

concentrations of dissolved N2 and Ar were found in 

ground water form the fractured crystalline rocks, and 

(he dissolved gas concentrations in water increased 

with depth at Mirror Lake.

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chlorofluorocarbons (CCI^F and CChFj) as hydrologic 

tracer and age-dating tools The alluvium and terrace 

system of central Oklahoma: Water Resources 

Research, v. 28. p. 2257-22X3.

Busenbcrg, Fairybiadcs, Weeks, F.P., Plummcr, L.N.. and 

Bartholcmay, R.C., 1993, Age dating ground water by 

use of chlorofluorocarbons (CCl^F and CCUF^) and 

distribution of chlorofluorocarbons in the unsaturatcd 

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J.M., 1980, Organic solute-mineral surface 

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Driscoll, C.T., Likens, G.H., Hcdin, L.O., Haton, J.S., and 

Borman, F.S., 1989, Changes in the chemistry of 

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Schlosscr, P., Michcl R.L., Shapiro, A.M., and Wood, 

W.W., 1996, Helium isotope analysis and tritiumhelium age dating the Mirror Lake basin, Grafton 

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Substances Hydrology Program Proceedings of the

156
14 

12 

10'

I I I I I

rMagnesium

50 2

o: 6

f LJ

C£ 40 

CO i

Ul^? 20

6^ °

' ' __I JK

OIj 

^ 20

15

10,

i Calcium

4*' M 2

*& * 2 ' **14 **

1 ' ' ' ' ' ' ' ' ' ' ' ' ' ' ' 

. Sulfate

51-

o

8000

UJ 7000 :

rd 6000

 5000

 4000

^: 3000

o

UJ

cc:

Q-2000

1000

I I I | I I I I | I I I I | I I I I | I I I I | I I I I | I I I I | I I I I | I I I T

46 3

*' , . '

 ., 2» '« 28 ,.

4» J7*7 . 2ljg% 44* 43,27 ' 3

 . | "^ 39 " 

?'." 

Normalized sulfur 

dioxide emissions 

to precipitation 

: in kilotons

: Sulfur dioxide

j emissions in kilotons,

I. Northeastern U.S.

f(U. S. Environmental

: Protection Agency, 1988)

Annual precipitaion 

in millimeters, 

Mirror Lake, N.H.

I 

A

8000 CO D

O 

7000 O

6000 -7

(/)

5000 z

O

00

4000 ^22UJ

3000 Ul 

Q

X

2000 

Q

cc:

1000

00

1940 1945 1950 1955 1960 1965 1970 1975 1980 1985 1990 1995

YEAR OR DATE OF RECHARGE

Figure 3A Relation of concentrations of magnesium, calcium, and sulfate in 

ground water to year of recharge based on chlorofluorocarbon-12 concentrations in ground water. Entries above symbols are sample numbers, listed in 

table 2.

Figure 3B. Amount of precipitation at Mirror Lake, estimates of sulfur dioxide 

emissions for the northeastern United States, and the sulfur dioxide emissions normalized to the precipitation at Mirror Lake.

technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resoruces Investigations Report 94-4015.

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Denver, J.M., Hamilton, P.A., Michel, R.L., and 

Coplen, T.B., 1993, Chlorofluorocarbons (CCl^F and 

CC1 2 F2 ) as age-dating tools and hydrologic tracers in 

shallow ground water of the Delmarva Peninsula,

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Research, v. 29, 3837-3860.

Elkins, J.W., Thompson, T.M., Swanson, T.H. Butler, J.H., 

Hall, B.D. Cummings, S.O. Fisher, D.A., and Roffo, 

A.G., 1993, Slowdown in the growth rates of 

atmospheric Chlorofluorocarbons 11 and 12-Resultof 

the Montreal Protocol: Nature, v. 364, p. 780-783.

Hermansson, H.P., Sjoblom, R., and Akerblom, G., 1991, 

Geogas in crystalline bedrock: Swedish National Board

157
for Spent Nuclear Fuel, Statcns Karnbranslc Namnd, 

SKN Report 52, p. 22.

Hester, N.E., Stephen, E.R., and Taylor, O.C., 1974,

Fluorocarbons in the Los Angeles basin: Journal of Air 

Pollution Control Association, v. 24, p. 591-595.

Likens, G.E., Bonnan, F.H., Eaton, J.S., and Johnson, N.M., 

1977, Biogcochemistry of a forested ecosystem: New 

York, N.Y., Springer-Verlag, 146 p.

Likens, G.E., Borman, F.H., Pierce, R.S., and Eaton, J.S., 

1985, The Hubbard Brook Valley, p. 9-39, in An 

ecosystem approach to aquatic ecology Mirror Lake 

and its Environments, G.E., Likens, cd.. New York, 

Springer-Verlag, 516 p.

Lovelock, J.E., 1971, Atmospheric fluorine compounds as 

indicators of air movements: Nature, v. 230, p. 379.

Lovlcy, D.R., and Woodward, J.C., 1992, Consumption of 

Frcon CFC-11 and CFC-12 by anaerobic sediments and 

soils: Environmental Science and Technology, v. 26, 

no. 5, p. 925-929.

Mazor, E., and Verhagen, B. Th., 1983, Dissolved ions, 

stable isotopes and noble gases in thermal waters of 

South Africa: Journal of Hydrology, v. 63 p. 315-329.

 , 1984, Dissolved ions, stable isotopes and noble gases 

in thermal waters of South Africa reply: Journal of 

Hydrology, v. 72, p. 201-203.

Plummcr, L.N., Michcl R.L., Thurman, E.M., and Glynn, 

P.D., 1993, Environmental tracers for age-dating young 

ground water, in Alley, W.M., cd., Regional groundwater quality, New York, Van Nostrand Rcinhold, p. 

255-297.

Reynolds, G.W., Hoff, J.T., and Gillham, R.W., 1990, 

Sampling bias caused by materials used to monitor 

halocarbons in groundwater: Environmental Science 

and Technology, v. 24, p. 135-142.

Roedder, Edwin, 1972, Composition of fluid inclusions, in 

Fleischer, Michael ed.. Data of geochemistry (6th ed.), 

U.S. Geological Survey Professional Paper 440-JJ, 

164 p.

Rosanski, K.., and Florkowski, T., 1979, Krypton-85 dating 

of groundwater, in Isotope Hydrology 1978, v. II, p. 

949-961, Vienna: International Atomic Agency, IAEASM-228-51.

Russell, A.D., and Thompson, G.M., 1983, mechanisms 

leading to enrichment of the atmospheric fluorocarbons 

CCl^F and CCI 2 F2 in ground water: Water Resources 

Research, v. 19, no. I, p. 57-60.

Shapiro, A.M., Wood, W.W., Busenbcrg, E., Durenkard, S., 

Plummer, L.N., Torgerscn, T., and Schlosser, P., 1996, 

A conceptual model for estimating regional fluid 

velocity in the bedrock of the Mirror Lake area, 

Grafton County, New Hampshire, in Morganwalp, 

D.W., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 93-4015.

Tolstikhin, I.N., and Kamensky, I.L., 1969, Determination 

of groundwater ages by the T-~ He method: 

Geochemistry International, v. 6, p. 810-811.

U.S. Environmental Protection Agency, 1988, Comparison 

of historic 802 and Nox emission estimates: EPA/600/ 

7-88/009A, PB89-10364/HSU.

Weeks, E.P., Earp, D.W., and Thompson, G.M., 1982, Use 

of atmospheric fluorocarbons F-l 1 and F-12 to 

determine the diffusion parameters of the unsaturated 

zone in the southern high plains of Texas: Water 

Resources Research, v. 18, p. 1365-1378.

Winters, T.C., 1984, Gcohydrologic setting of Mirror Lake, 

West Thornton, New Hampshire: U.S. Geological 

Survey Water Resources Investigations Report 84- 

4266, 61 p.

158
Helium Isotope Analysis and Tritium-Helium Age Dating 

in the Mirror Lake Basin, Grafton County,

New Hampshire

By S. Drenkard 1 , T. Torgersen 1 , R. Weppernig1 , K. Parley 1 , P. Schlosser1 , 

R.L. Michel2 , A.M. Shapiro2 and W.W. Wood2

Abstract

Ground-water samples from the U. S. 

Geological Survey's Mirror Lake fractured 

bedrock site in New Hampshire were analyzed for 

tritium (3H) and its decay product helium-3 ( 3He) 

as well as for helium-4 (4He) and neon (Ne). 

Parallel measurements of the radioactive 

mother/daughter pair H/3He can be used to 

determine a ground-water age ~ that is, the time 

elapsed since the water last equilibrated with the 

atmosphere. Preconditions for precise 

calculations by this method are separation of the 

tritiogenic 3He signal from other He components 

dissolved in water (atmospheric, nucleogenic and 

mantle 3 He) by use of 4He and Ne data. This 

separation technique is complicated for samples 

analyzed from the Mirror Lake site because of a 

large helium excess with an estimated He/ He 

ratio of about 1.65xlO~6. Water analysis from 

shallow wells yield reliable H/ He-based ages, 

but the error in the age calculation increases with 

depth because of an increase in excess He and the 

related errors in the associated corrections. 

Possible origins of the excess He are discussed, 

and mixing of waters (young, tritium and ^Heenriched with old, He-enriched) as an apparent 

source is ruled out. He isotope measurements of 

crushed rock samples indicate that the source of 

excess He is not the local bedrock. On the basis of 

the observed increase in 3H/3He age and presence

' Lament-Dohcrty Earth Observatory, Rt. 9W, Palisades, N.Y.

10964

"U.S. Geological Survey, National Center, Reston, Va.

of excess He with depth, the potential of the 4He 

concentration as an additional proxy age-dating 

tool is explored.

INTRODUCTION

An important concept for estimation of regional 

hydraulic properties beyond the space scale of 

borehole tests is knowledge of the travel time or 'age' 

of a water sample that is, the time since last 

equilibration with the atmosphere. Three methods are 

used for ground-water age dating in the range of 

several months to several decades: the ratio of 3H to 

3 He concentrations (Tolstikhin and Kamensky, 1969; 

Torgersen and others, 1978), chlorofluorocarbon (CFC) 

concentrations (Busenberg and Plummer,1991,1992),

etc

and krypton-85 ( Kr) decay (Smethie and others,

1 1

1992). The H/ He technique works well in porous 

media (Poreda and others, 1988; Schlosser and others, 

1988, 1989; Solomon and Sudicky, 1991) and is less 

vulnerable to local anthropogenic contamination than 

CFC dating. The 3 H/3He method has not been used 

extensively to study fractured-bedrock aquifers, 

although in principle it should be equally applicable. 

Age dating based on a combination of the CFC, 85Kr, 

and -*H/3 He methods has the potential to detect water 

mixing and, under favorable circumstances, even 

mixing rates. In addition, the He concentration can be 

used together with other geochemical tracers as a tool 

to determine general paths of ground-water flow. This

1 1

paper describes the preliminary results of H/ He 

dating and He isotope investigations in the wellcharacterized bedrock environment of the Mirror Lake 

watershed, New Hampshire. This investigation is part 

of the U.S. Geological Survey Toxic Substances 

Hydrology Program and was partially funded by the 

NATO and the 'Deutscher Akademischer Austausch 

Diensf.

159
SAMPLE COLLECTION AND 

MEASUREMENTS

Water samples for He-isotope and 3 H analysis 

were collected with those for other geochemical 

analyses (for details see Shapiro and others, 1996). For 

H, 1 L of water was collected in glass bottles and 

analyzed by low-level gas counting after electrolytic 

enrichment. Water samples for He isotope analysis 

were collected in pinch-off copper tubes (3/8-in. i. d., 

about 80 cm long) carrying about 40 g of water. A 

flow-control valve was used to prevent degassing 

during sampling.

The dissolved gases were extracted from water 

samples using a vacuum extraction system and 

transferred into flame-sealed glass ampoules. He 

isotope concentrations were measured by a dedicated 

mass spectrometer after separation of water vapor, neon 

(Ne) and other gases. Ne was measured in a quadrupole 

mass spectrometer. Typical precision of the He/ He 

ratio analysis is about ±1 to 2 percent. He in groundwater samples is measured routinely with a precision of 

±1 to 3 percent. Because of the high helium 

concentrations, most of the gas samples had to be split 

before injection into the mass spectrometer; this had 

significant impact on the overall precision of the 

measurements of samples with high Fie excess.

3H/3He-DATING METHOD

The H/ He-age-dating technique is based on 

the radioactive decay of atmospheric 3 H that enters 

ground-water during recharge and the accumulation 

with time of its daughter product He. The age ( I ) 

can be calculated from the equation

T1/2

In (2) In 1 +

tn

H

where: T , /2 = half life of 3 H (12.43 years)

[ Hetrj] = He concentration due to H decay 

[ 3 H ] = 3 H concentration corrected to the

day of sampling

Although the 3 H content in a sample can be 

measured directly, not all He ( Hetot) in a water 

sample is of tritiogenic origin (Torgersen and others, 

1978). In addition, there is an air saturation 

component, 3 Hesat, depending on the recharge 

temperature, as well as possible supersaturation 

caused by the dissolution of air bubbles ( Hea jr) 

during recharge. Occasionally, non-atmospheric He 

and He sources, He^.. and Hepx .,, are observed:

Hetot =

The air-related 3 He contribution and the

4

corresponding Heajr fraction, with the isotopic ratio 

of 1.384x10~6, can be calculated as long as the recharge temperature and the Ne content are known:

4i 

4i

'tot

He 4

Neair

tot = Hesat +Heair *Hecxc

"Heair = (He/Ne)air Ne,ot - Nesa ,

3 Heair = 1.384xl(r6 x (He/Ne)ajr x 

( Netot - Nesat ),

where Netot, Nesat and Neair are defined as above for 

He and the He/Ne ratio of air is 0.288 (for further 

details, see Schlosser and others, 1989).

Because 3 FI decay does not produce 4He, the 

difference between the measured He content and the 

calculated Heajr plus 4Hesat yields the He fraction 

derived from in situ Fie sources (4Hecxc ). If the 

3 He/4Fie ratio of that additional source can be 

determined, the - He contribution of the source can be 

calculated. The difference between the measured 3 He 

and calculated Heajr plus ~ Hecxc is the tritiogenic ~ He 

fraction (for further details see Torgersen and 

others, 1978, and Schlosser and others, 1989):

3 Hetri = 3 Hetot -( 3 He/4 He)air x[3 Hesat

+ (He/Ne)airx(Nctot-Nesat)]

- ( 3 He/4He)cxc x 4Hecxc

RESULTS AND DISCUSSION

The results of the water-sample analyses are 

listed in table 1. Most of the samples have Ne 

concentrations within a factor of two of saturation and, 

therefore, a relatively small additional air component. 

On the other hand, a significant number of samples 

contain excess He that must have been added to the 

water in situ. For the Mirror Lake Basin ground-water 

samples, the measured He data are plotted as a 

function of the sampling depth in fig. la. Even if it is 

assumed that ground water recharged during the period 

of maximum H fallout (concentration of about 800 

TU) that occured around 1963, H decay would produce 

less than 1.65x 10"' 2 ccSTP/g 3 Hc through 1992 

(ccSTP/g means cubic centimeters of gas at a pressure 

of 1 atmosphere and 0°C per gram of water). The Hesat 

fraction in the samples is about 6x10~ 14 ccSTP/g. 

Recharge-temperature assumptions and the measured 

Ne concentrations produce only small variations in 

3Hesat. Thus, an explanation of measured 3 He 

concentrations above about 1.71xlO" 12 ccSTP/g 

requires an excess HQ source.

160
Table 1. Tritium, helium, neon, and age data for ground-water samples from the Mirror Lake Basin (All data 

refer to the day of He sampling.)

[ccSTP/g, cubic ccnlimclers of helium (lie) or neon (Ne) gas at 0° (' and one atmosphere per gram of sample water; ll^xc , non-air 

helium component; R cxc Micelle ratio of the excess helium component; TU, tritium (Ml) unit; &, Ml and He samples were not 

collected at the same time. The Ml data were decay corrected to the lie-sampling date; +, age dating was not possible because of an 

overly high excess helium component; corresp.,, corresponding; %, per cent; log (neg. arg.), the argument is negaliv; the logarithm of a 

negaliv argument is nol defined]

Sample 

identification

C01 -18

CO 1 104-114 

FS1-17

FS1-25

FS 1-35*1

FS1-3502

FS2 134-149.3

FS3 - 25

FS3 60-75.1

FSE-32

FSE-43

H1 107-122

H1 178-193

IS-1 56-58.8

K1 -8

R1 -32

R1 65-80.1

R1 110-125.1

R1 147-162.1

T1 57-72

T1 155-157

TR1 170-190

TR1 -63#1

TR1 - 63 #2

TR2 151-170.1

Depth Sampling Measured 3ne/ 

in dale 3ne/-»He 4ne 

meters ralio error 

in %

5.2 12/3/90

33.2 6/16/92 

5.3 9/24/91

7.6 9/25/91

10.4 9/25/91

10.4 12/5/90

43.2 7/7/92

6.2 12/5/90

20.6 7/12/92

9.7 5/21/92

13.1 5/19/92

34.9 8/13/91

56.5 8/13/91

17.5 6/3/92

2.1 11/30/90

10.5 9/26/91

23.0 7/14/92

35.8 7/14/92

47.1 7/15/92

19.6 8/15/91

49.8 8/15/91

54.8 6/30/92

18.3 9/26/91

18.3 9/26/91

48.7 6/22/92

1.40E-06 1

7.33E-07 1.5 

1.43E-06 1.5

1.58E-06 1.5

2.30E-06 1.5

2.20E-06 1

1.63E-06 2

1.65E-06 1

1.83E-06 1.5

1.59E-06 1

1.62E-06 2.5

1.14E-06 1.5

1.11E-06 1.5

2.26E-06 1.5

1.38E-06 1

1.36E-06 1

1.48E-06 1

1.50E-06 1.5

1.50E-06 2

1.72E-06 1.5

1.66E-06 2

2.39E-06 1

1.40E-06 1.2

1.39E-06 1.2

1 .69E-06 2

4 He . 4He 3 He . 

in error in 

ccSTP/g in% ccSTP/g

5.43 E-08 1

2.30E-07 3 

6.11 E-08 2

6.84E-08 2

7.75E-08 2

8.51 E-08 1

4.71 E-06 15

7.32E-08 1

2.35E-07 3

1.29E-07 2

7.53E-06 20

2.53E-07 6

1.91 E-06 9

1.13E-07 2

3.74E-08 1

6.25E-08 1

3.79E-07 3

8.69E-07 4

1.31 E-06 5

1.27E-06 4

9.05E-06 20

1.12E-07 1

2.12E-07 3

1.87E-07 3

8.67E-06 20

7.63E-14

1.69E-13 

8.75E-14

1.08E-13

1.78E-13

1.87E-13

7.69E-12

1.21E-13

4.30E-13

2.06E-13

1.22E-11

2.89E-13

2.11E-12

2.56E-13

5.15E-14

8.48E-14

5.60E-13

1.31E-12

1.97E-12

2.19E-12

1.51E-11

2.69E-13

2.96E-13

2.60E-13

1.46E-11

Ne. 

in 

ccSTP/g

2.35E-07

2.70E-07 

2.40E-07

2.41 E-07

2.50E-07

2.67E-07

2.70E-07

2.81 E-07

2.75E-07

2.50E-07

2.13E-07

2.19E-07

2.90E-07

2.42E-07

1.62E-07

2.63E-07

2.96E-07

3.42E-07

3.04E-07

3.10E-07

2.08E-07

2.64 E-07

3.45E-07

3.12E-07

2.43E-07

Ne 4 Heexc< 

error in

in % ccSTP/g

2

5

3

3

3

2

25

2

5

4

33

5

25

2

2

2

8

10

15

25

33

2

3

3

33

1.74E-10

1.66E-07 

5.49E-09

1.25E-08

1.91 E-08

2.18E-08

4.65E-06

5.79E-09

1.69E-07

7.11E-08

7.48E-06

2.04E-07

1.84E-06

5.69E-08

4.19E-09

2.96E-10

3.07E-07

7.84E-07

1.24E-06

1.20E-06

9.00E-06

5.00E-08

1.26E-07

1.10E-07

8.62E-06

SH , 3H/3hle Co"esp- 3H/ 3 He Corresp. 

in age lor a9e age lor a9e

1.65. 10'6 '" 2.10-8 '" 

In years *"** in years Vaars

13.30 1.2 1

15.60 log(neg. arg.) 

10.95 & 2 1

15.60 4.5

26.56 & 12.6

26.56 & 12.2

0.07 +

18.00& 6.3

12.60 19.3

1430 4

21.90 +

16.60 log(neg

14.10 +

27.70 14.5

16.40 0

12.20 0

13.40

6.57

1.96

21.80

12.90

21.60 19

15.00 "-29"

15.00 "-22"

0.76 +

1

2

2

1

1

1

.arg.)

2

1

1

2

1.3

20 

6.3

11

16.3

16.5

 f

88

44

27

+

33

 f

23

2.7'

0

27

31

30

+

1

2 

1 5

1

2

2

1

4

3

4.5

2

1

2

2

2

2

In ground water contained in at least 10 yearold bedrock, such as that found in the Mirror Lake 

Basin, the in situ source of excess He is expected to 

be due to radioactive decay of thorium (Th) and 

uranium (U) in the rock matrix, producing He with 

a 3 He/4He ratio of about 2x1 (T8 . The calculated 

amount of excess He in the water samples is given 

in table 1. 4Hecxc is dominated by the external He 

source; this leads to a first-order increase in the 

measured 4He concentration with depth as well as 

an increase in 4Hecxc (see fig. Ib). From the 

measured 3 H concentrations as well as other 

geochemical data and the CFC concentrations 

(Busenberg and Plummer, 1996), it is known that 

the apparent age of most of the analyzed samples 

should not exceed 50 to 100 years. With an assumed 

concentration of 2.8 ppm U, a Th/U ratio of 3.8 and 

a bedrock porosity of 0.05, steady state 4He 

production and total release to the fluid over a 

period of 100 yrs years would yield 4Hecxc values of 

about 10~'° ccSTP/g in the ground water (Torgersen 

and Clarke, 1985). The deep groundwater samples

have 4He concentration in excess of this value, and, 

therefore, the obtained data can not be explained 

without additional in situ sources such as those 

found by Torgersen and Clarke (1985), Torgersen 

and Ivey (1985), Torgersen and others (1989), 

Martel and others (1989), Stute and others (1992), 

Hiyagon and Kennedy (1992), Marty and others 

(1993).

Another possible explanation for the high 4He 

concentrations might be mixing of very young water 

(containing 3 H and tritiogenic 3He) and very old water 

with high He concentrations. This scenario is less 

probable for ground-water samples from the drift-zone 

but could be a potential scenario for ground water in 

the bedrock because of the irregular fractures in the 

bedrock. This might explain the close similarity of 

ground-water data from TR1 170-190 to those of the 

shallower well TR1 63, and the large difference to 

those of TR2 151-170 (see fig. I), which is only about 

200 m away. To produce the measured high He 

concentrations in ground water and rather young 

apparent (mixing) ages, very old ground water (high

161
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u.

£3

CO

Q 

Z

~

0

w

CO

OSw

UJ

S

5

X

w

Q

It)

u

Q

<

 J

2UJ

CQ

Dt

IX

S

5

X

fcm

Q

u

10

20

30

40

50

60

/ U

0 i

10

20

30

40

50

60

70

. |

-

1

~

-

-

'-

-

-

~

-

-

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O

to

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' . 1 1 1 ,,.,,, 1

 m

m " 

o>

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1 «

TR1 63 " i

CT

3

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"o

;§i

1

TR1 170-190 3

i , l . . ,...,. I

ID' 13 10' 12

3 He CONTENT,

v'.. .

 mm

A- .

TR1 63

«

TR1 170-190

,,!,.! . , 1 , , 1 ,

. 7

.,,.,,., , ,

-

t I

FSE 43 ;

-

-

-

i i '_ m TR2 151-170 ;

*T=X=i -^

w '-

-_

10' 11

IN ccSTP/g

FSE 43 ;

" ~

_

-

\

i i ]

'i A 'i

TR2 151-170 !

-

i 1 » i i i i . i 1

ft C 

j rt~ O ^ /\ **

Figure 1a. Changes of 3He concentrations are 

plotted against depth for ground-water samples from the Mirror Lake Basin. The vertical 

line at 1.65x10" 12 ccSTP/g marks the limit of 

3He that would be accumulated from decay of 

3H in water with 800 TU recharged in 1963. The 

bars represent the best estimate for the range 

in 3 He error; for low 3He concentrations, the 

error bars are smaller or equal to the symbol 

size.

FSE 43 : well name and depth at which water 

was collected.

ccSTP/g : cubic centimeters of gas at 0"C and 

1 atmosphere per gram of water.

10 10 10"

4 He CONTENT, IN ccSTP/g

Figure 1b. Changes of 4He concentrations are

plotted against depth for ground-water samples

from the Mirror Lake Basin. The bars represent

the best estimate for the range in 4He error; for

low 4He concentrations, the error bars are

smaller or equal to the symbol size.

FSE 43 : well name and depth at which water

was collected.

ccSTP/g : cubic centimeters of gas at 0°C and

1 atmosphere per gram of water.

A

He content) must be mixed with very young water 

that recharged around the time of the bomb-derived 3 H 

peak (1963). Such a mixing scenario is shown in 

figure 2. The resulting 3 He/4He ratios of water samples 

can be derived for different initial 3 H concentrations 

(fig. 2, numbers left of the different curves) and 

mixing ratios (fig. 2, numbers below the curves 

indicate the mixing ratio of old and young waters). 

Although most drift-well 3 He/4He ratios could, in 

principle, be explained by mixing (except the water 

sample from well FSR 43), most of the deeper wells 

would require mixing with waters with an 

unrealistically high initial 3 H concentration. Therefore, 

mixing is an unlikely explanation for the observed He 

distribution in ground waters in the Mirror Lake Basin. 

A plot of the He contents as a function of the

He contents (see fig. 3) for the samples shows that the 

He supplied to the deep wells is dominated by a source 

that can be approximated by a He isotope ratio of 

1.65xlO~6 . This ratio is much higher than the 

wellknown and expected one for radiogenic He

(typically around 2x10"x ) and, thus, an 'exotic' source 

of He is required to explain the measurements. Two 

sources could be assumed:

(1) Although the local intrusive rocks are older

it

than 10 years, they might contain trapped 

primordial gas with a high 3 He/4He ratio normally 

found only in young mantle material.

(2) There is a local, young source of He from the 

mantle that is outgassing, this He is mixing with He 

produced in local bedrock and is transported in 

ground water into the Mirror Lake Basin where it 

mixes with younger surface water in the bedrock.

In order to address the first possibility, rock 

cuttings from two drill cores (from wells at different 

sides of the basin) at depths >100 m were analyzed for 

the He and He content of their main components 

(basalt, granite, pegmatite, and schist). Samples were 

crushed to release the gas trapped in vesicles and fluid 

inclusions, and 3 He and 4He contents were measured 

(table 2). The measured 3 I Ie/4He ratios for bedrock are

y ~j

Rrock~ 2-2x 10" to 6x 10 ; these ratios are significantly

162
water mixing ratio 

young : old

1:0 1:1 1:2

Figure 2. Plot of the He isotope ratio as a function of the He/Ne ratio for ground-water samples from the Mirror Lake Basin. Beside the 

measured ground-water-sample data is shown 

the result for a mixing of (1) 30-year-old water 

with only a tritiogenic and no radiogenic 3He 

component and various initial tritium-derived 

3He values, with (2) old, radiogenic, i.e. 4Heenriched water with a typical 3He/4He ratio of 

2x10 8 . Mixing ratios are given along the x-axis 

corresponding to the calculated data points 

above (indicated by '+' for the 0 TU curve), 

'plus tritiog. He': The initial tritium-derived 3He 

content if no mixing with old waters has yet 

occurred.

He/Ne RATIO

2.0 10

1.5 1 0

1.0 10'

5.0 1 0"

.-11

Equation of line :

3 He=-6.52x 10" 1 4ccSTP/g+1.65x 10" Vhe 

correlation coefficient = 0.999

2 1 0

.-6 410"" 610"" 810 

4 He CONTENT, IN ccSTP/g

1 10

Figure 3. Plot of measured 3He- as a function 

of 4He-concentrations for the Mirror Lake Basin 

ground-water samples. Air saturation for 10°C 

is about 5x10"8 ccSTP/g for 4He and 6x10" 14 

ccSTP/g for 3He. The plot suggests admixture 

of a He component with an isotope ratio of 

1.65x10"6 . The bars represent the best estimate for the ranges in 3He and 4He errors; for 

low He concentrations, the error bars are 

smaller or equal to the symbol size. 

ccSTP/g : cubic centimeters of gas at 0°C andl 

atmosphere per gram of water.

Table 2. Helium-3/Helium-4 ratios and Helium-4 contents 

of bedrock samples

[ccSTP, cubic centimeters of gas at OoC and 1 atmosphere; grock, 

grams of bedrock sample material; 3He, helium isotope with the mass 

3; 4He, helium isotope with the mass 4]

Sample 

origin

ISI/390-395

FS2/330-340

IS 1 7440-445

FS2/

FSE I; 3/1 95-200

FS El 3/242-250

IS 1 12 \ 5-220

FS2/300-3IO

Sample 

material

basalt

basalt

granite

granite

pegmatite

pegmatite

schist

schist

3He/4He 4He 

x10'7 ccSTP/grockx10'7

6.00±0.30

5.IO+

2.60+

2. 1 6+

4.73+

4.l()±

0.22+

2.64+

.25

.125

.11

.235

.21

.015

.14

1.29

2.06

3.20

2.55

18.53

14.99

27.68

3.54

lower than those observed in waters from deep wells 

(1.65xlO~6) and inconsistent with our hypothesis 

mentioned above. Therefore, local rocks are not a 

viable source for the excess He component observed 

in the water samples. The fact that the measured 

bedrock ^He/^He ratios are higher than the typical

o

radiogenic isotope ratio of about 2x10 might at least 

partly be a consequence of the rock crushing 

procedure. Only part of in situ radiogenic He may be 

released during the crushing procedure, whereas 

primordial gas trapped in relativly larger vesicles 

would be more efficiently released; in fact, the basalts 

have the highest observed ^He/ He ratios. It is also 

possible that the rock samples were incompletely 

cleaned of an adsorbed air component resulting in a 

release of He with the observed ratio of around 2x10~7 . 

This possibility will have to be verified by melting 

rock samples. Then radiogenic He contained in the 

rock matrix will be released. It should be noted that 

the He concentrations in the rock samples are up to 

2.7x10"6 ccSTP/g rock. This very high concentration,

163
drift wells : age for R =1.65x10 

bedrock wells: age for Rexc=l.65x10' 6

drift wells : age for Rexc=2xlO" 

bedrock wells: age for R =2xlO~ 8

10 20 30 40 

3 H/3 He-BASEDAGES, IN YEARS

together with the observed He isotope ratio, might be 

an indication of in situ radiogenic He production 

combined with an 'exotic' He source (preferentially 

high in 3 He) and He diffusion into the rock from the 

fluid. This possibility also could be tested by stepwise 

heating and extraction of He from the rock samples.

The second possible scenario for the 'exotic' He 

component observed in the Mirror Lake ground-water 

samples could be degassing from deep, local mantlederived rock. He-isotope studies at numerous sites suggest that the occurence of mantle He is related to tectonically active regions (O'Nions and Oxburgh,l983; 

Hooker and others, 1985; Oxburgh and others, 1986; 

Oxburgh and O'Nions, 1987; O'Nions and 

Oxburgh, 1988; Stute and others, 1992; Marty and others, 1992; Griesshaber and others, 1992). There is no 

recent volcanic activity in the Mirror Lake region and 

only limited recent seismic activity. However, observations in the Paris Basin (Marty and others, 1993) suggest that the timescale for the degassing of the mantle 

He component is longer than hypothesized by O'Nions 

and Oxburgh (1988). If this is the case, the 'exotic' He 

may originate from an earlier event, for example the 

breakup of the North Atlantic. A clearer understanding 

of the regional source of'exotic' He and its transport 

into the ground-water from the Mirror Lake Basin will 

emerge when deeper ground-water samples are available for analysis and a regional vertical flux of radiogenic and exotic He can be determined.

Although the source of the excess He is 

unknown, ~ HA He-based ages calculated using an 

excess-helium ratio of Rcxc= 1.65x1 CT6 are shown in 

table 1. Reasonable agreement of the 3H/^He-based 

ages with CFC-based ages (Busenberg and Plummer, 

1996) indicates that this approach is not unreasonable. 

However, if 4Hecxc exceeds 4Hea j r by more than a

Figure 4. Plot of excess helium (Heexc) as a 

function of 3H/3He based ground-water ages for 

the Mirror Lake Basin samples for the two 

assumptions of the 3He/4He ratio of the excess 

helium source, Rexc . It could be that the best 

assumption for Rexc varies from sample to sample and the plotted results of the two assumptions reflect only an age range. Note, that the 

4He concentrations (see fig.1B), and, at least 

tendentially, the 3H/3He-based ages (and CFCbased ages, not shown here) increase with 

depth in the aquifer.

factor of about 4 to 5, reliable dating is not possible. 

Because of the heterogeneity in the bedrock and the 

drift zone it can not be excluded that (for example) the 

drift-water samples have to be treated differently. 

Therefore 3H/3He-based ages calculated with 

Rexc=Rrad=2x 10"8 are also shown in table 1. The 

contribution of the two discussed He sources with 

Rcxc=1.65xlO~6 and Rcxc= Rrad=2xl(Tx , respectivly, 

cannot be determined and might vary. It should be 

noted that the resulting ages are very sensitive to 

changes in Rcxc if that ratio exceeds about IxlO"6 .

Because the calculated ground-water ages and 

the He content of most of the samples increase with 

depth, the He concentration itself may carry age 

information, as shown in figure 4 where the 

calculated 3 H/3 He-based ages are plotted against 

measured 4He concentrations. The shape of that curve 

is nearly exponential and may reflect a flux of He 

through the local bedrock. This hypothesis would be 

supported by more data points from deeper wells; by 

use of CFC data (Busenberg and Plummer, 1996); and 

additional age-dating tools, such as 85Kr-based age 

dating (Smethie and others, 1992) or flow-path 

analysis using other dissolved chemical compounds 

and the results of the hydraulic tests.

REFERENCES

Busenberg, E. and Plummer, N.L., 1991,

Chlorofluorocarbons (CC12 F2 and CCl^F): Use as an 

age-dating tool in hydrologic tracer studies in shallow 

groundwatcr systems, in Mallard, G.E., and Aronson, 

D.A., cds., US Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the Technical 

Meeting, Montcrcy, California, March 11-15 1991: 

U.S. Geological Survey Water Resources 

Investigations Report 91-4034, p. 1-6.

164
Buscnbcrg, E. and Plummcr, N.L., 1992, Use of 

Chlorofluoromcthancs (CCl^F and CCl2F2 ) as 

hydrologic tracers and age dating tools. Example- The 

alluvium and terraces system of Central Oklahoma: 

Water Resources Research, v. 28, p. 2257-2283.

Buscnberg, E., and Plummcr, L.N., 1996, Concentrations of 

chlorofluorocarbons and other gases in ground waters 

at Mirror Lake, New Hampshire, //; Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological 

Survey Toxic Substances Hydrology Program-- 

Proceedings of the Technical Meeting, Colorado 

Springs, CO, September 20-24 1993; US Geological 

Survey Water Resources Investigations Report 94- 

4015.

Chamberlain, C.P., and Sender, L.J., 1990, Heat-producing 

elements and the thermal and baric patterns of 

metamorphic belts: Science, v. 250, p. 763.

Griesshabcr, E., O'Nions, R.K., and Oxburgh, E.R.,1992, 

Helium and Carbon isotopes systematics in crustal 

fluids from the Eiffcl, the Rhine Grabcn and Black 

Forest: Chemical Geology, v. 99, p. 213-235.

Hiyagon H., and Kennedy, B.M., 1992, Noble Gases in 

CH4-rich gas fields. Alberta, Canada: Geochimica ct 

Cosmochimica Acta, v. 56, p. 1569-1589.

Hooker, P.J., Bcrtami, R., Lombardi, S., O'Nions, R.K., and 

Oxburgh, E.R., 1985, ^He anomalies and crust mantle 

interaction in Italy: Gcoehimica ct Cosmochimica 

Acta, v. 49, p. 2505-2513.

Martcl, D.J., Dcak, J., Dovcnyi, P.. Horvath, F., O'Nions, 

R.K., Oxburgh, E.R., Stegcna, L., Stutc, M., 1989, 

Leakage of helium from the Pannonian Basin: Nature, 

v. 343, p. 908-912.

Marty, B., O'Nions, R.K., Oxburgh, E.R., Martcl, D., and 

Lombardi, S., 1992, Helium Isotopes in Alpine 

Regions: Tcctonophysics, v. 206, p.71-78.

Marty, B., Torgcrscn, T., and Mcynicr, V., 1993, He Isotope 

fluxes and ground water ages in the Dogger aquifer. 

Paris Basin: Water Resources Research (in press).

O'Nions, R. K. and Oxburgh, E.R., 1983, Heat and He in the 

Earth: Nature, v. 306, p. 429-431.

O'Nions R.K. and Oxburgh, E.R., 1988, He, volatile fluxes 

and the development of continental crust: Earth and 

Planetary Science Letters, v. 90, p. 331-347.

Oxburgh, E.R., O'Nions, R.K., and Hill, R.I., 1986, Helium 

isotopes in sedimentary Basins: Nature, v. 324, p. 632- 

635.

Oxburgh, E.R. and O'Nions, R.K., 1987, Helium loss, 

tectonics and the terrestrial heat budget: Science, 

v. 237, p. 1583-1588.

Porcda, R.J., Ccrling, T.E., Salomon, O.K., 1988, Tritium 

and helium isotopes as hydrologic tracers in a shallow 

unconfincd aquifer. Journal of Hydrology, v. 103, p. 1-9

Schlosscr, P., Stutc, M., Dorr, H., Sonntag, C., and

Munnich, K.O., 1988, Tritium/3 Hc dating of shallow 

groundwater: Earth and Planetary Science Letters, 

v. 89, p. 353-362.

Schlosscr, P., Stutc, M., Sonntag, C., and Munnich, K.O., 

1989, Tritiogcnic 3Hc in shallow groundwater: Earth 

and Planetary Science Letters, v. 94, p. 245-256.

Shapiro, A.M., Wood, W.W., Buscnbcrg, E., Drcnkard, S., 

Plummcr, L.N., Torgcrscn, T., Schlosscr, P., 1996, A 

conceptual model for estimating regional fluid velocity 

in the bedrock of the Mirror Lake area, Grafton 

County, in Morganwalp, D.W., and Aronson, D.A., 

cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, CO, September 20-24 

1993; US Geological Survey Water Resources 

Investigations Report 94-4015.

Solomon, D.K., and Sudicky, E.A., 1991, Tritium and 

Hclium-3 Isotope Ratios for Direct Estimation of 

Spatial Variations in Groundwater Recharge: Water 

Resources Research, v. 27, p. 2309-2319

Smcthic, W.M. Jr., Solomon, D.K., Schiff, S.L., and

Mathicu, G., 1992, Tracing ground-water flow in the 

Bordcn aquifer using kryton-85: Journal of Hydrology, 

v. 130, p. 279- 297.

Stute, M., Sonntag, C., Deak, J., and Schlosscr, P., 1992, He 

in deep circulating ground water in the Great 

Hungarian Plain: Flow dynamics and crustal and 

mantle helium fluxes: Gcoehimica ct Cosmochimica 

Acta, v. 56, p. 2051-2067.

Tolstikhin, N.I. and Kamcnsky, l.L.,1969, Determination of 

groundwater ages by the T^Hc method. Geochemistry 

International, v. 6, p. 810-811

Torgerscn,T., Clarkc, W.B., and Jcnkins, W.J., 1978, The 

tritium-helium-3 method in hydrology, In: Isotope 

Hydrology (IAEA Wicn), v. II, p. 917-930.

Torgcrscn, T. and Clarkc, W.B., 1985, Helium accumulation 

in groundwater, I an evaluation of sources and the 

continental flux of crustal 4Hc in the Great Artesian 

Basin, Australia: Gcoehimica ct Cosmochimica Acta, 

v. 49, p. 1211-1218.

Torgcrscn, T. and Ivcy, G.N., 1985, Helium accumulation in 

groundwater, II a model for the accumulation of 

crustal 4Hc degassing flux: Geochimica ct 

Cosmochimica Acta, v.49, p. 2445-2452.

Torgcrscn, T., Kennedy, B.M., Hiyagon, H., Chiou, K.Y., 

Reynolds, J.H., Clarkc, W.B., 1989, Argon 

accumulation and the degassing flux of 40Ar in the 

Great Artesian Basin, Australia: Earth and Planetary 

Science Letters, v. 92, p. 43-56.

165
166
Observational, Experimental and Inferred Evidence for

Solute Diffusion in Fractured Granite Aquifers: 

Examples from the Mirror Lake Watershed, Grafton

County, New Hampshire

By Warren W. Wood1 , Alien M. Shapiro1 , Paul A. Hsieh2 , and Jerry B. Councell 1

Abstract

The role of solute diffusion between ground 

water and granite in the Mirror Lake drainage area 

was evaluated by direct observation, experiment, 

and inference. The outcrops display ubiquitous 

Liesegang bands associated with fractures that 

clearly indicate the activity of diffusion in this system in the past. Laboratory experiments determined 

that the effective diffusion coefficient for l37Cs was

1-5 ^

approximately 6 x 10 m /s in granite from Mirror 

Lake. The Cs penetrated to a depth of approximately 7 mm in 101 days, demonstrating the potential for rapid diffusion in this system. Porosities of 

32 granite samples averaged of 1.46 percent with a 

range of 1.07 to 2.32 percent. Measurements of carbon-isotope of ground water in the fractures suggest 

that calcite, identified in the granite, is dissolving 

and the bicarbonate generated is diffusing to the 

fractures; that is, a significant amount of the dissolved solids in the water in the fractures are derived 

from diffusion of weathering products from the rock 

matrix. These observations taken together are consistent with the interpretation that diffusion is a 

major process controlling solutes in this fractured 

granite aquifer.

INTRODUCTION

Diffusion can be an important mechanism controlling solute transport in ground water in fractured crystalline rocks even if ground water residence times in the 

fractures are relatively short. In order to evaluate effects 

of diffusion in a fractured rock system, we studied a 

natural, hydrologically and chemically unstressed 

igneous and metamorphic terrane in the Mirror Lake 

watershed near West Thornton, Grafton County, New 

Hampshire (42° 56' N, 71° 43' W). In this paper we first

'U.S. Geological Survey, Rcston, Va. 

2 U.S. Geological Survey, Mcnlo Park, Ca.

visually and petrographically describe the granite for evidence of diffusion, conducted laboratory experiments to 

determine diffusion coefficients and evaluated solute 

concentration in the fractures for evidence of diffusion.

The geology, and general hydrology of this site is 

described by Winter (1984). Low permeability glacial 

tills provides storage whereas the fractures in the underlying crystalline rock are the main conducts for regional 

ground-water flow. The residence time for flow in fractures in the upper 150 m of the crystalline rock is relatively short. Ninety six of the 100 ground-water 

samples collected from the fractured bedrock indicate 

residence times of less than 50 years on the basis of concentrations of chlorofluorocarbons, manufactured since 

the late I930's, and tritium from atmospheric testing of 

thermonuclear devices in the 1950's and early 1960's. 

These samples were collected from hydraulically isolated intervals in bedrock wells at depths ranging of 50 

to 150 m below the rock surface.

OBSERVATIONAL EVIDENCE FOR 

DIFFUSION

Observational evidence for diffusion in this 

system was obtained at the 1-93 (Interstate Highway 93) 

road cut on the eastern side of the Hubbard Book drainage area. At this exposure, numerous conspicuous ferric hydroxide "Liesegang bands" are formed parallel to 

many of the fractures and are known to form as a result 

of the diffusion processes (Berner, 1971). In this system, we hypothesize that the bands of ferric hydroxide 

are formed by the reaction of ferrous iron, produced by 

the weathering of the ubiquitous am-phiboles and 

biotites, with oxygen as it diffuses into the rock from 

the fracture. Analytical solution of the diffusion equation, with appropriate boundary and initial conditions, 

demonstrates that the ferric hydroxide bands can form 

in response to fluctuating oxygen concentration in the 

fracture. Furthermore on the basis of measured diffusion coefficients and observed concentrations of iron 

and dissolved oxygen the bands will form in times of 

tens to hundred of years. Thus, diffusion has operated

167
.,i these rocks on time scales that are of importance for 

toxic or radioactive-waste disposal or remediation.

EXPERIMENTAL EVIDENCE FOR 

DIFFUSION

Experimental evidence in support of diffusion 

was collected in a series of laboratory experiments on 

representative samples of granites, schists, and basalts 

from the Mirror Lake area. The diffusion coefficient 

was measure on 3-cm cubes cut from the unweathered 

part of large samples collected from fresh outcrops at 

the 1-93 road cut. One face of the cube was polished 

with fine 1200 mesh carborudum grit, washed in distilled water, dried in a vacuum oven at 50 °C for 3 days 

and weighed. The cubes were then placed in a CO2 - 

pressurized chamber for 1 week to replace the slightly 

soluble air with highly soluble CO2 . Samples were then 

placed in a distilled-water bath at room temperature for 

6 months and then, removed; their surfaces were then 

blotted dry and the sample weighed. Background radioactivity was measured using a sodium iodide crystal and 

multi-channel analyzer in the energy range of Cs 

(cesium-137). The five unpolished surfaces on each 

cube were covered with paraffin wax and placed in a 

solution containing ll7Cs with an activity of 26 disintegrations per milliliter per second.

After 101 days, three replicates of each rock type 

were withdrawn, the wax was removed, and the five 

waxed sides of the cube were lightly ground using a diamond surfaced lapidary wheel to remove any ll7Cs that 

might have penetrated between the wax and the rock. 

The rock was then dried in a vacuum oven at 50 °C for 

3 days, stopping the diffusion of l37Cs until the sample 

could be analyzed. Radioactivity of the samples was 

then measured and distance determined from the 

exposed polished face to the back face by means of a 

micrometer. The polished face of the sample was then 

lightly ground using a coarse, 100-mesh lapidary polishing wheel, measured with the micrometer to determine thickness of material removed, weighted and 

radioactivity measured. This procedure was repeated 

numerous times until there was no change in the activity 

of Cs with successive grindings.

Initially there was concern that polishing the face 

would create micro fractures that would artificially 

increase the diffusion rate. To evaluate this, we performed a replicate experiment similar to that described 

above only with pairs of cubes, one having a natural 

fracture face and the other a polished face. After 3 days 

of soaking, the cube with the polished face had greater 

depth penetration of l37Cs, however, after 10 days the 

penetration was the same in both cubes. Thus, we conclude that for our relatively long soaking period that we 

did not significantly biases the sample by polishing the

3 x 10" 13 m 2 /S

02468

DEPTH OF PENETRATION, IN MILLIMETERS

Figure 1. Graph showing the activity of cesium-137 observed 

in the rock (A) to activity in the solution (A0) as a function of the 

depth of penetration of the cesium-137 for granite from Mirror 

Lake after 101 days of exposure. Also illustrated is a modeled 

range of effective diffusion coefficients, in m2/s, that bracket the 

observed data.

exposed face. However, it can be seen in figure 1 that 

the activity ratio for penetration depths less than 3 mm 

are slightly higher than the theoretical curve suggesting 

some effect of polishing, but the effect is small and does 

not alter significantly the overall fit of the observed data 

to the theoretical curve.

To evaluate the effective diffusion coefficient 

from these experiments, consider the equation of mass 

conservation for a sorbing solute,

r>(OC

d\

where C is the aqueous phase concentration of the solute (M/L3 ), 6 is the porosity (L3 /L3 ), p is the bulk density of the solid (M/L3 ), s is the mass of the sorbed 

solute per unit mass of the solid phase (M/M), D is the 

effective diffusion coefficient (LYT), t is time (T), and 

x is the spatial coordinate (L). M, L and T denote units 

of mass, length and time, respectively. If the concentration of the sorbed solute is proportional to the 

aqueous phase concentration, s = kC, where k is the 

sorption coefficient (LVM), the equation of mass 

conservation can be rewritten as

ac 

at

Jl (12C

R6 dx 2 (2)

where R is the retardation coefficient (1+ pk/6).

Because diffusion does not penetrate far we may 

assume a semi-infinite domain. With an initial condition 

of C=0 and boundary conditions C(x=0,t)=C0 and dC/dx 

(x. »oo,t) = 0, the solution to the previous equation is

168
C(x,t) = C0 erfc[

L2V(Dt/R0)J (3)

where erfc is the complementary error function.

Because of the experimental procedure of drying 

the samples prior to measurements, the activity of Cs 

is measured on the solid only which is denoted as m = 

0C(x,t) + ps = 6RC(x,t). Using equation (3), the combined concentration of the fluid and solid phase concentration is

m(x,t) = 0RC erfc[~

|_2V(Dt/R0 ) (4)

Assuming the L17Cs activity is directly proportional to the l37Cs concentration, then the ratio of activities, A/A0, is equal to the ratio of concentrations, 

m(x,t)/0RCo. Because the sorption coefficient, k, is 

large, the i:?7Cs sorbed on the solid phase is large in 

comparison to that in solution. Thus, A0 is defined as 

the difference between the initial activity of the entire 

rock sample and the activity measured after the first 

grinding of the rock face. That is, it is assumed that the 

first section of the rock face is in equilibrium between 

the solute and the solid. The activity of the remaining 

sections of the rock are defined by the difference 

between the activities of the successive grindings of the 

rock sample. These values are associated with positions 

that are 1/2 the measured distance between successive 

sections of the removed rock.

Results of this diffusion experiment for a granite 

sample are shown in figure 1, where the relative activity, A/A0, of l37Cs is plotted as a function of the depth 

of penetration into the sample. The activity is expressed 

in disintegration's per second per gram of removed 

solid. The coefficient which best determined the spatial 

distribution shown in figure 1 is 6 X 10~ 13 m2/s. It can 

be seen that this value is well constrained within a narrow range of potential values. This value incorporates 

the effects of the porosity and tortuosity of the void 

space in the sample, as well as the retardation of 137Cs. 

This value is similar to the value of 1 X 10' 1 ' m2/s 

which was determined by Skagius and Neretnieks 

(1988) for 137Cs in a biotite gneiss from Sweden.

The effective diffusion coefficient, D/R0, for 

Cs in basalt was determined by the same method and 

was found to average 2 X 10~ 14 m2/s. The difference 

between the effective diffusion coefficient of the granite 

and basalt samples is related to the difference in porosity, tortuosity and retardation coefficient. The porosity 

of the granite is 4 to 5 times larger than that of the basalt 

and is the major factor in the differences.

Porosities were determined from the difference 

between wet and dry weights of the sample cubes

described above and assuming a bulk density for the 

samples [granite 2,640 kg/m3 ; basalt 2,850 kg/m3 ]. 

Average porosity for 32 samples of granite was 1 .46 

percent with a range from 1 .07 to 2.33 percent. The 

average porosity for 15 samples of basalt was 0.36 percent and ranged from 0.58 to 0.24 percent. Porosity of 

23 samples of schist averaged 1 .87 percent and ranged 

form 1 .07 to 3.36 percent.

INFERRED EVIDENCE OF DIFFUSION

Evidence of inferred diffusion in this system is 

shown by concentrations of bicarbonate, stable carbon 

isotopes, and presence of calcite in the rock. Alkalinity 

in this geochemical system is almost exclusively a function of the bicarbonate ion (HCO3 -) that formed by 

weathering reactions of carbonate and silicate minerals 

with soil carbon dioxide. For example, the weathering 

of calcite (CaCO3 ) to Ca++ (calcium) and bicarbonate 

(HCOr) carbonic acid (H 2CO3 ) is given by 

C02 + H 2O

CaCO3 + H2CO3 = 2HCO3 - (5)

We observed concentrations and isotopic signatures of bicarbonate that could only be the result of dissolution of carbonate minerals. Yet there is no calcite or 

other carbonate mineral presently in the fractures. If 

there were some at one time, they have been dissolved. 

Thus, we hypothesize that bicarbonate is generated by 

weathering of calcite in the rock and is diffusing from 

the rock matrix to the fracture.

Calcite was identified in small grains in the rock 

matrix by petrographic evaluation, scanning electron 

microscope (SEM) and energy dispersive X-ray analysis (EDAX ) of cores and outcrop samples. Total carbon analysis of a granite sample from the 1-93 road cut 

yielded 0. 1 2 percent carbon. It is assumed that this carbon is from the calcite identified by SEM and EDAX.

Evidence for dissolution of calcite is shown by 

the increase in isotopically heavy carbon with increasing alkalinity (fig. 2). Water in this system is generally 

thermodynamically under-saturated with respect to calcite below an alkalinity of approximately 125 mg/L 

thus, if calcite were present it would dissolve. The average stable carbon isotopes in five soil CO2 gas samples 

collected in the watershed had a 5 |:*C of -22 %o relative to that in belemnitella americana from the Cretaceous Pedee Formation (PDB). Because there is a 

fractionation factor of approximately 8 °/oo at the ambient temperature of this system, bicarbonate ions in the 

water will have a 8 I3C value of approximately -14 °/oo 

in equilibrium with a soil gas of -22 %o. Thus, weathering of silicate mineral by carbonic acid would yield a 

8 C in the water of -14 °/oo because there is no other

169
3UO L ' ' ' ' 1 '

p

250 |r

»-

. \

m j

Soil Gas

2 5 t > i 150 t"

5 1 L

3 g 100 r

§ d : 

< \ 50 ir

 Equilibrium with Calotte * ! 

-30 -25 -20 -15 -10

1 3.

-5

CARBON ISOTOPES, AS &' C

Figure 2. Graph showing alkalinity, as CaCOS as a function 

of stable carbon isotopes in water from 38 wells in the Hubbard Brook drainage basin. Also shown is the average carbon isotopic value of soil gas in the drainage basin and the 

approximate alkalinity concentration at which the water is in 

thermodynamic equilibrium with calcite.

source of carbon. However, if weathering of a carbonate mineral by carbonic acid occurs, there are two 

source of carbon: one from the carbonic acid and one 

from the carbonate mineral. Because carbonate minerIT I ->

als typically have a 8 C near zero, the C dissolved in 

the water as the bicarbonate ion reflects both sources of 

carbon as indicated in the above equation. Thus, at 

equilibrium, the fi'^C carbon isotopes determined in the 

water will have a value near -7 o/oo (one carbon atom 

from the mineral 8 C = 0 and one from the carbonic 

acidS l3C = 14).

Also consistent with diffusion of bicarbonate is 

the strong correlation between residence time determined from chlorofluorocarbon dating (Busenberg and 

Plummer, 1992) and concentration of alkalinity (fig. 3), 

that is, the younger water is thermodynamically undersaturated with respect to calcite, which would rapidly 

dissolve if exposed to this water. The fact that it has 

taken tens of years to approach equilibrium is consistent 

with transport by diffusion.

We propose the following generalized conceptual model for diffusion in this crystalline fractured system. Overburden water, low in total dissolved solids, 

alkalinity, pH and thermodynamically undersaturated 

with respect to calcite and saturated with dissolved oxygen enters the fractures. Oxygen and carbonic acid diffuses from the fracture into the rock matrix where 

biotites, pyroxenes, calcite, and feldspars are weathered 

consuming oxygen and carbonic acid and forming ferrous iron and bicarbonate ions among other products. 

Interaction of the iron diffusing from the rock matrix to 

the fracture with the oxygen diffusing from the fracture 

to the rock matrix, causes iron to precipitate as ferric

2000

1990

o

< 1980

oLU

oc

1970

1960

1950

0 50

ALKALINITY, AS

100

CaCO 3'

150 200 250

IN MILLIGRAMS PER LITER

Figure 3. Graph showing year of recharge, as a function 

of alkalinity for ground water from wells sampled in 1991 in 

the Mirror lake drainage basin.

hydroxide and form the observed Liesegang bands. 

Rate of weathering, diffusion rates and concentration of dissolved oxygen in the fracture control distance of the bands from the fracture. Bicarbonate 

generated by dissolution of calcite diffuses to the 

fracture giving the observed 8 C signature.

SUMMARY

In summary, three independent lines of evidence 

suggest significant solute diffusion in this system. Paleodiffusion in the form of ferric hydroxide Liesegang 

bands is consistent with assumption of diffusion into 

and out of the rock matrix. Laboratory determined diffusion coefficients and porosity values suggest that the 

observed concentration gradients will provide significant flux of solutes into and out of the rock in a relatively short time. And finally stable isotopes of carbon 

provide evidence of calcite dissolution while SF^M and 

petrographic analyses identified calcite in the granite.

REFERENCES CITED

Bcrncr, R.A., 1971, Principles of Chemical Scdimcntology: 

New York, McGraw-Hill, 240 p.

Busenberg, Kurybiades and Plummer, N.L., 1992, Use of 

chloroflurocarbons (CCI^ and CCbF^) as hydrologic 

tracers and age-dating tools: The alluvium and terrace 

system of central Oklahoma: Water Resources Research, 

v. 28, no. 9, p. 2257-2284.

Crank,.!., 1975, The mathematics of diffusion: Oxford, Claredon Press, 414 p.

Skagius, Kristina and Ncrctnicks, Ivars, 1988, Measurements 

of cesium and strontium diffusion in biotitc gneiss: 

Water Resources Research, v. 24, no. I, p. 75-84.

Winter, T.C., 1984, Gcohydrologic setting of Mirror Lake, 

West Thornton, New Hampshire: U.S. Geological Survey 

Water-Resources Investigations Report 84-4266, 61 p.

170
A Conceptual Model for Estimating Regional Ground

Water Velocity in Bedrock of the Mirror Lake Area,

Grafton County, New Hampshire

By Alien M. Shapiro1 , Warren W. Wood 1 , Eurybiades Busenberg 1 , Stefan Drenkard2 , 

L. Niel Plummer1 , Thomas Torgersen3 , and Peter Schlosser2

Abstract

Estimation of regional ground-water velocities requires a knowledge of the residence time of 

the ground water and the distance it has traveled. 

Ground-water residence times are estimated in the 

Mirror Lake area on the basis of concentrations of 

chlorofluorocarbons and the parent-daughter isotopes tritium and helium. The distance the ground 

water has traveled, however, cannot be identified 

in the Mirror Lake area because there is no single 

area of recharge to the bedrock and the heterogeneity of hydraulic properties of the bedrock indicate that a regional flow line cannot be 

conceptualized from successive down-gradient 

sampling locations. Measurements of alkalinity in 

water samples collected from fractures in the bedrock are positively correlated with ground-water 

age. The alkalinity is controlled by the concentration of bicarbonate ions, and carbon isotopes indicate that the bicarbonate concentration is the 

result of dissolution of calcite in the rock matrix. 

The diffusion of bicarbonate ions from the rock 

matrix to fractures is hypothesized as a second 

indicator of the residence time of ground water in 

the bedrock. A simple model of ground-water 

flow and the transport of bicarbonate ions in the 

bedrock is proposed to investigate the relationsamong ground-water velocity, residence time 

and alkalinity. The model consists of a single fracture in contact with a rock matrix from which the 

average ground-water velocity in the bedrock can 

be estimated.From this simple conceptual model,

'U.S. Geological Survey, Reston, Va.

2 Lamont-Doherty Earth Observatory, Department of 

Geochemistry, Palisades, N.Y.

 University of Connecticut, Department of Marine Sciences, 

Groton, Conn.

the length of paths of fluid movement in the bedrock and the average bedrock velocity can be estimated. This conceptual model, however, must be 

reconciled with other geochemical data that indicates that ground waters of various ages are mixing either naturally or as a result of ground-water 

sampling.

INTRODUCTION

In shallow ground-water systems, concentrations of chlorofluorocarbons (CFC's), and the parentdaughter isotopes tritium and helium can be used to 

estimate ground-water ages (Busenberg and Plummer, 

1992; Schlosser and others, 1988; Solomon and others, 1992). Ground-water ages estimated by these 

environmental tracers represent the residence time of 

the water since it was last in contact with the atmosphere. However, the distance the ground water has 

traveled within that time also must be identified to 

estimate the ground-water velocity. In the absence of a 

single, well-defined recharge location, additional 

hydrologic, geologic, and geochemical information 

must be used to infer the length of the ground-waterflow paths. In this article a simple model is conceptualized to estimate rates of ground-water flow using 

geologic and geochemical information, and the results 

of CFC and tritium-helium age dating of water samples collected from the glacial drift and the upper 200 

m of the bedrock in the Mirror Lake area in Grafton 

County, New Hampshire (fig. I).

SITE DESCRIPTION

The Mirror Lake watershed is located at the 

mouth of the Hubbard Brook valley in the southern 

part of the White Mountains of central New Hampshire (fig. 1). Its area is approximately 85 ha; Mirror 

Lake covers an area of approximately 15 ha. The elevation of Mirror Lake is approximately 213m above 

sea level; the highest elevations in the watershed

171
7 .,, 2 . ,.,,. ,c .

 Well site

« Well field

R1

FS3 ^^^TRI

ci/ TR2^

New /

Hampshire)

j m

Map 'area

/ \

 ^

FS2

H1

 C30 i:OIF£ET

300 e:: METERS

Figure 1. Location of the Mirror Lake in New Hampshire and 

location of bedrock wells in the Mirror Lake area.

(approximately 469 m) are located north and west of 

the lake. The bedrock in the Mirror Lake area is a sillimanite-grade schist that has been extensively intruded 

by granite and pegmatite and lesser amounts of basalt. 

Successive orogenies have caused folding and faulting 

resulting in extensive fracturing of the bedrock. Roadcuts exposed along Interstate Highway 93, east of Mirror Lake, show poorly connected fractures and 

irregular configurations and dimensions of schist, 

granite, pegmatite and basalt (Barton, 1996). In most 

areas of the Mirror Lake watershed, the bedrock is 

overlain by unconsolidated glacial deposits (drift) up 

to 50 m thick. In general, ground water flows from 

higher elevations north and west of Mirror Lake to 

lower elevations, where ground-water discharges, 

such as Mirror Lake, perennial streams in the vicinity 

of Mirror Lake, and the Pemigewasset River east of 

Mirror Lake. A more detailed discussion of the hydrogeology of the Mirror Lake area is provided by Winter 

(1984; 1985).

Wells in the bedrock and piezometers in the glacial drift have been installed in the Mirror Lake area to 

investigate regional ground-water flow (fig. I). Each 

bedrock well is drilled and cased through the glacial 

drift and the upper 3 m of the bedrock, and a 0.15-mdiameter open hole extends into the bedrock to the 

desired depth. Depths of bedrock wells in the Mirror 

Lake area range from 60 to 230 m. At most bedrockwell sites, one or more 0.05-m-diameter drift piezometers, each with a 0.6-m-long screened interval, were 

installed at various depths in the saturated part of the 

drift above the bedrock surface. A petal-cement basket

is installed above the screened interval and the annular 

space between the piezometer casing and the drill hole 

is grouted (Winter, 1984). Water-table wells in the 

drift (the screens of which intersect the water table) 

also were installed in the vicinity of most bedrock 

wells. Water levels have been recorded in water-table 

wells, drift piezometers, and in hydraulically isolated 

intervals of bedrock wells (Hsieh and others, 1996).

GROUND-WATER AGES

Water samples for geochemical and isotopic 

analysis (including concentrations of CFC's, tritium, 

and helium) were collected from selected fractured 

intervals in 14 bedrock wells and from a selected number of drift piezometers in the Mirror Lake area. In 

bedrock wells, water samples were collected by means 

of a piston pump located between packers that isolated 

a discrete, permeable interval of the well. In newly 

drilled bedrock wells, water samples were collected 

from isolated permeable intervals immediately after 

drilling and after the well was purged. In existing bedrock wells, the permeable intervals where samples 

were collected were isolated for up to a year by use of 

inflatable packers to prevent the short-circuiting of 

water from fractures in which hydraulic heads differed. Water samples also were collected from drift 

piezometers by means of a submersible or peristaltic 

pump. Of the more than 80 discrete bedrock and overburden intervals that were sampled, only 10 of the 

intervals yielded ground water containing no tritium 

which indicated that the ground-water ages exceeded 

45 years.

A preliminary interpretation of ground-water 

ages was made on the basis of concentrations of 

dichlorodifluoromethane (CFC-12), trichloro-fluoromethane (CFC- 11), and trichlorotrifluoroethane 

(CFC-113) (Busenberg and Plummer, 1996). Estimates of ground-water ages from concentrations of the 

parent-daughter isotopes tritium and helium were 

made for samples from drift piezometers and the shallowest intervals of the bedrock wells (Drenkard and 

others, 1996). The ground-water ages estimated from 

the tritium-helium and CFC analyses compare well 

with one another (Drenkard and others, 1996). However, the estimated ages do not correspond well with 

the presumed date of the maximum known atmospheric-tritium concentration. Both the tritium-helium 

and CFC interpreted ages appear to be approximately 

6 years younger than ground-water ages that would be 

estimated by using the maximum known tritium concentration as the date of the maximum atmospheric tri172
tium concentration (Busenberg and Plummer, 1996). 

This difference is still being investigated and may be 

the result of ground-water flow through the unsaturated zone, where the air phase can allow further equilibration with the atmosphere, or the difference may be 

a result of the mixing of waters of various ages (either 

naturally or during the collection of water samples). 

The ages of ground-water samples collected in 

July and August of 1992 from bedrock wells R-l, 

TR-2 and T-lare shown in figure 2. Younger groundwater ages were found at several deeper sampling 

locations. In addition, younger ground-water ages 

were found at locations closer to Mirror Lake which is 

an area of ground-water discharge. The variation in 

hydraulic head reflects changes in topography, implying that the direction of ground-water flow in the bedrock is toward Mirror Lake. In addition, the hydraulic 

head data suggest that the cross-section through the 

three bedrock wells in figure 2 could be hypothesized 

as a regional flow line. However, the ground-water 

ages indicate that fluid movement in the bedrock is a 

complex, three- dimensional process where a line of 

wells cannot be conceptualized as being along a 

regional flow line. The complexity of the flow regime 

is also supported by the variability of hydraulic properties of the bedrock (Hsieh and Shapiro, 1996), and 

the poor fracture connectivity that has been demonstrated in mapped fractures on bedrock outcrops and 

road cuts in the Mirror Lake area (Barton, 1996).

GROUND-WATER CHEMISTRY AND 

RESIDENCE TIME

Because recharge is areally distributed, the distance the ground water has moved to reach each sampling location is not identifiable. In addition, the 

distance between successive down-gradient sampling 

locations and the differences in ground- water ages at 

sampling points can not be used to estimate the 

ground-water velocity, because successive down-gradient sampling locations can not be interpreted as a 

regional flow line because of the heterogeneity of the 

hydraulic properties of the bedrock. In order to 

address the question of the length of flow paths associated with the sampling locations where ground-water 

ages have been estimated in the Mirror Lake area, we 

consider other geochemical evidence to indicate 

ground-water residence times.

Evidence indicates that diffusion from the rock 

matrix to fractures plays an important role in the 

chemistry of the ground water in the bedrock of the 

Mirror Lake area (Wood and others, 1996). The total

300

uj 250

LU

200

w

ujoc 150 *|2

LULL)

o? 100

co£

50

0 

-50

DRIFT

MIRROR LAKE

BEDROCK

0 100 200

HORIZONTAL SCALE,

IN METERS

EXPLANATION

T-1 BOREHOLE IDENTIFIER. 

BOREHOLE.

LOCATION OF CROWD- , 

»AT£R SAMPLE AND 

GROUND-WATER ACE

Figure 2. Ground-water ages (in years) estimated from CFC12 concentrations from water samples collected in July and 

August 1992 in bedrock wells in the Mirror Lake area.

dissolved solids concentration in water samples collected from overburden piezometers averages approximately 50 mg/L , whereas the total dissolved solids 

concentration in water collected from fractures in the 

bedrock is significantly higher, averaging approximately 250 mg/L. Residence times of less than 45 

years in the shallow bedrock are not sufficient for the 

dissolution of minerals characteristic of granite or 

schist along fracture faces to cause this increase in dissolved solids. Instead, it is hypothesized that water in 

the void space of the unfractured granite and schist is 

in chemical equilibrium with the associated minerals 

in the granite and schist. These equilibrium concentrations are significantly larger than the concentrations of 

dissolved solids in the water recharged through the 

overburden and into the fractures in the bedrock. Thus, 

the concentration gradient between fluid in the rock 

matrix and fluid in the fractures results in the mass flux 

of dissolved solids into the fractures.

The alkalinity of water samples collected from 

fractures in the bedrock offers additional evidence as to 

the role of diffusion in the chemistry of ground water 

(Wood and others, 1996). The alkalinity of water samples extracted from fractures in the bedrock is almost 

exclusively a function of the bicarbonate ion (HCOV ). 

As discussed in Wood and others (1996), isotopic signatures of carbon indicate that the bicarbonate ion is 

derived from the dissolution of carbonate minerals, yet 

there is no known evidence of calcite in fractures. 

However, calcite was identified in granite samples 

from cores and outcrops (Wood and others, 1996). The 

total carbon in granite samples from outcrops was 

0.12 percent of the sample weight. If all carbon is 

assumed to bound in calcite, then approximately 1 per173
cent of the rock matrix is calcite by weight. Wood and 

others (1996) also measured other properties that affect 

diffusion from the rock matrix, including porosity and 

the effective diffusion coefficient.

A comparison of alkalinity and ground water 

ages shows a positive correlation (fig. 3). Figure 3 

shows that water samples from the drift have low alkalinity and young ground water ages. Water samples 

collected from fractures in bedrock indicate that alkalinity increases with ground-water age, but the alkalinity is bounded by the saturated concentration of 

bicarbonate ions in the ground water. Therefore, we 

hypothesize that bicarbonate ions are diffusing from 

water in the rock matrix and into fractures, and the 

bicarbonate concentration is a function of the residence time associated with ground-water flow in the 

fractures. The estimate of the equilibrium bicarbonate 

concentration shown in figure 3 was calculated by 

means of the geochemical-analysis program PHREEQE (Parkhurst and others, 1980), whereby a representative solution of overburden water was reacted 

with calcite and a partial pressure of carbon dioxide 

calculated from the chemistry of the overburden water.

ESTIMATING THE LENGTH OF GROUNDWATER FLOW PATHS IN BEDROCK

In order to investigate the relation among alkalinity and ground-water velocity and residence time, 

we consider a simplistic and preliminary conceptual 

model of ground-water flow and the transport of bicarbonate ions through the bedrock (fig. 4a). The model 

consists of a single fracture. Ground water is advected 

through the fracture with a specified velocity. The 

bicarbonate concentration at the inlet to the fracture is 

representative of the overburden water (18 mg/L). We 

assume that there is no dispersion in the fracture; however, as the ground-water is advected along the fracture, bicarbonate ions diffuse from the rock matrix to 

the fracture because of the concentration gradient. The 

gradient depends on the bicarbonate concentration in 

the fracture and the position of a calcite-dissolution 

front in the rock matrix. Assuming that the dissolution 

of calcite in the rock matrix is not kinetically controlled, calcite dissolution will occur as a sharp front 

(Rubin, 1983; Willis and Rubin, 1987). Between the 

fracture and the calcite dissolution front, the bicarbonate concentration is controlled by the diffusion equation. At positions farther into the rock matrix, the 

bicarbonate concentration in equilibrium with solidphase calcite is assumed to prevail (122 mg/L).

V)

5

So"

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li I

2 <

? cc

. UJ

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Z DC

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< Q.

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160

140

120

100

80

60

40

20

ft

 , , , r , T , . - r , 1 T -, , -T . ^-. . | . . . r 

APPROXIMATE EQUILIBRIUM

WITH CALCITE 

i -.. 

2 

* 

0 

* * .

EXPLANATION

, " BEDROCK WELL

_ n

» DRIFT PIEZOMETER

o. .......................

10 20 30 40 

GROUND-WATER AGE, IN YEARS

50

Figure 3. Ground-water ages estimated from CFC-12 concentrations and alkalinity of water samples collected from 

bedrock wells and drift piezometers in the Mirror Lake area.

At the inlet to the fracture, additional calcite dissolution occurs because of the difference between the 

bicarbonate concentration of the overburden water and 

that of the water in the rock matrix. Thus, the calcitedissolution front will be farther from the fracture at the 

inlet than at locations farther down-gradient. The distribution of the bicarbonate concentration along the 

length of the fracture is a function of the ground-water 

velocity through the fracture. For an infinite velocity, 

the bicarbonate concentration along the fracture will 

be constant and equal to the concentration of the overburden water, because the advective flux through the

(A)

CALCITE DISSOLUTION FHONT

HCOj' DIFFUSION

AOCK-MATAIX

WATEfl IN EQUILIBRIUM

WITH CALCITE

(B)

. tc 

z u;

g5 100

E 5 

£ °- so i^|i 60

o*i « 

o z

~ 20

v = 1 X 10'4 METERS PER DAY ;

 If I

"I / v = 1 X 10-3 METERS PER DAY :

I ; j

T '^~~ ~ ........ " " -"Tl 'r ~

/ i ........---- i :

\f. y ""~ v = 1 X 10-2 METERS PER DAY | 

i i i :

200 400 600 800 1000 

DISTANCE FROM FRACTURE INLET, IN METERS

Figure 4. (a) Conceptual model of calcite dissolution and 

bicarbonate diffusion from a rock to a single fracture, 

(b) Bicarbonate concentration after 1000 years as a function of 

distance along a fracture for three ground-water velocities.

174
fracture will be large relative to the diffusive flux from 

the rock matrix. For a zero flow velocity, the bicarbonate concentration in the fracture will achieve the 

saturated bicarbonate concentration associated with 

the rock matrix. Velocities between these two 

extremes will yield distributions of bicarbonate concentration as a function of the distance along the fracture.

The distribution of the bicarbonate concentration along the fracture also depends on the time since 

the initial conditions. For example, if the initial calcite-dissolution front is assumed to coincide with the 

fracture face, and the initial bicarbonate concentration 

in the fracture is assumed to be the saturated bicarbonate concentration, the distribution of the bicarbonate 

concentration in the fracture will depend on the time 

after these initial conditions. For example, in figure 

4b, assuming the initial conditions discussed above, 

the bicarbonate concentration along a fracture is 

shown after 1,000 years for three values of the groundwater velocity in the fracture. These calculations were 

made using a sharp-interface model for calcite dissolution in the rock matrix as discussed in Willis and 

Rubin (1987). The bicarbonate concentration in the 

fracture was calculated by a one-dimensional finitedifference approximation for time-dependent advective transport subject to diffusion from the rock 

matrix. Longer elapsed times do not appreciably affect 

the distribution of the bicarbonate concentration along 

the fracture for the velocities of 10-4 and 10-2 m/d . 

For a velocity of 10-3 m/d and longer elapsed times, 

the position along the fracture where calcite dissolution occurs is shifted farther from the fracture inlet 

than that shown in figure 4b. This results in a different 

distribution of bicarbonate concentrations along the 

fracture, which, in turn affects the calculations of the 

length of the flow path. The sensitivity of the elapsed 

time to the calculation of the distance ground water 

has traveled along the fracture is being investigated to 

place bounds on the calculations.

From figure 4b, estimates of the flow-path 

length associated with a specific sampling location can 

be made on the basis of the ground-water age and the 

bicarbonate concentration measured at that sampling 

location. For example, a bicarbonate concentration of 

40 mg/L could be associated with each of the profiles 

of bicarbonate concentration that were generated from 

the three velocities in figure 4b. For each velocity, 

there is a unique distance from the fracture inlet where 

that bicarbonate concentration will be present. 

Because we assume that there is no dispersion in the 

fracture, the distance a parcel of ground-water moves

in the fracture also can be defined as the product of the 

ground-water velocity and the residence time measured by ground-water age dating. Because the conceptual model only considers transport in the bedrock, 

the measured residence time must be adjusted to 

account for transport through the unconsolidated glacial deposits; in the Mirror Lake area residence times 

in the drift range from 5 to 18 years (Busenberg and 

Plummer, 1996). The product of the (assumed) 

ground-water velocity and the (adjusted) residence 

time can be compared with the distance along the fracture shown in figure 4b for the (assumed) groundwater velocity and the measured bicarbonate concentration. A ground-water velocity is chosen such that 

the two calculations for the length of the flow path in 

the bedrock are equal. Preliminary results from 

selected sampling locations show that the length of 

flow paths in the bedrock range from 5 to 500 m with 1 *}

velocities of 10"' and 10 m/d. The calculation for all 

of the sampling locations and the sensitivity of these 

calculations to the travel time through the overburden 

and the assumed elapsed time since the initial conditions are currently being investigated.

Although this preliminary conceptual model 

allows us to investigate the significance of the groundwater velocity and ages in relation to bicarbonate concentrations, it is highly idealized and subject to uncertainties. The conceptual model assumes a constant 

velocity through an individual fracture, whereas, in 

reality transport in the bedrock occurs through a network of interconnected fractures of variable hydraulic 

properties. Also, the modeled bicarbonate distribution 

along the fracture is the result of an elapsed time since 

assumed initial conditions. The bicarbonate concentration differs with elapsed time and tectonic conditions 

where new fractures open and the flow regime changes 

in the bedrock over time. Furthermore, the correlation 

of the measured bicarbonate concentration and 

ground- water ages is not perfect and may not be 

assessable by the simple conceptual model considered 

in this article.

In addition, the conceptual model of flow and 

chemical transport in the bedrock only considers advective transport in the fracture. However, other geochemical data indicate the possibility of mixing of ground 

waters of various ages. For example, the maximum 

measured tritium concentration in ground-water samples is 35 TU. However, the maximum tritium concentration in 1992 should be approximately 300 TU on the 

basis of the peak tritium concentration in precipitation 

in the 1960's. Either the resolution of the sampling locations has not resolved the peak tritium concentration, or

175
the mixing of water of various ages reduces tritium 

concentrations. Also, excess helium concentrations 

reported by Drenkard and others (1996) are indicative 

of the mixing of older ground waters. The conceptual 

model discussed in this article needs to be reconciled 

with this geochemical evidence.

SUMMARY

A simple conceptual model of ground-water 

flow and the transport of bicarbonate ions in groundwater in the bedrock of the Mirror Lake area is proposed to estimate regional ground-water velocities. 

Concentrations of CFC's and the tritium- helium ratio 

are used to estimate the residence time of groundwater samples collected in the glacial drift and the 

upper 200 m of the bedrock. However, the distance the 

ground-water has traveled also is needed to identify 

the ground-water velocity. The complex flow regime 

in the shallow bedrock prohibits an independent estimate of the length of a flow paths in the bedrock. 

Alkalinity of ground-water samples (caused primarily 

by the diffusion of bicarbonate ions from the rock 

matrix) shows a positive correlation with groundwater ages. Thus, the bicarbonate concentration is 

used as second indicator of ground-water residence 

time in the bedrock. A simple conceptual model consisting of a single fracture in the rock matrix is used to 

investigate the relations among bicarbonate concentration and ground-water velocity and residence time in 

the bedrock. From this conceptual model, the average 

ground-water velocity in the bedrock can be identified 

from measured bicarbonate concentrations and estimated ground-water ages. This model is currently 

being applied to the data collected in the Mirror Lake 

area. However, the conceptual model must be reconciled with other geochemical information, which indicates that the mixing of ground waters of various ages 

is occurring either naturally or as a result of groundwater sampling.

REFERENCES

Barton, C. C., 1996, Characterizing bedrock fractures in 

outcrop for ground-water hydrology studies: An 

example from Mirror Lake, Grafton County, New 

Hampshire, in Morganwalp, D.W., and Aronson. D.A., 

cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Busenbcrg, E., and Plummcr, L. N., 1992, Use of

chlorofluorcarbons (CCI3F and CCI2F2) as hydrologic 

tracers and age-dating tools--The alluvium and terrace 

system of central Oklahoma: Water Resources 

Research, v. 28, no. 9. p. 2257-2283.

Buscnberg, E., and Plummcr, L. N., 1996, Concentrations 

of chlorofluorocarbons and other gases in ground 

waters at Mirror Lake, New Hampshire, in 

Morganwalp, D.W., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the Technical Meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Drenkard, S., Torgcrscn, T., Wcppcrnig, R., Farlcy, K., 

Schlosscr, P., Michcl, R., L., Shapiro, A. M., and 

Wood, W. W., 1996, Helium isotope analysis and 

tritium- helium age dating in the Mirror Lake basin, 

Grafton County, New Hampshire, in Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Hsich, P. A., Perkins, R. L.. and Roscnbcrry, D. O., 1996, 

Field instrumentation for multi-level monitoring of 

hydraulic head in fractured bedrock at the Mirror Lake 

site, Grafton County, New Hampshire, in Morganwalp, 

D.W., and Aronson, D.A., cds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Hsich, P. A., and Shapiro. A. M., 1996, Hydraulic

characteristics of fractured bedrock underlying the FSE 

well field at the Mirror lake site, Grafton County, New 

Hampshire, in Morganwalp, D.W.. and Aronson, D.A.. 

cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Parkhurst, D. L., Thorstcnsen, D. C., and Plummcr, L. N., 

1980, PHREEQE - A computer program for 

geochemical calculations: U.S. Geological Survey 

Water-Resources Investigations Report 80-96, 195 p.

Rubin, J., 1983, Transport of reacting solutes in porous 

media - Relation between mathematical nature of 

problem formulation and chemical nature of reactions: 

Water Resources Research, v. 19, no. 5, p. 123 I -1252.

Schlosser, P., Stutc, M., Dorr, H., Sonntag, C., and

Munnich, K. O., 1988, Tritium/^Hc dating of shallow 

ground water: Earth and Planetary Science Letters, 

v. 89, p. 353-362.

176
Solomon, D. K., Porcda, R. J., Schiff, S. L., and Cherry, J.

A., 1992, Tritium and hclium-3 as ground water age

tracers in the Bordcn aquifer: Water Resources

Research, v. 28, no. 3, p. 741-755. 

Willis, C., and Rubin, J., 1987, Transport of reacting solutes

subject to a moving dissolution boundary: Numerical

methods and solutions Water Resources Research,

v. 23, no. 8, p. 1561-1574. 

Winter, T. C., 1984, Gcohydrologic setting of Mirror Lake,

West Thornton, New Hampshire: U.S. Geological

Survey Water-Resources Investigations Report 84-

4266,61 p. 

Winter, T. C., 1985, Physiographic setting and geologic

origin of Mirror Lake, in Likens, G. E., cd.. An

Ecosystem Approach to Aquatic Ecology: New York, 

Springer-Vcrlag, p. 40-53.

Wood, W. W., Shapiro, A. M., and Hsich, P. A., 1996, 

Observational, experimental and inferred evidence of 

solute diffusion in granite rocks Examples from the 

Mirror Lake watershed, Grafton County, New 

Hampshire, in Morganwalp, D.W., and Aronson, D.A.. 

cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

177
178
Overview of Research at the Cape Cod Site:

Field and Laboratory Studies of Physical, Chemical, and

Microbiological Processes Affecting Transport

in a Sewage-Contaminated Aquifer

By Denis R. LeBlanc 1

Abstract

The Cape Cod Toxic-Substances 

Hydrology research site in Falmouth, Mass., is the 

focus for multidisciplinary field studies of the 

physical, chemical, and microbiological processes 

affecting transport in a sewage-contaminated sand 

and gravel aquifer. Past and current research at 

the site is summarized in a synopsis of 62 papers. 

Past research has examined the distribution and 

fate of contaminants in a sewage plume that is 

more than 4.5 kilometers long and originates from 

the Otis Air Base wastewater-treatment facility. 

Current research includes efforts to characterize 

specific physical, chemical, and microbiological 

processes affecting transport and fate of solutes, 

bacteria, and protozoa at the site. These processes 

are characterized by means of laboratory 

experiments, such as batch and column tests; field 

experiments, such as natural-gradient tracer tests; 

and investigations of the sewage plume. The 

results of the research show the importance of 

physical and chemical heterogeneity, limited 

dispersion, and chemical processes at the 

sediment-water interfaces in determining the fate 

of contaminants in this aquifer.

INTRODUCTION

The U.S. Geological Survey (USGS) Cape Cod 

Toxic-Substances Hydrology research site is located in 

the northern part of Falmouth, Mass. (fig. I). This site

U.S. Geological Survey, Marlborough, Mass.

is the focus for multidisciplinary field studies of the 

physical, chemical, and microbiological processes 

affecting the movement of contaminants in a sewage 

plume originating from Otis Air Base. It was selected 

for study as part of a nationwide program of the USGS 

to describe the transport and fate of contaminants in 

the subsurface. The purpose of this paper is to review 

past research efforts at the Cape Cod site and to 

introduce the research topics discussed in other papers 

in this proceedings. Earlier summaries of research at 

the Cape Cod site can be found in reports by LeBlanc 

(I984b), Franks (1987), Ragonc (1988), Mallard and 

Ragone (1989), and Mallard and Aronson (1991).

SITE DESCRIPTION

The study area is on a broad sand and gravel 

glacial-outwash plain that was formed during the last 

Pleistocene glacial retreat. The outwash plain slopes 

southward to Nantucket Sound and is pitted with many 

kettle holes, some of which contain ponds. The area 

contains several valleys that transect the plain from 

north to south. Most of these valleys do not contain 

streams but have wetlands at their southern ends.

The top 30 to 50 m (meters) of outwash are 

composed of stratified, medium to coarse sand with 

some gravel. The sand and gravel overlies fine sand 

and silt that locally contains lenses of clay, silt, sand, 

and gravel. These unconsolidated sediments overlie a 

crystalline (granodiorite) bedrock surface, which 

generally slopes to the southeast through the study 

area.

179
71°00

42°30'

MASSACHUSETTS

0 20 40 60 MILES

0 20 40 60 KILOMETERS

70° 34' 70° 32'

41° 38'

41° 37'

41° 36'

33'

~r

, _ _ OTIS AIR BASE

-t 50 nSC SEWAGE-DISPOSAL ^ i_j

EXPLANATION

30 -^ WATER-TABLE CONTOUR -- Shows altitude of water table above 

sea level, in feet. Arrows show direction of ground-water flow.

Figure 1. Location of study area, showing sewage plume, water-table contours, and tracer-test site.

180
On the basis of measured values for similar 

sediments on Cape Cod, LeBlanc (1984c) estimated 

that the horizontal hydraulic conductivity of the sand 

and gravel in the study area ranges from 60 to 90 m/d 

(meters per day). Results of an aquifer test conducted 

in the study area in 1984 indicate that the horizontal 

hydraulic conductivity of the sand and gravel may 

locally be as high as 120 m/d (Garabedian and others, 

1988). The horizontal hydraulic conductivity of the 

underlying fine sand and silt is estimated to be about 

one-tenth that of the sand and gravel (LeBlanc, 

1984a). The crystalline bedrock is assumed to have a 

very low hydraulic conductivity; therefore, the 

bedrock is considered to be the bottom of the regional 

ground-water flow system.

Ground water in the unconsolidated sediments 

is under unconfined (water table) conditions. The 

water table slopes toward the south at about 1.5 m/km 

(meters per kilometer) (fig. 1). Seasonal variations in 

recharge from precipitation produce an annual watertable fluctuation of 0.3 to 0.9 m; the highest levels are 

in the spring and the lowest are in the fall.

Ground-water recharge to the study area occurs 

primarily from precipitation and underflow from 

upgradient areas. Little surface-water runoff occurs 

because the sandy soils are very permeable. Estimated 

recharge to the aquifer is 0.5 meters per year, or about 

45 percent of the total precipitation (LeBlanc, 

1984a). Estimated rates of horizontal ground-water 

velocity in the sand and gravel range from 0.2 to 0.6 

m/d. These estimates are based on an average 

hydraulic gradient of 1.5 m/km, a horizontal hydraulic 

conductivity of 60 to 120 m/d, and a porosity of 30 to 

40 percent (LeBlanc, 1984c).

SEWAGE PLUME

Disposal of treated sewage onto infiltration sand 

beds at Otis Air Base (fig. 1) since 1936 has created a 

plume of contaminated ground water (LeBlanc, 1984 

b, c) that is 0.8 to 1.1 km (kilometers) wide, 23 m 

thick, and more than 4.5 km long (fig. 1). The plume 

moves to the south in the direction of ground-water 

flow and is overlain by up to 15 m of uncontaminated 

ground water derived from precipitation that recharges 

the aquifer. Part of the plume discharges to Ashumet 

Pond (fig. 1), which is located about 500 m from the 

disposal beds.

The plume of sewage-contaminated ground 

water is characterized by elevated concentrations of

dissolved solids, boron, chloride, sodium, phosphorus, 

ammonium, nitrate, detergents (LeBlanc, 1984c), and 

volatile organic compounds (VOC), including 

dichloroethene, trichloroethene, and tetrachloroethene 

(Thurman and others, 1984; Barber and others, 

1988). Boron, chloride, and sodium appear to be 

conservative and nonreactive constituents that are 

attenuated primarily by hydrodynamic dispersion. 

Phosphorus movement is greatly retarded by 

adsorption onto the sediments (LeBlanc, 1984c); 

colloidal precipitation of iron-phosphate compounds 

also may affect phosphorus transport near the disposal 

beds (Gschwend and Reynolds, 1987; Backhus and 

Gschwend, 1990; Backhus and others, 1993).

Ammonium is the predominant nitrogen species 

in the center of the plume within 1.5 km of the 

disposal beds (LeBlanc, 1984c). The distribution of 

ammonium is caused, in part, by adsorption onto the 

aquifer sediments, which retards the movement of 

ammonium (Ceazan and others, 1989). Beyond 

1.5 km, the predominant nitrogen species changes to 

nitrate. Nitrate cannot be detected in the center of the 

plume immediately downgradient of the disposal beds, 

even though nitrate concentrations in the sewage 

effluent are as high as 16 mg/L (milligrams per liter) 

(as nitrogen (N)), because microbially mediated 

denitrification has converted the nitrate to nitrogen gas 

(Smith and others, 1991b).

Detergent (methylene-blue-active subtances) 

concentrations exceed 0.5 mg/L from 0.9 to 3.0 km 

downgradient from the disposal beds. This 

distribution of detergents reflects the use of 

nonbiodegradable detergents during 1946-64 

(LeBlanc, 1984c; Thurman and others, 1986; Thurman 

and others, 1987). Although more than 90 percent of 

the biodegradable detergents are presently removed by 

the wastewater-treatment facility, the remaining 

detergents which enter the aquifer degrade slowly and 

are detected as far as 1.5 km from the disposal site 

(Field and others, 1992a, 1992b).

Elevated VOC concentrations are present in two 

zones in the study area. The source of the VOC zone 

immediately downgradient from the disposal beds 

may be unrelated to the sewage disposal because the 

VOC's are found beneath the sewage plume. 

However, the VOC zone 500 to 2,600 m downgradient 

from the disposal beds is thought to originate from the 

sewage-treatment facility because the VOC's are 

found within the sewage plume (Thurman and others, 

1984; Barber and others, 1988; Barber and others,

181
1992b). VOC concentrations in the downgradient 

zone exceed 50 micrograms per liter, which suggests 

that these compounds are mobile and not readily 

degraded in the sandy aquifer (Barber, 1988; Barber 

and others, 1988).

Bacterial population counts as large as 2 x 

106/mL (per mi Hi liter) are found in ground water near 

the disposal site; these counts decrease by an order of 

magnitude more than 1 km from the beds (Harvey and 

others, 1984). These numbers appear to correlate with 

the availability of degradable organic compounds; 

concentrations of dissolved organic carbon (DOC) 

decrease from 12 mg/L to less than 2 mg/L over the 

same distance (Thurman and others, 1986; Harvey and 

Barber, 1992; Metge and others, 1993). Assays of 

microbial activity have been made for ground water 

and aquifer sediments because more than 90 percent of 

the bacteria are attached to silt- and clay-sized 

particles (Harvey and George, 1987; Smith and Duff, 

1988). These assays show that rates of microbially 

mediated denitrification are greatest in water and 

sediment collected from a 1- to 2-m-thick zone near 

the top of the plume. The presence of the thin zone of 

elevated microbial activity demonstrates the need for 

sampling ground water and aquifer sediments at 

closely spaced vertical intervals (Smith and others, 

1991a.)

RESEARCH ON TRANSPORT 

PROCESSES

Past research efforts at the Cape Cod ToxicSubstances Hydrology research site have focused on 

defining and describing the extent of ground-water 

contamination in the sewage plume. Current research 

at the site includes efforts to characterize specific 

physical, chemical, and microbiological processes 

affecting the transport and fate of solutes and 

microorganisms in the aquifer. These processes are 

characterized using small-scale laboratory 

experiments, such as batch and column tests; 

intermediate-scale field experiments, such as naturalgradient tracer tests with transport distances of 1 to 

280 m; and large-scale investigations of the sewage 

plume.

Physical Transport

A major objective at the Cape Cod site has been 

to relate the dispersion of solutes to the heterogeneity

of the aquifer's hydraulic properties. A direct measure 

of the dispersion of solutes in the aquifer was obtained 

by a spatial-moments analysis of a large-scale naturalgradient tracer test (fig. 1) conducted during 1985-88 

(LeBlanc and others, 1991; Garabedian and others, 

1991). Bromide, a nonreactive tracer, was monitored 

with a three-dimensional array of about 10,000 

sampling points in an abandoned gravel pit as the 

tracer moved 280 m through the aquifer. The spatialmoments analysis indicates that the dispersivity is 

about 1.0 m in the direction of flow (longitudinal), 

about 0.02 m in the transverse horizontal direction, 

and about 0.002 m in the transverse vertical direction 

(Garabedian and others, 1991). The extremely 

detailed bromide distributions were also used by 

Knopman and others (1991) to test various samplingdesign strategies.

Application of a numerical model of densitydependent flow (LeBlanc and Celia, 1996) to the 

large-scale natural-gradient tracer test shows that the 

density difference between the ambient ground water 

and the tracer solution was sufficient to cause part of 

the downward movement of the tracer cloud observed 

during the test. The density-induced downward 

movement was most important during the first 37 days 

of transport when the density difference was greatest. 

Intermittent recharge from precipitation also caused 

part of the downward movement.

Hess and others (1992) used the theoretical 

stochastic transport equations of Gelhar and Axness 

(1983) to estimate aquifer macrodispersivity at the 

Cape Cod site from the statistical properties of the 

hydraulic-conductivity distribution. The statistics 

were obtained from the analysis of nearly 1,500 

measurements of hydraulic conductivity that were 

made using borehole-flowmeter tests and permeameter 

analyses of cores (Wolf and others, 1991) near the 

location of the large-scale tracer test. The range of 

estimated longitudinal macrodispersivity is 0.35 to 

0.78 m; this range is similar to the longitudinal 

dispersivity observed in the large-scale tracer test.

Reilly and LeBlanc (1996) conducted a wellpurging experiment to test the hypothesis that the 

spatial heterogeneity of hydraulic conductivity and 

water chemistry in the aquifer near a well can cause 

the chemical composition of the water sampled from 

the well to vary temporally as the well is pumped. 

They observed temporal trends in specific 

conductance and concentrations of ferrous iron and 

calcium during well purging which agree with the

182
hypothesis that the trends are due to flow and solute 

transport in the heterogeneous aquifer in the 

immediate vicinity of the well and not to the purging 

of standing water in the well. Morin and others (1988) 

used borehole geophysical logs to demonstrate that 

disturbance of the formation during drilling can affect 

the spatial heterogeneity of hydraulic conductivity 

adjacent to the well.

In addition to the above efforts to characterize 

aquifer properties and their effects on flow and solute 

transport at spatial scales of 1 to 280 m, recent efforts 

also have focused on large-scale characterization of 

the aquifer. Moench and others (1996) conducted an 

aquifer test at a partially penetrating well in the sand 

and gravel aquifer near the site of the large-scale 

natural-gradient tracer test and analyzed the waterlevel-drawdown data by type-curve matching. The 

analysis indicates that the horizontal hydraulic 

conductivity is 105 m/d and that the ratio of vertical to 

horizontal hydraulic conductivity is about 1:2, values 

that are consistent with estimates obtained from a 

stochastic analysis of detailed measurements of 

hydraulic conductivity made at the same site (Hess 

and others, 1992).

Masterson and Walter (1996) developed a 

preliminary regional-scale ground-water-flow model 

of western Cape Cod and used the model to examine 

the pathlines of contaminants emanating from the 

sewage-disposal beds. Modifications to aquifer 

properties used in the model resulted in similar 

simulated hydraulic-head distributions and rates of 

ground-water discharge to streams, but markedly 

different contaminant pathlines. An accurate 

representation of the regional hydrogeologic 

framework, especially the thickness and hydraulic 

conductivity of the fine sand and silt beneath the 

permeable sand and gravel, was needed to simulate the 

observed path of the sewage plume accurately, even 

though the simulated heads and discharge rates were 

insensitive to these aquifer properties in the model.

Chemical Processes

More than 30 tracer tests have been conducted 

to determine geochemical controls on reactive 

transport in the heterogeneous aquifer. Two reactive 

tracers, lithium and molybdate, were monitored as part 

of the 1985-88 large-scale natural-gradient tracer test 

and were found to be significantly retarded relative to 

bromide (LeBlanc and others, 1991). Adsorption of

lithium, a cation, occurs on the mineral surfaces and, 

more significantly, inside the weathered grains (Wood 

and others, 1990), where adsorption is controlled by 

diffusion into pores inside the grains. The diffusioncontrolled adsorption resulted in a skewed distribution 

of lithium, with higher concentrations near the leading 

edge of the solute cloud and lower concentrations in 

the trailing edge. Adsorption of molybdate (an 

oxyanion of molybdenum) was affected during the 

tracer test by the presence of the sewage plume, which 

caused variations with depth in pH and concentrations 

of phosphate, an oxyanion that competes with 

molybdate for adsorption sites (Stollenwerk and Kipp, 

1990).

The effect of hydrologic and geochemical 

processes on metal-ion transport was evaluated in a 

series of laboratory experiments and small-scale tracer 

tests conducted at the Cape Cod site (Davis and others, 

1996). The results of 12 natural-gradient tracer tests 

performed during 1988-92 using zinc, nickel, 

chromium, selenium, and ethylenediaminetetraacetic 

acid (EDTA) show that adsorption-desorption, 

aqueous-complexation, and oxidation-reduction 

reactions affect the transport of the reactive metals 

during transport through the uncontaminated, oxic 

zone and the sewage-contaminated, mildly reducing 

zone of the aquifer. The transport of chromium 

showed a marked dependence on its speciation and the 

chemical composition of the ambient ground water 

(Kent and others, 1994; Anderson and others, 1994). 

The results of the laboratory experiments and smallscale tracer tests were used to design a naturalgradient tracer test, which began in April 1993, to 

examine transport of a complex mixture of reactive 

metal ions (Davis and others, 1996). The large-scale 

test, in which eight tracers (bromide, chromium, zinc, 

copper, lead, nickel, potassium, and EDTA) were 

dissolved in 10,000 liters of water and injected as a 

pulse into the aquifer, will involve detailed 

observation of the tracer cloud as it moves as much as 

200 m through the sand and gravel.

The geochemical properties of the aquifer 

sediments that control metal-ion transport are being 

investigated in conjunction with the tracer 

experiments to determine their potential use as 

indicators of the spatial variability of metal adsorption 

in the aquifer. Fuller and others (1996) show that lead 

and zinc are adsorbed primarily by iron- and 

aluminum-oxide coatings on the surfaces of the quartz 

grains. The amount of adsorption, which varies by a

183
factor of two to four, is related to the amount of iron 

and aluminum that can be dissolved from the aquifer 

material by partial chemical extraction when the 

amounts that can be adsorbed and dissolved are 

normalized to surface area.

Hess and others (1996) compared the spatial 

variability of metal-ion adsorption to estimates of 

hydraulic conductivity based on the results of 

laboratory experiments on 375 sediment samples from 

14 boreholes at the tracer-test site. Zinc and lead 

adsorption were measured in batch experiments 

(Fuller and others, 1996), whereas hydraulic 

conductivity was estimated from grain-size 

distributions. There is a statistically significant, but 

weak, negative correlation between lead adsorption 

and hydraulic conductivity. Garabedian and others 

(1988) used a stochastic analysis to demonstrate that a 

negative correlation similar to that observed for lead 

adsorption increases the apparent dispersion of a 

sorbing solute.

Geochemical heterogeneity of the sand and 

gravel also affects the sorption of organic compounds. 

Barber and others (1992a) and Barber (1994) showed 

that sorption of chlorobenzenes to the Cape Cod 

sediments increases as the particle size decreases 

because the fine-grained sediments have a larger total 

surface area, a greater abundance of magnetic 

minerals, and higher concentrations of sediment 

organic carbon than coarse-grained sediments in the 

aquifer.

Adsorption of metals ions in the sewage plume 

(Rea and others, 1996) has prevented significant 

transport of the metals zinc, copper, and lead away 

from the sewage-disposal beds. The extent of 

adsorption is affected by vertical gradients of pH 

characteristic of the transition between the sewagecontaminated ground water and the uncontaminated 

ground water that overlies the plume (Smith and 

others, 199 la).

Adsorption to the aquifer sediments also has 

retarded the transport of phosphate in the sewage 

plume (Rea and others, 1996). Phosphate 

concentrations in the ground water near the disposal 

beds are similar to those measured in the treated 

sewage. Because phosphate sorbs strongly to the 

sediments, the high concentrations in ground water 

indicate the presence of a large reservoir of phosphate 

sorbed to the aquifer sediments. The reservoir of 

sorbed phosphate may be remobilized after sewage 

disposal is stopped, as planned in late 1995, and some

phosphorus may be transported in ground water that 

discharges to Ashumet Pond (fig. 1). A series of 

laboratory column experiments (Stollenwerk, 1996) 

indicates that the phosphate may desorb slowly as the 

sewage plume is flushed naturally from the aquifer. 

About 160 pore volumes were eluted from the column 

before phosphate concentrations decreased to levels 

generally considered by limnologists not to cause 

eutrophication of lakes (about 0.05 mg/L).

The above studies indicate that the geochemical 

environment greatly affects the transport and fate of 

reactive contaminants. An earlier study (Lee, 1991) 

had shown that the principal process affecting the 

ambient chemistry of the shallow, uncontaminated 

ground water in the study area is the carbon-dioxidecontrolled hydrolysis of sodium feldspar. Lee (1996) 

measured carbon dioxide concentrations in soil gases 

from the unsaturated zone at 20 sites. The 

measurements show that carbon dioxide 

concentrations vary significantly from site to site, 

depending on land use, and may cause related 

variations in ambient water chemistry over the study 

area.

Microbiological Processes

The microbial populations and their activity and 

transport in the aquifer have been characterized by 

several techniques. The use of small-scale tracer tests 

to measure microbial activity directly in the aquifer at 

a site where denitrification is occuring was tested 

using methane and hexafiuoroethane as tracers (Smith 

and others, 1991c). These dissolved gases were 

transported without retardation, but concentrations of 

methane apparently decreased because of 

biodegradation. Acetylene, a dissolved gas, was 

subsequently used as a tracer because it inhibits the 

final step in the denitrification process and allows in 

situ measurement of the denitrification rate (Brooks 

and others, 1996). The measured rates of 

denitrification were generally lower than those 

reported for aquatic sediments or measured in the 

laboratory with sediments from the Cape Cod site, 

suggesting that laboratory methods may overestimate 

the rate of denitrification in the aquifer. Brooks and 

others (1992) reported that measurements of 

denitrification in the laboratory may also be biased 

when chloramphenicol, a common amendment in 

microbial-activity assays, is used.

The denitrification studies suggest that 

biodegradation can remove contaminants from

184
aquifers. Smith and others (1994) identified several 

strains of bacteria that consume nitrate when hydrogen 

or formate are added to sediment samples collected 

from the zone of active denitrification in the sewage 

plume. The degradation of trichloroethylene was 

observed in sediment from the Cape Cod site that had 

been innoculated with genetically engineered bacteria 

(Krumme and others, 1993). Survival of the parental 

strain of the genetically engineered bacteria for more 

than 100 days after it had been injected into the aquifer 

indicates that laboratory strains of bacteria can survive 

for extended periods of time in the subsurface and may 

be effective agents for bioremediation (Krumme and 

others, 1994; Thiem and others, 1994).

Tracer tests conducted at the Cape Cod site 

show that bacteria can be transported significant 

distances through the aquifer (Harvey and others, 

1989). In one test, the bacteria moved about 7 m at the 

same rate as a nonreactive tracer, bromide, under a 

natural hydraulic gradient, although the concentrations 

of bacteria declined relative to the concentration of 

bromide. Adsorption and differential size exclusion 

(filtering) of bacteria from small pores can vary 

significantly in the heterogeneous aquifer sediments 

and greatly affect bacterial transport and attenuation 

(Harvey and Garabedian, 1991, 1992; Harvey and 

others, 1993). Geochemical conditions, such as the 

pH of the ground water and the presence of 

oxyhydroxide coatings on the sediment grains, also 

affect the transport of bacteria (Scholl and Havey, 

1992). In column experiments using sediment from 

the Cape Cod site, an increase in pH from 5.8 to 7.9 

resulted in a 70 percent decrease in bacterial 

attachment (Metge and others, 1996). Bacterial 

attachment was also affected by changes in the amount 

and type of DOC and the ionic strength and divalentanion concentration in the ground water. The pH also 

affects the attachment of bacteriophage (bacteriaspecific viruses) to the Cape Cod sediments (Kinoshita 

and others, 1993).

A survey of the abundance of protozoa in the 

sewage plume showed that their abundance typically 

varies from 10,000 to 100,000 per gram of sediment 

and is directly related to the concentration of DOC in 

the ground water (Kinner and Harvey, 1996). The 

protozoa consist largely of small (2-3 micrometerlong) flagellates. It had been assumed that the role of 

protozoa in the subsurface ecosystem would be similar 

to their role in other environments that is, as 

predators of bacteria. But their small size and

correlation with the concentration of DOC suggests 

that they feed directly on the organic compounds.

The transport of protozoa was examined by 

means of several small-scale, natural-gradient tracer 

tests (Harvey and others, 1996). The retardation and 

immobilization of the protozoa were several orders of 

magnitude greater than the retardation and 

immobilization of bacteria in the same experiments. 

The high degree of immobilization and retardation 

may be related to the surface chemistry of the protozoa 

and not to their size.

SUMMARY

Past research efforts at the Cape Cod ToxicSubstances Hydrology research site have focused on 

defining and describing the extent of ground-water 

contamination in a sewage plume emanating from the 

Otis Air Base sewage-treatment facility on Cape Cod, 

Mass. Current research at the site includes efforts to 

characterize specific physical, chemical, and 

microbiological processes affecting the transport and 

fate of solutes and bacteria in the aquifer. These 

processes are characterized by means of small-scale 

laboratory experiments, such as batch and column 

tests; intermediate-scale field experiments, such as 

natural-gradient tracer tests with transport distances of 

1 to 280 m; and large-scale investigations of the 

sewage plume.

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190
The Transport of Inorganic Contaminants in a 

Sewage Plume in the Cape Cod Aquifer, Massachusetts

By Brigid A. Rea1 , Douglas B. Kent2 , Linda C. D. Anderson2 , James A. Davis2 ,

and Denis R. LeBlanc3

Abstract

The active and abandoned sewage-disposal 

beds at the Massachusetts Military Reservation 

sewage-treatment plant are a major source of 

inorganic contaminants, such as zinc, copper, and 

phosphate, in the Cape Cod aquifer, Massachusetts. The distribution and mobilities of these 

chemical constituents around the sewage-treatment plant are strongly affected by geochemical 

processes; extensive adsorption results in confinement of the most intensive concentrations to the 

near-source region, including currently used and 

abandoned disposal facilities. Beyond the disposal facilities, zinc and copper movement continues to be controlled by sorption processes. Zinc 

and copper contamination was present at the same 

depths, although copper concentrations were 

much lower than those of zinc. Phosphate concentrations were high in the suboxic zone near the 

source and are controlled by adsorption to sediments. Phosphate concentrations in the anoxic 

zone were much lower and likely are controlled 

by ferrous phosphate solubility.

INTRODUCTION

Inorganic contaminants in the subsurface travel 

at widely variable rates. The mobility of contaminants 

is strongly dependent on their speciation, aquifer pH, 

redox reactions, and mineral composition of the aqui-

'U.S. Geological Survey, Boulder, Colo. 

2U.S. Geological Survey, Menlo Park, Calif. 

3 U.S. Geological Survey, Marlborough, Mass.

fer material (Davis and Kent, 1990; Sposito, 1984). In 

this paper, we report on the processes that affect the 

distribution of sewage-derived phosphate, zinc (Zn) 

and copper (Cu) in the ground water near the sewageeffluent disposal beds (also termed "sand beds") at the 

U.S. Geological Survey Cape Cod Toxic Substance 

Hydrology Research site (fig. 1). A contrast in relative 

rates of transport is made between the two cations (Zn 

and Cu) and phosphate (an oxyanion) on the basis of 

expected ion-adsorption phenomena. Cation adsorption generally increases with an increase in pH, 

whereas anion adsorption decreases with an increase 

inpH.

In a previous paper (Rea and others, 1991), we 

described a small plume of dissolved Zn in a transition 

zone between pristine recharge water (infiltrated precipitation) and suboxic sewage-contaminated water; 

the 2-m-thick plume extended approximately 350 m 

downgradient from the sewage disposal beds (fig. 1). 

We postulated that the mobility of Zn was controlled 

by sorption reactions, which are affected by pH and Zn 

concentration. Concentrations of dissolved Zn in 

equilibrium with sorbed Zn on the aquifer solids 

decrease with increasing pH; above pH 6.0, Zn concentrations in equilibrium with sorbed Zn are often 

below detection (Rea and others, 1991). At constant 

pH, the amount of Zn sorbed increases with increasing 

Zn concentration until the maximum sorptive capacity 

is achieved (Davis and Kent, 1990; Sposito, 1984). In 

the sewage-contaminated zone, Zn can move only 

after the available sorption sites are saturated with 

sorbed Zn and other metals that compete for sorption 

sites. In the transition region, decreased pH, resulting 

from the mixing with pristine ground water, enhances 

Zn mobility enough to cause the observed Zn plume 

(Rea and others, 1991).

191
41°38'30" -T

Sewage

Treatment

Plant

w^ N /u ^ Sludge 

 9/ -JA^; Drying Pad

large-scale 

tracer test 

MLS area

V M, <2 fo If / ¥W fr" £ \ /H p r( M rf* C *

\

Explanation 

military reservation boundary

- - township boundary

 monitoring well

multi-level sampler (MLS)

. Hallden and M. Kumler 8/6/93

Figure 1. Location map showing the Massachusetts Millitary Reservation Sewage Treatment Facility, and well sites.

Zn in the sewage effluent is likely derived from 

pipe corrosion (Foerstner and Van Lierde, 1983). This 

led us to look for elevated concentrations of other 

metal ions, in particular Cu and lead (Pb). Both Cu 

and Pb adsorb to minerals more strongly than does Zn 

(Sposito, 1984; Dzomback and Morel, 1990). Consequently, contamination of the aquifer by these metals 

is expected to be limited to an area near the source;

however, mobilization of these metals could occur in 

response to changes in the ground-water chemistry.

Phosphate contamination has been reported 

previously in the Cape Cod aquifer (LeBlanc, 1984; 

Gschwend and Reynolds, 1987). Recently, local communities have become concerned that the contribution 

of phosphate from the Massachusetts Military Reservation (MMR) sewage-treatment plant may be

192
responsible for algal blooms, fish kills, and a general 

decline in the water quality (eutrophication) of Ashuinet Pond (K-V Associates, Inc. and IEP, Inc., 1991). 

An improved understanding of the geochemical processes affecting the transport of phosphate in the aquifer and of the interaction of the sewage plume with 

Ashumet Pond is required to evaluate remediation 

strategies.

In addition to studying the transport of these 

chemical constituents, we examined the possible 

sources of contamination and their potential effect on 

ground-water quality.

SITE DESCRIPTION

Discharge of secondary-treated sewage effluent 

from the MMR sewage-treatment plant into a sand and 

gravel aquifer on Cape Cod, Massachusetts (fig. 1) has 

generated distinct zones of ground-water chemistry 

separated by steep vertical gradients in chemical composition (LeBlanc, 1984; Smith and others, 1991; 

Davis and others, 1991; Kent and others, 1993). Pristine ground water (that is, water derived from precipitation) overlies the sewage plume. The pristine ground

Table 1 . Description of sampling sites

water is oxic, has a pH of 4.5 to 5.5, and low concentrations of dissolved salts (Davis and others, 1991). 

The sewage plume consists of a suboxic rim and an 

anoxic core (Smith and Duff, 1988). The suboxic rim 

is characterized by dissolved oxygen (DO) concentrations less than 5 uM and mildly reducing conditions, a 

pH of 5.5 to 6.5, and moderate concentrations of dissolved salts (Davis and others, 1991; Kent and others, 

1993). The anoxic core has ferrous iron (Fe2+) and 

ammonium concentrations greater than 300 uM. Values of pH and dissolved salt concentrations in the 

anoxic core are similar to those in the suboxic zone 

(Gschwend and Reynolds, 1987; Smith and Duff, 

1988; Kent and others, 1993).

METHODS

Ground-water samples were collected from a 

number of multilevel samplers (MLS) and clusters of 

observation wells (Smith and others, 1991; LeBlanc 

and others, 1991) around the MMR sewage-treatment 

plant in September, 1992 (fig. 1, table 1). The DO concentration and pH of the samples were measured at 

each site in the field (Davis and others, 1993).

Site number and descriptor Well type Description of location

S318

(Active sand beds)

S438

(Abandoned sand beds)

S314

F230

S434

Multilevel sampler Secondarily-treated sewage effluent currently 

and observation well1 disposed in this location

Observation well

Observation well

Cluster of 

observation wells

Cluster of

(Inactive sludge drying pad) observation wells

Disposal of treated sewage effluent ceased in 

this location in 1984

Less than 10 m downgradient of site S318 in the 

active sand beds

Approximately 100 m downgradient of site S318 

in the active sand beds

Sludge from treatment plant dried on these pads 

prior to disposal in land fill

1 All observation wells: 5.1 cm diameter with 0.6-m-long screens.

193
Ground-water samples were filtered (0.4 |um), preserved by acidification to pH 2 (with trace-metal grade 

HC1), and shipped overnight to Menlo Park, California, for analysis. These samples were analyzed for 

metals, boron, and phosphate by inductively coupled 

argon-plasma spectrometry. Some samples were analyzed in the field for phosphate using the standard colorimetric method (Franson, 1985), which is sensitive 

only to monomeric species. The difference in phosphate concentration determined by the two methods 

differed by an average of less than 1 percent (r2 of 

0.997), except at the lowest concentrations where the 

colorimetric method has a large uncertainty. This 

showed that dissolved phosphate was present as 

monomeric species, often called "orthophosphate."

RESULTS

Ground-water Chemistry

The ground-water chemistry at sites near the 

sewage-treatment plant depends on the historical use 

of the overlying sediments and the proximity of the 

sites to the source of contamination (table 1). At the 

active sewage-disposal sand beds, the concentration of 

DO in the ground water decreased from the water table 

(13.9 m above sea level and approximately 11.6 m 

below land surface) to a depth of about 9.5 m above 

sea level; all DO measurements from altitudes below 

9.5 m were near zero (site S318, fig. 2). Dissolved 

boron (B) concentrations exceeded 20 |uM in all 

depths examined (B is used as a indicator of sewagecontaminated water (LeBlanc, 1984)). The presence 

of B showed that sewage-contaminated ground water 

extended upward to the water table. Ground water at 

this site consistently has significant concentrations of 

DO in the sewage-contaminated zone (Smith and others, 1991; Rea and others, 1991). The concentration 

of sewage-derived elements in the ground water at an 

abandoned sand bed fell within the ranges of those in 

the ground water beneath the active bed.

The vertical zonation of ground-water chemistry 

typical of sewage plume was much better developed at 

well site F230 than at the active sand beds; site F230 is 

approximately 100 m downgradient from the active 

sand beds (fig. 3; table 1). The presence of some B 

(and phosphate) just below the water table indicates 

that ground water at this depth is a mixture of sewagecontaminated and pristine ground water because both 

B and phosphate are below detection in pristine

ground water. In contrast with the active sand beds 

(site S318), dissolved oxygen concentrations at site 

F230 were low throughout the sewage-contaminated 

zone (where B concentrations exceeded 10 |uM; see 

fig. 3). Ground water from the sampling point near sea

") i

level was anoxic and had an Fe concentration of 23

|LlM.

High concentrations of DO in the ground water 

from the well cluster in the inactive sludge-drying pad 

indicate that the chemistry of the ground water at the 

site is similar to that of pristine ground water (table 2). 

Concentrations of phosphate and B were very low 

throughout this well-oxygenated profile.

Distribution of Zinc

Zn concentrations in the active sewage disposal 

beds agree well with Zn concentrations in groundwater samples collected in August 1990 by Rea and 

others (1991); the uppermost ground-water sample 

had a Zn concentration of almost 2.3 |uM at each sampling (fig. 2). The concentration of dissolved Zn generally decreased with depth but remained above the 

concentrations found in the pristine (background) 

ground water. Ground water at the water table in the 

abandoned sand beds also contained concentrations of 

Zn well above those in pristine ground water (table 2). 

At sites 10 and 100 m downgradient from the abandoned sand beds (site S314 and site F230, respectively), dissolved Zn concentrations were high in 

comparison with those observed beneath the active 

sand beds (site S318) (fig. 3, table 2). Some contamination from the submersible pump used to withdraw 

water samples from the screened wells may have 

occurred. However, good agreement between Zn concentrations in ground water collected with submersible 

pump and peristaltic pump (in which the water never 

contacts metal) were obtained at the active sand beds 

(table 1, fig. 2). This point requires further investigation.

Concentrations of Zn at the inactive sludge-drying pad, which is upgradient of the disposal beds, 

decreased with depth (table 2). This indicates some 

contamination of the sediments from past sludge-drying practices. Current practices involve drying sludge 

in lined beds.

194
PHOSPHATE CONCENTRATION,

15

MICROMOLAR 

100 200 300

DISSOLVED OXYGEN

CONCENTRATION, MICROMOLAR

0 50 100 150 200 250

0.5 1.0 1.5 2.0

ZINC AND COPPER

CONCENTRATION. MICROMOLAR

0 10 20 30 40 50

BORON CONCENTRATION. 

MICROMOLAR

Figure 2. Vertical profiles of ground-water chemistry in one of the active sewage disposal beds (site S318). (A) Dissolved 

orthophosphate concentration ( ) and pH (B) as a function of depth. (B) Dissolved zinc (V) and dissolved copper (A) as a 

function of depth. (C) Dissolved boron (O) and dissolved oxygen (+) as a function of depth.

PHOSPHATE CONCENTRATION, 

MICROMOLAR

5.0 10

ZINC AND COPPER 

CONCENTRATION, MICROMOLAR

DISSOLVED OXYGEN 

CONCENTRATION, MICROMOLAR

100 200 300

10 20

BORON CONCENTRATION, 

MICROMOLAR

Figure 3. Vertical profiles of ground-water chemistry of site F230, approximately 100 meters downgradient of the active 

sewage disposal beds. (A) Dissolved orthophosphate concentration ( ) and pH (EJ) as a function of depth. (B) Dissolved 

zinc (v) and dissolved copper (A) as a function of depth. (C) Dissolved boron (O) and dissolved oxygen (*) as a function of 

depth. Error bars on sample points show screened interval on well.

195
Table 2. Chemical characteristics of ground water in the vicinity of MMR sewage-treatment plant

[Altitude measured in meters; all concentrations are units of micromolar except pH, which is on National Bureau of 

Standards pH scale; <, below detection limit of instrument]

Location

and

well site

Altitude Phosphate Zinc

above

sea level1

Copper Boron Dissolved

Oxygen

pH

Inactive sludge-drying pad

S434-13 13.21 4.20

S434-25 10.46 18.1

S434-75 -4.89 <3.2

Active sand bed

S318-36 10.24 177

Downgradient of active sand bed

S314-35 13.93 121

Abandoned sand bed

S438-41 12.52 116

2.78 

.61

.69

5.31

1.30

0.26 

<.06 

<.06

.84

1.5 

1.1

31.5

1.17 27.0

1.02 30.1

200

170

320

16

63

94

5.17

5.58

5.65

5.59

5.30

5.75

'Water table at 13.9 m above sea level.

Distribution of Copper and Lead

Dissolved Cu concentrations beneath the active 

sand beds were 0.5 to 1 jo,M throughout the vertical 

profile below the water table (fig. 2). Similar values 

for Cu were found in the ground water beneath the 

abandoned sand beds and 10 m downgradient from the 

sewage-disposal beds (table 2). At site F230, concentrations of Cu were only slightly greater than background Cu concentrations in the aquifer.

Concentrations of dissolved Pb were below the 

detection limit (less than 0.1 joM) in all samples.

Distribution of Phosphate

High concentrations of phosphate were 

observed in the ground water at all depths beneath the 

active sand beds (fig. 2, table 1). Phosphate concentrations in the ground water at this site were similar to 

those measured in a sample of the effluent (170 uM). 

Phosphate adsorbs strongly to minerals, but adsorption 

decreases with increasing pH (Sposito, 1984; Sigg and 

Stumm, 1980; Goldberg and Sposito, 1986a, 1986b). 

Thus, we expect that the observed concentrations of 

phosphate also are supported by a large reservoir of

adsorbed phosphate. A similar reservoir must exist in 

the abandoned sand beds because a high concentration 

of phosphate still was detected in the ground water 

(table 2) nearly 10 years after sewage disposal at that 

site had ended.

A distinct phosphate profile was observed in the 

ground water at site F230 (fig. 2). The phosphate concentration of the nearly pristine ground water near the 

water table was low; in the suboxic zone, high concentrations of phosphate, similar to those in the active 

sand beds, were observed. Low phosphate concentrations were measured in the anoxic ground water, similar to those reported earlier (Gschwend and Reynolds, 

1987).

DISCUSSION

The results of this study provide insight into the 

development of the vertical zonation of ground-water 

chemistry that characterizes the sewage plume. High 

concentrations of DO introduced to the upper region 

of the sewage-contaminated zone during infiltration of 

the sewage effluent (fig. 2) decreased during the first 

100 m of transport to concentrations characteristic of 

the suboxic zone (fig. 3). Low concentrations of DO

196
affect the redox conditions of the aquifer which, in 

turn, can cause changes in the mobility of various ions.

Zn and Cu are strongly adsorbed to sediments at 

the pH values observed at the sewage-disposal beds 

(Rea and others, 1991; Coston and others, 1992) indicating that a large reservoir of these elements must be 

adsorbed to the sediments. Lower concentrations and 

less movement of Cu downgradient compared to that 

of Zn are consistent with the greater adsorptivity of Cu 

(although complexation with organic material could 

enhance Cu mobility). The absence of dissolved Pb 

may result from the even greater adsorptivity of Pb 

compared to that of Cu at the pH in the ground water 

beneath the sewage-disposal beds (Davis and Kent, 

1990; Sposito, 1984). Alternatively, contamination by 

to Pb may be less extreme than that by Cu. Extractions 

of core material collected from the vicinity of the sewage-treatment plant, currently underway, will enable 

us to quantify the concentration of sorbed metals with 

improved accuracy. The contribution of metals from 

abandoned facilities (effluent-disposal beds and 

sludge-drying pads) is likely; leakage of effluent from 

the sewage treatment plant also is possible. Further 

investigation is required to determine all sources of 

input.

Phosphate transport in the suboxic zone probably is controlled by adsorption. In the anoxic zone, 

phosphate transport is closely linked to the availability 

of Fe2+ and the solubility of Fe2+-phosphate minerals 

(Gschwend and Reynolds, 1987). Extractions of core 

material will help evaluate the geochemical processes 

affecting phosphate mobility. Additional sampling 

downgradient is required to understand geochemical 

processes affecting phosphate transport.

SUMMARY AND CONCLUSIONS

The Cape Cod aquifer has been contaminated 

with Zn, Cu, and phosphate by a sewage plume emanating from the MMR sewage-treatment facility. As a 

result of the extensive adsorption of these solutes to 

aquifer solids under the prevailing chemical conditions of the ground water, the greatest amount of contamination is confined to a region near the source of 

contamination, including abandoned facilities. The 

geochemical processes controlling the mobility of Zn, 

Cu, and phosphate are strongly coupled to those processes that give rise to the steep vertical chemical gradients in the sewage plume. In the suboxic zone, 

mobility is controlled by adsorption processes, which

depend on pH, the concentration of the adsorbing species, and the concentrations of other solutes that compete for adsorption sites. The greater adsorptivity of 

Cu compared to that of Zn results in the observed 

decreases in mobility of Cu. Phosphate mobility in the 

suboxic zone is controlled by adsorption. In the 

anoxic zone, dissolved phosphate concentrations are 

likely controlled by the solubility of Fe2+ phosphate.

REFERENCES

Coston, J.A., Fuller, C.C., and Davis, J.A., 1992, The search 

for a geochemical indicator of lead and zinc sorption in 

a sand and gravel aquifer, Falmouth, Massachusetts, in 

Kharaka, Y.K., and Maest, A.S., eds., Water-Rock 

Interactions, WRI-7, v. 1, Balkema, Rotterdam, the 

Netherlands, p. 805-808.

Davis, J.A., Kent, D.B., Rea, B.A., Maest, A.S., and Garabedian, S.P., 1993, Influence of redox environment and 

aqueous speciation on metal transport in ground 

water preliminary results of tracer injection studies, 

in Alien, H., Perdue, E.M., and Brown, D., eds., Metals 

in Groundwater: Chelsca, Mich., Lewis Publishers, 

p. 223-273.

Davis, J.A., Kent, D.B., Rea, B.A., Garabedian, S.P., and 

Anderson, L.D., 1991, Effect of the geochemical environment on heavy-metal transport in ground water, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. Geological Survey Toxic Substances Hydrology Program 

Proceedings of the technical meeting, Monterey, 

California, March 11-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 

91-4034, p. 53-62.

Davis, J.A., and Kent, D.B., 1990, Surface-complexation 

modeling in aqueous geochemistry, in Hochclla, M.F., 

and White, A.F., eds., Mineral-Water Interface 

Geochemistry: Washington, D.C., Mineralogical 

Society of America, p. 177-260.

Dzomback, D.A. and Morel, F.M.M., 1990, Surface complexation modeling Hydrous ferric oxide: New 

York, John Wiley and Sons Ltd.

Focrstner, U. and Van Lierde, J.H., 1983, Trace metals in 

water purification processes, in Foerstner, U., and Wittmann, G.T.W., eds., Metal pollution in the aquatic 

environment: New York, Springer-Verlag, p. 324-359.

Franson, M., 1985, Standard methods for examination of 

water and wastewater: Washington, D.C., American 

Public Health Organization.

Goldberg, S. and Sposito, G., 1986a, A chemical model of 

phosphate adsorption by soils, I Reference oxide 

minerals: Soil Science Society of America Journal, 

v. 48, p. 772-778.

197
Goldberg, S. and Sposito, G., 1986b, A chemical model of 

phosphate adsorption by soils, II Noncalcareous 

soils: Soil Science Society of America Journal, v. 48, 

p. 779-783.

Gschwend, P.M. and Reynolds, M.D., 1987, Monodisperse 

ferrous phosphate colloids in an anoxic groundwater 

plume: Journal of Contaminant Hydrology, v. 1, no. 3, 

p. 309-327.

Kent, D.B., Davis, J.A., Anderson, L.C.D., and Rea, B.A., 

1991, Transport of zinc in the presence of a strong 

complexing agent in a shallow aquifer, Cape Cod, 

Massachusetts, in Mallard, G.E., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geolgoical Survey Water-Resources Investigations Report 91-4034, p. 78-83.

Kent, D.B, Davis, J.A., Anderson, L.C.D., Rea, B.A., and 

Waite, T.D., 1994, Transport of chromium and selenium in the suboxic zone of a shallow aquifer Influence of redox and adsorption reactions: Water 

Resources Research, v. 30, no. 4, p. 1099-1114.

K-V Associates, Inc., and IEP, Inc., 1991, Ashumet Pond, 

Falmouth/Mashpee, Massachusetts Volume 1, A 

diagnostic/feasibility study: Falmouth, Mass., 158 p.

LeBlanc, D.R., 1984, Sewage plume in a sand and gravel 

aquifer. Cape Cod, Massachusetts: U.S. Geological 

Survey Water-Supply Paper 2218, 28 p.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W., 

Quadri, R.D., Stoltenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural-gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 Experimental design and observed tracer movement: Water 

Resources Research, v. 27, no 5, p. 895-910.

Rea, B.A., Kent, D.B., LeBlanc, D.R., and Davis, J.A.,

1991, Mobility of zinc in a sewage-contaminated aquifer, Cape Cod, Massachusetts, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 88-95.

Sigg, L., and Stumm, W., 1980, The interaction of anions 

and weak acids with the hydrous goethite (a-FeOOH) 

surface: Colloids and Surfaces, v. 2, p. 101-117.

Smith, R.L. and Duff, J.H., 1988, Denitrification in a sand 

and gravel aquifer: Applied and Environmental 

Microbiology, v. 54, no. 5, p. 1071-1078.

Smith, R.L., Harvey, R.W. and LeBlanc, D.R., 1991, Importance of closely spaced vertical sampling in delineating 

chemical and microbiological gradients in ground 

water studies: Journal of Contaminant Hydrology, v. 7, 

no. 3, p. 285-300.

Sposito, G., 1984, The surface chemistry of soils: New 

York, Oxford University Press, 234 p.

198
Potential Long-Term Effects of Phosphate Contamination 

in a Sand and Gravel Aquifer, Cape Cod, Massachusetts

By Kenneth G. Stollenwerk1

Abstract

Phosphate in the Ashumet Valley sewage 

plume, located near Falmouth, Massachusetts, 

could increase the rate of eutrophication in 

Ashumet Pond. A series of column experiments 

has been initiated to identify the geochemical 

processes affecting transport of phosphate in this 

aquifer. The principal focus of these experiments 

is to determine the amount and rate of phosphate 

desorption from aquifer sediment after sewage 

disposal is stopped. The first set of column 

experiments simulated the oxic zone that 

surrounds the core of the plume. Results indicate 

that the sediment has a relatively high capacity to 

adsorb phosphate and that adsorption is 

kinetically controlled. The rate of desorption is 

slow. Approximately 160 pore volumes were 

eluted before the phosphate concentration 

decreased to levels considered not to cause 

eutrophication; half the phosphate was still 

adsorbed on the sediment. A one-dimensional 

solute transport code simulated the column data 

reasonably well; however, a kinetic term needs to 

be included in the model to improve accuracy.

INTRODUCTION

Part of the Ashumet Valley sewage plume from 

the Massachusetts Military Reservation sewageinfiltration beds is discharging to Ashumet Pond 

(fig. 1). Evidence of the discharge includes increased 

specific conductance and concentrations of boron and 

phosphate (PO4) in water samples collected from 

wells near the pond, from minipiezometers set in the

U.S. Geological Survey, Denver, Colo.

pond bottom, and from near-shore surface water 

(LeBlanc, 1984).

Ashumet Pond is phosphorus (P) limited. The 

dissolved PO4 concentration in Ashumet Pond (1991) 

averages 0.32 to 0.64 |j,mol/L, although localized areas 

of higher concentration have been measured near 

shoreline sources (K-V Associates, 1991). Generally, 

PO4 concentrations greater than 0.64 |imol/L are 

considered to cause eutrophication (Vollenweider, 

1968). The condition of Ashumet Pond is considered 

to be mesotrophic with signs of eutrophication (K-V 

Associates, 1991). Approximately 30 percent of the 

current annual PO4 input to the pond is from the 

sewage plume. Computer simulations predict that by 

the mid 1990's the sewage plume will contribute more 

than 70 percent of the PO4 load to the pond (K-V 

Associates, 1991).

Because of the potential for eutrophication, the 

National Guard has decided to begin remediation 

efforts to prevent future discharge of PO4 to the pond. 

Discharge of sewage at this site is scheduled to 

terminate in 1995. The National Guard is funding an 

investigation by the U.S. Geological Survey to obtain 

the hydrogeologic and geochemical information 

needed to evaluate methods for preventing discharge 

of PO4 in the aquifer upgradient of Ashumet Pond. 

Objectives of this study include (1) delineation of the 

PO4 plume between the sewage infiltration beds and 

Ashumet Pond; (2) description of ground-water flow 

between the sewage infiltration beds and Ashumet 

Pond; (3) identification of the geochemical processes 

affecting PO4 transport; and (4) predict future 

transport of PO4 after cessation of sewage discharge.

Phosphate in the plume is attenuated by sorption 

to aquifer sediment. As of the last major sampling, in 

1988, the leading edge of the sewage plume was about 

4.5 km from the infiltration beds, whereas the leading 

edge of the PO4 plume had only traveled 0.8 km 

(fig. 1). The amount and rate of PO4 released from the

199
71*00

73*00' 

I

72*00'

I

2*30' -/

MASSACHUSETTS

0 20 40 60 MILES

| I I______I

0 20 40 60 KILOMETERS

70° 34' 33' 70° 32'

. ___ - OTIS AIR BASE

-± 50 ___ -SEWAGE-DISPOSAL ^ u*j

AREA OF

PHOSPHORUS-CONTAMINATED, 

GROUND WATER

41° 38'

41° 37'

41° 36'

AREA OF

SEWAGE-CONTAMINATED

GROUND WATER

CRANBERRY 

BOGS

2000 FEET

I____I

500 METERS

EXPLANATION

 30 -^ WATER-TABLE CONTOUR -- Shows altitude of water table above sea

level. Contour interval 5 feet. Arrows show direction of ground-water flow.

OBSERVATION WELL--Used to delineate phosphorus-contaminated area

Figure 1. Water-table map showing location of sewage and phosphorus plumes, 1988. Phosphorus plume is defined by 

dissolved concentration greater than 1 |amol/L (Modified from LeBlanc, 1984).

200
sediment after sewage discharge is stopped could have 

a major effect on the remediation strategy. One 

method of evaluating the release of PO4 from aquifer 

sediment is through use of laboratory column experiments designed to simulate the flow and geochemical 

conditions in the plume as closely as possible. Different geochemical environments exist within the aquifer 

as a result of mixing of freshwater recharge with the 

sewage plume (Smith and others, 1991). This paper 

describes the results from a column experiment representing the partially oxic conditions that surround the 

core of the plume.

METHODS

Experimental

Replicate columns, 30 cm long by 2.5 cm inside 

diameter, were filled with the less than 2-mm-size 

fraction of sediment collected from the unsaturated 

zone of the aquifer above the sewage plume. This size 

fraction accounted for 90 percent by weight of the 

sediment sample. Pre-sewage-plume conditions were 

established in each column by rinsing the sediment 

with several pore volumes of uncontaminated ground 

water (UGW) collected from the aquifer. A bromide 

tracer was used to estimate the dispersivity for 

transport simulation. After the bromide was eluted 

from the columns, influent was switched to sewagecontaminated ground water (SCGW) collected from 

the oxic part of the plume. The SCGW contained 

15.8 famol/L PO4 and had a higher pH and dissolved 

solids content than did UGW (table 1). Sewagecontaminated ground water was eluted through the 

column until the PO4 concentration in effluent 

approached the influent concentration. The influent 

then was changed to UGW to evaluate the rinse-out of 

PO4 from the columns. Separate peristaltic pumps 

were used to drive the solutions through each column 

at a velocity of 0.4 m/d, the velocity of ground water 

in the aquifer (LeBlanc and others, 1991). Effluent 

was collected in a fraction collector.

Phosphate was analyzed by the 

phosphomolybdate colorimetric procedure (Fishman 

andFriedman, 1985). Sulfate, chloride, nitrate, nitrite, 

calcium, magnesium, potassium, sodium, and 

ammonium concentrations were analyzed by ion 

chromatography.

Table 1. Chemical composition of ground water used in the 

column experiment

[Concentrations in u.mol/L except pH which is in National Bureau of 

Standards pH scale; <, below detection limit]

Concentration

Solute Uncontaminated 

ground water

Na

K

Ca

Mg

CI

SO4

NO3

HCO3

PO4

pH

250

20

22

33

353

3I

<

98

<

5.6

Sewage-contaminated 

ground water

1 450

2I5

298

1 93

874

323

758

492

I5.8

'7.2

'Laboratory measured pH immediately after receiving sample was 

6.3. Degassing of COj caused pH increase to 7.2. No attempt was made to 

readjust pH for column experiments.

Solute Transport Model

The solute transport code HYTEQ (Kool, 1990) 

was used to simulate the reaction and transport of PO4 

in the column experiment. HYTEQ is a coupling of 

the HYDRUS one-dimensional solute-transport code 

(Kool and van Genuchten, 1990) and the MINTEQA2 

geochemical speciation code (Allison and others, 

1990). MINTEQA2 contains the diffuse-layer 

surface-complexation model that was used to simulate 

PO4 sorption. Input parameters for the diffuse-layer 

model are listed in table 2. The concentration of 

sorption sites was determined by potentiometric 

titration of sediment with hydrogen ion (H+ ) to a pH 

of 3.5. Equilibrium constants for the three phosphate 

reactions represent best fits to sorption data from batch 

experiments, which were conducted under the same 

chemical conditions used in the column experiments.

EXPERIMENTAL RESULTS

Phosphate was completely removed from the 

first 45 pore volumes of SCGW (fig. 2). A rapid 

increase in concentration was observed initially, 

followed by a much slower approach to complete 

breakthrough. Breakthrough curves for both columns 

were similar for the first 120 pore volumes. The 

greater concentration of PO4 in column 2 between 

pore volumes 65 and 80 reflects a period of time when

201
Table 2. Surface-complexation model parameters.

[FeOM° , neutral charged reactive surface site; LogK'nt, log of the intrinsic complexation constant]

Sediment properties

Sediment concentration - 4,600 grams per liter 

Surface area - 0.33 square meters per gram 

Concentration of complexation sites - 6,400 fimol/L

Surface complexation reactions LogKint

(1)FeOH°+ H + = FeOH2+

(2) FeOH° = FeCT + H +

(3) FeOH° + 3H+ + PO43 - = FcH 2 PO4° + H 2 0

(4) FeOH° + 2H+ + PO43 - = FeHPO^ + H 2 0

(5) FeOH° + H + + PO43 ' = FePO42 - + H2O

6.1

-7.3 

27.8 

21.6 

16.5

the average velocity was 20 percent greater in column 

2 than it was in column 1. As a result, there was less 

time for PO4 to react with sediment in column 2. At 

pore volume 96, the pump driving water through 

column 2 began to malfunction, resulting in very slow 

flow velocities. Accordingly, contact time in column 2 

increased, and more PO4 was removed from solution 

than would have been had the pump not malfunctioned.

The dependence of dissolved PO4 concentration 

on time of contact between PO4 and sediment is well 

established (see, for example, Lijklema, 1980) and 

indicates a kinetic control on PO4 sorption. Martin 

and others (1988) present convincing evidence for the 

formation of crystallites of griphite (Fe3Mn2 

(PO4)2 5 (OH)2 ) on the surface of naturally occurring 

goethite. They concluded that a mono-layer of PO4 

was rapidly adsorbed initially, followed by slow 

reprecipitation. The manganese (Mn) was supplied by 

Mn oxide impurities in the goethite. In the absense of 

Mn, an iron phosphate hydroxide mineral phase would 

form. On the other hand, Willett and others (1988) 

present equally convincing evidence that PO4 rapidly 

adsorbs on all available sites on fresh hydrous ferric 

oxide (HFO). The rate of PO4 removal from solution 

depended on the time required to diffuse to less 

accessible sites within the porous HFO structure. The 

different conclusions of these two studies may be a 

function of differences in experimental conditions and 

materials.

Sediment from the sewage-contaminated aquifer 

of this study contains amorphous to crystalline iron 

oxides. Both of the above mechanisms might 

contribute to the slow reaction of PO4 observed in the 

columns. Although recrystallization of a PO4-mineral 

phase has not yet been investigated, there is direct

evidence for diffusion. Breakthrough curves for 

molybdate from previous experiments using 

identically prepared columns could only be simulated 

by a diffusion mechanism (Stollenwerk and Kipp, 

1990). Whatever the mechanism, it is clear that flow 

velocities in column experiments should be 

representative of those in the aquifer to reflect the 

chemical and physical processes that operate in the 

aquifer.

Phosphate in effluent from both columns never 

attained the influent concentration of 15.8 |j,mol/L. 

Column 1 was switched to UGW at pore volume 115, 

and column 2 was switched at pore volume 160. PO4 

concentration in column 1 effluent was 13 |j,mol/L; in 

column 2 it was 15 (jmol/L. As UGW began to elute 

from both columns, a brief spike in PO4 concentration 

was observed (fig. 2). The spike could have been the 

result of an ion-exchange reaction, although the 

concentrations of other anions, such as sulfate, are less 

in UGW than in SCGW. Also, PO4 adsorbs much 

more strongly than do the other anions in UGW and 

SCGW. The spike also could represent an initial 

temporary equilibrium between PO4 that was loosely 

adsorbed to the exterior of grains and PO4-free UGW.

Phosphate in effluent from both columns rapidly 

declined at first, then changed to an asymptotic rate of 

decline. Desorption in column 1 began at pore volume 

115; however, the PO4 concentration did not decrease 

to the mesotrophic/eutrophic concentration boundary 

of 0.64 |amol/L until pore volume 270. The shape of 

the desorption curve for column 2 is similar. The 

amount of PO4 adsorbed or desorbed was quantified 

by keeping track of the mass balance for each aliquot 

collected from the column. A total of 60.8 and 

63.9 jimol were sorbed by sediment in columns 1 and 

2, respectively. Only about 45 percent had been 

desorbed by pore volume 270.

202
1 I

COLUMN 1 PHOSPHATE 

- COLUMN 2 PHOSPHATE 

COLUMN 1 pH

160 

PORE VOLUME

Figure 2. Breakthrough curves for phosphate from columns 1 and 2 and for pH from column 1. (C0 is the initial 

influent concentration of phosphate = 15.8 nmol/L; * represents the pore volume where influent was switched to 

uncontaminated ground water containing no phosphate).

EXPERIMENTAL 

SIMULATED - pH 5.5 

SIMULATED - pH 7.0

160 

PORE VOLUME

200 240 280 320

Figure 3. Experimental and simulated breakthrough curves for phosphate from column 1.

203
Also of interest are the pH profiles. Only the pH 

of effluent from column 1 is plotted in figure 2; the pH 

of column 2 effluent was similar. Although influent 

was switched to SCGW of pH 7.2 at pore volume 0, 

the pH of column effluent remained near 5.6 for 

almost 20 pore volumes before rapidly increasing to 

7.2. Apparently, the sediment has a buffering capacity 

for H+. Prior equilibration of the sediment with lowpH UGW saturated the oxide-mineral surface sites 

with H+. Desorption of H+ or ion exchange with 

cations in the SCGW maintained a low pH.

The reverse behavior was observed in the rinseout data. When pH 5.6 water was reintroduced to the 

columns, pH remained high. In fact, at pore volume 

275, the pH of effluent from both columns was still 

greater than 6. These data would seem to indicate that 

oxide minerals were removing H+ from solution. 

Effluent concentrations of calcium and potassium 

remained higher than the influent, indicating 

desorption, perhaps by ion exchange with H+. Future 

column experiments will be designed to address the 

pH question.

SIMULATION RESULTS

Simulation results with respect to column 1 

predict that the concentration of PO4 should rapidly 

increase to the influent concentration at pore volume 

35 (fig. 3). This is sooner than the observed 

breakthrough, which began at pore volume 43. The 

model also underpredicted the amount of PO4 

adsorbed. The equilibrium constants for PO4 sorption 

are based on data from batch experiments that were 

conducted for only 48 hours; thus, the time for 

intragranular diffusion was limited. Phosphate in the 

column experiment had a much longer time to diffuse 

to reaction sites. A kinetic term added to the transport 

model should improve the simulation.

Desorption of PO4 was simulated at two pH 

values. The pH 7 simulation was conducted at the 

average measured pH of column effluent between pore 

volumes 120 and 280 (pH 7). This simulation predicts 

that PO4 should be rapidly rinsed from the column in 

less than 20 pore volumes. The other simulation was 

conducted at the pH of the influent water (pH 5.5). 

Results exactly matched the actual desorption curve 

for PO4 with the exception of the initial concentration 

spike. This match is likely fortuitous, because the 

actual distribution of pH in the column should be 

somewhere between the values used in these

simulations. Near the influent end of the column, 

water would be expected to be in equilibrium with 

sediment at a pH of 5.5. Reactions would still be 

affecting pH at the effluent end of the column, thereby 

maintaining an increased pH. A more accurate 

simulation would take into account the changing pH 

within the column; however, this simulation has not 

yet proved feasible with the HYTEQ code. The pH 

could only be fixed at one value for a simulation.

SUMMARY

Evidence from column experiments indicates 

that transport of PO4 in the oxic zone of the sewage 

plume should be retarded by sorption. As a result, a 

reservoir of sorbed PO4 might exist between the 

sewage infiltration beds and Ashumet Pond. Data 

from column experiments indicate that a relatively 

large pulse of PO4 could initially elute from the 

aquifer after cessation of sewage discharge. 

Following this spike, PO4 will desorb and 

concentrations in ground water will decline slowly. 

Dissolved PO4 could remain at concentrations greater 

than the eutrophication limit of 0.64 ^irnol/L for many 

years. Column data also indicate that approximately 

half the sorbed PO4 might be irreversibly adsorbed to 

the sediment.

A major question that needs to be answered 

before considering remediation strategies concerns the 

amount of PO4 currently in the aquifer. The column 

experiments were run until complete breakthrough of 

PO4 was achieved. This condition might not represent 

the existing condition in the aquifer; thus, there might 

still be a substantial amount of sorption capacity left to 

intercept the PO4 plume. Furthermore, this first set of 

column experiments represents the lower end of the 

PO4-concentration range observed in one part of the 

aquifer. Additional experiments need to be done in the 

upper concentration range of the oxic zone, and 

conditions in the anoxic core of the plume need to be 

simulated. Finally, accurate prediction of the potential 

effect of PO4 on Ashumet Pond will require 

simulation of the PO4 plume in three dimensions. 

Initial simulations of one-dimensional column 

experiments are relatively accurate; however, some 

modifications to the model are necessary to improve 

accuracy.

204
REFERENCES

Allison, J.D., Brown, D.S., and Novo-Gradec, K.J., 1990, 

MINTEQA2/PRODEFA2, a geochemical assessment 

model for environmental systems Version 3.0 user's 

manual: Athens, Ga., U.S. Environmental Protection 

Agency, 106 p.

Fishman, M.J., and Friedman, L.C., eds., 1985, Methods for 

determination of inorganic substances in water and 

fluvial sediments: U.S. Geological Survey Techniques 

of Water-Resources Investigations, book 5, chap. Al, 

p. 489-491.

K.-V Associates, Inc., 1991, Ashumet Pond,

Falmouth/Mashpec, Massachusetts Volume 1, A 

diagnostic/feasibility study: Falmouth, Mass., 158 p.

Kool, J.B., 1990, HYTEQ: Computer model for onedimensional flow and multi-component solute 

transport with chemical equilibrium speciation, 

Version 1.0: HydroGeoLogic Inc., Herndon, Virg.

Kool, J.B., and van Genuchten, M.Th., 1990, HYDRUS: 

One-dimensional variably saturated flow and transport 

model, including hysteresis and root water uptake. 

Version 3.3: HydroGeoLogic Inc., Herndon, Virg.

LeBlanc, D.R., 1984, Digital modeling of solute transport in 

a plume of sewage-contaminated ground water, in 

LeBlanc, D.R., ed., Movement and fate of solutes in a 

plume of sewage-contaminated ground water, Cape 

Cod, Massachusetts U.S. Geological Survey Toxic 

Waste Ground-Water Contamination Program: U.S. 

Geological Survey Open-File Report 84-475, p. 11-45.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W., 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, no. 5, p. 895-910.

Lijklema, Lambertus, 1980, Interaction of orthophosphate 

with iron (III) and aluminum hydroxides: 

Environmental Science and Technology, v. 14, no. 5, 

p. 537-541.

Martin, R.R., Smart, R.St.C., and Tazaki, Kazue, 1988, 

Direct observation of phosphate precipitation in the 

goethite/phosphate system: Soil Science Society of 

America J., v. 52, p. 1492-1500.

Smith, R.L., Harvey, R.W., and LeBlanc, D.R., 1991, 

Importance of closely spaced vertical sampling in 

delineating chemical and microbiological gradients in 

groundwater studies: Journal of Contaminant 

Hydrology, v. 7, no. 3, p. 285-300.

Stollenwerk, K.G., and Kipp, K.L., 1990, Simulation of 

molybdate transport with different rate-controlled 

mechanisms, in Melchior, D.C., and Bassett, R.L., eds., 

Chemical Modeling of Aqueous Systems II: 

Washington, American Chemical Society Symposoum 

Series No. 416. chap. 19, p. 243-257.

Vollenweider, R.A., 1968, The scientific basis of lake and 

stream eutrophication, with particular references to 

phosphorous and nitrogen as eutrophication factors: 

Organization for Economic Cooperation and 

Development, Paris, France, Technical Report 

DAS/DSI/68-27, 182 p.

Willet, I.R., Chartres, C.J., and Nguyen, T.T., 1988,

Migration of phosphate into aggregated particles of 

ferrihydrite: Journal of Soil Science, v. 39, p. 275-282.

205
206
Background Aqueous Chemistry and Effects of Carbon 

Dioxide Variations in Recharge in Shallow Ground Water 

in a Glacial Outwash Aquifer, Cape Cod, Massachusetts

By Roger W. Lee1

Abstract

Previous geochemical research at the 

Otis Air Base Site, Cape Cod, Massachusetts, 

determined that the principal geochemical process 

developing background water chemistry of 

shallow ground water is CO2-controlled 

hydrolysis of sodium feldspars. Preliminary 

geochemical modeling, however, demonstrated 

that CO2 sources should vary laterally over the 

project area. Field analyses of unsaturated zone 

gases showed variations in available CO2 for 

recharge waters. These variations depend on land 

use and vegetative cover in the area of groundwater recharge. On the basis of measurements of 

CO2 concentrations in unsaturated-zone gases at 

20 sites, CO2 dissolved in recharge water can 

range from about 0.035 to 1.000 millimoles per 

liter in the project area. The various land-use 

practices and local surface features in and around 

Otis Air Base (residential subdivisions, golf 

courses, large cultivated grass fields, kettle ponds, 

and woodlands) are associated with variations in 

the amount of CO2 in the unsaturated zone.

INTRODUCTION

Recent research in ground water at the Otis Air 

Base research site, Cape Cod, Massachusetts, has 

focused on hydrologic, geochemical, and microbiological processes in and around a sewage plume 

in the glacial outwash aquifer. The initial research

'U.S. Geological Survey, Austin, Texas

identified the mineral-water interactions and the role 

of CO2 in the uncontaminated parts of the aquifer. The 

term "background" is used to describe ground water 

not contaminated by the sewage effluent. This term is 

not to be confused with "natural" or "pristine" ground 

water, which implies that ground water has not been 

affected by manmade changes to land use. Background water can be affected by fertilizers, septic-tank 

effluent, road salting, or other land uses in the area of 

study.

Research into the background geochemistry 

around Otis Air Base has been described from groundwater samples collected in 1988 and 1989 (Lee, 1991). 

The principal reaction determined for the shallow 

ground water is the CO2-controlled hydrolysis of 

sodium feldspar. Geochemical mass-transfer models 

were developed using the computer code PHREEQE 

(Parkhurst and others, 1980) to test hypotheses of 

chemical evolution of shallow ground water against 

actual chemical data. The purpose of this paper is to 

characterize CO2 content and seasonal variations in 

the unsaturated zone under different land uses. 

Samples for analysis of CO2 concentration in gases of 

the unsaturated zone were collected at 20 locations for 

different seasons during 1991-92.

BACKGROUND AQUIFER CHEMISTRY

The chemistry of background shallow ground 

water is determined by the chemistry of recharge 

water and the interaction of that recharge water with 

minerals and gases in the unsaturated and saturated 

zones. Development of conceptual and quantitative 

models provides insight necessary to improve the 

knowledge and understanding of the geochemical processes occurring within the aquifer.

207
Mineralogy

The water-table aquifer consists of 90 to 200 

feet of glacial outwash sand and gravel that overlies a 

thick glacial till deposit (Oldale, 1969). The area of 

interest is in a valley (locally referred to as Ashumet 

Valley) that extends from the southern part of Otis Air 

Base to as far south as the cranberry bogs (fig. 1). The 

valley is bounded by sand and gravel valley ridges on 

the west and on the east. Ground water generally flows 

from north to south parallel to the axis of the valley 

(fig. I). Thorough discussions of the hydrology and 

geology of the site are given by 

LeBlanc (1984) and Garabedian and LeBlanc (1991).

Minerals in the glacial outwash sands and 

gravels are described by Wood and others (1990) and 

Barber and others (1992). The sediments are wellsorted fine- to medium-grained sands with interspersed gravel and cobble zones. The sands contain 

principally quartz with small amounts of feldspar 

(albite, orthoclase, and microcline). Quartz, plagioclase, and orthoclase constitute about 95 percent by 

weight of the mineralogy with accessory minerals of 

glauconite, goethite, biotite, muscovite, magnetite, and 

minor amounts of ferroaluminosilicates and oxides 

(Barber and others, 1992). Clay minerals are sparse, 

with trace amounts (less than 5 percent by weight) of 

kaolinite, muscovite, chlorite, and some smectite. 

Magnetite is present in some areas, as are ferruginous, 

manganiferous, and organic coatings on gravels and 

sand grains.

Ground-Water Chemistry

Water chemistry and the distribution of solutes 

associated with the sewage plume are described by 

LeBlanc (1984) and Lee (1991). Background water 

chemistry is dilute and generally contains less than 

150 mg/L dissolved solids (table 1). Hydrochemical 

facies range from no dominant cations or anions to 

sodium bicarbonate- or sodium chloride-dominated 

water. Waters that contain large concentrations of 

dissolved solids (greater than 100 mg/L) and are dominated by sodium and chloride ions indicate possible 

contamination from winter road-salting practices 

(D.R. LeBlanc, U.S. Geological Survey, oral 

commun., 1988). Ground water containing sewage 

effluent has substantial amounts of dissolved boron 

(up to 400 |J.g/L); whereas background water generally 

contains less than 50 jig/L dissolved boron (LeBlanc,

1984). All ground-water samples used in this study 

contain less than 50 |J.g/L dissolved boron and are considered uncontaminated by the sewage effluent. 

Samples from sites F294, F357, F358, F373, and F433 

in map view appear to be from the contaminant plume, 

but, in reality, are from uncontaminated water above 

the sewage plume (fig. 1). The plume is overlain by up 

to 45 ft of uncontaminated ground water derived from 

precipitation that recharges the aquifer. Incongruent 

dissolution of feldspars to kaolinite has been shown to 

be a reaction consistent with observed water chemistry 

(Lee, 1991)--

(albite) 

2NaAlSi3O8

(kaolinite)

2CO2 + 11H2O =>

+4H4SiO4 (1)

If a homogeneous mineral assemblage and 

uniform recharge chemistry are assumed, then the 

chemical samples collected at increasing depths in the 

aquifer at all sites would be expected to show an 

increase in the chemical evolution of water resulting 

from increased contact time with the aquifer matrix. 

At some of the well sites, however, concentration of 

dissolved solids, which can indicate the degree of 

chemical evolution, decreases with depth (Lee, 1991).

Reaction Model

Based on observed chemical data, aquifer 

mineralogy, and chemical composition of recharge, it 

is possible to construct a reaction sequence for the 

chemical evolution of background water in the glacial 

sand and gravel aquifer. The extent of chemical 

reaction of silicate hydrolysis is determined from the 

acid content of the recharge water, which is a function 

of the amount of CO2 entering the ground-water 

system (Busenberg and Clemency, 1976).

Variations in chemical data are probably 

attributable to nonuniform distributions of CO2 in 

recharge water over the project area. This condition 

can occur because of areas of thin soil cover over the 

quartz sands and the effects of local differences in 

land cover, such as bare or poorly vegetated areas, 

mixed hardwood and conifer woodlands, large grassy 

areas on Otis Air Base and local parks, and cultivated 

grassy areas in suburban landscaping and golf 

courses. Variations in land cover have been shown to 

affect water chemistry in the unsaturated zone (Moss, 

1990). These surficial features in the study area could 

affect CO2 production in the unsaturated zone

208
2'30* -j

73-00' TZ'OO- 

/-" -- -- -- -- 

MASSACHUSETTS

0 20 40 60 MILES

| I I____I

0 20 40 60 KILOMETERS

70" 34' 33'

41° 38'

41° 37'

41° 36' 

I S315* I

, _____- OTIS AIR BASE 

-± 50 -"^_ SEWAGE-DISPOSAL

AREA OF 

CONTAMINATED 

GROUND WATER

CRANBERRY 

BOGS

EXPLANATION

 30 -4 WATER-TABLE CONTOUR -- Shows altitude of water table above sea

level. Contour interval 5 feet. Arrows show direction of ground-water flow.

F412 OBSERVATION WELL

Figure 1 . Water-table map and locations of background water samples for the area of investigation.

209
ro_t

o

Table 1 . Well numbers, depths, and water-chemical data for background water samples from the Cape Cod site (Ch

per liter unless otherwise specified)

Depth 

below 

land 

surface 

Well no. Site no. (feet)

1

2

3

4

5

6

1

8

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

S3 13.038

S3 13.060

S3 13.080

S315.149

S31 6.082

S316.134

F234.099

F235.094

F238.106

F240.095

F24 2.077

F290.09 1

F294.064

F294.109

F299.020

F357.079

F357.099

F358.049

F358.089

F373.060

F373.ll 3

F4 12.042

F4 12.064

F4 12.078

F4 12.091

F41 2.108

F418.089

F418.103

F41 8.122

F418.141

F433.064

F433.090

F433.104

F436.036

F436.060

F436.076

Mean

38

60

80

149

82

134

99

94

106

95

77

91

64

109

20

79

99

49

89

60

113

42

64

78

91

108

89

103

122

141

64

90

104

36

60

76

Temperature 

<°C)

9.0

10.0

10.0

13.0

11.0

11.0

10.0

10.5

10.0

9.5

10.0

10.0

10.0

9.5

9.0

9.5

12.2

10.5

9.5

13.0

12.2

10.0

9.5

9.5

9.5

9.5

9.5

9.5

9.5

9.5

9.0

9.0

9.5

10.0

13.0

9.0

PH 

(units)

5.70

5.60

5.90

6.08

6.50

6.50

5.70

5.60

5.70

5.80

5.80

6.20

5.80

6.30

5.40

5.60

5.60

5.50

6.00

5.68

6.31

5.50

5.80

5.50

5.50

5.70

5.80

6.20

6.40

6.40

5.50

5.90

6.30

5.60

5.76

6.70

5.88

Calculated 

dissolved 

solids

37

65

47

61

139

64

64

66

70

36

38

61

56

75

59

63

45

25

33

53

100

28

81

170

182

120

54

72

65

63

44

71

78

31

54

81

68

Ca

0.93

2.40

2.60

2.17

27.00

3.90

3.90

4.20

4.50

1.50

1.50

4.40

1.40

8.80

7.70

1.80

1.79

1.10

1.20

1.75

8.82

0.66

0.54

3.30

4.40

3.20

3.70

6.70

5.10

4.40

1.40

2.60

5.50

0.88

2.70

12.00

4.22

Mg

0.89

3.00

2.30

1.10

7.10

2.80

2.80

1.90

4.10

1.80

1.90

2.00

1.80

4.30

1.80

3.20

2.06

0.85

0.96

1.66

3.36

1.20

1.00

4.50

5.90

3.50

3.20

3.00

2.30

2.10

1.80

2.40

2.40

1.20

1.81

2.90

2.52

Na

9.30

12.00

6.30

7.94

14.00

6.70

9.60

7.20

8.20

5.80

6.30

8.50

15.00

14.00

5.40

13.00

6.85

4.00

5.70

10.50

15.30

4.40

25.00

52.00

54.00

34.00

7.20

10.00

9.70

9.30

8.80

17.00

16.00

5.80

10.70

10.00

12.93

K

0.80

1.60

0.80

0.52

5.30

4.40

1.40

1.60

0.90

0.50

0.60

0.80

0.70

1.10

1.10

0.90

0.60

0.90

0.50

0.68

1.42

0.50

0.90

1.70

2.10

0.90

0.80

1.00

1.20

1.00

0.90

1.30

1.50

1.00

0.71

1.70

1.23

HCO3

3.48

3.01

3.18

11.98

35.78

10.02

5.73

3.29

4.08

4.62

6.68

10.48

2.05

24.01

2.79

4.04

9.99

0.75

3.57

5.99

40.65

0.87

2.00

0.50

2.06

3.88

2.34

7.98

14.37

13.65

2.19

8.53

18.11

1.74

6.99

33.67

6.92

S04

5.7

21.0

15.0

7.7

22.0

16.0

17.0

25.0

27.0

5.4

7.3

6.8

4.8

4.9

9.1

13.0

5.3

4.3

5.8

6.5

6.0

6.0

19.0

5.3

7.4

7.1

4.6

5.9

7.2

6.1

7.0

9.0

10.0

6.0

4.5

13.0

9.8

Cl

10.0

12.0

6.8

12.0

37.0

8.6

11.0

12.0

9.8

10.0

9.2

9.0

24.0

14.0

9.0

16.0

11.2

7.8

7.0

14.6

16.2

9.3

27.0

94.0

100.0

60.0

16.0

21.0

9.7

9.4

13.0

23.0

12.0

9.6

19.5

12.0

19.2

NO3

2.08

3.54

2.44

1.42

1.82

5.76

5.76

3.37

3.99

0.71

0.44

13.70

0.01

0.44

17.20

4.87

3.30

0.00

1.82

5.75

9.29

0.00

0.53

1.86

0.00

2.30

8.86

10.60

6.64

10.10

4.03

3.94

8.41

0.00

1.50

2.39

4.14
and the subsequent flux of water and CC>2 through the 

unsaturated zone to the water table.

Significant vertical variations of chemical 

constituents can indicate one or more components 

of inhomogeneity of the chemical system. This 

inhomogeneity complicates development of a single 

quantitative mass-transfer model. The simulation 

process is further complicated by the difficult task of 

collecting chemical data along true flow paths in a 

water-table aquifer for comparison with results of 

geochemical mass-transfer models.

CO2 IN THE UNSATURATED ZONE

In order to improve the understanding of CO2 

variability in recharge water in the project area, it was 

necessary to analyze for CC>2 concentrations in the 

unsaturated zone for the different land uses. Because 

much of the production of CC>2 occurs in the root zone 

of the soil and subsoil, the C(>> concentration can vary 

significantly from season to season and with depth 

within the unsaturated zone (Reardon and others, 

1979; Wood and Petraitis, 1984).

Land Use

Although the project area contains parks, 

residential areas, golf courses, a wildlife sanctuary, 

Otis Air Base, and other land uses, four principal types 

of land cover affect CC>2 production in the unsaturated 

zone. These are: areas where vegetative cover is 

sparse to absent, such as gravel pits and utility rightof-ways; mixed hardwood and conifer woodlands; 

grassy areas of parks, and much of Otis Air Base; and 

landscaped residences, and golf courses (fig 2).

Methods

Twenty sites in the project area were selected to 

sample and analyze soil gases (fig 2). Four were 

selected to show typical analytical results (fig. 3a-3d). 

Gas samples were withdrawn through a 1 /2-inchdiameter hollow probe driven to a specified depth. 

Concentration profiles of CC>2 in unsaturated zone gas 

were obtained at each site to depths of up to 12 ft. 

Samples were collected over five consecutive seasons 

represented by the months of November 1991, 

February 1992, May 1992, July 1992, and November 

1992. Samples were collected by first withdrawing the

unsaturated-zone gases at a rate of 1 L/min with a 

peristaltic pump through intake slots near the base of 

the probe set to the specified depth. A constant value 

could usually be obtained by pumping for about 

1 minute per ft of probe. Samples were withdrawn 

through the silicon tubing attached to the probe using a 

50-mL plastic syringe with needle. The sample was 

immediately injected into a field gas chromatograph, 

which was equipped with a thermal conductivity 

detector. Peak separation of the CC>2 component of the 

gas was achieved using a micro-column packed with 

the HayesSep A and following the conditions of 

temperature and flow rate recommended in the user's 

manual for the equipment. The chromatograph contains an internal reservoir that accurately subsamples 

and passes the gas through the analytical column. 

Analytical precision was ± 1 percent and accuracy was 

+ 2 percent. Concentration of CC>2 was measured in 

parts per million, volume/volume. Calibration of the 

instrument was achieved using commercial standard 

gases.

Results of Gas Sample Analyses

Selected sites representative of results of the 

sampling for four significant land uses (Site a, grassy 

field; Site K, gravel pit; Site O, woodland; Site Q, golf 

course) are shown in figures 3a-d. Ranges of CC>2 concentrations given below represent approximate ranges 

of values in the subsurface from the actual data at each 

site. The smallest concentrations of C(>> were in the 

gravel-pit areas where CC>2 ranged from about 500 to 

2,000 ppm. Much of the vegetation had been removed 

at these locations, and the generation of CC>2 at the 

root zone was probably limited. The woodland areas 

contained the next smallest concentrations of CC>2 

about 2,000 to 8,000 ppm. Woodlands were characterized by a moderate density of hardwood-conifer mix 

of trees, and low density ground cover with some 

decaying vegetation and mulch. CC>2 data from the 

woodlands overlapped CC>2 data for uncultivated 

grassy areas about 6,000 to 8,000 ppm. Uncultivated 

grassy areas were covered with slow-growing grasses 

that were maintained by occasional mowing but were 

not watered or fertilized. The largest concentrations of 

CC>2 were found in the cultivated grassy areas on the 

golf courses 10,000 to 50,000 ppm. Hybrid grasses, 

high growth density, intense watering, and fertilization 

are conducive to elevated production of CC>2 in the 

root zone.

211
70*34' 70*33' 70*32'

41*39'

41*38'

41*37'

41*36'

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F433

o

2000 FEET

500 METERS

EXPLANATION

yiy^^ff- "-r 

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AIR BASE

SEWAGE FACILITY

GRASSY FIELD

RESIDENTIAL

POND

GOLF COURSE

WETLAND

GRAVEL PIT

WOODLAND

S GAS SAMPLING LOCATION

WELL LOCATION

F238

Figure 2. Land-use map in the vicinity of Otis Air Base, Cape Cod, Massachusetts, showing well cluster locations and 

unsaturated-zone-gas sampling locations.

212
a. Site K--Gravel Pit b. Site O-W

DEPTH, IN FEETBELOW 

LAND SURFACE SURFACE

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 ^B A

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A FEBRUARY, 1992 

O MAY. 1992 

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CO2 CONCENTRATION, IN PARTS PER 

MILLION (VOLUME/VOLUME)

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3

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CO2 CONCENTRATIO

MILLION (VOLUME/

Figure 3. Selected sites of the four principal land uses in the vicinity of Otis Air Base showing concentrations of carbon dioxide 

NJ 

CO
Changes in CO2 concentrations in gases of the 

unsaturated zone also were variable with depth, 

depending on the season and the land use. Concentrations of CC>2 in the unsaturated zone were greater than 

concentrations of CC>2 in the atmosphere. At gravel-pit 

areas (fig. 3a), changes in CC>2 concentrations with 

depth below the root zone (0 to 2 ft) were small to zero 

with increasing depth. At woodland sites (fig. 3b), 

CC>2 concentrations increased with increasing depth 

during all the seasons sampled. At uncultivated grassy 

field sites (fig. 3c), CC>2 concentrations decreased 

slightly with increasing depth. The CC>2 concentrations were usually largest in the upper part of the 

unsaturated zone because of the production of CC>2 in 

large concentrations in the root zone beneath the 

grasses. At golf-course sites (cultivated grasses) 

(fig. 3d), the largest variations in CC>2 concentration 

were found. Concentrations of C(>> were observed to 

increase with depth, not change with depth, or 

decrease with depth, depending on the season of the 

year. These effects are presumed to result from photosynthetic activity of the cultivated grasses and production of CC>2 in the shallow root zone. At site Q 

(fig. 3d), CC>2 concentration decreased with increasing 

depth, and relatively large CC>2 concentrations were 

measured during the February 1992 sampling. These 

results reflect late winter growth of the golf-course 

grasses. Another possibility is ice that formed in the 

low-lying area of the sample site may have inhibited 

escape of CC>2 to the atmosphere from the unsaturated 

zone.

Small or no increases in CC>2 concentration with 

increasing depth indicate an unsaturated zone at near 

steady-state with CC>2 production from the vegetative 

cover. An increase in CC>2 concentration with depth 

implies lower CC>2 production in the root zone at the 

time of sampling than at some recent (within hours) 

previous time. A decrease in CC>2 concentration with 

depth implies that CC>2 production in the root zone is 

greater at the time of sampling than at some recent 

previous time. The response time of gas concentrations in the unsaturated zone to large amounts of CC>2 

production in the root zone likely depends on the distance from the source of CC>2, the rate of gaseous diffusion, and the rate of downward movement of 

moisture. Although no time of response to CC>2 production in the root zone was determined for the various sites, the probable amount of time would be on the 

order of hours to days; achievement of true steadystate CC>2 concentrations in the unsaturated zone 

would be unlikely.

Gradients in CC>2 concentration in the 

unsaturated zone also depended on the season. During 

active vegetative growth for example, in July 1992 

CC>2 production in the soil zone developed decreasing 

CC>2 concentrations from the root zone to the water 

table. During dormant periods, especially in February 

1992, the gradient was reversed, indicating that larger 

amounts of CC>2 may be dissolved by unsaturated-zone 

water and carried to the water table during periods of 

vegetative growth and root respiration in summer than 

during dormant periods in winter. CC>2 may then be 

lost from the aqueous phase through the unsaturated 

zone to the atmosphere when the gradient in the unsaturated zone is reversed during dormant periods in 

winter. The net amount of CC>2 that enters recharge 

water is determined by the amount of CO2 generated 

in the soil and root zone less the amount of CC>2 lost to 

the unsaturated zone and the atmosphere. Thus, CO2 

concentration in recharge water in the project area is 

variable, depending to a large extent on the vegetative 

cover in the recharge area.

Application of Henry's law to analytical data of 

unsaturated-zone gases for the purpose of calculating 

aqueous CC>2 in recharge at 10 °C (typical temperature of shallow ground water, LeBlanc, 1984), produces the following approximate ranges of 

concentrations of CC>2 dissolved in recharge water for 

each of the four land usesgravel pits 

woodlands 

grassy fields 

golf courses

0.035-0.150 mmol/L 

0.150-0.350 mmol/L 

0.250 - 0.350 mmol/L 

0.400- 1.000 mmol/L

The initial reaction model developed by Lee 

(1991) assumed consistent recharge-water chemistry, 

aquifer mineralogy, and generation of CO2 during percolation of recharge through the unsaturated zone. In 

order to improve the comparison between the background chemical data and model results for solutes, 

pH, and CC>2, it was necessary to vary amounts of CC>2 

available for chemical reaction in the computer code 

PHREEQE (Parkhurst and others, 1980). The 

chemistry of recharge water, CC>2 from decay of 

organic matter in the aquifer, and mineral mass 

transfers were held constant for each of the six models 

(table 2). Only CC>2 in recharge water was changed in 

each of the models. CC>2 model input to recharge 

water ranged from 0.035 to 1.00 mmol/L in order to 

simulate the CC>2 in unsaturated zones for each of the 

four land uses described above. The model results are 

presented in table 2; the results cover most of the 

observed ranges of concentrations of solutes and show 

chemical evolution of pH and total CO2-

214
Table 2. PHREEQE mass-transfer simulations [Concentrations in millimoles per liter unless otherwise 

specified]

RECHARGE WATER 

 Woodlands 

C02

pH (units)

Ca

Mg

Na

K

Cl

S04

N03

0.038

0.037

0.0076

0.251

C02

pH (units)

Ca

Mg

Na

K

Cl

S04

N03

SiC-2

Gravel Pits

Model 1 Model 2 

0.150 

5.15

Golf Courses 

Model 5 Model 6

0.035

5.48

Grassy Fields 

Model 3 Model 4

0.250 0.350 0.400 1.00 

5.04 4.97 4.94 4.74 

0.036 0.036 0.036 

0.036 0.036 0.036 

0.205 0.205 0.205 

0.011 0.011 0.011 

0.216 0.216 0.216 

0.072 0.072 0.072 

0.004 0.004 0.004

Add Minerals to above Solution in 15 Steps 

NaCl (halite) 0.300 NaAlSi308 (albite) 

KAlSi308 (orthoclase) 0.075 CaAl2Si208 (anorthite) 

Mg5Al2Si3Oio (chlorite) 0.900 C02 (carbon dioxide)

Subtract (precipitate) Mineral from Solution at Each Step 

Al2Si205 (OH)4 (kaolinite) total subtracted in step 15

FINAL RESULTS

0.935 1.050 1.150 1.250 1.300 1.900 

6.60 6.49 6.40 6.34 6.31 6.05 

0.086 0.086 0.086 

0.061 0.061 0.061 

0.430 0.430 0.430 

0.036 0.036 0.036 

0.241 0.241 0.241 

0.072 .072 0.072 

0.004 0.004 0.004 

0.455 0.455 0.455

215
Figure 4 graphically depicts the chemical 

evolution of shallow ground water from the approximate ranges of CO2 in recharge water for the four 

major land uses, as well as the actual chemical data 

for the 36 wells in the project area. With minor 

exceptions, the chemical data for wells generally are 

within the range of values prescribed by the models, 

indicating the principal conceptual reaction process of 

CO2-controlled feldspar hydrolysis to kaolinite. Also, 

it is possible to estimate from the graph which type of 

land use and mechanism for CO2 generation was 

prominent in the recharge area for each well sampled.

Figure 2 shows the principal land uses in the 

area of Otis Air Base, site locations of wells sampled, 

and site locations of samples of unsaturated zone gas 

samples. The land use in the vicinity of a particular 

water well likely will affect CO2 concentration in the 

shallow ground water as a result of recharge through 

that land use. Different land uses upgradient may 

produce significantly different CO2 concentrations in 

recharge, that may, in turn, affect chemical reactions 

along deeper flow paths in the downgradient direction. 

The result of this process would be small vertical gradients of concentrations of dissolved solids in ground 

water, increasing in some places and decreasing in 

others. For example, water from wells at site S313 

show a slight increase in concentrations of dissolved 

solids from 38 ft to 60 ft below land surface. A 

decrease in concentration of dissolved solids was

noted from 60 ft to 80 ft below land surface, however 

(table 1). This indicates an upgradient source of 

recharge to the deepest well (80 ft) through a land use 

that produces concentrations of CO2 lower than 

recharge to shallower wells. Poorly vegetated areas or 

woodlands in the vicinity of S313 could be in the 

recharge area for well 3 (fig. 4). Recharge to wells 1 

and 2 probably passed through greater CO2-producing 

land uses such as grassy areas (fig. 4) on the sewage 

treatment facility.

CONCLUSIONS

Background ground water at the Otis Air Base 

site, Cape Cod, Massachusetts, contains small concentrations of dissolved solids (less than 150 mg/L). Solutes found having elevated concentrations in the study 

area are sodium, bicarbonate, and chloride. Sodium 

and chloride increases have likely resulted from the 

application of road salt for deicing roadways. The 

chemistry of ground water appears to be dominated by 

CO2-controlled weathering of silicates, principally 

sodium feldspar, with the formation of kaolinite. 

Increases in depth in the aquifer do not always correlate with increases in concentrations of dissolved solids as might be expected for a relatively homogeneous 

chemical system. PHREEQE mass transfer modeling, 

based principally on incongruent feldspar dissolution,

Q.

7.0

6.5

6.0

5.5

5.0

4.5

«* * » 

 . *12 * » 3* * * *4 5* **«*<? * *

* O D ^1 .*** \

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19 , " ****** \

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MODEL 

RESULTS

" .35 

,**20

 

25 31

GRAVEL " 

PITS GRASSY 

FIELDS

1 7

 BACKGROUND WELL AND

IDENTIFICATION NUMBER (table 1)

WOODLANDS GOLF COURSES

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0

TOTAL INORGANIC CARBON CONCENTRATION, 

IN MILLIMOLES PER LITER

Figure 4. Observed and simulated (PHREEQE model) relations between total inorganic carbon concentrations and pH for 

selected land uses.

216
has demonstrated that variable CC>2 uptake by 

recharge water in the unsaturated zone is one process 

that could account for some of the chemical differences observed in background water in the glacial outwash aquifer. Land use, such as bare or poorly 

vegetated areas, mixed hardwood and conifer woodlands, and grassy areas on Otis Air Base, in local 

parks, in suburban landscaping, and on golf courses, 

affect the production of CO2 and its flux from the 

unsaturated zone into the aquifer. Measured amounts 

of CC>2 in gases of the unsaturated zone underlying the 

principal land uses and vegetative covers indicate 

large variations in the amount of CC>2 available to 

recharge waters. The amount of unsaturated-zone C(>> 

that dissolves in recharge water ranges from about 

0.035 to 1.00 mmol/L in the study area. The amount of 

CO2 available to react with feldspar minerals in the 

aquifer determines, in part, the concentrations of 

dissolved constituents in the ground water. This 

relation can account for areas where concentrations of 

dissolved solids in ground water decrease with 

increasing depth when sodium and chloride ions from 

road salting are not involved.

REFERENCES

Barber, L.B., II, Thurman, E.M., and Runnells, D.D., 1992, 

Geochemical heterogeneity in a sand and gravel 

aquifer Effect of sediment mineralogy and particle 

size on the sorption of chlorobenzenes: Journal of 

Contaminant Hydrology, v. 9, no. 1, p. 35-54.

Busenberg, E., and Clemency, C.V., 1976, The dissolution 

kinetics of feldspars at 25 °C and 1 atm CC>2 partial 

pressure: Geochimica Cosmochimica Acta, v. 40, 

p. 41-49.

Garabedian, S.P., and LeBlanc, D.R., 1991, Overview of 

research at the Cape Cod site Field and laboratory 

studies of hydrologic, chemical, and microbiological 

processes affecting transport in a sewage-contaminated 

sand and gravel aquifer, in Mallard, G.E., and Aronson, D.A., eds., U.S. Geological Survey Toxic Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91 4034, p. 1 9

LeBlanc, D.R., 1984, Sewage plume in a sand and gravel 

aquifer, Cape Cod, Massachusetts: U.S. Geological 

Survey Water-Supply Paper 2218, 28 p.

Lee, R.W., 1991, Background aqueous chemistry of shallow 

ground water in a sewage-contaminated glacial 

outwash aquifer, Cape Cod, Massachusetts, in Mallard, 

G.E., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program-Proceedings of 

the technical meeting, Monterey, California, March 

11-15, 1991: U.S. Geological Survey Water-Resources 

Investigations 91-4034, p. 96-101.

Moss, P.D., 1990, Acid attenuation in the unsaturated zone 

of a sandstone aquifer: Proceedings, Geological 

Society of America, Dallas, Texas, p. A339.

Parkhurst, D.L., Thorstenson, D.C., and Plummer, L.N., 

1980, PHREEQE-A computer program for geochemical calculations: U. S. Geological Survey WaterResources Investigations 80-96, 210 p.

Plummer, L.N., Jones, B.F., and Truesdell, A.H., 1976, 

WATEQF-A FORTRAN IV version of WATEQ, 

a computer program for calculating chemical equilibrium in natural waters: U.S. Geological Survey WaterResources Investigations 76-13, 61 p.

Oldale, R.N., 1969, Seismic investigations on Cape Cod, 

Martha's Vineyard, and Nantucket, Massachusetts, and 

a topographic map of the basement surface from Cape 

Cod Bay to the islands: U.S. Geological Survey 

Professional Paper 650-B, p. B122-B127.

Reardon, E.J., Allison, G.B., and Fritz, P., 1979, Seasonal 

chemical and isotopic variations of soil CO2 at Trout 

Creek, Ontario: Journal of Hydrology, v. 43, 

p. 355-371.

Wood, W.W., and Petraitis, M. J., 1984, Origin and distribution of carbon dioxide in the unsaturated zone of the 

Southern High Plains of Texas: Water Resources 

Research, v. 26, no. 9, p. 1193-1208.

Wood, W.W., Kraemer, T.F., and Hearn, P.P., Jr., 1990, 

Intragranular diffusion, an important mechanism 

influencing solute transport in clastic aquifers?: 

Science, v. 247, p. 1569-1572.

217
218
Heavy-Metal Transport in a Sand and Gravel Aquifer 

with Variable Chemical Conditions, Cape Cod,

Massachusetts

By James A. Davis 1 , Douglas B. Kent 1 , Jennifer A. Coston 1 , andKathryn M. Hess2

Abstract

An overview is presented of a field-based 

research program that is examining the significance of chemical speciation and variable aquifer 

chemistry on the transport of toxic metals in 

ground water. Natural-gradient tracer tests and 

laboratory experiments with subsurface materials 

have been used to identify the most important 

chemical reactions that influence the transport of 

chromium(VI) and metal-ethylenediaminetetraacetic acid (EDTA) complexes in the aquifer at 

the Cape Cod Toxic-Substances Hydrology 

Research site. The field experiments are being 

conducted in an uncontaminated, recharge zone 

and a mildly reducing, sewage-contaminated zone 

of the shallow, sand and gravel aquifer. The 

results of several years of research at the site have 

been used to design a large-scale tracer test at the 

site (begun in April 1993) involving eight tracers 

(bromide, chromium, zinc, copper, lead, nickel, 

EDTA, and potassium). The new experiment will 

be the most complex and detailed investigation of 

multispecies, multireaction transport ever conducted in the field. The pertinent results of previous small-scale tracer tests are reviewed, and the 

objectives and expected results of the new tracer 

test are discussed.

'U.S. Geological Survey, Menlo Park, Calif. 

2U.S. Geological Survey, Marlborough, Mass.

INTRODUCTION

Chemical reactions can have a major effect on the 

transport of inorganic solutes (Yeh and Tripathi, 1991; 

Liu and Narasimhan, 1989). The reactivity of these 

solutes depends strongly on chemical speciation. 

Species that precipitate or are extensively adsorbed are 

highly retarded, whereas species that are soluble and 

weakly adsorbed can be mobile in ground water (Davis 

and others, 199la). For example, zinc (Zn) has been 

shown to be extensively retarded in ground water when 

present as a free (uncomplexed) cation (von Gunten and 

others, 1991; Rea and others, 1991). In contrast, the 

formation of weakly adsorbing, anionic complexes 

greatly increases the mobility of Zn (Kent and others, 

1991,1992). Other metals, including Cu(II) (copper, in 

the plus two oxidation state) and a variety of 

radionuclides, have been shown to be mobile in ground 

water as a result of their presence as weakly adsorbing 

organic complexes (Jacobs and others, 1988; von 

Gunten and others, 1991; Killey and others, 1984). 

Changes in pH along ground-water flow paths can 

cause shifts in metal speciation from soluble to 

sparingly soluble species. In addition to complexation 

and acid-base reactions, oxidation-reduction (redox) 

reactions may cause changes in speciation that have a 

marked effect on transport. For example, aqueous 

Cr(VI) species (chromium, in the plus six oxidation 

state) are generally soluble and relatively weakly 

adsorbing; hence, they can be mobile in ground water. 

However, Cr(III) species, if uncomplexed, should be 

highly retarded in ground water because they are 

extensively adsorbed or insoluble (Kent and others, 

1994). Other inorganic contaminants subject to changes 

in speciation due to redox reactions include selenium 

(Se), sulfur, arsenic, uranium, nitrogen, Cu, iron (Fe), 

and manganese (Mn).

219
Although considerable advances are being made 

in the development of mathematical algorithms (Yeh 

and Tripathi, 1991; Rubin, 1990), many practical 

problems remain if reactive-transport models are to be 

applied to ground-water systems in predictive 

simulations. For example, redox reactions are 

generally included in models as half-reactions, so that 

one can either impose a constant activity of the 

aqueous electron, or, by including both oxidation and 

reduction reactions and considering the conservation 

of electrons, one can allow the model to compute the 

redox equilibrium state and the resultant electron 

activity. In applying a model to a ground-water 

system, however, neither of these approaches may be 

satisfactory. Particular solutes of interest may not 

equilibrate with an electron activity that is estimated 

from measuring the species of major redox couples. 

Calculations assuming a conservation of electrons 

would require that all electron transfer reactions are 

accounted for in the model; this may be difficult when 

immobile organic matter or another reductant 

associated with the porous medium is the electron 

donor. When neither of these approaches apply, an 

important component of reactive-transport modeling 

will be detailed site characterization and laboratory 

studies that allow an identification of the important 

chemical reactions that need to be considered. This is 

the approach we have emphasized in our research on 

the transport of reactive metals. Adsorption reactions 

also require special considerations. Some reactivetransport models have relied on empirical parameters, 

such as distribution coefficients or isotherms 

determined in laboratory studies, to describe 

retardation of solutes during transport. While this 

approach may be satisfactory under constant chemical 

conditions, the surface complexation modeling 

approach will provide better results under conditions 

of variable chemistry (Davis and Kent, 1990). 

Adsorption of inorganic solutes is affected by 

competive adsorption processes, solute speciation 

changes, and other factors under conditions of variable 

chemical conditions. In the surface complexation 

approach, adsorption is considered as a set of one or 

more chemical reactions involving an aqueous species, 

a surface site of the porous medium, and protons. In 

principle, reactions are required for each adsorbing 

solute; however, laboratory experiments may be used 

to minimize the number of reactions needed for a 

transport modeling application.

The purpose of this paper is to provide an 

overview of the research on multispecies transport at 

the Cape Cod site, including a discussion of the 

objectives of a new large-scale field experiment that 

was begun in April 1993. The new experiment is 

designed to evaluate the effects of speciation and 

variable aquifer chemistry on metal transport in 

ground water to a greater degree of complexity than 

has been done previously. To set the context of the 

research, the paper includes a review of the findings of 

previous research at the site that were pertinent to the 

design of the new field experiment.

SITE DESCRIPTION

The U.S. Geological Survey (USGS) has been 

conducting research for several years at a site in 

Falmouth, Mass., on Cape Cod (fig. 1). The research 

site is in an area of sand and gravel outwash that has 

been contaminated by land disposal of treated sewage 

for more than 50 years. A contaminant plume more 

than 5 km long has formed as a result of the sewage 

disposal (LeBlanc, 1984; Barber and others, 1988; 

Hess and others, 1991). The sewage plume sinks as it 

travels southward with the ground water. Average 

ground-water velocities determined from 

measurements of hydraulic conductivity, porosity, and 

hydraulic gradient range from 0.25 to 0.5 m/d 

(Garabedian and others, 1991; LeBlanc and others, 

1991).

In 1984-85, the USGS installed a spatially dense 

well field in an abandoned gravel pit at this site to 

begin three-dimensional studies of solute transport in a 

field setting. A large number (640) of multilevel 

sampling devices (MLS) were installed, arranged in 

71 rows which extend about 300 m downgradient from 

the first row. In preparation for the project described 

here, an additional 100 MLS were installed in the field 

during 1992. Each multilevel sampling device 

includes 15 sampling ports spaced 25 to 76 cm apart 

vertically to provide vertical profiles of tracer 

concentrations. In total, there is an array of over 

11,000 sampling points downgradient from the 

injection wells, allowing detailed and statistically 

significant studies of transport to be conducted.

The aquifer contains two zones with distinct 

chemical characteristics (Kent and others, 1994). The 

uncontaminated recharge zone has high concentrations 

of dissolved oxygen (DO), low pH (5.1), and low 

concentrations of dissolved salts (about

220
42° 30'-/

70° 00'

100 FEET 

-i 

30 METERS

EXPLANATION

 13.45 WATER-TABLE CONTOUR, AUGUST 2, 1985 -- 

Shows altitude of water table. Contour interval 

0.05 meters. Datum is sea level.

Figure 1. Location of large-scale natural-gradient tracer test. (Modified from LeBlanc and others, 1991, fig. 4.)

221
30 jj,S/cm). The sewage-contaminated zone is suboxic 

(mildly reducing) with a higher pH (6.1) and high concentrations of dissolved salts (about 300 (iS/cm). 

Fe(II) and dissolved sulfide are both below detection 

limit in the region of the anoxic zone penetrated by the 

MLS's; dissolved Mn concentrations range from 5 to

BACKGROUND

Large-scale Tracer Test

The site was used during 1985-87 for a detailed 

natural-gradient study of bromide, lithium, and 

molybdate transport (LeBlanc and others, 1991; 

Garabedian and others, 1991). The tracer test was 

designed to examine the relationship between fieldscale dispersion and aquifer heterogeneity developed 

by Gelhar and Axness (1983). Major conclusions of 

the study were as follows: (1) longitudinal mixing was 

the dominant dispersion process; (2) transverse 

horizontal and vertical dispersion were relatively 

small; and (3) horizontal displacement of the injected 

solute cloud was accurately predicted using estimates 

of the hydraulic conductivity, porosity, and measured 

hydraulic gradient (Garabedian and others, 1991). 

Solute concentrations were highly variable and 

difficult to predict on a small scale but the average 

characteristics (the spatial moments) could be 

predicted. Molybdate transport depended on many 

factors, including pH and the concentration of other 

anions (for example, phosphate and sulfate), and the 

sorption isotherm was nonlinear (Stollenwerk and 

Grove, 1987).

Small-scale, Multispecies Tracer Tests

The large-scale tracer test conducted in 1985-87 

characterized the hydrology of the site in great detail. 

This comprehensive work and the dense array of MLS 

have made it possible to conduct more complex 

geochemical transport experiments with reactive 

tracers. The Solute Partitioning project (USGS, 

Menlo Park, Calif.) has been conducting field- and 

laboratory-scale experiments to study the transport of 

solutes for which speciation is important (Kent and 

others, 1994, 1995, 1991, 1992; Anderson and others, 

1994; Davis and others, 1991 b). Our objectives are to: 

(1) identify the important chemical processes affecting

transport, (2) examine the relations among chemical 

and physical factors affecting transport, and (3) 

stimulate the development of coupled flow and 

geochemical models that describe the transport of 

these solutes.

Twelve small-scale natural-gradient tracer tests 

were performed during 1988-92 to study the transport 

of Zn, nickel (Ni), ethylenediaminetetraacetic acid 

(EDTA, a metal-complexing agent), Cr(VI), and 

Se(VI) in oxic and suboxic zones of the shallow sand 

and gravel aquifer at the Cape Cod site (Kent and 

others, 1992, 1994, 1995; Davis and others, 1991b). 

In each test, approximately 400 or 600 L of ground 

water were withdrawn from an MLS into an acidwashed polyethylene bag; the tracers were dissolved 

therein, and the "spiked" ground water was injected 

into several ports of the same or another MLS. 

Bromide ion (Br) was used as a conservative tracer. 

Breakthrough of the tracers was monitored at samplers 

downgradient from the injection MLS. Typically the 

movement of the tracers was monitored for several 

weeks following the injection; the travel distance for 

bromide ion was usually less than 20 meters during the 

experiments. Methods for the measurement of tracers 

are given in Kent and others (1994).

Cr transport showed a marked dependence on its 

speciation and the chemical composition of the 

aquifer. In the oxic zone, Cr(VI) transport was 

retarded relative to Br and there was little evidence for 

reduction (Kent and others, 1995). This was likely due 

to adsorption of Cr(VI) onto minerals such as Fe and 

Mn oxides, which is favored in the oxic zone by the 

low pH and low concentrations of anions that compete 

for adsorption sites, such as sulfate and phosphate. In 

the suboxic zone, Cr transport was increasingly 

attenuated with distance transported, but was only 

slightly retarded. The attenuation was due to 

reduction of Cr(VI) to Cr(III), which is immobile 

(Kent and others, 1994). Analysis of the speciation of 

dissolved Cr showed that Cr(VI) was the only mobile 

species. Cr(VI) transport was only slightly retarded in 

the suboxic zone because Cr(VI) adsorption was 

decreased by the higher pH and greater concentrations 

of competing sorptive anions present. The amount of 

Cr(VI) reduced along a flowpath appeared to be 

limited by the rate of diffusive transport to reductive 

sites present in the porous medium. Prediction of the 

amount of Cr reduced from thermodynamic 

considerations alone was not possible.

222
Transport of Zn- and Ni-EDTA complexes were 

also affected by the chemical characteristics of the 

site. In the oxic zone, both Zn and Ni transport were 

retarded. Dissolution of iron and aluminum oxides 

(enhanced by the low pH in this zone) resulted in 

dissociation of Zn from EDTA, with exchange for Fe 

and Al, which form very strong complexes with 

EDTA. From laboratory experiments, it is known that 

Zn2+ adsorbs extensively on the aquifer sediments in 

the absence of EDTA (Davis and others, 1993), so the 

transport of Zn2+ would be significantly retarded after 

dissociation from EDTA. In contrast, the rate of 

NiEDTA dissociation was slow. Even though the local 

conditions probably favored exchange of Fe for Ni in 

EDTA complexes at chemical equilibrium, this 

reaction did not occur to a significant extent.

At the suboxic site, Zn and Ni transport were 

only slightly retarded, and the breakthrough curve for 

total EDTA showed the same degree of retardation 

(Kent and others, 1992). Thus, both Zn and Ni were 

transported as EDTA complexes. In the transition 

zone between the oxic and suboxic zones, Zn 

concentrations increased downgradient from the 

injection MLS, as excess EDTA induced the 

desorption of Zn contamination that was adsorbed on 

the porous medium (Kent and others, 1991). The 

source of Zn contamination on the porous medium in 

the transition zone is the sewage effluent plume (Rea 

and others, 1991).

NEW LARGE-SCALE TEST WITH 

REACTIVE TRACERS

A new large-scale field experiment was begun 

on April 21, 1993, to study in detail the effects of 

chemical speciation on reactive transport. 

Approximately 10,000 L of native ground water were 

pumped into a clean plastic tank in which the chemical 

tracers were mixed with the water. The tracers added 

to the ground water were Br, potassium, Cr(VI), Zn, 

Ni, Cu, and lead (Pb), and EDTA. The spiked ground 

water was pumped into three ports each of six customdesigned MLS over an 8-hour period. All MLS 

surrounding and within the tracer cloud ~ that is, 

spatial "snapshots" ~ were sampled shortly after the 

injection and are now being sampled on a monthly 

basis until November 1993. The spatial sampling will 

then be performed on a bimonthly basis until July 

1994. Approximately 40,000 water samples will be 

collected for analysis of tracer concentrations over the

duration of the project. Water samples will be 

analyzed after each sampling trip to determine the 

concentrations of all chemical tracers as well as 

several other elements. Dissolved oxygen, pH, and 

other water quality characteristics will be determined 

in selected field measurements at the time of water 

sampling.

The tracer test will exploit the vertical gradient 

in ground-water chemistry at the site to study transport 

under variable chemical conditions. Small-scale tracer 

tests conducted previously at the site have shown that 

metal transport is controlled by a different set of 

chemical reactions in the oxic and suboxic zones. 

Therefore, to show the importance of speciation and 

chemical conditions on metal transport, tracers were 

injected in the new test over a larger vertical interval 

that includes both zones. This vertical interval also 

has a significant pH gradient that will affect the 

adsorption of metal ions. The top of the tracer cloud 

will move within the oxic zone, and the lower part of 

the tracer cloud will move within the suboxic zone. In 

the transition zone, at the interface between these two 

zones, is the well-defined region of zinc contamination 

derived from sewage effluent. In this zone, the Zn 

contamination will affect metal transport via 

competitive processes (for both adsorption and 

aqueous complexation).

The data collected will allow detailed 

characterization of the metal transport behavior, and 

the effects of chemical processes on metal transport 

will be quantified. Chromium(VI) transport is 

expected to be controlled by sorption reactions in the 

upper part of the tracer cloud and redox reactions in 

the lower part of the tracer cloud. Lead, copper, zinc, 

and nickel mobility in the aquifer will be controlled by 

several chemical factors, including the concentration 

of EDTA, concentrations of other cations competing 

for coordination with the ligand, and the abundances 

of adsorbing mineral phases in the aquifer. Because of 

differences in the thermodynamics and rates of 

dissociation of metal-EDTA complexes, it is expected 

that Pb transport will be attenuated first in the suboxic 

zone, followed by Zn and then Cu. Ni probably will 

not be dissociated from EDTA in the suboxic zone 

during the test, and its transport will be controlled by 

adsorption of the Ni-EDTA complex only. In the zone 

of zinc contamination, zinc will be desorbed by 

complexation with any excess EDTA and readsorbed 

farther downgradient as dispersion lowers the 

concentration of the ligand.

223
The research will allow observations of metal 

transport over a significant distance (up to 200 m) and 

time (up to 1.5 years), depending on the degree of 

attenuation and retardation during transport. The data 

collected will then allow a comparison of chemical 

reaction parameters (for example, sorption coefficients 

and redox reaction rates) in solute-transport models 

with those that have been measured in laboratory 

experiments and small-scale tracer tests at the site. The 

rate of Cr(VI) reduction observed in laboratory 

experiments suggests that slow processes may be more 

important in a large-scale tracer test. A large-scale 

tracer test also will allow an examination of the effect 

of dispersion on chemical speciation and transport. 

The physical heterogeneity of the aquifer has made it 

difficult to model transport in small-scale experiments. 

The greater transport distance and larger dimensions 

of the tracer cloud envisioned for the large-scale tracer 

test will improve the quality of mass balances for 

tracers and make it possible to consider certain 

simplifying assumptions within a solute-transport 

model. In addition to these objectives, the research 

will identify general methods and approaches for 

successful prediction of metal transport and stimulate 

the development of transport models that account for 

the effects of variable chemistry on solute mobility.

REFERENCES

Anderson, L.D., Kent, D.B., and Davis, J.A., 1994, Batch 

experiments characterizing the reduction of Cr(VI) 

using suboxic material from a mildly reducing sand 

and gravel aquifer: Environmental Science and 

Technology, v. 28, p. 178-185.

Barber, L.B., II, Thurman, E. M., and Schroeder, M.P., 

1988, Long-term fate of organic micropollutants in 

sewage-contaminated groundwater: Environmental 

Science and Technology, v. 22, p. 205-211.

Davis, J. A., Fuller, C.C., Coston, J. A., Hess, K. M. and 

Dixon, E.. 1993, Spatial heterogeneity of geochemical 

and hydrologic parameters affecting metal transport in 

groundwater: Environmental Research Brief, U. S. 

Environmental Protection Agency, Ada, OK, 

EPA/600/S-93/006, 49 p.

Davis, J.A., and Kent, D.B., 1990, Surface complexation 

modeling in aqueous geochemistry, in Hochella, M.F., 

and White,, A.F., cds., Mineral-Water Interface 

Geochemistry: Mineralogical Society of America, 

Washington, D. C., Reviews in Mineralogy, v. 23, 

p. 177-260,

Davis, J.A., Kent, D.B., Garabedian, S.P., and Anderson, 

L.C.D., 199 la, Effect of the geochemical environment 

on heavy metal transport in ground water, in Mallard, 

G.E., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Monterey, California, March 11 - 

15, 1991: U. S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 53-62.

Davis, J.A., Waite, T.D., Kent, D.B., and Anderson, L.C.D., 

1991b, Reduction and transport of chromium(VI) 

under mildly reducing conditions in a sand and gravel 

aquifer, in Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U. S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 72-77.

Garabedian, S.P., LeBlanc, D.R., Gelhar, L.W., and Cclia, 

M.A., 1991, Large-scale natural-gradient tracer test in 

sand and gravel. Cape Cod, Massachusetts, 2 

Analysis of spatial moments for a nonreactive tracer: 

Water Resources Research, v. 27, p. 911 -924.

Gelhar, L.W., and Axness, C.L., 1983, Three-dimensional 

stochastic analysis of macrodispersion in aquifers: 

Water Resources Research, v. 19, p. 161-180.

Harvey, R.W, George, L.H., Smith, R.L., and LeBlanc, 

D.R., 1989, Transport of microsphercs and indigenous 

bacteria through a sandy aquifer Results of naturaland forced-gradient tracer experiments: Environmental 

Science and Technology, v. 23, p. 51-56.

Harvey, R.W., Smith, R.L., and George, L., 1984, Effect of 

organic contamination upon microbial distributions 

and heterotrophic uptake in a Cape Cod, Mass., 

Aquifer: Applied and Environmental Microbiology 

v. 48, p. 1197-1202.

Hess, K.M., Wolf, S.H., Celia, M.A., and Garabedian, S.P., 

1991, Macrodispersion and spatial variability of 

hydraulic conductivity in a sand and gravel aquifer, 

Cape Cod, Massachusetts: Environmental Research 

Brief, Environmental Protection Agency, Ada, OK., 

p. 9, EPA/600/M-91/005.

Jacobs, L.A., von Gunten, H.R., Keil, R., and Kuslys, M., 

1988, Geochemical changes along a river-groundwater 

infiltration flow path: Geochimica et Cosmochimica 

Acta, v. 52, p. 2693-2706.

Kent, D.B., Davis, J.A., Anderson, L.C.D., and Rea, B.A., 

1995, Transport of chromium and selenium in a 

pristine sand and gravel aquifer Role of adsorption 

processes: Water Resources Research, v. 31, no. 4, 

p. 1041-1050.

Kent, D.B., Davis, J.A., Anderson, L.C.D., Rea, B.A., and 

Waite, T.D., 1994, Transport of chromium and 

selenium in the suboxic zone of a shallow aquifer 

Influence of redox and adsorption reactions: Water 

Resources Research, v. 30, p. 1099-1114.

224
Kent, D.B., Davis, J.A., Anderson L.D. and Rca, B.A.,

1991. Transport of zinc in the presence of a strong 

complexing agent in a shallow aquifer on Cape 

Cod, Massachusetts, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 78-83. 

Kent, D.B., Davis, J.A., Rea, B.A., and Anderson, L.D.,

1992. Ligand-enhanced transport of strongly adsorbing 

metal ions in the ground-water environment, in 

Kharaka, Y., and Maest, A.S., eds., Water-Rock 

Interactions: Rotterdam, The Netherlands, p. 805-808.

Killey, R.W.D., McHugh, J.O., Champ, D.R., Cooper, E.L., 

and Young, J.L., 1984, Subsurface cobalt-60 migration 

from a low-level waste disposal site: Environmental 

Science and Technology, v. 18, p. 148-157.

LeBlanc, D.R., 1984, Sewage plume in a sand and gravel 

aquifer, Cape Cod, Massachusetts: U. S. Geological 

Survey Water-Supply Paper 2218, 28 p.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W., 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural-gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, p. 895-910.

Liu, C.W., and Narasimhan, T.N., 1989, Redox-controlled 

multiple-species reactive chemical transport, 1 

Model development: Water Resources Research, v. 25, 

p. 869-882.

Rca, B.A., Kent, D.B., LeBlanc, D.R., and Davis, J.A., 

1991, Mobility of zinc in a sewage-contaminated 

aquifer, Cape Cod, Massachusetts, in Mallard, G. E.

and Aronson, D. A., eds., U. S. Geological Survey 

Toxics Substances Hydrology Program Proceedings 

of the Technical Meeting, Monterey, California, March 

11-15, 1991: U. S. Geological Survey WaterResources Investigations Report 91-4034, p. 88-95.

Rubin, J., 1990, Solute transport with multisegment, 

equilibrium-controlled reactions A feed forward 

simulation method: Water Resources Research, v. 26, 

p. 2029-2055.

Smith, R.L., and Duff, J.H. 1988, Denitrification in a sand 

and gravel aquifer: Applied and Environmental 

Microbiology, v. 54, p. 1071-1078.

Smith, R.L., Harvey, R.W., and LeBlanc, D.R., 1991b, 

Importance of closely spaced vertical sampling in 

delineating chemical and microbiological gradients in 

groundwater studies: Journal of Contaminant 

Hydrology, v. 7, p. 285-300.

Stollenwerk, K. G., and Grove, D. B., 1987, Naturalgradient tracer test in sand and gravel 

nonconservative transport of molybdenum, in Franks, 

B. J., ed., U.S. Geological Survey Program on Toxic 

Waste Ground-Water Contamination, Proceedings of 

the Third Technical Meeting, Pensacola, Florida, 

March 23-27, 1987: U. S. Geological Survey OpenFile Report 87-109, Chap. B, p. 17-22.

von Gunten, H.R., Karametaxas, G., Krahenbuhl, U., 

Kuslys, M., Giovanoli, R., Hoehn, E., and Keil, R., 

1991, Seasonal biogeochemical cycles in riverborne 

groundwater: Geochimica et Cosmochimica Acta 

v. 55, p. 3597-3609.

Yeh, G.T., and Tripathi, VS., 1991, A model for simulating 

transport of reactive multispccies components Model 

development and demonstration: Water Resources 

Research, v. 27, p. 3075-3094.

225
226
Evaluation of Geochemical Indicators of Metal 

Adsorption in a Sand and Gravel Aquifer, Cape Cod,

Massachusetts

By Christopher C. Fuller1 , Jennifer A. Coston1 , James A. Davis 1 , and Eleanor Dixon 1

Abstract

The geochemical properties of an aquifer 

sand that control metal-ion adsorption were 

investigated to determine their potential use as 

indicators of the spatial variability of metal 

adsorption in the aquifer. Over the length of a 

4.5-meter long core, lead (Pb) and zinc (Zn) 

adsorption varied by a factor of two and four, 

respectively. Pb2+ and Zn2"1" are adsorbed 

primarily by iron and aluminum oxide coatings on 

quartz-grain surfaces. Upon normalization to 

surface area, both Pb2+ and Zn2"1" adsorption were 

significantly correlated with the amount of iron 

and aluminum dissolved from the aquifer material 

by partial chemical extraction. The use of partial 

chemical extraction combined with surface-area 

measurement as a potential indicator of metal 

adsorption in other aquifers needs to be tested.

INTRODUCTION

Modeling the transport of metals requires use of 

accurate adsorption coefficients to describe metal ion 

sorption onto surfaces of aquifer solids. In natural 

systems, metal adsorption is commonly thought to be 

controlled by iron (Fe) and aluminum (Al) oxide 

coatings on particle surfaces (Davis and Kent, 1990). 

Metal adsorption by pure oxide phases is well 

described by surface-complexation models (Dzombak 

and Morel, 1990). However, determining the 

appropriate type and surface density of adsorption

U.S. Geological Survey, Menlo Park, Calif.

sites can cause difficulties in applying surfacecomplexation models to mixed assemblages of 

mineral surfaces comprising natural systems (Davis 

and Kent, 1990). In addition, the spatial variability in 

adsorption properties within an aquifer, like hydraulic 

conductivity (Hess and others, 1992), can be of 

importance in describing solute transport and need to 

be accounted for (Robin and others, 1991).

A study of the spatial heterogeneity of metal-ion 

adsorption and hydraulic conductivity at a site on 

Cape Cod has recently been completed (Davis and 

others, 1993). The objectives of the study were to 

determine the predominant metal adsorbing surface in 

the aquifer solids, to quantify the spatial variability of 

metal adsorption and hydraulic conductivity, and to 

determine if a geochemical property of the aquifer 

material can be used as an indicator of the surfaces 

controlling lead (Pb2+) and zinc (Zn2+) adsorption. 

An analysis of the spatial variability of metal 

adsorption and hydraulic conductivity is included in 

this volume (Hess and others, 1995).

This paper summarizes the results of the 

evaluation of geochemical indicators of metal 

adsorption in the aquifer and their potential use for 

describing the spatial variability of lead and zinc 

adsorption in the aquifer sand. For example, if the 

variability in metal ion adsorption is caused by a 

greater reactivity with the surfaces of a secondary 

mineral phase or coating, the partial chemical 

extraction results could describe the variability of the 

reactive phase. The properties of the aquifer sand 

evaluated include Pb2+ and Zn adsorption, surface 

area, grain-size distribution, and partial chemical 

extraction.

227
SITE DESCRIPTION

Aquifer material was collected at the U.S. 

Geological Survey Toxic Substances-Hydrology 

Research site near Falmouth, Massachusetts (fig. 1). The

site is a shallow, unconfined aquifer of glacial outwash 

(LeBlanc and others, 1991). Sediments in the aquifer 

consist predominantly of medium to coarse sand, with 

localized zones of gravel, fine sand, and silt. The aquifer 

has an average porosity of 0.39 and an average ground73»00'

42* 30'-/

TRANSECT CORES

F415C38

F415C40,

F415C41 -Jj 

F415C42* 

F415C43*

F415C53* 

F415C66 '

30 FEET

10 METERS

F415C39

F415C54

EXPLANATION

 CORING SITE LOCATION AND IDENTIFIER 

Figure 1. Site map with coring locations, Cape Cod, Massachusetts.

228
water velocity of 0.4 m/d. The amount of sedimentary 

organic carbon in the sand and silt fractions is between 

0.01 percent and 0.05 percent of the dry weight (Barber 

and others 1992). The sand is primarily quartz (95 

percent) with minor amounts of alkali feldspars and 

ferromagnetic minerals (Coston and others, 1992).

MINERAL SURFACES THAT CONTROL 

METAL ADSORPTION

9-4- 9+ Adsorption of Pb andZn by the aquifer sand is 

less dependent on pH than is typically observed for pure 

oxides (Coston and others, 1992). The adsorption edges 

span 2 pH units compared to 1 unit for pure oxides 

(Davis and Kent, 1990). Adsorption of Pb2+ and Zn2+ on 

the highly weathered, coarse aquifer sand was dominated 

by quartz grains (95 percent by weight) in the sand 

(< 1,000 fj,m diameter). The reactivity of the natural 

quartz surfaces was attributed to 10- to 300-nm thick 

coatings on the sand grains that are composed of surface 

precipitates of complex mixed oxides and silicates 

containing high concentrations of Fe and Al. The 

coatings are derived from the weathering of feldspars and 

other accessory minerals in the aquifer. A decrease in 

Pb2+ adsorption was observed with a diamagnetic sand 

treated with hydroxylamine hydrochloride to remove Feand Al-oxide coatings. These results indicate that partial 

chemical extractions could be useful as indicators of 

metal adsorption.

METHODS

Collection of Aquifer Material

Cores for the spatial heterogeneity study (Fuller 

and others, 1991) were collected using a wirelinepiston core barrel and plastic core liners (Zapico and 

others, 1987) and were frozen until used. Three 

consecutive 1.5-m long cores were collected from 

each bore hole. The material from one transect core, 

F415 C38 (fig. 2), was used for this study. Cores were 

subdivided into 10-cm-long vertical subsamples, dried 

in a laminar flow hood, and sieved to remove grains 

greater than 1,000-^m diameter.

Determination of Variability of Metal 

Adsorption by Batch Experiments

For each vertical subsample, Pb and Zn2+ 

adsorption were measured in three batch experiments 

using a sand to artificial ground water (AGW) ratio of

94- 94- 50 g/L. Pb and Zn were added simultaneously 

from a combined stock solution to yield total 

concentrations of 20 and 5 jiM, respectively. Each of 

the three batch experiments for a subsample was 

designed to attain a different pH value after 48 hours 

of reaction (pH 5.1, 5.3, and 5.5), by addition of 

predetermined aliquots of 0.04M NaOH or 0.04M 

HNO}. The 48-hour reaction period was chosen on 

the basis of adsorption kinetic experiments (Davis and 

others, 1993). The amount of metal adsorption was 

calculated from the change in the dissolved metal 

concentration. Metal concentrations were determined 

by inductively coupled plasma atomic emission 

spectrometry (ICAP-AES). Adsorption results for 

each of the pH values for each vertical subsample 

were normalized to pH 5.3 using the Kurbatov 

equation with a macroscopic proton coefficient of one 

(Kurbatov and others, 1951; Honeyman and Leckie, 

1986), and averaged.

Characterization of Aquifer Material for 

Indicators of Metal Adsorption

Three partial-chemical-extraction methods were 

used to determine if Pb2+ and Zn2 "1" adsorption 

correlate with the amount of Fe, Al, and manganese 

(Mn) extracted from the aquifer material. Partial 

chemical extractions of aquatic sediments and aquifer 

materials are commonly used to estimate the 

partitioning of trace metals among secondary phases 

such as amorphous Fe, Al, and Mn oxides (Tessier and 

others, 1979; Chao, 1984; Robinson, 1984/85; Tessier 

and others, 1989).

Sand samples were treated with hydroxylamine 

hydrochloride (0.25 M NH2OH»HC1 in 0.25 M HC1) at 

50 °C for 0.5 and 72 hours. Hydroxylamine-HCl (HA) 

extracts iron by reductive dissolution at low pH (Chao 

and Zhou, 1983). The HA extraction is not selective for 

Fe; for example, Al and Mn also were extracted from the 

sand. Al dissolved by HA could be caused by partial 

dissolution of feldspars and other aluminosilicate phases 

as well as aluminum oxide coatings on grain surfaces. 

The HA extraction for 0.5 hours was chosen because it 

has been reported to be the most effective method for 

selective dissolution of amorphous Fe oxyhydroxides 

with minimal dissolution of crystalline Fe oxyhydroxides 

(Chao and Zhou, 1983). Dissolved Fe concentrations 

increased over time in HA extracts of a 250 to 500-j.im 

diamagnetic fraction of the sand, reaching a constant 

concentration after 48 hours of extraction. However, 

dissolved Fe increased through 96 hours of extraction of

229
UJhUJ

UJ 

CC

O 

O

13

UJ 

UJ _J

< 19 LU 12

w

UJ

O

CD

Q 10

PERCENTAGE OF LEAD ADSORBED 

0 20 40 60 80 100

Q| , , , lTmr

_ © LEAD

0 0.1 0.2 0.3 0.4 

LEAD ADSORBED, IN MICROMOLES PER GRAM

PERCENTAGE OF ZINC ADSORBED

0 10 20 30 40 50 

13 ~i i i i i n

UJ 

UJ _j

< UJ 10 12

UJ

O

CD

-M cc 11

UJ 

UJ

Q 10

UJ

CC

O 9

O

<3>2INC

0 0.01 0.02 0.03 0.04 0.05 

ZINC ADSORBED, IN MICROMOLES PER GRAM

Figure 2. Lead and zinc adsorption in micromoles per gram of sand and percentage of total metal as a function of core 

interval altitude (meters above sea level) in core F415 C38. Horizontal error bars represent one standard deviation of the 

three averaged adsorption values. Vertical error bars represent the depth range of core intervals.

an unfractionated sand. The continued release of Fe is 

probably due to dissolution of ferromagnetic mineral 

grains. As a result, a 72-hour HA extraction was used to 

represent dissolution of crystalline Fe-oxide surface 

coatings from the quartz grain surfaces. Five grams of 

sand were extracted with 25 mL of HA in polycarbonate 

centrifuge tubes. Samples were centrifuged at 

10,000 rpm for 10 minutes and the resulting supernatant 

analyzed by ICAP-AES.

Sand samples were also extracted at 25°C for 24 

hours with a dithionite-citrate (DC) solution (0.08M 

Na2S2O4 in 0.2M ammonium citrate) buffered at pH 

8.5. The low-temperature DC extraction at pH 8.5 

was used to extract the sand because it is effective in 

dissolving crystalline Fe coatings with less dissolution 

of Al at this pH (Stephen Short, Australian Nuclear 

Science and Technology Organization, Australia, oral 

commun., 1992). Iron oxides are extracted by 

reductive dissolution by dithionite with Fe 

complexed in solution by citrate.

The total adsorption site density of mineral 

surfaces can be estimated from the surface area 

determined by gas adsorption (Davis and Kent, 1990). 

Single-point surface areas were measured by nitrogen 

gas adsorption on samples from core F415 C38. 

Grain-size distributions were determined on the less 

than 1,000-u.m split of each vertical subsample by dry 

sieve analysis (American Society for Testing 

Materials, 1986).

RESULTS AND DISCUSSION

The variability of Pb2+ and Zn2+ adsorption in 

core F415-C38 (fig. 2) was typical of others in the 

transect (Davis and others, 1993). Pb2+ adsorption 

varied by about a factor of two over the length of the 

core. Zn2+ adsorption was less than that observed for 

Pb2+, but varied by about a factor of three or four.

The surface area of subsamples from core 38 

varied by a factor of two, with no apparent trend with

230
Table 1. Range in concentrations of iron, aluminum and 

manganese in core F415 C38 from 0.5- and 72-hour 

hydroxylamine and 24-hour dithionite-ammonium citrate 

extractions.

[[iinol/g, micromoles per gram; nmol/irr, micromoles per square meter; 

11 A, hydroxylamine hydrochloride; DC, dithionite-ammonium citrate; 

Max, maximum concentration; Min, minimum concentration; Al, aluminum: Fe, iron; Mn, manganese]

Hxtraclion Fe Al Mn

0.5-hour HA

|.imol/g

jamol/m2

72-hour HA

jamol/g

|.imol/m

DC

l^mol/g

(amol/m

Min

0.4

1.2

12

33

5.6

24

Max

2.4

10

19

86

11

52

Min

.7

2.6

11

33

1.0

3.7

Max

2.4

9.9

21

81

1.9

7.9

Min

.02

.05

.11

.30

.03

.12

Max

.16

.79

.20

1.5

.12

.60

depth. The tenth-percent!le grain-size diameter, d )0, 

ranged from 0.174 to 0.362 mm in this core; d ]0 did 

not appear to change as a function of depth. Generally, 

the surface area decreased as d| 0 increased but did not 

appear to vary linearly with d|Q. Instead, the relation 

between surface area and d ]0 might be better described 

by a parabolic function (Davis and others, 1993).

The range in concentrations of extracted Fe, Al 

and Mn in this core is shown in table 1. No apparent 

trend with depth was observed for any of the extracted 

metals. In units of micromoles per g, Al was 

significantly correlated with Fe in the three 

extractions, despite the greater selectivity of DC for Fe 

over Al. A linear correlation between two variables is 

considered significant if the correlation coefficient (r) 

is greater than the critical value of r for the 5 percent 

level of significance (Sokal and Rohlf, 1973). The 

strong correlation of Fe and Al in all three extractions 

indicates that the source of both metals is largely the 

dissolution of mixed-oxide coatings on the grain 

surfaces. If the source was from dissolution of 

separate minerals instead, such as a mixture of 

magnetite and feldspar, it seems unlikely that the 

abundances of these minerals would covary 

throughout the core. Extracted Mn was strongly 

correlated with Fe in both the 0.5- and 72-hour HA 

extractions, but not in the DC extraction. Extracted 

Fe, Al, or Mn by the three methods did not correlate 

with surface area. The lack of correlation is probably

due to the fact that the surface area is indicative of the 

entire surface area of the samples, including surfaces 

other than those containing the extracted elements.

Potential Use of Geochemical Properties as 

Indicators of Metal Adsorption

Linear regressions of the Pb2"1" adsorption data in 

units of micromoles per of gram of sand did not yield 

significant correlations with d )0 (in mm) or surface 

area (in m /g). This result indicates that Pb2"1" 

adsorption occurs at specific areas on the sand surface 

that are more reactive, whereas the surface area is a 

measurement of the entire surface area. Pb2"1" 

adsorption also was not correlated with 0.5-hour HAextractable Fe, Al, or the sum of Fe and Al in units of 

micromoles per gram. The 0.5-hour HA extraction 

should dissolve amorphous coatings of Fe and Al 

oxides that are more reactive with respect to 

adsorption than crystalline oxides (Davis and Kent, 

1990). However, significant correlations were 

observed for linear regressions of Pb2"1" adsorption 

with 72-hour HA Al, and the sum of Fe and Al and 

with the DC Fe, Al, and the sum of Fe and Al (table 2). 

Because the molar ratio of Pb2"1" adsorbed to Mn 

extracted was greater than 1 for all extractions, Mn 

was not used as a variable for correlation analysis with *"% i

Pb . No significant correlations were found between 

Zn2"1" adsorption (in units of micromoles per gram) and 

surface area, grain size, or any elements determined in 

the three extraction methods.

Because the batch adsorption experiments were 

conducted using conditions of equal sand mass per 

volume of AGW, the adsorption and extraction data 

were normalized per unit surface area (micromoles per 

square meter) to eliminate this factor as a variable. 

Significant positive correlations between Pb2+ and 

Zn adsorption and 0.5 hour HA, 72-hour HA, and 

DC extracted Fe, Al, and the sum of Fe and Al were 

obtained upon normalizing to surface area (fig. 3). 

Pb2"1" (table 2) and Zn2"1" adsorption were more highly 

correlated with the 72-hour HA- and DC-extracted Fe 

and Al than with the 0.5-hour HA extraction. Because 

Fe and Al are highly correlated with each other in all 

extractions, Pb2"1" or Zn2+ adsorption each correlate 

well with extractable Fe, Al, and the sum of Fe and Al, 

when normalized to surface area. Therefore, the 

extraction data do not allow differentiation between 

metal adsorption onto Al or Fe phases.

231
Table 2. Summary of regression analysis for lead adsorption at pH 5.3 for subsurface samples of 

coreF415C38

, micromoles per gram; nmol/m\ micromoles per meter squared; n-2, degrees of freedom; r, linear correlation 

coefficient; Pb, lead; Fe, iron; Al, aluminum]

DC 

UJ hUJ

^

UJ

cc

CD O 

CC CO

CO

Q LLJ 

< _J

UJ o

-1 ^

O

cc

o

1.5 -

1.0 -

0.5 -

Dependent 

variable

Independent 

variable n-2 r rcd slope intercept

fjmol/g Pb, Fe, Al

Pb

Pb 

Pb 

Pb 

Pb 

Hmol/m2 Pb, Fe, Al

Pb

Pb

Pb

Pb

Pb

Pb 

Pb 

Pb 

Pb

72 hr Ala

72 hr Fe+Al a 

DCFeb 

DCAIb 

DC Fe+Alb

0.5-hr Fec

0.5-hr Al c

0.5-hr Fe+Al c

72-hr Fca

72-hr Ala

72-hr Fe+Ala 

DCFch 

DC Alh 

DC Fc+Alb

19 0.60

19 .52 

21 .63 

21 .62 

21 .64

24 .60

24 .76

24 .73

19 .86

19 .88

19 .89 

21 .80 

21 .87 

21 .82

.55 .006

.55 .003 

.55 .012 

.53 .082 

.53 .Oil

.50 .045

.50 .060

.50 .030

.55 .009

.55 .010

.55 .005 

.53 .017 

.53 .111 

.53 .015

.144

.145 

.121 

.112 

.115

.718

.533

.595

.365

.336

.326 

.308 

.279 

.293

a 72-hour hydroxylamine hydrochloride extraction at 50 °Ch24-hour dithionite-ammonium citrate extraction . 

C0.5-hour hydroxylamine hydrochloride extraction at 50 °C. 

dCritical value of r for I percent level of significance.

i i i i i i i , . | , 1 < 1

jfck&fc

A^-<5k * I

| 1 . 1 . 

H^P J(^^

' ' ' ' 

^

1 i i i i

20 40 60 80 100

EXTRACTED IRON CONCENTRATION 

IN MICROMOLES PER SQUARE METER

Figure 3. Lead adsorption (micromoles per meter squared) as function of 72-hour hydroxylamine hydrochloride extracted 

iron (micromoles per meter squared) for core F415 C38. Error bars present one standard deviation of averaged adsorption 

value and uncertainty in surface area and iron determinations.

232
SUMMARY

The significant correlations between metal 

adsorption and the extracted Fe and Al (DC or 72-hour 

HA) suggest that extractable metals, combined with 

surface area measurements, could be used as an 

effective geochemical "indicator" of the potential 

variability of metal adsorption in this aquifer. Because 

of the many potential artifacts and time-intensive 

nature of adsorption experiments conducted with 

natural materials, simple laboratory characterization 

methods are needed to provide a first approximation of 

the adsorptive reactivity of site-specific materials at 

contaminated field sites. Although these 

measurements are not more accurate than carefully 

controlled adsorption measurements, the measurement 

of simple geochemical indicators could be preferable 

in site-assessment or other nonresearch-related 

applications. However, the usefulness of the indicator 

discovered in this study needs to be tested at several 

other sites before it can be considered an effective 

characterization tool.

REFERENCES

American Society for Testing Materials, 1986, Annual Book 

of ASTM Standards: Soil, Rock and Building Stones. 

American Society for Testing Materials, Philadelphia, 

Pa., 560 p.

Barber, L.B., II, Thurman, E.M, and Runnells, D.D., 1992, 

Geochemical heterogeneity in a sand and gravel 

aquifer Effect of sediment mineralogy and particle 

size on the sorption of chlorobcnzenes: Journal of 

Contaminant Hydrology, v. 9, p. 35-54.

Chao, T.T., 1984, Use of partial dissolution techniques in 

geochemical exploration: Journal of Geochemical 

Exploration, v. 20, p. 101-135.

Chao, T.T., and Zhou, L., 1983, Extraction techniques for 

selective dissolution of amorphous iron oxides from 

soils and sediments: Soil Science Society of America 

Journal, v. 47, p. 225-232.

Coston, J.A., Fuller, C.C., and Davis, J.A., 1992, The search 

for a geochemical indicator of lead and zinc adsorption 

in a sand and gravel aquifer, Falmouth, Massachusetts, 

USA, in Kharaka, Y.K., and Macst, A.S., eds.. 

Proceedings of the 71 '1 International Symposium on 

Water Rock Interaction: Balkema, Rotterdam, p. 41- 

44.

Davis, J.A., and Kent, D.B., 1990, Surface complexation 

modeling in aqueous geochemistry, in Hochella, M.F., 

and White A.F., cds., Mineral-Water Interface 

Geochemistry: Mineralogical Society of America, 

Washington, D.C., p. 177-260.

Davis, J.A., Fuller, C.C., Coston, J.A., Hess, K.M., and 

Dixon, E., 1993, Spatial heterogeneity of geochemical 

and hydrologic parameters affecting metal transport in 

groundwater: Environmental Research Brief, U.S. 

Environmental Protection Agency, Ada, OK., 

EPA/600/S-93/006

Dzombak, D.A. and Morel, F.M.M., 1990, Surface

complexation modeling hydrous ferric oxide: New 

York, John Wiley, 393 p.

Fuller, C.C., Coston, J.A., Hess, K.M., and Davis, J.A., 

1991, Spatial variability of geochemical and 

hydrologic properties of subsurface materials in a sand 

and gravel aquifer. Cape Cod, Massachusetts, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, Calif, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigation 

Report 91-4034, p. 84-87.

Hess, K.M., Wolf, S.H., and Celia, M.A., 1992, Largescale natural-gradient tracer test in sand and gravel. 

Cape Cod, Massachusetts, 3, Hydraulic conductivity 

variability and calculated macrodispcrsivities: Water 

Resources Research, v. 28 no. 8, p. 2011- 2027.

Hess, K.M., Davis, J.A., Fuller, C.C., Coston, C.C., and 

Dixon, E., 1996, Spatial variability of metal-ion 

adsorption in a glacial outwash aquifer. Cape Cod, 

Massachusetts, in Morganwalp, D.W., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Honeyman, B.D., and Leckie, J.O., 1986, Macroscopic 

partitioning coefficients for metal ion adsorption: 

Proton stoichiometry at variable pH and adsorption 

density, in Davis, J.A., and Hayes, K.F., eds., 

Geochemical Processes at Mineral Surfaces, 

American Chemical Society, Washington, D.C.: 

American Chemical Society Symposium Series no. 

323, p. 162-190.

Kurbatov, M.H., Wood, G.B., and Kurbatov, J.D., 1951, 

Isothermal adsorption of cobalt from dilute solutions: 

Journal of Physical Chemistry, v. 55, p. 1170-1182.

LeBlanc, D. R., Garabedian, S.P., Hess, K.M., Gelhar, L.W., 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural-gradient tracer test in sand 

and gravel aquifer, Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, p. 895-910.

233
Robin, J.J.L., Sudicky, E.A., Gillham, R.W., and 

Kachanoski, R. G., 1991, Spatial variability of 

strontium distribution coefficients and their correlation 

with hydraulic conductivity in the Canadian Forces 

Base Borden Aquifer: Water Resources Research, v. 

27, p. 2619-2632.

Robinson, G.D., 1984/85, Sequential chemical extractions 

and metal partitioning in hydrous Mn-Fe-oxide 

coatings: reagent choice and substrate composition 

affect results: Chemical Geology, v. 47, p. 97-112.

Sokal, R. R., and Rohlf, F.J., 1973, Introduction to

Biostatistics: San Francisco, W. H. Freeman and Co., 

368 p.

Tessier, A., Campbell, P.G.C., and Bisson, M., 1979,

Sequential extraction procedure for the speciation of 

paniculate trace metals: Analytical Chemistry, v. 51, 

p. 844-851.

Tessier, A., Carignan, R., Dubreuil, B., and Rapin, F., 1989, 

Partitioning of zinc between the water column and the 

oxic sediments in lakes: Geochimica ct Cosmochimica 

Acta, v. 53, p. 1511-1522.

Zapico, M.M., Vales, S., and Cherry, J.A., 1987, A wirelinc 

piston core barrel for sampling cohcsionlcss sand and 

gravel below the water table: Ground Water 

Monitoring Review, v. 7, p. 74-82.

234
Spatial Variability of Metal-Ion Adsorption and Hydraulic 

Conductivity in a Sand and Gravel Aquifer, Cape Cod,

Massachusetts

By Kathryn M. Hess 1 , James A. Davis2 , Christopher C. Fuller2, and Jennifer A. Coston2

Abstract

The spatial variability of metal-ion 

adsorption and hydraulic conductivity was 

assessed from the results of laboratory 

experiments on 375 sediment samples collected 

from 14 boreholes in a glacial outwash, sand and 

gravel aquifer on Cape Cod, Massachusetts. Zinc 

and lead adsorption were measured in batch 

experiments. Hydraulic conductivity was 

estimated on the basis of grain-size distributions. 

Mean lead adsorption is greater than mean zinc 

adsorption; mean hydraulic conductivity is 

similar to that measured previously in this aquifer. 

All three properties vary significantly within the 

aquifer; variability in zinc adsorption is greater 

than variability in lead adsorption. A strong 

positive correlation is observed between the two 

metal-ion adsorptions. There is a statistically 

significant, but small, negative correlation 

between lead adsorption and hydraulic 

conductivity and a statistically insignificant 

negative correlation between zinc adsorption and 

hydraulic conductivity. Vertical correlation scales 

obtained by fitting an exponential model to 

experimental semivariograms equaled 0.10- 

0.26 meters for the hydraulic-conductivity and 

adsorption data sets. Similar vertical correlation 

scales were determined in an earlier study of the 

variability of hydraulic conductivity measured by 

use of permeameter and flowmeter tests. A

'U.S. Geological Survey, Marlborough, Mass. 

2 U.S. Geological Survey, Menlo Park, Calif.

horizontal correlation scale could not be identified 

in this study, possibly because the horizontal 

spacing between boreholes was large relative to 

the correlation scale, and the number of samples 

was small.

INTRODUCTION

The dispersion of reactive solutes, such as zinc 

and lead, during transport in the subsurface is greater 

than the dispersion of nonreactive solutes. 

Heterogeneity in the physical and geochemical 

properties of aquifer sediments and correlations 

among these properties are thought to control fieldscale dispersion of reactive species (Garabedian and 

others, 1988; Kabala and Sposito, 1991; MirallesWilhelm, 1993). A negative correlation between 

hydraulic conductivity (K) and adsorption has 

generally been assumed; surface area and, therefore, 

adsorption tend to increase as grain size and, therefore, 

K decrease. Several detailed investigations of the 

spatial structure of the hydraulic-conductivity 

distribution in aquifers have been reported (Sudicky, 

1986; Hess and others, 1992; Rehfeldt and others, 

1992). To date, however, few attempts have been 

made to quantify spatial variability in adsorption and 

correlation among physical and geochemical 

properties that affect adsorption; Robin and others 

(1991) found a statistically significant negative 

correlation between hydraulic conductivity and 

distribution coefficient for 1,279 core samples from 

the Base Borden aquifer in Ontario, Canada. The 

ultimate objectives of our research are to improve our 

understanding of the processes controlling metal-ion 

adsorption and to explain the magnitude of dispersion 

of reactive species observed in ground-water transport;

235
investigating the spatial structure of the variability in 

adsorption and hydraulic conductivity and the 

correlations among these properties is one part of this 

research. In this paper, we report initial results of our 

investigation of spatial variability in zinc adsorption 

(Zn-ads) and lead adsorption (Pb-ads) and hydraulic 

conductivity in a sand and gravel aquifer on Cape Cod, 

Massachusetts. This research is supported by the R. S. 

Kerr Environmental Research Laboratory of the U.S. 

Environmental Protection Agency and the U.S. 

Geological Survey (USGS) Toxic Substances 

Hydrology Program.

This study was conducted using sediments 

collected at the USGS Toxic Substances Hydrology 

Research Site on Cape Cod, Massachusetts (fig. 1). 

The aquifer at the site is shallow, unconfined, and 

composed of glacial-outwash sediments. The 

unconsolidated sediments consist predominately of 

medium to coarse sand, interbedded with gravel and 

fine sand. Typically less than 1 percent of the 

sediments are fine grained (less than 0.0625 mm size 

fraction), and the amount of sedimentary organic 

carbon is less than 0.1 percent by dry weight (Barber 

and others, 1992). The aquifer has an average porosity 

of 0.39; average ground-water velocity within the 

aquifer is 0.4 m/d (LeBlanc and others, 1991).

The cores were collected in the vicinity of the 

array of multilevel samplers (fig. 1) that was installed 

for a large-scale, natural-gradient tracer test conducted 

during 1985-88 to investigate macrodispersion in the 

aquifer (Garabedian and others, 1991; LeBlanc and 

others, 1991; Stollenwerk, 1987). This array is 

currently being used to investigate the transport of 

lead and zinc, as well as other reactive and nonreactive 

solutes, in a second large-scale test (Davis and others, 

1995, this proceedings). Variability in K was 

previously investigated at this site by use of 

permeameter tests of cores (Wolf and others, 1991) 

and flowmeter tests conducted in long-screened wells 

(Hess and others, 1992).

METHODS

In this study, 14 continuous cores of subsurface 

material (fig. 1) were collected from the sand and 

gravel aquifer. These cores were analyzed in the 

laboratory for grain-size distribution and adsorption 

properties.

Core Collection

Cores were collected from an approximate 

vertical interval of 4 m. Individual cores were up to 

1.5 m long and 5 cm in diameter. Cores were collected 

in plastic core liners using a wireline-piston core 

barrel (Zapico and others, 1987). As many as four 

cores were collected in each of the 14 boreholes. 

Cores were frozen until analyzed. Cores were 

collected from eight boreholes positioned 1 to 4 m 

apart along a 16-m-long transect (fig. 1) oriented 

approximately parallel to the current direction of 

ground-water flow and the general paleodirection of 

the flow of the streams that deposited the outwash 

sediments. These transect cores were used to assess 

spatial correlation of metal-ion adsorption and K. 

Additional widely spaced cores were collected from 

six boreholes (fig. 1), positioned an average of 30 m 

apart, to analyze variability in geochemical and 

hydrologic properties and to assess how representative 

the transect cores are of a larger portion of the aquifer 

(spatial stationarity within the aquifer properties).

Laboratory Analyses

Each core was sectioned into subsamples 

approximately 10 cm long for laboratory analysis. 

Samples were dried in a laminar-flow hood and sieved 

to remove grains larger than 1 mm in diameter. The 

less-than-1 mm fraction was then divided into two 

parts with a riffle splitter. One part was used for batch 

adsorption experiments; the other part was used for 

grain-size analysis. The largest fraction was removed 

to reduce the possibility of obtaining 

nonrepresentative subsamples for the batch 

experiments.

Grain-size distributions on the less-than-1 mm 

fraction were determined using a dry-sieve analysis in 

the USGS Sediment Laboratory in Harrisburg, Pa. 

The weight of the greater-than-1 mm fraction was 

included in the final grain-size data set. Wolf and 

others (1991) reported agreement between K 

calculated from grain-size distributions using the 

relation developed by Hazen (1893) and K measured 

in permeameter tests of cores from this site. 

Therefore, Hazen's relation was used to estimate K 

values for the samples in this investigation 

236
42° 30-/

73° 00' 72° 00' xMASSACHUSETTS

42 00

70° 00'

0 20 40 60 KILOMETERS

TRANSECT CORES

F415C38

F415C40, /415C39

F415C41 ^JJ 

F415C42* 

F415C43*

F415C53* 

F415C66 

\

30 FEET

10 METERS

Multilevelsampler array

100 200 300 FEET

30 60 90 METERS

F415C54

EXPLANATION

LOCATION OF CORE AND IDENTIFIER

Figure 1. Location of cores and array of multilevel samplers within the gravel pit at the Cape Cod Toxic SubstancesHydrology Research Site.

Lead and zinc adsorption were measured in 

batch experiments using aquifer samples and artificial 

ground water. Fuller and others (1996) and Davis and

where K is hydraulic conductivity, in cm/s at !0°C; 

A is a constant to correct for temperature 

(1. 157 cm/(s mm2 ) at 10 °C); and </ lo is the grain 

diameter, in mm, that is larger than 10 percent of all 

the grains in the sample, by weight (IOth percentile).

others (1994) report detailed methodologies for these 

experiments.

237
SUMMARY OF LABORATORY RESULTS

Figure 2 presents the estimated K values for 

core F415C38 (see fig. 1 for location) and illustrates 

the variability of A' values. Estimates of A" were made 

for 375 sediment samples. The geometric mean K of 

0.073 cm/s (table 1) is only slightly less than the mean 

of 0.1 cm/s reported by Hess and others (1992) based 

on borehole flowmeter tests, and by LeBlanc and 

others (1991) based on a large-scale tracer test at this 

site. The variance of the natural logarithm of K(\nK) 

of 0.138 is similar to that reported by Hess and others 

(1992) based on the permeameter tests conducted on 

cores from this site. This variance is slightly less than 

the variance of 0.24 reported for the flowmeter tests 

(Hess and others, 1992). If only the 268 samples from 

the transect are considered, the mean is 0.071 cm/s and 

the variance is 0.141 The similarities between the 

mean and variance of the hydraulic conductivities of 

the transect cores and those of all cores suggest that

the transect cores represent a larger portion of the 

aquifer than that covered by the transect cores and that 

spatial stationarity may exist in the K distribution.

A total of 374 and 371 samples were analyzed 

for lead and zinc adsorption, respectively. Results of 

the adsorption batch experiments for core F415C38 

are shown in figure 2 of Fuller and others (1996). 

Results are reported in micromoles of solute adsorbed 

per gram of sediment (umol/g) at a pH of 5.3 and a 

sand-to-water ratio of 50 g/L. The mean Pb-ads of 

0.225 ujnol/g is significantly greater than the mean 

Zn-ads of 0.011 umol/g (table 1). This result was 

expected because lead has a greater affinity for 

surfaces than does zinc. However, the variance of the 

Zn-ads data is unexpectedly greater than that of the 

Pb-ads data. As is the case with the K data, spatial 

stationarity in the adsorption-data sets is suggested if 

the transect results are compared to results from all 

cores.

Table 1. Statistical summary of hydraulic-conductivity, leadadsorption, and zinc-adsorption data sets

ALTITUDE, IN METERSABOVE SEALEVEL

^. ?S <° o o

11111 ' ' ' "'i 1 ' ' '

§

a ® 

5 5

5 ^

S s ^ 

- $ 5 ~

* ' 3

Q 

5

Hydraulic Lead adsorp- Zinc 

conductivity tion adsorption 

(centimeters (micromoles (micromoles 

per second) per gram) per gram)

Number of values 375 374 371 

Minimum .007 .091 0 

Maximum .261 .371 .028 

Geometric mean .073 .225 .011

Variance of natural .138 .019 .212 

logarithm

The data sets are truncated at zero and positively 

skewed, suggesting that the data are lognormally 

distributed. The log-normality of K is widely accepted 

(Freeze, 1975); the lognormal nature of the adsorption 

data is speculated here. A rigorous test of normality is 

not attempted here because of the correlated nature of 

the data sets and the limited amount of data.

Correlations Among Aquifer Properties

A significant positive correlation is found

10-2 10'

HYDRAULIC CONDUCTIVITY, 

IN CENTIMETERS PER SECOND

Figure 2. Vertical distribution of hydraulic conductivity 

estimated from grain-size data using the relation of Hazen 

(1893).

others (1996) and Davis and others (1994) describe the 

mineral-surface properties that control the adsorption 

of these metal ions; they attribute the zinc and lead 

adsorption variability to variability in the surface area 

and the composition of the metal-oxide coatings on the 

sediment grains.

238
Table 2. Pearson correlation coefficients between hydraulic 

conductivity, lead adsorption, and zinc adsorption

[P-valucs arc reported in parentheses. P-values greater than 0.05 are 

considered statistically insignificant]

Hydraulic Lead Zine 

conductivity adsorption adsorption

Hydraulic 

conductivity

Lead adsorption

Zinc adsorption

1

-0.291 

(().()()() 1)

-.0X6 

(0.097)

1

0.515 1 

(0.0001)

Spatial Variability of Aquifer Properties

A geostatistical analysis allows quantification of 

the spatial correlation structure of discrete spatial data 

(Journel and Huijbregts, 1978). The basic tool of the 

analysis is the experimental semivariogram, which is 

calculated as the mean-squared differences between 

sample values at specific separation distancesn(h)

I/O (*)

(2)

A negative correlation between adsorption and 

K has been suggested as a contributor to the enhanced 

dispersion of reactive solutes observed in the field 

relative to the dispersion of nonreactive solutes 

(Garabedian and others, 1988); that is, zones of low K 

may have greater adsorption capacities that increase 

dispersion of the reactive species. The dependence of 

the metal-ion adsorption data on the K data measured 

in this investigation is shown in figure 3. A 

statistically significant, but small, negative correlation 

is identified for Pb-ads and K (table 2). However, a 

statistically insignificant (p > 0.05) negative 

correlation is calculated for Zn-ads and K. The greater 

variance in the Zn-ads data compared to that in the Pbads data is shown in figure 3.

where y is the semivariogram, h is the mean separation distance between observations, n^,) is the number 

of data pairs separated a distance /?, and Yt is the spatially distributed data being analyzed. In practice, the 

semivariogram is constructed so that h is the mean distance between points within a range of distances, 

known as a lag class. Data exhibiting spatial correlation will have variogram values that are small at small 

separation distances and that increase to an asymptotic 

value equal to the variance of the data set at greater 

distances.

Variogram analyses were made on data from the 

transect cores to characterize spatial correlation within 

the hydraulic-conductivity and adsorption data sets. 

Variograms were constructed in the vertical and

EC

UJ

DC 

UJ 

Qu

</)

UJ

_l

o

5 

oDCo 

5 

z

z

o

CL

cc

O

Q

< 

Ul

0.3

0.2

0.1

0

, , , , | 1 1 I 1 | I 1 . ! | 1 1 1 I | 1 1 I I | . I 1 I ^

* LEAD

+ ZINC

_ 

X xx 8X x x + x

x *&* *$*/ * X x

 ^^3$k xj? 1? X x X -

4 4- *+ x ^<^ xx x

4 4* 4 V x 4

4- 4-4- + + x

4-4-+ 

4 4H-t44--H4-H4- 4H- 4-4-4- 

44- 

4-4- 4- 

4-4

+

4-4-H444+H-4- t Hill 4-4-4-

4- 4 44H+WW- 444 Mi II K 4-4-44-

4- 4-4- -H44 t HttHI 44- + 4- 

 44- + 4H+»4»H- *- 

4- 4f^+H- 4f 4- 4

4 4- -HH- -H- 4 4 4 +

444 4-

-1- 41-4 4 4 4 41-4

444

i i i i 1 ,1 i t i 1 i i i i 1 i L_J 1 1 1 1 1 1 1 1 1 1 1 

U.UH

0.03

0.02

0.01

0

0 0.05 0.10 0.15 0.20 0.25 0.30

DC 

UJ

DC 

UJ 

O.

01_J 

O 'S. 

O 

cr 

o

qP

G.cr 

o to

Q

N

HYDRAULIC CONDUCTIVITY, IN CENTIMETERS PER SECOND

Figure 3. Relation between lead and zinc adsorption determined from batch experiments and hydraulic conductivity 

estimated from grain-size data.

239
horizontal directions along the transect. This transect 

is about 6 m from and approximately parallel to the 

transect of cores and boreholes used by Hess and 

others (1992) to characterize correlation within the 

hydraulic-conductivity distribution in this aquifer. 

The variogram analyses were conducted with the 

natural-logarithm transforms of data (ln/C InPb-ads, 

and InZn-ads) for the 268 transect core samples.

In the variogram analyses, the lag classes were 

based on the sample spacing. In the vertical direction, 

samples were, on average, 0.1 m long, and sampling 

was usually continuous within boreholes. Therefore, 

the minimum lag in the vertical direction was set equal 

to 0.1 m. Each lag-class calculation included at least 

100 data pairs in the vertical direction.

In the horizontal direction, sampling was not 

continuous and was limited to discrete boreholes. The 

minimum spacing between boreholes was 

approximately 1 m; a mean distance of 1.4 m was used 

for the smallest lag in the horizontal semivariograms 

to ensure a minimum of 50 data pairs in each lag class. 

Vertical semivariograms are better defined than 

horizontal semivariograms because of the continuity 

in sampling in the vertical direction and the 

discontinuity in sampling in the horizontal direction.

A negative-exponential form was chosen to 

model the experimental semivariograms 

InPb-ads, and InZn-ads, respectively. These values are 

similar to the vertical correlation scales for \nK (0.18- 

0.38 m) identified by Hess and others (1992). The 

differences in vertical correlation scale reported here 

are probably not significant considering that K was 

estimated on the basis of grain-size distribution and 

that the laboratory determinations of adsorption were 

made with some degree of error (see fig. 2 in Fuller 

and others, 1996). The similarity in vertical 

correlation scales for ln/C InPb-ads, and InZn-ads is 

interesting and warrants further investigation; 

however, the similarity in correlation scales does not 

necessarily imply causality.

The horizontal semivariograms show a lack of 

correlation at all separation distances investigated (1.4 

to 12 m). Therefore, no models were fit to these 

horizontal semivariograms. Hess and others (1992) 

reported horizontal scales for \nK of 2.9 to 8 m for this 

site. Horizontal scales could not be resolved in this 

study for InAT, as well as for the adsorption data sets, 

suggesting that the horizontal sampling density at 

small separation distances was not adequate to resolve 

small-scale correlation. Additional data may be 

necessary to determine the horizontal correlation 

scales of the adsorption properties.

r, (3)

where y is the variogram, .v 2 is the variogram sill, h is 

the separation distance between data points, and A, is 

the correlation scale. The sill is the asymptote to 

which the model approaches as the separation distance 

increases. The correlation scale is a length over which 

the data display significant spatial correlation. Both 

the sill and the correlation scale can be varied to fit the 

model to the experimental semivariograms. The sill is 

typically set to approximately the sample variance; the 

correlation scale is, therefore, the primary model 

parameter to be determined through fitting the model 

to the experimental semivariogram. This scale controls the rise of the model to the asymptote and quantifies the spatial correlation of the data.

The vertical semivariograms show greater 

correlations (smaller variogram values) at smaller 

separation distances than at larger distances. The 

vertical semivariogram for InZn-ads is shown in 

figure 4. The vertical correlation scales identified by 

the fitted models are 0.1, 0.15, and 0.26 m for InAT,

CONCLUSIONS

More than 370 laboratory batch experiments 

show that lead and zinc adsorption vary significantly 

within the sand and gravel aquifer on Cape Cod, 

Massachusetts. A statistically significant, but small 

negative correlation between lead adsorption and 

hydraulic conductivity was observed. This type of 

correlation could explain, in part, enhanced dispersion 

of reactive species observed in the field relative to the 

dispersion of nonreactive solutes. The correlation 

between zinc adsorption and hydraulic conductivity, 

however, is not statistically significant. Vertical 

correlation scales determined for metal-ion adsorption 

are similar to those determined for hydraulic 

conductivity in this and previous studies. Finally, 

horizontal correlation scales could not be resolved for 

the adsorption and hydraulic-conductivity data sets, 

possibly because of the large horizontal spacing 

between data used in this study.

240
0.20

°'15

ac 

o 

o E 0.10

0.05

0.0 0.5 1.0 1.5 2.0

SEPARATION DISTANCE, IN METERS

Figure 4. Vertical semivariogram of the natural logarithm of the zinc adsorption data from the eight transect cores (locations 

shown on fig. 1). The point values are the experimental semivariogram (eq. 2); the solid line is the negative-exponential 

model (eq. 3) fit to the experimental semivariogram. The model fit has a sill value of 0.17 and a correlation scale of 

0.26 meters.

REFERENCES

Barber, L. B., II, Thurman, E. M., and Runnclls, D. D.,

1992, Gcochcmical heterogeneity in a sand and gravel 

aquifer Effect of sediment mineralogy and particle 

size on the sorption of chlorobcnzcncs: Journal of 

Contaminant Hydrology, v. 9, no. I, p. 35-54.

Davis, J. A., Fuller, C. C., Coston, J. A., Hess, K. M., and 

Dixon, E., 1994, Spatial heterogeneity of gcochcmical 

and hydrologic parameters affecting metal transport in 

groundwatcr: Ada, Oklahoma, Robert S. Kcrr 

Environmental Research Laboratory, U.S. 

Environmental Protection Agency, EPA/600/S93/006, Environmental Research Brief, 22 p.

Davis, J. A., Kent, D. B., Coston, J. A., and Hess, K. M., 

1996, Heavy-metal transport in a sand and gravel 

aquifer with variable chemical conditions. Cape Cod, 

Massachusetts, //; Morganwalp, D.W., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Freeze, R. A., 1975, A stochastic-conceptual analysis of 

one-dimensional groundwatcr flow in nonuniform 

homogeneous media: Water Resources Research, 

v. II, no. 5, p. 725-741.

Fuller, C. C., Coston, J. A., Davis, J. A., and Dixon, E., 

1996, Evaluation of gcochemical indicators of metal 

adsorption in a sand and gravel aquifer, Cape Cod, 

Massachusetts, in Morganwalp, D.W., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program--Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Garabcdian, S. P., Gclhar, L. W , and Cclia, M. A., 1988, 

Large scale dispersive transport in aquifers: Field 

experiments and reactive transport theory: Cambridge, 

Mass.: Ralph M. Parsons Laboratory Report 315, 

Department of Civil Engineering, Massachusetts 

Institute of Technology, 290 p.

Garabcdian, S. P., LcBlanc, D. R., Gclhar, L. W., and Cclia, 

M. A., 1991, Large-scale natural-gradient tracer test in 

sand and gravel. Cape Cod, Massachusetts, 2 

Analysis of spatial moments fora nonrcactivc tracer: 

Water Resources Research, v. 27, no. 5, p. 911-924.

241
Hazen, A., 1893, Some physical properties of sands and 

gravels: Massachusetts State Board of Health, 24th 

Annual Report.

Hess, K. M., Wolf, S. H., and Celia, M. A., 1992, Largescale natural gradient tracer test in sand and gravel, 

Cape Cod, Massachusetts, 3 Hydraulic-conductivity 

variability and calculated macrodispersivities: Water 

Resources Research, v. 28, no. 8, p. 2011-2027.

Journel, A. G., and Huijbregts, C. J., 1978, Mining 

Geostatistics, London: Academic Press, 600 p.

Kabala, Z. J., and Sposito, G., 1991, A stochastic model of 

reactive solute transport with a time-vary ing velocity 

in a heterogeneous aquifer: Water Resources 

Research, v. 27, p. 341-350.

LeBlanc, D. R., Garabedian, S. P., Hess, K. M., Gelhar, L. 

W., Quadri, R. D., Stollenwerk, K. G., and Wood, W 

W., 1991, Large-scale natural-gradient tracer test in 

sand and gravel. Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, no. 5, p. 895-910.

Miralles-Wilhelm, F., 1993, Stochastic analysis of sorption 

and biodcgradation in three-dimensionally 

heterogeneous aquifers, unpublished PhD dissertation, 

Department of Civil Engineering, Massachusetts 

Institute of Technology.

Rehfeldt, K. R., Boggs., J. M., and Gelhar, L. W., 1992, 

Field study of dispersion in a heterogeneous aquifer, 

3 Geostatistical analysis of hydraulic conductivity: 

Water Resources Research, v. 28, no. 12, p. 3309-3324.

Robin, J. J. L., Sudicky, E. A., Gillham, R. W, and 

Kachanoski, R. G., 1991, Spatial variability of 

strontium distribution coefficients and their correlation 

with hydraulic conductivity in the Canadian Forces 

Base Borden Aquifer: Water Resources Research, 

v. 27, p. 2619-2632.

Sudicky, E. A., 1986, A natural gradient experiment on 

solute transport in a sand aquifer Spatial variability 

of hydraulic conductivity and its role in the dispersion 

process: Water Resources Research, v. 22, no. 13, 

p. 2069-2082.

Stollenwerk, K. G., and Grove, D. B., 1987, Naturalgradient tracer test in sand and gravel Nonconservative transport of molybdenum, in Frank, B. J., 

cd., U.S. Geological Survey Program on Toxic Waste-- 

Ground-Water Contamination, Proceedings of the 

third technical meeting, Pensacola, Florida, March 23- 

27, 1987: U. S. Geological Survey Open-File Report 

87-109, B-17-B-22.

Wolf, S. H., Celia, M. A., and Hess, K. M., 1991,

Evaluation of hydraulic conductivities calculated from 

multiport-pcrmcamcter measurements: Ground Water, 

v. 29, no. 4, p. 516-525.

Zapico, M. M., Vales, S., and Cherry, J. A., 1987, A

wireline piston core barrel for sampling cohcsionlcss 

sand and gravel below the water table: Ground Water 

Monitoring Review, v. 7, no. 3, p. 74-82.

242
Small-Scale Tracer Tests Applied to the Measurement of

in situ Denitrification Rates in a Sewage-Contaminated

Aquifer, Cape Cod, Massachusetts

By Myron H. Brooks 1 , Richard L. Smith 1 , and Stephen P. Garabedian2

Abstract

Small-scale natural-gradient tracer tests were 

used to measure in situ rates of denitrification in a 

nitrate-contaminated aquifer. Tracer solutions 

contained bromide, a conservative tracer, and 

acetylene, an inhibitor of nitrous oxide reductase. 

Breakthrough of the tracer solution and production 

of nitrous oxide were monitored at multilevel 

samplers approximately 10 meters downgradient 

from the injection multilevel sampler for a period 

of up to 40 days. Calculated rates of denitrification 

from two tracer tests were 620 and 410 nanomoles 

N2O per liter of aquifer per day. These rates are on 

the low end of the range of reported rates in 

aquatic sediments, and are similar, but lower than, 

previous rate measurements made at the Cape Cod 

site by using flask and whole-core incubation techniques. Results of this study suggest that incubation techniques using aquifer sediments might 

slightly overestimate in situ denitrification rates in 

ground water. Peak concentrations of nitrous oxide 

lagged 2 days behind the peak concentrations of 

the conservative tracer in both tests. Work is 

currently underway to examine possible mechanisms for this result, and to model it by using a 

modified one dimensional advection-dispersion 

model containing a Michaelis-Menten term for 

nitrous oxide production.

INTRODUCTION

Nitrate (NO3~) is one of the most widespread contaminants in shallow ground water in North America 

and Europe (Gillham and Cherry, 1982). Nitrate is an

'U.S. Geological Survey, Boulder, Colo. 

2 U.S. Geological Survey, Marlborough, Mass.

undesirable byproduct of numerous activities including 

fertilizer application, livestock production, and sewage 

treatment, and tends to be persistent in ground water 

despite its predicted instability at near neutral pH 

(Howard, 1985). This persistence is a problem if the 

NO3 "-contaminated aquifer is a source of drinking 

water. Elevated concentrations of NO3 in drinking 

water can cause methemoglobinemia in infants (Comley, 

1945), and have been implicated in the increased 

incidence of cancer (Jensen, 1982) and hypertension 

(Morton, 1971) in adults. Concern over the increased 

frequency of ground-water contamination by NO3 " has 

heightened interest in techniques that remove NO3 " 

from ground water. One remediation technique uses 

biological denitrification, a process carried out by bacteria in which NO3~ can serve as an electron acceptor 

and can ultimately be reduced to dinitrogen (N2 ). The 

reduction proceeds stepwise from NO3 ~ through nitrite 

(NO2~), nitric oxide, and nitrous oxide (N2O) to N2 ; 

each step is catalyzed by a separate enzyme. Geochemical evidence from several studies has indicated that 

denitrification occurs in ground water (Gillham and 

Cherry, 1978; Edmunds and Walton, 1983). Measurements of denitrification rate have been made indirectly 

by using incubations with aquifer core material (Slater 

and Capone, 1987; Smith and Duff, 1988; Smith and 

others, 199la). In one study, denitrification was 

directly measured in situ (Trudell and others, 1986) by 

using a drive point to inject, and then sample, a NO3~ 

containing tracer solution in a denitrifying zone 

approximately 3 m below land surface.

Remediation of NO3 " contamination employing 

denitrifying bacteria has been attempted using both 

"pump and treat" (van der Hoek and Klapwijk, 1987) 

and in situ techniques (Mateju and others, 1992). In situ 

remediation using denitrification is particularly attractive because the ultimate product (N2) is harmless. 

Most proposed in situ remediation plans involve stimulating existing populations of denitrifying bacteria

243
through the addition of carbon substrates. Although 

there have been successes with this approach (Mateju 

and others, 1992), little is known about the native bacterial populations that are used to catalyze the reduction of NO3 ~ or the physical and chemical factors that 

control their activity in situ. Information of this type 

will be of value to scientists and engineers designing 

NO3 " remediation programs that use biological denitrification. The purposes of this study were to test the feasibility of using small-scale natural-gradient tracer 

tests to quantify denitrification rates in a nitrate-contaminated aquifer, and to compare those rates to previously measured rates obtained by using aquifer core 

material incubated in flasks in the laboratory. Acetylene (C2H2), an inhibitor of nitrous oxide reductase (the 

enzyme that catalyzes the reduction of N2O to N2 ), was 

added to a ground-water solution containing a conservative tracer that was injected into a zone of active denitrification at the Cape Cod study site. Subsequent 

production of N2O was monitored downgradient, and 

the N2O produced was used to calculate an in situ rate 

of denitrification. The results of these experiments represent new information about the denitrification process in ground water and provide insight into the 

functioning of one of the enzymes that catalyzes the 

process. The purpose and scope of this paper are to 

describe the results of the tracer test experiments and to 

compare these results with other studies of groundwater contamination.

MATERIALS AND METHODS 

Study Site

The study site is a shallow sand and gravel aquifer 

located on Cape Cod, Mass. Contamination at this site 

is the result of more than 50 years of land disposal of 

secondarily treated sewage (LeBlanc, 1984). Naturalgradient tracer tests were performed at site F473 (fig. 1), 

which consists of an array of 21 multilevel samplers, 

each having 15 evenly spaced (vertical spacing = 25 

cm) sampling ports at altitudes of 8.56 to 12.09 m 

above sea level. One sampler located upgradient of the 

remaining 20 (F473MO-11) served as the injection 

sampler. The remaining samplers were arranged in 

three rows approximately 4,7, and 10 m downgradient 

of the injection sampler; each row was aligned roughly 

perpendicular to the direction of ground-water flow 

(fig. I)-

Tracer Injection and Sampling

Small-scale natural-gradient tracer tests were performed in May and June 1991 and August and September 

1992. Approximately 100 L of ground water was withdrawn from a single sampling port in the injection multilevel sampler by using a peristaltic pump outfitted with 

gas impermeable tubing. The sampling port, located 

10.08 m above sea level, was located within a previously 

identified zone of active denitrification. The ground water 

was pumped into a gas impermeable bladder that had 

been flushed seven times with N2 . The bladder contained 

12.88 g of sodium bromide that had been dissolved in 

approximately 1 L of anoxic, distilled water. Several 

liters of C2H2 were added to the headspace in the bladder 

and allowed to equilibrate for 1 hour to create a saturated 

solution of this highly soluble gas. Undissolved C2H2 was 

then vented from the bladder and the injectate solution 

was pumped back down the injection sampler port from 

which it was originally taken. The pumping rate for the 

injection was approximately 1.4 L per minute. Samples of 

the injectate solution were taken at the outset, approximate halfway point, and end of the injection period for 

NO3 , NO2 , Br", N2O, and C2H2 . Samples for these constituents were then collected daily for 33 days (1991) and 

40 days (1992) at selected multilevel samplers downgradient to monitor the production of N2O as the tracer cloud 

moved through the aquifer. Samples (60 mL) for analysis 

of NO3% NO2", and Br" concentrations were filtered in the 

field (0.45 urn). Samples for analysis of NO3 and NO2 

were preserved by freezing. Samples for Br analysis 

were normally analyzed on the day of collection. Samples 

for N2O and C2H2 analyses were collected in 20-mL plastic syringes using positive pressure supplied by the peristaltic pump. A 22-gauge needle was then attached to the 

syringe and 15 mL of sample was injected through the 

stopper of a 30-mL serum bottle. Prior to sample introduction, the serum bottles received 200 uL of 12.5 N 

NaOH to eliminate biological activity, were capped with 

butyl rubber stoppers, and crimp sealed.

Analytical Techniques

Br was analyzed by ion selective electrode. NO3" 

and NO2" were determined colorimetrically by using a 

flow-injection autoanalyzer. N2O and C2H2 were analyzed by gas chromatographic headspace analysis. N2O 

was quantified with an electron capture detector; C2H2 

was quantified with a flame-ionization detector. Concentrations in water were calculated from the headspace concentrations by using empirical solubility relationships.

244
RESULTS AND DISCUSSION

The composition of tracer solutions for the 1991 

and 1992 experiments is shown in table 1. With the 

exception of the variation in the contaminant NO3 concentration, the composition of the two tracer solutions 

was virtually identical for the two experiments. The 

mass of NaBr (12.88 g) added to the injectate solution

was calculated to produce a 100 ppm Br - solution if 

brought to a final volume of 100 L. Observed injectate 

Br concentrations can be used to calculate the actual 

volume of tracer solution for each test; they were 124.6 L 

in 1991 and 119.3 L in 1992.

Breakthrough curves of Br , C2H2 , and N2O for the 

1991 and 1992 experiments at multilevel sampling 

ports approximately 10 m downgradient from the

 - ,-v ~

41* 38' 

41* 37' 

EXPLANATION

 15-4 WATER-TABLE CONTOUR--SKawaltltud* 

4 of IMIW UM* above MI tovd, In mttcn. 

Arrows chow draction ot ground-water flow.

F473 , SAMPLING SITE

tii'

= 4

gF473M 

gF473M2-14

-47;*:-;; 

gF47.'

E

sf

n

1-13

44-14

JT473M6-13 ^

I73M6-14

473MJ-11

-473M4- 

473M4-12

 

11

1

3''

0 !

rw-n

Feet

Map oriented to Magnetic North 

Origin at well F343-3i

EXPLANATION

ffl Multilevel sampler 

A Coring nl«

F473 TRACER-TEST SITE, CAPE COD, MASSACHUSETTS 

UNITED STATE GEOLOGICAL SURVEY MARCH 1991

Figure 1. Ground-water study site on Cape Cod, Mass., showing location of the contaminant plume and the multilevel sampler 

array used in this study.

245
Table 1. Injectate solution composition for small-scale natural-gradient tracer tests 

performed at site F473, May and June 1991 and August and September 1992 

[ppm, parts per million]

Year Bromide 

(ppm)

Acetylene

(millimoles per

liter)

Nitrous oxide

(micromoles per

liter)

Nitrate-N 

(ppm)

Nitrite-N 

(ppm)

1991

1992

79.4

83.0

6.1

6.2

9.17

9.16

15.3

24.2

0.53 

.42

injection sampler are shown in figures 2 and 3. Maximum breakthrough was not observed at the same sampler for the two tests. The curves in figure 2 for the

1991 experiment were constructed from data collected 

at sampling site F473M6-13 (fig. 1). The curves for the

1992 experiment were constructed from data collected 

at sampling site F473M6-15 (fig. 1). The sampling port 

where maximum breakthrough occurred in the 1991 

was also 25 cm lower than the sampling port where 

maximum tracer breakthrough in the 1992 experiment 

was observed. Travel distance from the injection port 

to the sampling port was about 9.5 m in the 1991 experiment and about 10.5 m in the 1992 experiment. 

Dimensionless breakthrough curves (observed concentration divided by injectate concentration) for Br" and 

C2H2 were virtually identical in each test (figs. 2 and 3). 

C2H2 was added as an inhibitor of nitrous oxide reduc0.3

tase, but it also functioned as a second conservative 

tracer. Maximum concentrations of Br ~ and C2H2 

occurred 21 days after injection in 1991 (fig. 2), and 28 

days after injection in 1992 (fig. 3). More of the tracer 

was recovered in 1992 (peak dimensionless concentration = 0.41) than in 1991 (peak dimensionless concentration = 0.24). Differences in travel times and 

maximum conservative tracer concentrations observed 

for the two tracer tests are most likely related to a slight 

difference in hydraulic gradient that existed between 

the two experiments.

N2O production was similar for both tests. The 

peak N2O concentration in 1991 was 42.5 urn, whereas 

the peak concentration observed in 1992 was 38.9 jim. 

Background N2O concentrations were similar for both 

experiments (table 1, figs. 2 and 3). Maximum denitrification rates were calculated by dividing the amount of 

N2O produced (peak concentration minus the injectate

-40

0.0

DAYS SINCE INJECTION

Figure 2. Breakthrough curves of bromide, acetylene, and 

nitrous oxide at one multilevel sampling port at sampler 

F473M6-13 for an acetylene-block natural-gradient tracer test 

performed during May and June 1991. Bromide and acetylene 

values are normalized to the injectate concentration.

20 30 

DAYS SINCE INJECTION

Figure 3. Breakthrough curves of bromide, acetylene, and 

nitrous oxide at one multilevel sampling port at sampler 

F473M6-15 for an acetylene-block natural-gradient tracer test 

performed during August and September 1992. Bromide and 

acetylene values are normalized to the injectate concentration.

246
concentration) by the number of days from injection 

until the maximum N2O concentration was observed 

and then correcting for the average porosity of the 

aquifer. Rates for the two experiments were 620 nmol 

N2O (L aquifer)-' d' 1 in 1991 and 410 nmol N2O 

(L aquifer)' 1 d"' in 1992. These rates, which are a first 

approximation of actual in situ rates because they do 

not account for dispersion, represent the lower end of 

reported denitrification rates from aquatic sediments 

(Knowles, 1982). The calculated rates are similar to, 

but lower than, rates measured previously at the Cape 

Cod site by use of flask and whole core incubation techniques (Smith and Duff, 1988; Smith and others, 1991), 

suggesting that these techniques may slightly overestimate in situ rates. The larger of the two rates 

measured in this study would completely remove 14 mg 

N/L NO3 from the aquifer in 314 days if the process 

were a zero order reaction at all NO3" concentrations. 

This is in contrast to the results of in situ experiments 

performed by Trudell and others (1986). In that study, 

a tracer solution containing 13 mg N/L NO3' was 

injected into a zone containing virtually no NO3% and 

enough NO3 was denitrified within 356 hours so that 

the NO3 lost could be measured directly. The authors 

reported an increase in bicarbonate that was in stoichiometric agreement with the predicted amount for heterotrophic denitrification, and noted that organic 

carbon present in the aquifer was adequate to denitrify 

large amounts of NO3 ~ in the aquifer. At the Cape Cod 

site, ground water containing 15 to 24 mg N/L NO3" was 

present in the zone where the experiment was performed, and previous work had demonstrated that denitrification there was limited by the lack of available 

organic carbon (Smith and Duff, 1988). The amount of 

NO3 ' denitrified in 672 hours was detectable only 

because of the increased analytical sensitivity provided 

by the acetylene-block technique; NO3 ~ concentrations 

in the tracer clouds were unchanged during the experiments (data not shown).

In both tracer tests the peak N2O concentrations 

occurred 2 days after the peak conservative tracer 

concentrations occurred (figs. 2 and 3). Attempts to 

model these results using a modified one-dimensional 

advection-dispersion model containing a MichaelisMenten term for N2O production have been unsuccessful at replicating the timing and shape of the N2O 

breakthrough curves. The modeling approach being 

used is similar to the one used by Smith and others 

(1991b) to simulate a consumption process (methane 

oxidation) during similar natural-gradient tracer experiments at the Cape Cod site. One possible explanation 

for the lag between the peaks of the conservative tracer 

breakthrough curves and the N2O breakthrough curves 

is that nitrous oxide reductase binds N2O in the presence of C2H2 . An early paper on the acetylene-block 

phenomenon reported that the inhibition was noncompetitive (Yoshinari and others, 1976), indicating that 

the mechanism of inhibition does not involve C2 H2 

binding to the active site of the enzyme. If N2O did bind 

to all available enzyme active sites as the tracer cloud 

moved through the aquifer, no net N2O production 

would occur until all of the active sites of nitrous oxide 

reductase in both attached and free-living bacteria were 

saturated with N2O. This could result in the breakthrough curve characteristics observed in the two tests. 

Laboratory experiments with a denitrifying pure 

culture are being conducted to test this hypothesis. If 

this phenomenon can be demonstrated, an attempt will 

be made to incorporate it into the modified advectiondispersion model.

SUMMARY

The feasibility of measuring ground-water denitrification rates in situ by using natural-gradient tracer 

tests was demonstrated. Br - was used as the conservative tracer and C2H2 was added to inhibit the activity of 

nitrous oxide reductase. Production of N2O was monitored at downgradient sampling ports, and the N2O 

produced as the tracer cloud moved through the aquifer 

was used to calculate an in situ denitrification rate. 

Rates from replicate experiments were similar and in 

the low end of the range of rates reported for soils and 

aquatic sediments. The rates were similar to rates 

reported previously for the Cape Cod site in studies that 

employed flask and whole core incubation techniques. 

The shape and timing of N2O breakthrough curves 

relative to the conservative tracers was different than 

expected. It is hypothesized that this result may be 

related to the noncompetitive inhibition of nitrous 

oxide reductase by C2H2 . Experiments are underway to 

test this hypothesis, and to model this result using a 

modified one-dimensional advection-dispersion model 

containing a Michaelis-Menten term for N2O production.

247
REFERENCES

Balderston, W.L., Sherr, B., and Payne, W.J., 1976, Blockage by acetylene of nitrous oxide reduction in Pseudomonas perfectomarinus: Applied and Environmental 

Microbiology, v. 31, p. 504-508.

Comley, H.H., 1945, Cyanosis in infants caused by nitrates 

in well water: Journal of the American Medical Association, v. 29, p. 112-116.

Edmunds, W.M., and Walton, N.R.G., 1983, The Lincolnshire limestone Hydrogeochemical evolution over at 

ten-year period: Journal of Hydrology, v. 1, p. 201-211.

Gillham, R.W., and Cherry, J.A., 1978, Field evidence of 

denitrification in shallow ground water flow systems: 

Water Pollution Research in Canada, v. 13, p. 53-71.

Gillham, R.W., and Cherry, J.A., 1982, Contaminant migration in saturated unconsolidated geologic deposits: 

Geological Society of America Special Paper 189, 

p. 31-62.

Howard, K.W.F., 1985, Denitrification in a major limestone 

aquifer: Journal of Hydrology, v. 76, p. 265-280.

Jensen, O.M., 1982, Nitrate in drinking water and cancer in 

northern Jutland, Denmark, with special reference to 

stomach cancer: Ecotoxicology and Environmental 

Safety, v. 6, p. 258-267.

Knowles, R., 1982, Denitrification: Microbiological 

Reviews, v. 46, p. 43-70.

LeBlanc, D.R., 1984, Sewage plume in a sand and gravel 

aquifer: U.S. Geological Survey Water-Supply Paper 

2218, 28 p.

Mateju, V., Cizinska, S., Krejci, J., and Janoch, T., 1992, 

Biological water denitrification~A review: Enzyme 

and Microbiological Technology, v. 14, p. 170-183.

Morton, W.E., 1971, Hypertension and drinking water, a 

pilot statewide ecological study in Colorado: Journal of 

Chronic Diseases, v. 23, p. 537-545.

Slater, J.M., and Capone, D.G., 1987, Denitrification in 

aquifer soil and nearshore marine sediments influenced 

by groundwater nitrate: Applied and Environmental 

Microbiology, v. 53, p. 1292-1297.

Smith, R.L., and Duff, J.H., 1988, Denitrification in a sand 

and gravel aquifer: Applied and Environmental Microbiology, v. 54, p. 1071-1078.

Smith, R.L., Howes, B.L., and Duff, J.H., 199la, Denitrification in nitrate-contaminated groundwater Occurrence in steep vertical geochemical gradients: 

Geochimica et Cosmochimica Acta, v. 55, p. 1815- 

1825.

Smith, R.L., Howes, B.L., and Garabedian, S.R., 1991b, In 

situ measurement of methane oxidation in groundwater 

by using natural-gradient tracer tests: Applied and 

Environmental Microbiology, v. 57, p. 1997-2004.

Trudell, M.R., Gillham, R.W., and Cherry., J.A., 1986, An 

in situ study of the occurrence and rate of denitrification in a shallow unconfined sand aquifer: Journal of 

Hydrology, v. 83, p. 251-268.

van der Hoek, J.P., and Klapwijk, A., 1987, Nitrate removal 

from ground water: Water Research, v. 21, p. 989-997.

Yoshinari, T., and Knowles, R., 1976, Acetylene inhibition 

of nitrous oxide reduction by denitrifying bacteria: Biochemical and Biophysical Research Communications, 

v. 69, p. 705-710.

248
Overview of Research on the Distribution and Role of 

Protozoa in an Organically Contaminated Aquifer at

Cape Cod, Massachusetts

By Nancy E. Kinner1 and Ronald W. Harvey2

Abstract

The distribution, nature, and role of 

protozoa that inhabit contaminated and 

uncontaminated ground water were investigated 

at the U.S. Geological Survey (USGS) Cape Cod 

Toxic-Substances Hydrology Research Site. New 

information has included identification of large 

protozoan populations (up to 105/gram dry 

weight) in anoxic zones of the aquifer; a lack of 

correlation between dissolved oxygen 

concentrations and protozoan abundance or 

degree of encystment; a strong relation among 

concentrations of dissolved organic carbon, 

phosphate, sulfate, nitrate, numbers of free-living 

bacteria (FLB) and protozoa; and a ratio of FLB 

to protozoa that is 10-100 times lower than that 

typically found in other aqueous environments. 

An improved direct-counting procedure has 

allowed more accurate enumeration of protozoa in 

sandy, aquifer sediments than was previously 

possible.

INTRODUCTION

In 1987, protozoa were found at a variety of 

locations in the contaminant plume at the U.S. 

Geological Survey (USGS) Cape Cod ToxicSubstances Hydrology Research Site (fig. 1). In 1990, 

the National Science Foundation (NSF) funded a 3- 

year study (NSF Grant # BCS 9012183 awarded to the

Dept. of Civil Engineering, Univ. of New Hampshire, 

Durham, N.H.

2 U.S. Geological Survey, Boulder, Colo.

University of New Hampshire) involving the 

distribution and role of the protozoa in the aquifer. 

This paper presents an overview of the preliminary 

results obtained during this collaborative project 

between the University of New Hampshire, the USGS, 

and the Natural History Museum (London).

METHODS

A new application of epifluorescent direct 

counting was used to enumerate 4',6-diamidino-2- 

phenylindole (DAPI)-stained protozoa recovered from 

core material. Estimates of the number of protozoa 

present based on use of the epifluorescent method 

were comparable to those based on use of the 

Darbyshire liquid media most-probable-number 

(MPN) technique (Darbyshire and others, 1974), but 

differed from those of the Singh solid-media MPN 

method (Singh, 1946). Both the Darbyshire and Singh 

methods were originally developed for enumeration of 

protozoa in topsoil and recently have been applied to 

subsurface samples. During the Cape Cod protozoa 

study, the epifluorescent technique has been used 

because it has the advantages of low counting 

variability and avoids media/culturing selectivity. 

However, all cores also have been analyzed by the 

Darbyshire method because that method gives an 

indication of the percentage of the protozoan 

population that is encysted.

RESULTS AND DISCUSSION

A survey of the aquifer was conducted to 

determine the distribution of protozoa throughout the 

4.5-km long plume of sewage-contaminated ground 

water that originates at Otis Air Base (fig. 1). Cores

249
71'Ofl'

I I i . ___J_-- --^

MASSACHUSETTS

20 40

0 20 40 60 KILOMETERS

70* 34' 33*

41' 38'f 

41' 37M 

EXPLANATION

 15 -I WATER-TABLE CONTOUR - Showt .Mud* 

of wnrt*r table abov* MB tevri, ki m«t«r*. 

Arrow* show direction of ground-water flow.

Figure 1. Location of down-well incubation studies, U.S. Geological Survey Cape Cod Toxic Substances Hydrology 

Research Site (modified from Smith and others, 1991, fig. 1).

and porewater samples were collected at a pristine 

control site and along longitudinal (12 sites) and 

vertical (3 sites with 4-5 depths per site) transects 

through the plume. In oxic and anoxic zones in the 

contaminant plume, protozoan abundance typically 

varied between 104 and 105/gdw (gram dry weight)

sediment with 3 to 40 percent of the population 

encysted. In the alkylbenzene sulfonate (ABS)- 

contaminated part of the plume and at the pristine site, 

protozoan abundance was lower (10 -10 /gdw). There 

was no correlation between dissolved oxygen (DO) 

concentration and protozoan abundance or percent

250
encystment, indicating that the protozoa are not 

excluded from anoxic regions of the aquifer as was 

once believed. Indeed, their numbers are relatively 

high within the non-ABS parts of the plume. 

Consistently, more than 70 percent of the population 

appears to be trophic, suggesting that protozoa may 

play an active role in the microbial community of the 

contaminant plume. More than 99 percent are small 

flagellates (2-3 (im in length). Few amoebae and no 

ciliates were detected.

Principal component factors were used to assess 

significant multivariate relations among 

biogeochemical characteristics of the plume and the 

protozoan population. There was a strong relation 

among the variables dissolved organic carbon (DOC), 

PO4-3 , SO4-2 , NO-3 , FLB and protozoa. This relation 

indicated that protozoa and FLB may be using DOC 

(less than 0.45 m filtration fraction). Although a 

relation between DOC and FLB was expected and 

shown previously for the Cape Cod site (Harvey and 

Barber, 1992), the direct relation between protozoa 

populations and DOC concentration was surprising. 

There was only a limited inverse relation between the 

populations of FLB and protozoa indicating a weak 

predator-prey relation. The ratio of FLB to protozoa at

1 7

each site in the plume ranged from 10 to 10 ; 

normally this ratio is ~103 in aqueous environments 

when protozoa are preying on FLB (Fenchel, 1987). 

The survey data suggest the aquifer protozoa (2-3 (am) 

are either feeding on high-molecular-weight DOC 

directly and (or) on a limited size range of FLB.

When the NSF grant was written in 1989, it was 

assumed that protozoa would have a role in the 

organically contaminated subsurface similar to the one 

they play in most other environments as predators 

upon bacteria. As predators upon FLB, protozoa have 

been shown to have a positive effect on bacterial 

degradation of organic compounds in aqueous and 

topsoil ecosystems because they release nutrients and 

cytoplasmic constituents necessary for bacterial 

growth and they keep bacterial numbers low and in the 

exponential growth phase. In order to explore the 

potential predator-prey relation, laboratory columns 

(10 cm long) containing established communities of 

attached bacteria (106-107/gdw) and 2 to 3 (im 

protozoa (104-105/gdw) from the Cape Cod site were 

fed an injectate containing DAPI-stained, cultured 

bacteria (105 cells/mL, 1 (im cell length) that were 

isolated from the aquifer. The injectate also contained 

carboxylated microspheres (1.5-1.7 (am diameter) and

a conservative tracer (bromide, 20 mg/L). There was 

also a population of unstained FLB (106 cells/mL) 

present in the injectate. A similar injection was 

repeated after the column was sterilized. Fewer 

DAPI-stained FLB were expected to elute from the 

column (Cout/Cjn ) in the protozoa-populated 

(nonsterile) column because of predation. However, 

this was not observed and is again in agreement with 

the theories of Fenchel (Fenchel, 1986 and 1987) that 

the size ratio of protozoa:FLB must be on the order of 

10:1 for predation to occur. In the column studies, this 

size ratio was less than 5:1. As with the survey data, 

these results suggest that the 2 to 3 (am protozoa in the 

plume may be (1) selectively grazing on small FLB 

(0.2 to 0.3 (im), most of which exist within the first 

0.08 km of the plume (Harvey and Garabedian, 1991); 

and (or) (2) directly using high-molecular-weight 

DOC.

In experiments currently being conducted at the 

Cape Cod site, flow-through incubation chambers 

(100,000 molecular-weight cutoff, sterile dialysis 

membranes) are being used to study the relation 

between protozoa and FLB. Controls in these 

experiments include membranes containing sterile 

water and DAPI-stained glutaraldehyde-fixed (killed) 

aquifer FLB. Initial chamber tests using 0.8 to 1 (.im 

ground-water isolates suggest an equilibrium 

abundance or carrying capacity of ~106 FLB/mL 

porewater. In separate experiments, initial stocks of 

FLB (104, 105 , 107 cells/mL) converged to ~106 

cells/mL after 1 week in situ incubations. Interestingly, 

this is similar to the abundance of FLB found in situ in 

the area of the contaminant plume (F513, 0.45 km 

downgradient from the sewage treatment plant) where 

the study is being conducted (Harvey and Barber, 

1992). Subsequent experiments now being conducted 

involve in situ incubations of aquifer protozoa and 

FLB. In these experiments, population changes are 

being monitored over time. Comparison of these 

results with those of bacteria alone will help determine 

if there is an interaction between the aquifer FLB and 

protozoa.

ACKNOWLEDGMENTS

We thank D.M. Shay, C.A. Gruden, A.L. Bunn, 

and M. Kazmierkiewicz-Tabaka of the University of 

New Hampshire (Durham, New Hampshire) for their 

laboratory work with protozoa. We also thank David 

Metge, Larry Barber and Richard Smith (USGS;

251
Boulder, CO), respectively, for their analyses of 

bacteria, DOC, and ammonia samples and the New 

England District Office of the USGS for their 

logistical support.

REFERENCES

Darbyshire, J.F., Wheatley, R.E., Greaves, M.P., and Inkson, 

R.H.E., 1974, A rapid micromethod for estimating 

bacterial and protozoan populations in soil: Revue 

d'Ecologie et de Biologic du Sol, v. 11, p 465-475.

Fenchel, T., 1987, Ecological physiology: feeding. 

Ecology of Protozoa The Biology of Free-Living 

Phagotrophic Protists: Madison, Wise., Science 

Technology, p. 32-52.

Fenchel, T, 1986, Protozoan filter feeding: Progress 

Protistology, v. 1, p. 65-113.

Harvey, R.W. and Barber, L.B., 1992, Associations of freeliving bacteria and dissolved organic compounds in a 

plume of contaminated groundwater: Journal of 

Contaminant Hydrology v. 9, no. 1, p. 91 -104.

Singh, B.N., 1946, A method of estimating the numbers of 

soil protozoa, especially amoebae, based on their 

differential feeding on bacteria: Applied Biology, 

v. 33, p. 112-119.

Singh, B.N., 1955, Culturing soil protozoa and estimating 

their numbers in soil, in Kevan, D.K.M., ed., Soil 

Zoology, Proceedings, Second Easter School in 

Agricultural Science: New York, Academic Press, 

p. 403-411.

Smith, R.L., Harvey, R.W., and LeBlanc, D.R., 1991, 

Importance of closely-spaced vertical sampling in 

delineating chemical and microbiological gradients in 

groundwater studies: Journal of Contaminant 

Hydrology, v. 7, p. 285-300.

252
Transport of Indigenous Protozoa in a Sandy Aquifer,

Cape Cod, Massachusetts

By Ronald W. Harvey1 , Nancy E. Kinner2 ,

David W.

Abstract

Transport of flagellates in sandy, organically 

contaminated aquifer sediments was investigated 

in a small-scale (2 m travel distance) naturalgradient tracer test in Cape Cod. The flagellates 

(average cell size, 2-3 um), which appear to be 

common in the Cape Cod aquifer, were grown in 

Cerophyl, labeled with hydroethidine (a vital 

eukaryotic stain), and coinjected into aquifer 

sediments along with a conservative tracer 

(bromide). Rates of immobilization in undisturbed 

aquifer sediments were more than two orders of 

magnitude greater than those observed earlier for 

indigenous ground-water bacteria. The high rate 

of immobilization appeared to be related to the 

surface chemistries of the flagellates and not their 

size. Retardation for the flagellates was two- to 

three-fold greater than that generally observed for 

indigenous groundwater bacteria. Apparent 

dispersion was also significantly greater for the 

flagellates than for the bacteria and was much 

closer to that observed earlier for highly surfaceactive, microbial-sized, carboxylated-latex 

microspheres. The data suggest that flagellates are 

probably transported slowly through the aquifer.

INTRODUCTION

The goals of the protozoa research program at 

the U.S. Geological Survey (USGS) Cape Cod ToxicSubstances Hydrology Research Site in Falmouth,

Dan MacDonald2 , Amoret L. Bunn2 , and 

Metge1

Massachusetts, are to learn more about the 

distribution, abundance, population characteristics, 

grazing efficiency, transport behavior, and role in 

subsurface contaminant degradation of protozoa in 

ground water. This paper discusses investigations 

designed to improve understanding of the subsurface 

transport of the abundant (up to 10 /gdw [grams dry 

weight]) flagellates that inhabit the Cape Cod aquifer. 

Funding for this work was provided by NSF grant 

#BCS-9012183 awarded to the University of New 

Hampshire (UNH). The objective of the experiments 

discussed herein was to gain information on the degree 

of retardation, immobilization, and apparent 

dispersion of the stained flagellates under in situ 

conditions. A second objective was to compare 

flagellate transport with that observed for indigenous 

bacteria in earlier experiments at the Cape Cod site 

(Harvey and others, 1989; Harvey and Garabedian, 

1991) and with particulate tracers (protozoa-sized 

microspheres) in the laboratory (Harvey and others, 

1993).

'U.S. Geological Survey, Boulder, Colo. 

Dept. of Civil Engineering, Univ. of New Hampshire, 

Durham, N. H.

METHODS

The test involved an unconfined, sandy, outwash 

aquifer that has been contaminated by a 4.5 km-long 

plume resulting from on-land disposal of treated 

sewage (fig. 1). Flagellates were isolated from aquifer 

sediments sampled from a depth of 8.5-10.1 m in the 

contaminated zone near the tracer-test site. The 

flagellates were grown in nutrient media (Page, 1988) 

and labeled with the vital stain, hydroethidine (Harvey 

and others, 1993). Ground water (100 L) containing 

labeled flagellates (2.2 x 107/L) and bromide 

(165 mg/L) was added slowly to the aquifer, by use of 

a peristaltic pump, at depths of 9.0 and 9.3 m at

253
0 20 40 80 KILOMETERS

70»34' 70*37

41* 38' 

4f 37 

EXPLANATION

 15 -£ WATER-TABLE CONTOUR - Show* Mtud* 

of water tabte i/bov* tov*l, In imttrt. 

Arrows show direction of ground-water flow.

Figure 1. Location of natural-gradient tracer test with hydroethidine-labeled flagellates at the U.S. Geological Survey Cape 

Cod Toxic-Substances Hydrology Research Site (modified from Smith and others, 1991, fig. 1).

USGS well M4-15 (fig. 1). The injection well is 

within a large (20 by 220 m) array of 15-port multilevel samplers (MLS)(LeBlanc and others, 1991) and 

within 100 m of where several in situ experiments 

were conducted to assess the subsurface transport of 

microbial-sized microspheres and indigenous bacteria

(Harvey and others, 1989; Harvey and Garabedian, 

1991; Harvey and others, 1993). Ground-water samples (500 mL) were collected daily at wells M4- 

15A and M5-15 in order to capture breakthrough of 

both bromide and stained-flagellates 1 and 2 m downgradient from the point of injection.

254
Bromide was analyzed using a specific-ion 

electrode and labeled flagellates were enumerated by 

epifluorescence microscopy. Samples of ground water 

(50-250 mL) were fixed (1 % final concentration) with 

glutaraldehyde that had been buffered in 0.001 M 

cacodylate (pH 7) and filtered onto 25 mm- 

(diameter), 0.8-jim pore size, black polycarbonate 

filters. Enumerations of hydroethidine-labeled 

flagellates were made under incident ultraviolet light 

by use of an epifluorescence microscope that was 

fitted with a 400 nm dichroic mirror, 330 to 380 nm 

excitation filter, and 420 nm barrier filter.

RESULTS

Results of the field transport experiment are 

given in figures 2 and 3, which depict the 

dimensionless concentration histories (breakthrough 

curves) of labeled protozoa and bromide at wells M4- 

15A and M5-15 (about 1 and 2 m downgradient from 

the point of injection). At 1 m downgradient 

(figure 2), the peak bromide concentration was 

approximately the same as that present in the injectate 

(that is, C/C0 = 1.0), which suggests that well M4-15A 

lies squarely along the trajectory of the injectate cloud. 

The breakthrough curve for bromide was 

characterized by a single peak, which occurred 

approximately 2 days after injection. Relative to 

bromide, the center of mass for the stained flagellate^

was retarded by longer than 1 week. Also, the 

breakthrough curve for the protozoa had two distinct 

peaks of approximately equal magnitude and an 

elongated tail. The initial peak arrived at well M4- 

15A 1 day earlier than the bromide peak. The second 

peak arrived 3 days after the bromide peak.

At 2 m downgradient (figure 3) from the point 

of injection, peak bromide concentration was less than 

30 percent of that present in the injectate as a result of 

dispersion and a downgradient sampling coordinate 

that may have been slightly off center with respect to 

the trajectory of the center of bromide mass. The peak 

in bromide concentration appeared to pass well M5-15 

about four days after injection. Again, the 

breakthrough curve for the stained protozoa was 

characterized by multiple peaks with a center of mass 

that was substantive!y retarded (at least 1 week) 

relative to bromide. Although bromide concentrations 

returned to background levels at well M5-15 

approximately 12 days after injection, substantive 

numbers (more than 4 x 104/L) of stained-protozoa 

were still observed 45 days into the test.

DISCUSSION

Substantial retardation factors (RF = 2-3) for 

ground-water flagellates has been observed in flowthrough columns packed with sieved (0.5 to 1.0 mm 

grain size) aquifer sediments (Harvey and others, in

0.3

0.0*

5 10 15 

DAYS SINCE INJECTION

Figure 2. Dimensionless concentration histories of 

bromide and hydroethidine-labeled flagellates at U.S. 

Geological Survey well M4-15A, following injection into 

sandy aquifer sediments 1 meter upgradient. Depth of 

collection was 12.3 meters. Error bars represent the 

standard error among replicate microscopic fields during 

flagellate enumeration.

5 10 15 45 

DAYS SINCE INJECTION

Figure 3. Dimensionless concentration histories of 

bromide and hydroethidine-labeled flagellates at U.S. 

Geological Survey well site M5-15, following injection into 

sandy aquifer sediments 2 m upgradient. Depth of 

collection was 12.3 meters. Error bars represent the 

standard error among replicate microscopic fields during 

flagellate enumeration.

255
1995). In contrast, peak breakthrough of stained, 

indigenous bacteria in previous forced- and naturalgradient tracer tests occurred coincidentally with or 

slightly before the bromide peak (Harvey and others, 

1989; Harvey and Garabedian, 1991). However, 

substantial retardation has been observed for various 

types of microspheres and has been found to increase 

with the reactivity of their surfaces (Harvey and 

others, 1989). Retardation of colloids in the Cape Cod 

aquifer is also affected by physical heterogeneity 

(Harvey and others, 1993).

The rate of flagellate immobilization (almost 

100 fold per meter of travel distance) is several orders 

of magnitude greater than that observed for the 

indigenous bacteria traveling through these same 

sediments (Harvey and Garabedian, 1991). However, 

it does not appear that straining (physical entrapment 

within pores that are smaller than the limiting 

dimension of the suspended particle) plays a very 

significant role in the immobilization of the flagellates, 

judging from grain-size distribution of aquifer 

sediments in the area of the tracer test (Barber, 1990) 

and the prerequisite ratio of cell diameter to critical 

pore size of 1.5 (see, for example, Matthess and 

Pekdeger, 1985). Optimal size for transport of 

spherical colloids with buoyant densities of 1.05 

appears to be approximately 2 jim, judging from 

predictions made with a physical model and the results 

of column experiments involving 0.45 to 4.8 jim 

microspheres (Harvey and others, 1993). It is more 

likely that most of the immobilization is caused by 

sorption, although sorption of ground-water flagellates 

in the presence of aquifer mineral surfaces has not 

been investigated.

The comparatively large apparent dispersion 

exhibited by the flagellates relative to that of bromide 

suggests that the chemical (diffusive) and 

hydrodynamic components of dispersion do not 

adequately describe the spreading of the protozoan 

peak at either well M4-15A or M5-15. A large 

disparity between the apparent dispersion of the 

conservative tracer and of the microbes was not 

observed for bacteria. However, large degrees of 

apparent dispersion were observed in earlier tracer 

experiments involving highly reactive carboxylated 

microspheres (Harvey and others, 1989; Harvey and 

others, in press). The role of motility in the apparent 

dispersion of the stained flagellates is not known and 

was beyond the scope of our experiment. However, 

substantive transport was observed only in the

direction of flow and where there was significant 

breakthrough of bromide.

CONCLUSION

The cultured ground-water flagellates used in 

this experiment appear to be immobilized very quickly 

in the presence of aquifer sediment grain surfaces, 

even though the small (approximately 2 p,m) 

flagellates may be near the optimal size for advective 

transport in this habitat. The tendency of the 

flagellates to become immobilized at grain surfaces 

likely results in transport characteristics that are 

significantly different than those observed for the 

indigenous bacteria that may serve as a food source. 

Relative to the bacteria, the flagellates are subject to 

substantial retardation and apparent dispersion during 

downgradient transport through aquifer sediments. 

Additional research is needed on the role of motility 

and encystment in the mobility of protozoa in 

contaminated aquifers.

ACKNOWLEDGMENT

We thank John Loveland (University of 

Colorado) and others for reviews of the manuscript.

REFERENCES

Barber, L.B., 1990, Geochcmical heterogeneity in a glacial 

outwash aquifer Effect of particle size and 

mineralogy on sorption of non-ionic organic solutes: 

Boulder, Colorado, University of Colorado, Ph.D. 

thesis, 224 p.

Harvey, R.W., George, L.H., Smith, R.L., LeBlanc, D.R., 

1989, Transport of microspheres and indigenous 

bacteria through a sandy aquifer Results of natural 

and forced-gradient tracer experiments: Environmental 

Science and Technology: v. 23, no. 1, p. 51-56.

Harvey, R.W., and Garabedian, S.P., 1991, Use of colloid 

filtration theory in modeling movement of bacteria 

through a contaminated sandy aquifer: Environmental 

Science and Technology, v. 25, no. 7, p. 178-185.

Harvey, R.W., Kinner, N.E., MacDonald, D., Metge, D.W., 

and Bunn, A., 1993, Role of physical heterogeneity in 

the interpretation of small-scale laboratory and field 

observations of microorganism, microsphcre, and 

bromide transport through aquifer sediments: Water 

Resources Research, v. 29, no. 8, p. 2713-2721.

256
Harvey, R.W., Kinner, N., Bunn, A., MacDonald, D., 1993, 

Transport of protozoa through an organically 

contaminated sandy aquifer, in Stanford, J., and 

Simons, J., eds., Proceedings of the First International 

Conference on Ground Water Ecology, Tampa, Fla., 

April 26-29: Huntsville, American Water Resources 

Association, p. 111-118.

Harvey, R.W, Kinner, N.E., Bunn, A., MacDonald, D., and 

Metge, D., 1995, Transport behavior of groundwater 

protozoa and protozoan-sized microspheres in sandy 

aquifer sediments: Applied and Environmental 

Microbiology, v. 61, no. 1, p. 209-217.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W, 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W, 

1991, Large-scale natural-gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, no. 5, p. 895- 910.

Matthess, G. and Pekdeger, A., 1985, Persistence and 

transport of bacteria and viruses in groundwater-a 

conceptual evaluation: Journal of Contaminant 

Hydrology, v. 2, p. 171-188.

Page, F.C., 1988, A new key to fresh water and soil 

gymnamoebae: Ambleside, Cumbria, U.K, Fresh 

Water Biological Association.

Smith, R.L., Harvey, R.W., and LeBlanc, D.R., 1991, 

Importance of closely-spaced vertical sampling in 

delineating chemical and microbiological gradients in 

groundwater studies: Journal of Contaminant 

Hydrology, v. 7, no. 3, p. 285-300.

257
258
Use of Static Column Experiments to Identify Factors

Affecting Bacterial Attachment in Contaminated Aquifer

Sediments From Cape Cod, Massachusetts

By David W. Metge1 , Ronald W. Harvey1 , George R. Aiken 1 , and Larry B. Barber, II 1

Abstract

The mineralogy of, sediment in, and 

geochemistry of, ground water from the Cape Cod 

contaminated aquifer have been found to affect 

bacterial attachment and transport through aquifer 

sediments. Static column experiments, under 

simulated aquifer conditions, were used to identify chemical factors (for example, pH, the concentration of dissolved organic carbon [DOC], the 

presence and concentration of divalent ions and 

other competing ligands, and ionic strength) 

which can affect bacterial transport in the subsurface. pH strongly affected bacterial attachment in 

uncontaminated sediments; an increase in pH 

from 5.8 to 7.9 resulted in a 70 percent drop in 

fractional bacterial attachment. Also, attachment 

of bacteria to aquifer sediments was substantially 

affected by changes in the amount and type of 

DOC from both contaminated and uncontaminated ground water. Finally, ionic strength and 

divalent ions, such as sulfate, were found to affect 

bacterial attachment. The experiments indicate 

that competitive and (or) synergistic interactions 

can operate under different pH conditions. Thus, 

different geochemical conditions can affect cellgrain surface interactions by competing with bacteria for binding sites and (or) by modifying grain 

or cell-surface characteristics.

INTRODUCTION

Bacterial transport through contaminated 

ground water is receiving increasing focus, because

'U.S. Geological Survey, Boulder, Colo.

bioremediation is often advocated as a solution to 

contamination problems. Introduction of genetically 

engineered or contaminant-adapted bacteria to 

contaminated aquifers can be a more suitable approach 

than engineering approaches, such as "pump-andtreat," which might not be effective in handling low 

levels of ground-water contamination. Bacterialtransport studies (Harvey and others, 1989; Scholl and 

Harvey, 1992) have shown that bacterial attachment 

and the extent of transport can be affected by a number 

of complex physical and chemical factors, the most 

important of which are pH, ground-water chemistry, 

the amount of sediment-associated organic and metaloxide material, and sediment mineralogy.

Static column experiments, designed to simulate 

aquifer conditions, were performed to determine the 

effect of specific chemical factors on bacterial 

attachment and transport characteristics. The goal of 

these experiments was to determine how bacterial 

attachment in contaminated and uncontaminated 

ground water (over a pH range of 3-9) was affected by 

(1) the amount and type of DOC; (2) ionic strength; 

and (3) sulfate (SO4~2) concentrations. Results 

presented here show how differences in ground-water 

geochemistry affect bacterial attachment to aquifer 

sediments, and how contributions of geochemical 

constituents are affected by pH.

MATERIALS AND METHODS 

Site Description and Sample Collection

Samples were collected from the Cape Cod 

Toxic-Substances Hydrology Research site. The study 

site is a sand and gravel aquifer located on Cape Cod, 

Mass. Disposal of secondary-treated sewage in 

infiltration beds since 1936 has resulted in an 

extensive plume of contaminated ground water

259
7r«r

MASSACHUSETTS

20 *0

0 20 40 CO KHOMfTEftS

33' 70* 32*

41* 38'H41»3T

CTIS AIM BASE ' 

SEWAQE-OISPOSM.

CGW sit*, SDW314-51 

Water samples.

__ ^^ 1C

Figure 1. Location of sampling sites, Cape Cod Toxic-Substances Hydrology Research site, Cape Cod, Mass, (modified 

from LeBlanc.1984).

260
(LeBlanc, 1984). Contaminated ground water (CGW) 

was collected from a well (S314-51) located in the 

core of the plume and about 30 m downgradient from 

the sewage outfall. Water samples from this well were 

collected at a depth of 15.5 m (about 6.4 m below the 

water table). Uncontaminated ground water (UGW) 

and sediments were collected from well F350-13, 3 

km downgradient from the sewage infiltration beds at 

a depth of 3.9 m (about 2.7 m below the water table). 

This depth is well above the plume of contaminated 

ground water. Sediment samples were collected with a 

piston-type core barrel using a hollow-stem auger 

(Zapico and others, 1987). Locations of sampling 

sites are shown in figure 1. Samples were maintained 

at 4° C until processed. Ground-water chemistries 

were analyzed according to published methods 

(Brinton and others, written communication; 

Garbarino and Taylor, 1979). Summaries of these 

chemical data are provided in table 1.

Table 1 . Summary of chemistry data for contaminated and 

uncontaminated ground water from Cape Cod Toxic-Substances 

Hydrology Study Site

Chemical property

Na+ concentration ( I )

Cl- concentration (2)

SO4-2 concentration! 2)

DOC concentration (3)

Specific Conductance

Ionic Strength (I) (4)

pH

Uncontaminated 

ground water

8.4 mg/L

1 1. 8 mg/L

7. 1 mg/L

0.9 mg/L

62 nS/cm

0.8 x 1 0-3

4.9

Contaminated 

ground water

47.3 mg/L

1 2.3 mg/L

28.2 mg/L

3.9 mg/L

488 (.iS/cin

4.5 x 1 0-3

5.9

(1) Ionic species measured with inductively-coupled atomic-emission spectroscopy.

(2) Ionic species measured by ion chromatography.

(3) DOC concentration measured by carbon analysis.

(4) Calculated from totals of ionic species.

Experimental Procedures

Static column experiments were performed by 

the method of Scholl and Harvey (1991, 1992). The 

method simulates aquifer conditions but avoids 

applicability problems inherent with stirred-batch 

systems; all bacteria used in the experiments were 

indigenous bacterial populations from CGW. The 

following modifications were made. Fifteen grams of 

sieved sediment (0.5-1.0 mm grain size) were used in 

all experiments. The number of bacteria loaded into 

each column was kept constant (70 x 106 cells/mL,

4.7 x 106 cells/g sediment) in all experiments. These 

were significantly greater than what was used 

previously (20-40 x 106 cells/mL, l-4x 106 cells/g 

sediment). These modifications facilitated comparison 

between experimental treatments and improved 

resolution of the effects of geochemical conditions. 

All experiments were performed in a refrigerated 

incubator(10°± PC).

Ground waters from an uncontaminated well 

and from a well within the contaminant plume were 

used to examine how contaminant concentrations in 

the plume affect bacterial attachment to aquifer 

sediments. Untreated, uncontaminated and untreated, 

contaminated ground waters were used (over a range 

in pH of 3-9) in static column experiments to assess 

bacterial attachment in contaminated and 

uncontaminated parts of the aquifer.

Static-column experiments were performed with 

two types of treated water to examine how the 

concentration and type of DOC present in untreated 

CGW affect bacterial attachment to aquifer sediments. 

For the first experiment, contaminated ground water, 

adjusted to pH 7.0, was passed through a column 

packed with XAD-8, a nonionic macroporous resin, to 

remove the hydrophobic neutral fraction of DOC from 

CGW (Aiken and others, 1992). Water for the second 

experiment was prepared by ultraviolet (UV)- 

oxidation of DOC in CGW using a quartz flowthrough cell containing an 185 nm UV light source. 

Contaminated ground water (1.2 L) was recirculated 

through the UV unit for 16 hours (~10° C) with a 

peristaltic pump operating at approximately 4 

mL/min. All DOC concentrations were determined by 

carbon analyzer.

The degree of fractional bacterial attachment to 

sediments was examined by changing ionic strength, 

DOC, and SO4'2 in CGW and UGW. The effect of 

ionic strength on bacterial attachment in the presence 

and absence of naturally occurring DOC was 

examined by using two treatments to UGW. For the 

first experiment, specific conductance was increased in 

UGW by NaCl amendment to approximate specific 

conductance levels in CGW (488 ^S/cm). For the 

second experiment, UV-oxidized UGW was amended 

with NaCl for comparative purposes. We examined 

elevated concentrations of sulfate in contaminated and 

uncontaminated ground water to determine the effect 

of sulfate on bacterial attachment.

261
Uncontaminated ground water was amended with sulfate to concentrations found in untreated CGW (28 

mg/L). For comparative purposes, CGW was UV-oxidized for 16 hours to remove virtually all DOC and 

used to determine how sulfate content of CGW 

affected bacterial attachment.

RESULTS AND DISCUSSION

Overall Effects of Contaminants on Bacterial 

Attachment

Attachment experiments were performed with 

indigenous bacterial populations from CGW, using 

untreated UGW or CGW (over a pH range of 3-9), and 

sediments from an uncontaminated area of the aquifer. 

The pH strongly affected fractional bacterial 

attachment in the presence of untreated UGW. A 70 

percent drop (from 92 to 21 percent) accompanied a 2- 

unit increase in pH (pH 5.8 to 7.9). This increase in 

attachment under acidic pH conditions may be 

analogous to observations of increased sorption of 

dissolved organic materials to oxide surfaces with a 

decrease in pH (Davis, 1982). The effects of pH on 

bacterial attachment to sediments in untreated CGW 

was less clear. Fractional bacterial attachment varied 

by only 16 percent (from 67.4 to 83.7 percent) over the 

pH range of 3.5 to 8.8. Substantially greater 

attachment occurred under slightly acidic to alkaline 

conditions (pH 6.5-8.0) and less attachment occurred 

under acidic conditions than in untreated UGW over 

the same pH range. These observations corroborate 

results seen with similar aquifer material and 

conditions (Scholl and Harvey, 1992). The results 

indicate that DOC or the inorganic chemical matrix 

within CGW suppresses bacterial attachment under 

acidic to mildly acidic conditions, but promotes 

attachment under neutral to alkaline conditions.

Effects of DOC On Bacterial Attachment

The degree of fractional bacterial attachment to 

aquifer sediments was affected by slight changes in 

concentration and type of DOC present in CGW. 

When compared to the degree of bacterial attachment 

observed for untreated CGW over the pH range of 3.5- 

6.0, removal of the hydrophobic-neutral fraction of 

DOC (about 3 percent of the total DOC) from CGW 

by passage through a nonionic macroporous ion

exchange resin (at pH 7.0) increased bacterial 

attachment under acidic to mildly acidic conditions. 

However, the same relative fractional attachment as 

untreated CGW was observed at pH 6.1, indicating 

that the hydrophobic neutral fraction of CGW may 

suppress attachment under acidic pH conditions. 

When virtually all DOC from CGW was removed by 

UV oxidation (97 percent of DOC removed), a 20 to 

45 percent decrease in fractional bacterial attachment 

above pH 5.5 resulted, indicating that DOC other than 

the hydrophobic neutral fraction of DOC in CGW 

promoted bacterial attachment under alkaline 

conditions. It is not clear whether the DOC alters 

grain surfaces and (or) changes bacterial-surface sites. 

Other studies on adsorption of DOC to grain surfaces 

have indicated that DOC masks positive surface 

charges associated with iron oxyhydroxide coatings on 

grain surfaces (Davis, 1982). Results from our 

experiments indicate that, if masking of mineral 

surface charge by DOC occurs, the mechanism can 

promote or suppress attachment under differing pH 

conditions.

Effects of DOC / Ionic Strength on Bacterial 

Attachment

The pattern of bacterial attachment in ionicstrength-adjusted UGW was substantially different in 

the presence and absence of naturally occurring 

DOC. The specific conductance of UGW was raised 

8-fold (from 62 to 480 j^S/cm) and ionic strength was 

doubled (from 8 x 10'4 to 1.6 x 10"3 ) by addition of 

NaCl to UGW. Bacterial attachment to sediments in 

ionic-strength-adjusted UGW in the presence of DOC 

was increased (over the pH range 6-9) when compared 

to levels of bacterial attachment in untreated CGW. 

However, ionic strength adjustments to UGW (UVoxidized) in the near-absence of naturally occurring 

DOC had less of an effect on bacterial attachment 

compared to levels observed in UGW that was only 

altered with respect to ionic strength. This indicates 

that naturally occurring and anthropogenically derived 

DOC can affect bacterial attachment in different ways.

Effect of Sulfate on Bacterial Attachment

Elevated concentrations of 804 can also play a 

significant role in bacterial attachment to aquifer

262
sediments. The effect of elevated SO4~2 

concentrations was assessed by amending UGW 

(7.1 mg/L SO4"2) with Na2SO4 to concentrations 

comparable to those in CGW (28 mg/L SO4'2 ). A 

second system was prepared by destroying DOC in 

CGW by UV oxidation for 16 hours. This was done 

because the effects of SO4"2 in CGW could be masked 

by levels of DOC present in untreated CGW. Sulfate 

amendment to UGW decreased fractional bacterial 

attachment to aquifer sediments in the range of pH 

4.5-6 relative to the extent of bacterial attachment seen 

with unamended UGW. DOC removal from CGW 

decreased bacterial attachment to levels similar to 

those of sulfate-amended UGW, suggesting that SO4 

might compete with bacteria for sediment-grain 

binding sites and (or) alter cell-surface characteristics.

SUMMARY

We found that low (part-per-million) *\

concentrations of DOC and SO4 concentrations and 

modest (several-fold) changes in ionic strength can 

affect bacterial attachment to aquifer sediments. 

However, pH appears to be the primary factor 

governing bacterial attachment and transport. 

Competitive, synergistic, and inhibitory mechanisms 

appear to be affected by pH. More experiments are 

planned with natural and human-derived DOC (for 

example, humic acids and analogs of contaminant 

ionic and nonionic surfactants) to define the 

mechanisms governing bacterial attachment to aquifer 

sediments in the presence and absence of organic 

material. Also, the influence of divalent cations, such 

as Mg+2, Ca+2, and anions, such as SO4"2 , and their 

interaction with DOC to affect bacterial attachment 

deserve closer examination. These experiments were 

done with indigenous bacterial communities that are a 

heterogeneous mix of populations, each with different 

cell-surface characteristics that can differ in how they 

are affected by the plume geochemistry. Ultimately, 

we will apply results from the static-column 

experiments to field tracer tests to determine how 

contaminants affect bacterial transport.

REFERENCES

Aiken, G.R., McKnight, D.M., Thorn, K.A., and 

Thurman, E.M., 1992, Isolation of hydrophilic 

organic acids from water using nonionic macroporous 

resins: Organic Geochemistry, v. 18, p. 567-573.

Davis, J.A., 1982, Adsorption of natural dissolved organic 

matter at the oxide/water interface: Geochimica et 

Cosmochimica Acta, v. 46, p. 2381-2393.

Garbarino, J.R., andTaylor, H.E., 1979, An inductivecoupled plasma atomic-emission spectromctric method 

for routine water quality testing: Applied 

Spectroscopy, v. 33, p. 220-226.

Harvey, R.W., George, L.H., Smith R.L., and LeBlanc, 

D.R., 1989, Transport of microspheres and indigenous 

bacteria through a sandy aquifer: results of natural- and 

forced-gradient tracer experiments: Environmental 

Science and Technology, v. 23, p. 51 -56.

LeBlanc, D.R., 1984, Sewage plume in a sand and gravel 

aquifer, Cape Cod, Massachusetts: U.S. Geological 

Survey Water-Supply Paper 2218, 28 p.

Scholl, M.A., and Harvey, R.W., 1991, Laboratory studies 

of bacterial sorption to aquifer sediments from the 

Cape Cod toxic-substances hydrology research site, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Program in Toxic Substances 

Hydrology Program-Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91 -4034, p. 152-154.

Scholl, M.A., and Harvey, R.W., 1992, Laboratory investigations on the role of sediment surface and groundwater chemistry in transport of bacteria through a 

contaminated sandy aquifer: Environmental Science 

and Technology, v. 26, p. 1410-1417.

Zapico, M.M., Vales, S., and Cherry, J.A., 1987, A wireline piston core barrel for sampling cohesionless sand 

and gravel below the water table: Ground Water, v. 5, 

p. 74-82

263
264
An Experiment to Quantify Temporal Variability of

Water Samples Obtained from Screened Wells,

Cape Cod, Massachusetts

By Thomas E. Reilly 1 and Denis R. LeBlanc2

Abstract

The chemical composition of water sampled 

from some screened wells has been observed to 

exhibit sustained temporal variability. Results 

from a field experiment, undertaken on August 12, 

1992, in the area of the contaminant plume from 

the Otis Air Base sewage-disposal sand beds, 

support the recent hypothesis that the temporal 

variability can be attributed to the flux along the 

well screen reflecting the spatial heterogeneity in 

hydraulic conductivity and water chemistry of the 

aquifer near the well. The experiment consisted of 

withdrawing water from a well with a screened 

interval approximately 40 ft long and sampling 

over time during a 5-hour period. Water samples 

also were collected from multilevel samplers 

surrounding the well before, during, and after the 

pumping of the well. Selected constituent concentrations and properties measured in samples 

collected from the discharging well showed different temporal trends during the 5-hour test. For 

example, the ferrous iron concentration decreased, 

the calcium concentration increased, and specific 

conductance remained relatively constant. The 

different trends are due to the different distributions of the constituents around the well as documented by analysis of water samples collected by 

the set of multilevel samplers. The observed trends 

are in agreement with the hypothesis and apparently are due not to the purging of the water standing in the well but to the flow and transport of the 

water and chemical species in the heterogenous 

aquifer in the immediate vicinity of the well.

'U.S. Geological Survey, Reston, Virg. 

2 U.S. Geological Survey, Marlborough, Mass.

INTRODUCTION

Water that enters a well through a well screen is 

derived from different layers of porous material. These 

layers are present at different scales and can be characterized by different physical properties, such as hydraulic 

conductivity, and chemical properties. Layer boundaries 

defined by one property may not coincide with those 

defined by a different property. The water withdrawn 

from the well is a composite of the waters present in 

these different layers along the length of the screen. The 

length of well screens can span vertical distances of as 

little as 1 foot or as much as 100 or more feet. The rate of 

water withdrawn from each layer is variable and depends 

on the hydraulic conductivity, screen placement, pumpintake placement, and other factors.

Wilson and Rouse (1983), Rivett and others 

(1990), Gibs and Imbrigiotta (1990), Reilly and Gibs 

(1993), and others have observed that chemical concentrations of various constituents can change in the 

water discharging from a well over the time of sampling. Thus, questions arise as to what the concentration of the constituent is in the ground water, or what 

water is actually being sampled. Some of the observed 

temporal changes are due to the removal of water that 

had been residing in the well casing prior to sampling; 

but some of this variation appears to persist. The persistent temporal variations in the concentrations may 

be due to both the chemical and hydraulic-conductivity 

distributions in that part of the aquifer that contributes 

water to the well.

The variability of hydraulic conductivity in 

ground-water systems is widely recognized. Wolf and 

others (1991) and Hess and others (1992) recently 

showed considerable variation in hydraulic conductivity, on the scale of centimeters, in the sand and gravel 

aquifer at Cape Cod, Mass. Geochemical investigations, also conducted on Cape Cod, Mass. (Smith and 

others, 1991), indicated the necessity of collecting 

closely spaced samples for chemical measurements 

because of the vertical variability in chemistry over 

small distances.

265
The purpose of this paper is to describe an experiment performed at Cape Cod, Mass., to examine the 

hypotheses that (1) the observed concentrations of constituents in the water sampled from a screened well are 

due to the flux of these constituents into the well as 

determined by the hydraulic conductivity and chemical 

distribution near the well, and (2) the changes in concentration observed from samples collected continually from a long screened well can be interpreted to aid 

in understanding the hydraulic and chemical properties 

of the aquifer that is the source of water to the well.

DESCRIPTION OF THE EXPERIMENT

The experiment was designed to evaluate the 

chemistry of the water discharged from a well screened 

in a physically and chemically heterogenous aquifer. 

Well F453-63 was selected because it was known that 

the screen penetrated both the oxic and anoxic zones of 

the sewage plume from the Otis Air Base sewagedisposal sand beds (fig. 1). The ground-water chemistry differs in the oxic and anoxic zones, thereby providing a heterogeneous chemical environment along the 

screened interval of well F453-63. Three multilevel

71°00'

42°30'

73°00' 72°00'

/

MASSACHUSETTS 

/ BOSTO

42°00 

0 20 40 60 MILES

)T\

0 20 40 60 KILOMETERS

70°34' 33' 70°32'

41°38'

41°37'

I OTIS AIR BASE

SEWAGE-DISPOSAL

SAND BEDS

500 METERS

I

F45 3

EXPLANATION

WATER-TABLE CONTOUR -- Shows altitude 

of water table above sea level, in feet. 

Arrows show direction of ground-water flow.

SCREENED WELL USED FOR PURGING 

EXPERIMENT

Figure 1. Ground-water study site on Cape Cod, Massachusetts, 

showing the location of the contaminant plume and the screened 

well used in the well-purging experiment.

266
(A) Screened Well

MLS #2

Multilevel -""^

sampler (MLS) 

1 5 ports

3.3 ft spacing

1V1L

S

^o

c

i

ttz.

« » 

2.9ft

!

r43J-oj Land

67 

-60

-50 

Wate

-40

-20

I Altitude, i feet abov

i " 1" sea level

-0

--10

NOT TO SCALE

(B)

Approximate direction 

of ground-water flow

MLS#1

MLS #3 MLS #2

-\_ \ __ Screened Well

O O O

036 feet

Figure 2. Well and multilevel sampler layout for well-purging 

experiment: (A) cross-sectional view, (B) plan view.

samplers were installed near well F453-63 to describe 

the chemical distribution in the immediate vicinity of 

the screened well (fig. 2). The hydraulic conductivity 

distribution of the aquifer along the screen in well 

F453-63 ranges from 77 to 958 ft/d (Kathryn Hess, U.S. 

Geological Survey, written comm., 1993) as determined 

by the flowmeter method of Hess and others (1992).

Well F453-63 is constructed of 2-inch insidediameter polyvinyl chloride (PVC) pipe with four sections of 10-ft-long flush-joint threaded 2-inch insidediameter PVC screens. The screen extends from an altitude of 4.7 to 43.7 ft above sea level and the initial 

water-table altitude was 44.33 ft above sea level at the 

time of the test. The volume of water residing in the 

casing and screened section of the well is approximately 0.943 ft3 . The multilevel samplers installed in 

the vicinity of well F453-63 had 13 or 14 ports in the 

saturated zone. Each multilevel sampler consists of 15 

color-coded polyethylene tubes (0.25-inch outside 

diameter). The tubes run from land surface down the

inside of a 1.25-inch inside-diameter PVC casing and 

out holes spaced every 3.3 ft drilled through the PVC. 

The open, down-hole end of each tube is screened with 

a fine nylon fabric.

The experiment was undertaken on August 12, 

1992. Well F453-63 was sampled over time, and the 

multilevel samplers were used to document the chemical distribution in the ground water in the immediate 

vicinity of the well. A variable rate submersible pump 

was set 0.5 ft below the static water level in well F453-63. 

The continuous discharge from the screened well was 

measured 15 times during the 5-hour experiment using 

a graduated cylinder; discharge ranged from 0.047 to 

0.051 ft3/min, indicating that a volume equivalent to 

one casing volume is discharged about every 20 minutes. 

Water-level measurements were made before pumping, 

just prior to the end of the test, and a half hour after the 

end of the test. Drawdown was measured to be 0.02 ft 

in the pumped well after 5 hours of pumping.

267
The long-screen well was pumped for 5 hours 

and sampled at 5 minutes after discharge began, then 

every 10 minutes for the first 2 hours, and finally 

every 15 minutes for the remaining 3 hours. Water 

samples were collected from the three multilevel 

samplers before and after the test to observe any 

changes in constituent concentration in the immediate vicinity of the well screen. Multilevel sampler 2 

was also sampled halfway through the experiment. 

All water samples were analyzed in the field for 

specific conductance, pH and ferrous iron (Fe+2) concentration. The samples were subsequently analyzed 

for concentrations of boron, calcium, chloride, iron 

(Fe total), phosphorous, potassium, magnesium, 

manganese, sodium, zinc, and nitrogen species, including 

nitrous oxide, ammonium, nitrite and nitrate.

DATA FROM EXPERIMENT FOR 

SELECTED CONSTITUENTS

The distribution with depth of specific conductance, and concentrations of ferrous iron, calcium, 

chloride, and nitrous oxide (N2O) from multilevel sampler 2 during the test is shown in figure 3. The vertical 

distributions of the specific conductance and the chemical constituents are all different. The vertical variability in concentration is significant and concentrations 

change over short vertical distances. This vertical variability is expected because of the presence of the sewage plume, which is intersected by the screen of well 

F453-63. The specific conductance and four chemical 

constituents were chosen for collection because of their 

expected different distributions along the well screen.

o 20 40 60

CHLORIDE CONCENTRATION, 

IN MILLIGRAMS PER LITER

"0 200 400 600

SPECIFIC CONDUCTANCE, IN

MICROSIEMENS PER CENTIMETER

10 20 

NITROUS OXIDE, 

IN MICROMOLAR

30

J50

g

a 

< 400

w 

on

^ O n* 301 > CQ

<

S 2

S 10

< °0 10 20 30 

FERROUS IRON CONCENTRATION, 

IN MILLIGRAMS PER LITER

'0 5 10 15 20

CALCIUM CONCENTRATION,

IN MILLIGRAMS PER LITER

Figure 3. Concentration distribution with depth of specific conductance, ferrous iron, calcium, 

chloride, and nitrous oxide at multilevel sampler 2 after approximately 2.5 hours into the test.

268
Almost identical specific conductance and concentration distributions were found in all three multilevel 

samplers, indicating that the specific conductance and 

constituent concentrations are relatively uniform laterally (at least on the scale of feet) although they are very 

variable vertically.

The water discharging from the screened well was 

collected and analyzed as described above. The specific conductance and concentrations of ferrous iron, 

calcium, chloride, and nitrous oxide as measured during the course of the experiment are shown in figure 4. 

The concentrations of these selected constituents over 

time show some variability; for example, ferrous iron 

decreased, calcium increased, and chloride remained 

relatively constant.

IMPLICATIONS OF THE EXPERIMENTAL 

DATA ON WELL-PURGING CRITERIA

Reilly and Gibs (1993) simulated a hypothetical 

system to illustrate that the composition of water discharging over time from a physically and chemically 

heterogeneous aquifer can change over time depending 

on the distribution of the chemical constituents of the 

water and the hydraulic conductivity of the aquifer. 

They showed that changes over time are not necessarily due to the evacuation of the stagnant water in the 

well casing, but can also be due to the transport of the 

chemical constituents through the aquifer to the well. 

The results to date at the Cape Cod well-purging experiment corroborate these simulations.

PH ->

1^36

8^34

g§30

ttJ

H400

w.

fe-340

£w320 

UK

1000 1200 1400 1600 

MILITARY CLOCK TIME, AUGUST 12,1992

7

1000 1200 1400 1600 

MILITARY CLOCK TIME, AUGUST 12,1992

§g«

02

1000 1200 1400 1600 

MILITARY CLOCK TIME, AUGUST 12,1992

10

SS|s

*2

oo

, 2

1 1000 1200 1400 1600

MILITARY CLOCK TIME, AUGUST 12,1992

1200 1400 1600 

MILITARY CLOCK TIME, AUGUST 12,1992

Figure 4. Changes in specific conductance and concentrations of calcium, chloride, ferrous 

iron, and nitrous oxide over time in water pumped from well F453-63.

269
When a constituent or property has little variation 

over the screened interval, such as for the specific conductance and chloride concentration at well F453-63, 

the concentration or value of the property in the water 

discharging from the well changes little or not at all 

over time once the standing water in the well is 

removed. However, when the chemical distribution is 

variable along the screened interval, as was the case for 

ferrous iron, calcium and nitrous oxide, the concentration of the constituent in the water discharging from the 

well does not necessarily remain constant as the well is 

pumped. The concentration of ferrous iron in the 

ground water in the screened interval of the well 

remained near or below the detection limit from the 

water table to about half way down the screened interval and then rapidly increased to about 20 mg/L for the 

bottom half of the screened interval (fig. 3). The corresponding temporal trend in the concentration of ferrous 

iron in the water pumped from the screened well continuously decreased from about 8 to 6 mg/L during the 

5-hour test. In contrast, the concentration of calcium in 

the ground water in the screened interval of the well 

decreased from about 15 mg/L over the top half of the 

screen to about 5 mg/L over the bottom half of the 

screened interval (fig. 3). The concentration of calcium 

continuously increased over time from about 7 to 9 mg/L 

during the 5-hour test (fig. 4). The nitrous oxide concentration in the ground water has a peak concentration 

of 20 fiM in the middle of the screened interval (fig. 3). 

The trend of the nitrous oxide concentration from the 

pumped well over the 5-hour test is unclear (fig. 4).

It is important to note that a volume of water equal 

to the amount of water stored in the well is discharged 

approximately every 20 minutes. The observed temporal trends for the ferrous iron and the calcium concentration are persistent over the 5-hour test, during which 

time approximately 15 casing volumes were evacuated. 

The mixing of the water in the well bore will also affect 

the temporal distribution of concentration to a degree, 

but the persistence of the observed trends and the hypothetical numerical experiments of Reilly and Gibs 

(1993) indicate that the primary cause of these trends is 

a result of flow and transport within the heterogeneous 

aquifer system. Thus, the observed trends in water 

chemistry over time are not due to the purging of the 

water standing in the well but rather are due to the flow 

and transport of the water and chemical species in the 

aquifer in the immediate vicinity of the well.

The data obtained from multilevel samplers 1, 2 

and 3 indicate that the water chemistry was relatively 

constant laterally around the well in each layer, even 

though the vertical variation in water chemistry 

between layers can be great. The amount of water discharged from the screened well over the entire 5-hour 

test period was approximately 14.4 ft3 . This volume of 

water would be removed from a cylinder of aquifer 

material (assuming a porosity of 0.4) that extends about 

0.5 ft away from the well screen. The water at the 

multilevel samplers did not move much and is representative of the temporal and areal variation in water 

chemistry around the well. Thus, the temporal trends 

observed in the water discharging from the pumped 

well are probably due primarily to the vertical variations in water chemistry and the hydraulic conductivity 

of the aquifer. Flowlines will converge on the layers of 

high hydraulic conductivity and transport solutes 

vertically into different layers as the well is pumped. 

This mixing of waters of differing composition can 

cause the water chemistry in the water mixture in the 

discharging well to change over time until a steady 

transport field is developed.

FUTURE STUDY PLANS

Work is currently progressing on the numerical 

simulation of the experiment. The simulation of the 

transport processes in the screened zone will attempt to 

reproduce quantitatively the observed temporal concentration trends in the discharge water from the 

screened well. The solute-transport simulation of the 

ground-water system, which will use cylindrical coordinates, will incorporate the estimated hydraulic conductivity distribution and the chemical distribution 

determined in this experiment. This simulation will 

enable the mass of the differently distributed chemical 

species entering the well screen to be estimated quantitatively. The additional process of transport within the 

well bore itself also will be evaluated. This simulation 

can either corroborate or prove invalid the hypothesis 

that the observed trends are indeed due to the physical 

and chemical heterogeneity in the aquifer.

SUMMARY

The chemical composition of water pumped from 

some screened wells has been observed to exhibit 

sustained temporal variability. Results from a field

270
experiment, in the area of a contaminant plume from 

the Otis Air Base sewage-disposal sand beds, support a 

recent hypothesis that such temporal variability can be 

attributed to the flux along the well screen reflecting the 

spatial heterogeneity in hydraulic conductivity and 

water chemistry of the aquifer near the well. The experiment consisted of withdrawing water from a well with 

a screened interval approximately 40 ft long and sampling over time during a 5-hour period. Water samples 

also were collected from multilevel samplers surrounding the well before, during, and after the pumping of 

the well. All water samples were analyzed in the field 

for specific conductance, pH and ferrous iron. The 

samples were subsequently analyzed for boron, calcium, chloride, iron (Fe total), phosphorus, potassium, 

magnesium, manganese, sodium, zinc, and nitrogen 

species, including nitrous oxide, ammonium, nitrite, 

and nitrate.

The data collected from this experiment show that 

the chemical composition of water discharging from a 

screened well in a physically and chemically heterogeneous aquifer can have persistent variations over time. 

The concentrations of selected constituents in the water 

sampled over time from the discharging well show 

trends that appear to be associated with the vertical distribution of these constituents in the aquifer. The chemical data collected during the experiment corroborates 

results of earlier work that indicate that these trends 

depend on the distribution of the chemical constituents 

and the hydraulic conductivity around the well.

REFERENCES

Gibs, Jacob, and Imbrigiotta, T.E., 1990, Well-purging 

criteria for sampling purgeable organic compounds: 

Ground Water, v. 28, no. 1, p. 68-78.

Hess, K.M., Wolf, S.H., and Cclia, M.A., 1992, Large-scale 

natural gradient tracer test in sand and gravel, Cape 

Cod, Massachusetts, 3 Hydraulic conductivity variability and calculated macrodispersivities: Water 

Resources Research, v. 28, no. 8, p. 2011-2027.

Reilly, T.E., and Gibs, Jacob, 1993, Effects of physical and 

chemical heterogeneity on water-quality samples 

obtained from wells: Ground Water, v. 31, no. 5, 

p. 805-813.

Rivett, M.O., Lerner, D.N., and Lloyd, J.W., 1990, Temporal variations of chlorinated solvents in abstraction 

wells: Ground Water Monitoring Review, v. 10, no. 4, 

p. 127-133.

Smith, R.L., Harvey, R.W., and LcBlanc, D.R., 1991, 

Importance of closely spaced vertical sampling in 

delineating chemical and microbiological gradients 

in groundwater studies: Journal of Contaminant 

Hydrology, v. 7, no. 3, p. 285-300.

Wilson, L.C., and Rouse, J.V., 1983, Variations in water 

quality during initial pumping of monitoring wells: 

Ground Water Monitoring Review, v. 3, no. 1, 

p. 103-109.

Wolf, S.H., Celia, M.A., and Hess, K.M., 1991, Evaluation 

of hydraulic conductivities calculated from multiportpermeameter measurements: Ground Water, v. 29, 

no. 4, p. 516-52.

271
272
Preliminary Type-Curve Analysis of an Aquifer Test in an 

Unconfined Sand and Gravel Aquifer, Cape Cod,

Massachusetts

By Allan F. Moench 1 , Denis R. LeBlanc2, and Stephen P. Garabedian2

Abstract

An aquifer test was conducted in a sand 

and gravel, glacial-outwash deposit on Cape Cod, 

Massachusetts. A partially penetrating well was 

pumped at a constant rate of 320 gallons per 

minute for 72 hours. Changes in hydraulic head 

were observed at 20 locations through observation wells and piezometers distributed both radially and vertically within the saturated zone of the 

aquifer. The test was analyzed using Neuman's 

analytical model for flow to a partially penetrating 

well in a water-table aquifer. On the basis of geophysical well logs, cores, and other investigations 

conducted in the local area, the aquifer was initially assumed to have a saturated thickness of 

80 feet for the purpose of analysis. However, 

drawdown observed in a piezometer located 

108 feet below the water table, and an initial 

analysis, revealed a complete lack of conformity 

between theoretical and observed drawdown in 

piezometers located near the pumped well. Also, 

the computed horizontal hydraulic conductivity 

was twice that obtained by independent studies. 

Ongoing studies of the regional stratigraphy indicate that the saturated thickness of the aquifer is 

actually closer to 160 feet than to 80 feet. By 

increasing the assumed saturated thickness to 

160 feet, the theoretical drawdown was found to 

match observed drawdown for most of the test 

period at nearly all points of observation. Results

'U.S. Geological Survey, Menlo Park, Calif. 

2 U.S. Geological Survey, Marlborough, Mass.

indicate that there is a remarkable degree of 

aquifer homogeneity with respect to flow at the 

scale of the test. The aquifer transmissivity and 

specific yield were determined to be 38 feet 

squared per minute and 0.23, respectively, and the 

ratio of vertical to horizontal hydraulic conductivity was estimated to be about 1:2.

INTRODUCTION

As part of an effort to quantify the hydraulic 

properties of the unconfined, sand and gravel, glacialoutwash aquifer at the Cape Cod Toxic-Substances 

Hydrology Research site near Falmouth, Massachusetts, a 72-hour aquifer test was carried out in August, 

1990. The test was run by pumping a well at a constant rate and by observing the resulting changes in 

hydraulic head at locations that differ in distance and 

azimuth from the pumped well and in depth below the 

water table. In this paper transmissivity, horizontal 

and vertical hydraulic conductivity, and specific yield 

are estimated by type-curve analysis using the 

Neuman (1974) analytical model for flow to a partially 

penetrating well in a water-table aquifer.

WELL-AQUIFER CONFIGURATION AND 

DATA COLLECTION

Although the unconsolidated sediments in the 

test area are approximately 300 ft thick (LeBlanc, 

1984), studies indicate that clean, medium to coarsegrained, high-permeability glacial outwash overlies 

fine-grained, relatively low-permeability material at a 

depth of about 80 ft below the water table (Hess and 

others, 1992). The pumped well (well F507-080),

273
which was installed in July 1990 for this aquifer test 

and a concurrent forced-gradient tracer test, was 

drilled by cable-tool methods to a depth of about 80 ft 

below land surface. Located within an abandoned 

gravel pit, the well is 338 ft south and 76 ft west of the 

origin of coordinates (well no. F343-036 in figure 1) 

on a magnetic-north oriented grid (LeBlanc and 

others, 1991). An 8-inch inside-diameter (i.d) polyvinyl chloride (PVC) casing was installed with an 8-inch 

i.d PVC screen along the bottom 47 ft of the well. 

Backfill consisted only of natural collapse material 

and cuttings from the hole. The top and bottom of the 

screen were located 13.2 and 60 ft, respectively, below 

the initial water table, which was approximately 19 ft 

below land surface. Observation piezometers and 

wells were constructed by auguring to prescribed 

depths and installing 2-inch i.d. PVC casings with 

2-ft-, 9-ft-, or 39-ft-long PVC screens. Changes in 

hydraulic head were observed at 20 locations (see 

table I and figure 1) and at the pumped well.

In August 1990, well F507-080 was pumped 

at a rate of 320 gal/min for 72 hours. Discharge water 

was diverted through fire hose to the Otis Air Base 

sewage-infiltration beds located about 500 to 600 ft 

upgradient (north) of the test area. The rate of well 

discharge was monitored by a manometer and orifice 

at the discharge point and by noting the time required 

to fill a 55-gal drum at the end of the fire hose. Adjustments to the wellhead valve were made as necessary to 

maintain a constant flow rate. Water levels in piezometers and observation wells analyzed in this paper were 

measured manually using a steel tape. In addition, 

pressure transducers were used at 12 of the locations. 

Drawdown in the pumped well was monitored with an 

electric tape. Just prior to the end of the test the drawdown in the pumped well was 4.2 ft.

RESULTS AND ANALYSIS

This analysis was performed using the analytical model of Neuman (1974) for flow to a partially 

penetrating well in a water-table aquifer. Model 

parameters are defined in table 2. Only manual measurements of hydraulic head are used in the analysis as 

very early measurements are not needed. To perform 

the analysis, head measurements were converted to 

drawdown, and time was divided by the square of the 

radial distance between the pumped well and the point 

of observation. Composite time-drawdown plots (h as 

a function of t/r2 ) were then drawn for type-curve

analysis. The procedure described by Moench (1993) 

was used to generate computed values for the type 

curves. Assuming an initial saturated thickness of 

80 ft, close agreement was obtained between theoretical type curves and drawdown observed at distant 

locations, but drawdown at locations close to the 

pumped well bore no resemblance to the theoretical 

type curves. The transmissivity and specific yield 

obtained from the analysis were 40 ft /mm and 0.14, 

respectively. The ratio of vertical to horizontal 

hydraulic conductivity, K7/KP was determined to be 

1:5. The horizontal hydraulic conductivity based on 

the assumed saturated thickness of 80 ft was 

0.50 ft/min. This is about twice that obtained in prior 

studies (Hess and others, 1992). The value of the ratio 

K7/Kr was also less than that obtained by other 

investigators (Hess and others, 1992).

Observations made during this aquifer test 

indicate that the change in lithology at a depth of 80 ft 

is not sufficiently sharp to justify the assumption of an 

impermeable base to the aquifer at that depth. A plot 

of drawdown measured in two piezometers located 

about 14 ft from one another in the horizontal direction and about 90 ft from the pumped well is shown in 

figure 2. One piezometer is situated about 20 ft above 

and the other about 26 ft below the hypothesized base 

of the glacial outwash. Drawdown in both piezometers occurs in direct response to pumping at well F507- 

080. Thus, a saturated thickness of 80 ft was deemed 

inappropriate for this analysis.

Because of (1) the lack of agreement 

between the drawdown at observation points near the 

pumped well and the theoretical type curves; (2) the 

presence of drawdown in an observation piezometer 

28 ft below the assumed base of the aquifer; and 

(3) the fact that the calculated hydraulic properties 

are not in complete accord with those from prior 

studies, an additional analysis was performed under 

the assumption that the initial saturated thickness 

was 160 ft. This value for saturated thickness correlates with a significant change in lithology from 

coarse to very fine sands based on work cited by 

LeBlanc (1984, fig. 5) and recent regional geologic 

studies by John Masterson (USGS, oral commun., 

1993). Results of the analysis follow.

Composite time-drawdown plots for the 

piezometers and observation wells indicated in 

table 1 are shown in figures 3 and 4. Superimposed 

on the data are the type curves that resulted in the 

"best" fit to the data, as determined by visual

274
41 "30'

IlllllllB ,

Cape"

ilSlf^lRfaiiliJgip^isHiiii;;-, Cod """"""""""""""""""

MAGNETIC 

NORTH

F343-36

F376-3

85-32

F434-60 

FJ381-56

F5Q7-8Q 7

F505-59 

F505-80 

F505-32

F384 -33

F347-3

100 200 feet

25 50 meters

Figure 1. Plan view of the positions of the pumped well (F507-080) and observation wells and piezometers in the study area, 

Cape Cod, Massachusetts.

275
Table 1. Locations of observation piezometers in study area, Cape Cod, Massachusetts 1

[ft, feet; zD |, zD2 , and ZD normalized distances from bottom of aquifer to bottom of observation-well screen, top of observation-well screen, and middle 

of piezometer, respectively; r/b, normalized radial distance from pumped well to point of observation; b, saturated thickness]

Well 

number

F505-032

F505-059

F505-080

F504-032

F504-060

F504-080

F383-032

F383-061

F383-082

F383-129

F384-033

F381-056

F347-031

F434-060

3 F450-061

F476-061

F478-061

3 F385-032

3 F377-037

3 F376-037

radial 

distance 

(ft)

23.9

19.5

21.6

46.6

49.8

53.1

93.0

92.9

94.8

96.7

137.3

159.8

225.7

38.6

66.3

65.6

101.3

224.6

85.1

227.6

depth2 

(ft)

10.7

30.6

58.4

9.6

30.0

57.5

12.1

39.9

61.8

107.8

15.8

20.0

14.8

2.0

1.7

2.2

2.2

10.0

13.3

13.2

screen z D1 z D2 

length 

(ft)

2

9 0.75 0.81

2

2

9 .76 .81

2

2

2

2

2

2

2

2

39 .74 .99

39 .74 .99

39 .74 .99

39 .74 .99

2

2

2

ZD

0.93

.63

.93

.63

.92

.74

.61

.32

.90

.87

.90

.93

.91

.91

(r/b)

0.022

.015

.018

.085

.097

.11

.34

.34

.35

.36

.74

1.00

1.99

.058

.17

.17

.40

1.97

.28

2.02

1 Dimensionless ratios are computed assuming a saturated

thickness of 160 feet. 

2 Depth below the initial water table to the

top of the screen. 

3 Data from this well not used in the analysis.

inspection. Also shown for reference purposes is the 

Theis type curve with dimensionless time defined 

with respect to specific yield. To avoid confusion 

the type curves are plotted without displaying the 

coordinate axes. The type curves are plotted on the 

same scale as the composite time-drawdown plots 

and the match points are located at a dimensionless 

drawdown of h D=l and dimensionless time of t D =1 

(see table 2 for definitions of dimensionless parameters). Transmissivity and specific yield are calculated from the definitions of h D and tDy, and from 

the value of the match points in the figures.

The type curves were produced for each 

piezometer using the dimensionless ratios listed 

in table 1. The parameters D and dD, needed to 

specify the location of the pumped-well screen, 

were 0.38 and 0.082, respectively. Given that the 

saturated thickness is known (or assumed), the 

only parameter requiring trial-and-error adjustment 

is the ratio KZ/KP which is needed in the model to 

obtain p. A value of 10"2 was used for a in the

276
Table 2. Dimensionless expressions and notations used in report

K

r 

S 

S 

T 

t 

z

dimensionless expressions:

tDy Tt/r2Sy

nD 47rT(hi-h)/q0

0 K2 r2/Krb2

a S/Sy

dD d/b

ZD 

Z DI

ZD2

notation: 

b

d

h

z/b 

Zi/b 

z2/b

initial saturated thickness of aquifer, L

vertical distance from initial water table to top

of pumped-well screen, L

elevation of water table above base of aquifer, L

initial elevation of water table above base of

aquifer, L

hydraulic conductivity in the vertical direction, LT'1

hydraulic conductivity in the horizontal direction, LT'1

vertical distance from initial water table to bottom

of pumped-well screen, L

pumping rate, L3 T"1

radial distance from center of pumping well, L

storativity

specific yield

transmissivity, L2!1 " 1

time since start of pumping, T

vertical distance above bottom of aquifer, L

vertical distance from bottom of aquifer to bottom

of observation-well screen, L

vertical distance from bottom of aquifer to top of

observation-well screen, L

computations, but because no attempt was made to 

evaluate storativity, any value of a less than 10~2 

could have been used. The value of the ratio Kz/Kr 

that best matches all the drawdown data simultaneously was determined to be approximately 0.5. The 

"best" overall match point appears to be located at 

hr h=0.09 ft and 1/^=6x10^ min/ft2 . On the basis of 

this match point, the transmissivity is calculated to be 

38 ft2/min, the specific yield is 0.23, the horizontal 

hydraulic conductivity is 0.24 ft/min, and the vertical 

hydraulic conductivity is 0.12 ft/min.

DISCUSSION

The fact that intermediate-time and late-time 

drawdown curves fall close to the type curves for a 

single value of K^/K,. is an indication that the aquifer is 

remarkably homogeneous to flow at the scale of this 

test about 20 to 200 feet. The specific yield is about 

what is expected for clean, well-graded sands. The 

estimated horizontal hydraulic conductivity is about 

the same as the geometric mean of values reported by 

Hess and others (1992), who used a flowmeter to

277
UJ 

UJ 

b_

0.5

0.4

0.3

o

Q 0.2

0.1

0.0

piezometer

F383-129 

F383-082

.,;."

/* 

V" /

*"V>+4t

 H. 44f

+ 4 

+* . 4

, ++++

 I-M-

-I -I

10 100

TIME, IN MINUTES

1,000 1 0,000

Figure 2. Drawdown at two depths in piezometers near the pumped well (F507-080), Cape Cod, Massachusetts.

estimate the vertical distribution of horizontal hydraulic conductivity at 16 wells. The estimated ratio of 

vertical to horizontal hydraulic conductivity is not 

inconsistent with the value of 1:1.2 obtained by Hess 

and others (1992) by statistical methods. The results 

are also nearly identical to values reported by 

LeBlanc and others (1988) from an aquifer test conducted about 7,000 ft downgradient. Not all the data 

are in complete accord with the type curves. The 

responses in piezometers near the water table generally exhibit greater drawdown at early and intermediate time than expected theoretically. It is possible 

that these observations can be partially explained by 

the mechanism of delayed drainage discussed by 

Nwankwor and others (1992). The overall match 

point could be fine tuned with additional effort by 

assuming different values of saturated thickness, but 

it is unlikely that the final computed hydraulic properties would change substantially. Also, because of 

the geologic evidence in support of the chosen value

of saturated thickness, there would be little justification in altering this parameter in the model.

CONCLUSION

Preliminary analysis of drawdown data 

collected during a 72-hour aquifer test at the Cape 

Cod Toxic-Substances Hydrology Research Site 

was found to yield unambiguous estimates of 

transmissivity and specific yield. Estimates of 

vertical and horizontal hydraulic conductivity are 

consistent with estimates from previous independent studies. The remarkable agreement between 

observed drawdown and the analytical model indicates not only that the aquifer is reasonably homogeneous to flow but also that the theory adequately 

describes three-dimensional, axisymmetric flow in 

clean, well-sorted, glacial-outwash aquifers of the 

type found at the Cape Cod site.

278
UJ 

UJ

O 

Q

o:

0.1

0.02

2

...OTheis type 

curve for

piezometer:

F505-032 

F505-059 

F505-080

data 

o

A

o

type 

curves

0.1

0.02

2

piezometer:

F504-032 

F504-060 

F504-080

data 

o

A

o

type 

curves

0.1

0.02

piezometer:.

F434-060 

F476-061 

F478-061

data

o 

A 

O

type 

curves

0.001 0.01 0.1 1 10

t/r2 , IN MINUTES PER FOOT SQUARED

Figure 3. Composite plot of drawdown data and type curves for a) piezometers F505-032, F505-059, and F505-080 located 

about 20 feet from the pumped well (F507-080, b) piezometers F504-032, F504-060, and F504-080 located about 50 feet from 

the pumped well, and c) screened observation wells F434-060, F476-061, and F478-061 located, respectively, 39,66, and 

101 feet from the pumped well, Cape Cod, Massachusetts.

279
0.6

0.1

UJ 

UJ

O

0.01 

0.6

Theis type 

curve for

type 

piezometer: data curves

F383-032 

F383-061 

F383-082 

F383-129

o

A

O

D

0.1

0.01

b

type 

piezometer: data curves

-O<3000

// F384-033 o 

//Theis type F38V-056 A 

curve for F347-031 O 

toy

0.0001 0.001 0.01 0.1 1 

t/r2 , IN MINUTES PER FOOT SQUARED

Figure 4. Composite plot of drawdown data and type curves for a) piezometers F383-032, F383-061, F383-082, and F383- 

129 located about 94 feet from the pumped well (F507-080), and b) piezometers F384-033, F381-056, and F347-031 located, 

respectively, 137,160, and 226 feet from the pumped well, Cape Cod, Massachusetts.

280
REFERENCES

Hess, K.M, Wolf S.H., and Celia M.A., 1992, Large-scale 

natural gradient tracer test in sand and gravel, Cape 

Cod, Massachusetts, 3 Hydraulic conductivity 

variability and calculated macrodispersivities: Water 

Resources Research, v. 28, no. 8, p. 2011-2027.

LeBlanc, D.R., 1984, Sewage plume in a sand and gravel 

aquifer, Cape Cod, Massachusetts: U.S. Geological 

Survey, Water-Supply Paper 2218, 28 p.

LeBlanc, D.R., Garabedian, S.P., Quadri, R.D., Morin, 

R.H., Teasdale, W.E., and Paillet, F.L., 1988, Hydrologic controls on solute transport in a plume of 

sewage-contaminated ground water, in Ragone, S.E., 

ed., U.S. Geological Survey Program on Toxic 

Waste Ground-Water Contamination Proceedings 

of the second technical meeting, Cape Cod, Mass., 

October 21-25, 1985: U.S. Geological Survey OpenFile Report 86-481, p. B7-B12.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W., 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 Experimental design and observed tracer movement: Water 

Resources Research, v. 27, no. 5, p. 895-910.

Moench, A.F., 1993, Computation of type curves for flow to 

partially penetrating wells in water-table aquifers: 

Ground Water, v. 31, no. 6, p. 966-971.

Neuman, S.P., 1974, Effect of partial penetration on flow in 

unconfined aquifers considering delayed gravity 

response: Water Resources Research, v. 10, no. 2, 

p. 303-312.

Nwankwor, G. I., Gillham, R.W., van der Kamp, G., and 

Akindunni, F.F., 1992, Unsaturated and saturated flow 

in response to pumping of an unconfined aquifer 

Field evidence of delayed drainage: Ground Water, 

v. 30, no. 5, p. 690-700.

281
282
Numerical Simulation of Downward

Movement of Solutes During a Natural-Gradient Tracer

Test in Sand and Gravel, Cape Cod, Massachusetts

By Denis R. LeBlanc1 and Michael A. Celia2

Abstract

A numerical, finite-element, solute-transport 

model was used to test the hypothesis that the 

downward movement of a tracer cloud observed 

during a natural-gradient tracer test in sand and 

gravel on Cape Cod, Mass., was caused, in part, by 

density-induced sinking. The tracer solution, 

which included the nonreactive tracer, bromide, 

was 0.1 percent denser than the ambient ground 

water. The center of mass of the bromide cloud 

moved vertically downward about 3.2 meters 

during 237 days of transport. The model simulated 

density-dependent flow and solute transport along a 

two-dimensional vertical section 25 meters high 

and 136 meters long aligned with the direction of 

ground-water flow. Transport of the bromide cloud 

was simulated for a period of 237 days divided into 

191 time steps. The temporal pattern of recharge 

applied to the top boundary of the model was 

determined from daily precipitation and estimates 

of evapotranspiration. On the basis of an analysis 

of spreading of the bromide cloud during the 

tracer test, dispersivity was increased with time in 

the simulation asymptotically from 0.05 to 

0.96 meters. The simulated downward movement 

after 237 days was 2.1 meters (about two thirds of 

the observed movement). The simulation showed 

that density-induced downward movement was 

most important during the first 37 days of

'U.S. Geological Survey, Marlborough, Mass. 

2 Dept. of Civil Engineering and Operations Research, 

Princeton Univ., Princeton, N.J.

transport when the density difference between the 

ambient ground water and the tracer cloud was 

greatest. Earlier work indicated that the 

difference between the observed and simulated 

downward movement may be due, in part, to 

simulation in only two dimensions of the threedimensional flow that occurred around the tracer 

cloud as it moved downward through the ambient 

ground water. Additional simulations during this 

and earlier studies showed how the amount of 

downward movement is affected by the size and 

shape of the initial bromide cloud; aquifer 

properties, such as dispersivity and anisotropy of 

hydraulic conductivity; and the type of boundary 

used to represent the water table in the model.

INTRODUCTION

A natural-gradient tracer test was conducted in 

1985-88 in a sand and gravel aquifer on Cape Cod, 

Mass. (fig. 1), to measure the dispersion of solutes and 

relate the dispersion to the spatial variability of 

hydraulic conductivity (LeBlanc and others, 1991). 

The test began in July 1985 with the injection of 7.6 

cubic meters of tracer solution into the unconfined 

aquifer. The solution contained a nonreactive tracer, 

bromide, and two reactive tracers, lithium and 

molybdate. Movement of the tracers under a natural 

hydraulic gradient was observed by periodic collection 

of water samples from a three-dimensional array of 

9,840 sampling points.

The location of the center of mass of the 

bromide cloud was computed for each of 16 sampling 

rounds by the method of spatial moments (Garabedian

283
71°00'

73°00 72 C 00-

42°30'-/L

42 00'

MASSACHUSETTS

20 40 60 MILES 41

0 20 40 60 KILOMETERS

Tracer-test site

Figure 1. Location of Cape Cod natural-gradient tracer-test 

site.

and others, 1991). The vertical position of the center 

of mass (one component of the first moment) showed 

an initial period during which the tracer cloud moved 

downward. The downward movement during the first 

237 days of the test (fig. 2) was 3.2 m (meters), which 

is about 70 percent of the total downward movement 

observed during the entire 51 1-day experiment.

LeBlanc and others (1991) noted that the 

vertical flow caused by accretion of areal recharge 

could account for only part of the initial downward 

movement of the tracer cloud. Ground-water flow is 

nearly horizontal in the aquifer, and the estimated 

recharge is insufficient to cause all the observed drop. 

They hypothesized that the additional downward 

movement was caused by the density difference 

between the ambient ground water and the tracer 

solution. The estimated density of the tracer solution 

was 1.00037 g/cm3 (grams per cubic centimeter), 

whereas the estimated density of the ambient, 

uncontaminated ground water was 0.999409 g/cm .

Although the tracer solution was only about 

0. 1 percent denser than the ambient ground water, 

LeBlanc and Celia (1991) used analytical and 

idealized numerical models to demonstrate that the 

density contrast may have been sufficient to cause at 

least part of the observed downward movement. The 

preliminary modeling also suggested that the amount 

of downward movement would be affected by the 

magnitude of the density difference, the dispersivity, 

and the size of the initial tracer cloud. The purpose of 

this paper is to describe the use of a numerical model 

of the field-scale tracer test to test the hypothesis that 

density caused a significant part of the downward 

movement observed during the experiment.

NUMERICAL SIMULATION OF THE 

TRACER TEST

SUTRA (Voss, 1984), a two-dimensional finiteelement model that simulates solute transport and 

dispersion in a density-dependent flow field, was used 

to simulate the first 237 days of the tracer test, when 

the effect of the density difference on downward 

movement of solutes was expected to be greatest. The 

numerical simulation described below used a 

rectangular grid and assumed that solute concentration 

and density were linearly related. Also, the 

dispersivity components were assumed to be 

independent of the direction of flow (Voss, 1984, 

p. 48-50).

Description of Simulation

The grid was a two-dimensional vertical section 

encompassing an area 25 m high, the saturated thickness of the sand and gravel (LeBlanc and others, 

1991), and 136 m long, the approximate distance 

traveled by the leading edge of the tracer cloud in 

237 days (Garabedian and others, 1991). This area was 

subdivided into 8,004 rectangular elements arranged in 

69 rows and 116 columns. Garabedian and others

so 100 150 

DAYS

200 250

RECHARGE, DAILY IN

CENTIMETERS 

A 3ro o>

lljl.ll ,

B

Ih. . . i

50 100 150 200 250

DAYS

Figure 2. Simulated and observed vertical position of the 

tracer cloud (A) and daily recharge applied to the top 

boundary of the model (B) during the first 237 days of the 

tracer test.

284
(1991) observed that longitudinal dispersivity 

increased asymptotically during the test to a value of 

0.96 m, whereas transverse vertical dispersivity was 

less than 0.002 m. Based on these observations, the 

longitudinal dispersivity used in the model was 

increased asymptotically from 0.05 to 0.96 m and 

transverse dispersivity was held constant at 0.005 m as 

the simulated cloud moved through the aquifer. To 

keep the Peclet number less than 4 and the transverse 

spacing less than 10 times the transverse dispersivity 

in the zone through which the solute cloud passed 

(Voss, 1984, p. 232-233), the length of the elements 

increased from 0.2 to 3.5 m in the direction of flow, 

and the height of the elements increased from 0.15 m 

in the area through which the cloud passed to 6.0 m 

below the transport zone.

Aquifer characteristics were similar to those 

reported in LeBlanc and others (1991) and Hess and 

others (1992). An ambient, steady, horizontal velocity 

of 0.42 m/d (meters per day) was established by 

specifying a hydraulic conductivity of 110 m/d, an 

effective porosity of 0.39, an appropriate fluid flux 

along the left boundary of the model based on an 

estimate of flow from areas upgradient from the 

modeled area, and appropriate pressures along the 

downgradient (right side) of the model. A value of 

0.83 was used for the ratio of vertical to horizontal 

hydraulic conductivity (Hess and others, 1992).

The top and bottom of the model were specified 

as zero-flux boundaries. About 60 cm of recharge 

from precipitation was added intermittently during the 

simulation to the top row of elements (fig. 2). The 

amount and temporal distribution of recharge were 

determined from a water balance calculated by the 

method of Thornthwaite and Mather (1957) from 

estimates of evapotranspiration rates and 

measurements of precipitation at a weather station 1.8 

kilometers from the site.

The simulation began with the instantaneous 

release of a solute cloud that was 1.8 m high and 3.4 m 

long. The center of the cloud was initially located 5.6 

m from the left boundary of the model and 1.3 m from 

the top boundary. The initial tracer cloud had a 

dimensionless relative concentration (concentation in 

the ground water divided by the concentration in the 

initial tracer solution) of 1.0 and a fluid density of 

1.00037 g/cm3 , whereas the ambient ground water, 

inflow along the left boundary, and recharge from 

precipitation had a relative concentration of 0.0 and a 

fluid density of 0.999409 g/cm3 . The solute cloud was 

simulated for 237 days with 191 time steps that

increased during the simulation period from 0.2 to 

4.5 days. Small time steps were used early in the 

simulation to minimize numerical dispersion (Voss, 

1984, p. 234). The spatially constant longitudinal 

dispersivity was increased in a stepwise fashion from 

0.05 to 0.96 m by stopping the simulation at 

appropriate times, entering the new value of 

dispersivity, and continuing the simulation.

Spatial moments of the simulated concentration 

distributions were calculated for each time step by 

numerical integration of concentrations using bilinear 

interpolation across the model elements. The 

moments of the simulated distributions were then 

compared to the moments of the observed 

concentrations reported by Garabedian and others 

(1991). In this study, the first moment in the vertical 

direction is used to show the vertical position, or the 

altitude, of the center of mass of the simulated and 

observed tracer clouds.

Simulated Movement of the Tracer Cloud

The simulated downward movement of the 

tracer cloud is compared to the observed downward 

movement in figure 2. The simulated downward 

movement after 237 days was 2.1 m, or about two 

thirds of the observed movement. About 40 percent 

(0.9 m) of the simulated downward movement 

occurred during the first 37 days of the simulation. 

Because there was no precipitation during this period, 

no recharge was added to the model and all of the 

downward movement was caused by the density 

difference between the tracer cloud and the ambient 

ground water. The rate of downward movement was 

greatest early in the test, when the density difference 

was largest, and decreased with time as dilution of the 

tracer cloud by dispersion decreased the density 

difference.

After 37 days of transport, the additional 

downward movement caused by the density difference 

was small compared to the downward movement 

caused by recharge from precipitation. The simulated 

downward movement between 37 and 237 days (1.2 

m) is consistent with the total recharge applied to the 

top row of the model (0.45 m) during this period and a 

porosity of 0.39 (LeBlanc and others, 1991). The 

downward movement because of recharge is most 

evident between 37 and 45 days after the start of the 

tracer test, when recharge from two intense storms 

caused the simulated center of mass to move 

downward about 0.4 m.

285
DISCUSSION

The difference between the observed (3.2 m) 

and simulated (2.1 m) downward movement may be 

due, in part, to simulation of the tracer test in two 

dimensions. LeBlanc and Celia (1991) used an 

analytical model (Yih, 1965) to show that densityinduced downward movement predicted by twodimensional models may be about 25 percent smaller 

than that obtained from three-dimensional models. A 

two-dimensional model constrains movement of water 

to a two-dimensional plane, and the solute body 

cannot sink as rapidly as it would if the ambient 

ground water were free to move in any direction.

The difference between the observed and 

simulated downward movement may also be due, in 

part, to inaccurate specifications of aquifer properties 

and tracer-test conditions. LeBlanc and Celia (1991) 

demonstrated that the downward movement is very 

sensitive to the rate at which concentrations in the 

tracer cloud are diluted by dispersion. If the 

asymptotic value of longitudinal dispersivity (0.96 m) 

reported by Garabedian and others (1991) had been 

used for the entire simulation period, dispersion would 

have quickly decreased maximum concentrations and 

limited the amount of vertical movement. A 

dispersivity that increases from an initially small value 

to the asymptotic value is needed to represent 

accurately the relation between density-induced 

sinking and dilution by dispersion. LeBlanc and Celia 

(1991) also showed that the rate of downward 

movement is sensitive to the anisotropy of hydraulic 

conductivity. The amount of downward movement 

decreases as the aquifer becomes more permeable in 

the horizontal direction than in the vertical direction 

because flow occurs preferentially in the direction of 

highest permeability.

The size and shape of the initial tracer cloud 

also affect the amount of downward movement. 

LeBlanc and Celia (1991) showed that high 

concentrations persist longer in larger clouds than in 

smaller clouds; as a result, the centers of mass of 

larger clouds tend to sink farther than those of small 

clouds with the same initial density. Several 

preliminary simulations during this study 

demonstrated, however, that clouds that are short 

horizontally tend to sink farther than clouds that are 

long horizontally because the ambient ground water 

moves more easily around the shorter clouds as the 

clouds move downward because of the density 

difference.

The density difference directly affects the rate of 

downward movement. The initial densities of the 

tracer cloud and the ambient ground water were 

estimated from the concentrations of dissolved solids 

in the uncontaminated ground water and the tracer 

solution. The effects of sewage contamination 

beneath the site and chemical reactions involving 

reactive solutes in the tracer cloud (LeBlanc and 

others, 1991) on the density difference were not 

considered in this analysis.

Several preliminary simulations during this 

study demonstrated that the specification of model 

boundaries also affects the amount of downward 

movement. The rate of downward movement 

increases when the top boundary of the model is 

represented as a specified-pressure boundary rather 

than as a zero-flux boundary. As the tracer cloud 

sinks, the ambient water is replaced more easily by 

flow across the specified-pressure boundary than by 

flow up and around the sinking cloud when a zero-flux 

boundary is used. The proper representation of the 

water table as a specified-pressure or zero-flux 

boundary needs further examination.

SUMMARY

A numerical, finite-difference, solute-transport 

model was used to test the hypothesis that the 

downward movement of a tracer cloud observed 

during a natural-gradient tracer test in sand and gravel 

on Cape Cod, Mass., was caused by density-induced 

sinking. The model included an asymptotically 

increasing dispersivity, an anisotropic hydraulic 

conductivity, and intermittent recharge from 

precipitation.

The simulated downward movement after 237 

days was 2.1 meters, which is about two thirds of the 

observed downward movement (3.2 m). The densityinduced downward movement was most important 

during the first 37 days of transport when the density 

constrast between the ambient ground water and the 

tracer cloud was greatest.

The difference between the simulated and 

observed downward movement may be due, in part, to 

the simulation of the test with a two-dimensional 

model, which was shown by LeBlanc and Celia (1991) 

to predict less downward movement than a threedimensional model. The difference between the 

observed and simulated downward movement may 

also be due, in part, to inaccurate specifications of

286
aquifer properties, such as dispersivity and anisotropy 

of hydraulic conductivity; and tracer-test conditions, 

such as the size and shape of the initial tracer cloud 

and the type of boundary used to represent the top of 

the model.

REFERENCES

Garabedian, S.P., LeBlanc, D.R., Gelhar, L.W., and Celia, 

M.A., 1991, Large-scale natural-gradient tracer test in 

sand and gravel, Cape Cod, Massachusetts, 2 

Analysis of spatial moments for a nonreactive tracer: 

Water Resources Research, v. 27, no. 5, p. 911-924.

Hess, K.M., Wolf, S.H., and Celia, M.A., 1992, Large-scale 

natural gradient tracer test in sand and gravel, Cape 

Cod, Massachusetts, 3 Hydraulic-conductivity 

variability and calculated macrodispersivitics: Water 

Resources Research, v. 28, no. 8, p. 2011-2027.

LeBlanc, D.R., and Celia, M.A., 1991, Density-induced 

downward movement of solutes during a naturalgradient tracer test, Cape Cod, Massachusetts, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 10-14.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gclhar, L.W.. 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural-gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, no. 5, p. 895-910.

Thornthwaite, C.W., and Mathcr, J.R., 1957, Instructions 

and tables for computing potential evapotranspiration 

and the water balance: Ccntcrton, N.J., Drexel 

Institute of Technology, Publications in Climatology, 

v. 10, no. 3, 311 p.

Voss, C.I., 1984, A finite-clement model for saturatcdunsaturated, fluid-dcnsity-dcpendent ground-water 

flow with energy transport or chemically reactive 

single-species solute transport: U.S. Geological 

Survey Water-Resources Investigations Report 84- 

4369, 409 p.

Yih, Chia-Shun, 1965, Dynamics of nonhomogcneous 

fluids: New York, MacMillan, 301 p.

287
288
Effects of Changes in Aquifer Properties on Simulated 

Fluid-Particle Pathlines, Cape Cod, Massachusetts

By John P. Masterson1 and Donald A. Walter1

Abstract

A ground-water-flow model is being 

developed by the U.S. Geological Survey (USGS) 

to estimate the paths of contaminants from the 

Massachusetts Military Reservation and, in particular, to determine whether these contaminants 

will affect public-supply wells, ponds, streams, 

and coastal embayments.

Preliminary model results determined by 

use of the regional-scale flow model of western 

Cape Cod developed by the USGS indicate that 

the pathlines of contaminants emanating from the 

infiltration beds of the Massachusetts Military 

Reservation sewage-treatment plant do not coincide with the known extent of the contaminant 

plume. Modifications to aquifer properties simulated by the model result in a similar head distribution and stream-discharge rates for the aquifer, 

yet markedly different contaminant pathlines. 

Therefore, calibration simply based on head and 

stream-discharge matching is insufficient for 

transport analysis. Accurate calibration might 

require the incorporation of contaminant information to use ground-water-flow models as tools for 

predicting plume migration.

INTRODUCTION

Recent investigations at the Massachusetts 

Military Reservation (MMR) indicate that contaminated ground water is migrating from the reservation 

from suspected contamination sites (ABB Environmental Services Inc., 1992). The National Guard

'U.S. Geological Survey, Marlborough, Mass.

Bureau Installation Restoration Program began investigating potential problems related to suspected past 

releases of toxic and hazardous material at the MMR 

in 1986. Several questions remain unanswered concerning ground-water supply and contamination near 

the MMR. These questions include (1) what is the 

predicted path of contaminants from sources on the 

base; (2) what is the source of water to specific wells, 

ponds or streams; and (3) how will proposed activities, 

such as water-supply development or ground-water 

withdrawals for remediation affect ground-water flow 

and associated contaminant migration?

The answers to these questions require an 

understanding of the complex interaction of factors 

affecting ground-water flow, including recharge, 

ground-water withdrawals, aquifer structure and properties, and hydrologic boundaries, such as ponds and 

streams. Ground-water-flow models that can integrate 

the effect of these factors are valuable tools for 

addressing these questions.

The USGS, in cooperation with the National 

Guard Bureau, is developing a three-dimensional 

ground-water-flow model using MODFLOW 

(McDonald and Harbaugh, 1988) with a particle-tracking routine, MODPATH (Pollock, 1989), to investigate 

ground-water flow and advective transport near the 

MMR. In a recently completed investigation, Masterson and Barlow (1994) developed a regional-scale 

flow model that encompassed the flow cell of the western part of Cape Cod (including the MMR) to determine the effects of future ground-water withdrawals 

on ground-water flow. This regional flow model provides an opportunity to examine the effects of changes 

in aquifer properties on particle pathlines near the 

MMR and the boundary conditions necessary for the 

development of the more refined MMR model. This 

paper presents preliminary results of the effects of 

changes in aquifer properties on particle pathlines.

289
HYDROGEOLOGIC SETTING

The location of the study area is shown in figure 

1. The area of interest is located on a broad southward-sloping glacial outwash plain known as the 

Mashpee Pitted Plain. The outwash plain is bounded 

by moraine to the north and west, by an adjacent outwash plain to the east, and by Vineyard Sound to the 

south. There are numerous collapse structures and 

kettle ponds within the outwash plain. The outwash 

consists of glacial silts, sands and gravels. The sediments were deposited during the Pleistocene epoch as 

part of a delta deposited into a large, proglacial lake 

that formed to the south of the retreating Laurentide 

ice sheet. These glacial deposits can be divided into 

top-set, fore-set, and bottom-set deposits (Byron 

Stone, U.S. Geological Survey, oral commun., 1993). 

Top-set deposits consist of glaciofluvial outwash 

deposits of coarse sand and gravel. The underlying 

fore-set deposits are glaciolacustrine sediments that 

consist primarily of medium to fine sand. Bottom-set 

deposits are glaciolacustrine deposits consisting of 

fine sand and silt. Grain size of all three depositional 

units decreases to the south with increasing distance 

from the sediment source. The deltaic sediments are 

underlain locally by glacial till consisting primarily of 

silt and clay. A hydrogeologic section through the glacial sediments is shown in figure 2. The upper 70 to 

200 feet of sediment are distal top-set and proximal 

fore-set deposits that consist of stratified medium to 

coarse sand and gravel. Underlying these deposits are 

bottom-set deposits consisting of fine sand and silt. 

Thickness of the bottom-set deposits ranges from 125 

to 200 feet in the study area; these deposits are underlain by glacial till. The unconsolidated sediments are 

underlain by granodiorite bedrock.

The flow system of western Cape Cod is 

bounded by saltwater to the north, south, and west, 

and by a regional ground-water divide to the east. Precipitation is the sole source of recharge to the groundwater system. The ground-water system is bounded at 

depth by bedrock, which is assumed to have a much 

lower permeability than the overlying sediments. 

Ground-water discharges primarily to streams and 

coastal saltwater embayments. The flow system is 

characterized by radial ground-water flow from a 

water-table mound to the north of the study area that 

has a maximum head of approximately 70 ft above sea 

level.

The direction of regional ground-water flow is 

approximately southward in the study area. Groundwater flow is locally influenced by kettle ponds that 

create a flow-through condition in which shallow 

ground-water discharges to the upgradient ends of the 

ponds and recharges the aquifer at the downgradient 

ends of the ponds. Ground water deeper in the flow 

system flows beneath the ponds. Some ponds, such as 

Johns and Coonamessett Ponds, have outlets that 

allow for limited surface outflow. The streams located 

in western Cape Cod are primarily gaining streams 

that receive their water from kettle-pond outflow and 

ground-water discharge.

Estimated hydraulic conductivities for the 

sand and gravel deposits typically range from 125 to 

250 ft/d with values as high as 380 ft/d. Horizontal 

hydraulic conductivities of the fine sands and silts 

range from 30 to 125 ft/d (Masterson and Barlow, 

1994).

METHODS AND RESULTS

The original three-dimensional steady-state 

ground-water-flow model developed for western Cape 

Cod (Masterson and Barlow, 1994) consists of five 

layers that extend from the water table to the bedrock 

surface with a horizontal grid spacing of 1,320 by 

1,320 ft. A particle-tracking routine, MODPATH 

(Pollock, 1989), was coupled with this flow model to 

determine the effects of changes in aquifer properties 

on one of the plumes emanating from the MMR. This 

plume (Ashumet Valley plume) is the result of secondary-treated effluent discharged to the MMR sewagetreatment-plant infiltration beds since 1936. The 

plume extends approximately 12,500 ft south of the 

reservation boundary and is approximately 3,500 ft 

wide and 80 ft thick (ABB Environmental Services 

Inc., 1992).

In the model analysis, 25 particles were started 

at the water table in each of the two nodes that contain 

the sewage-treatment-plant disposal beds. Fifty particles were used because they were found to delineate 

pathlines emanating from the infiltration beds adequately.

Particle pathlines from the original model 

following the known path of the plume near the source 

are shown in figure 3; the eastern-most particles 

discharge to Ashumet Pond. However, away from the 

source, pathlines bend to the southeast, beneath the 

Childs River and Johns Pond and discharge to the 

Quashnet River. This configuration does not coincide 

with the known extent of the plume in 1988 (fig. 3a)

290
Figure 1. Study area on southwestern Cape Cod, Massachusetts, showing location of section A-A'.

291
FEET 

300

200

-400

2.000 4.000 6,000 8,000 METERS 

VERTICAL GREATLY EXAGGERATED

EXPLANATION

MEDIUM-COARSE SAND AND GRAVEL _Y_ 

FINE SAND AND SILT 5

- WATER TABLE

CLAY, TILL AND OTHER FINE-GRAINED 

SEDIMENTS

BEDROCK

WELL AND IDENTIFIER

1

Figure 2. Hydrogeologic section A-A' showing location of observation wells in the study area, Cape Cod, Massachusetts.

(ABB Environmental Services Inc., 1992). In crosssection (fig. 4a), particle pathlines are shown to 

descend to the lowest active layer in the flow system 

(240 ft below sea level) before discharging upward to 

the Quashnet River. The known plume extent in vertical profile has a maximum depth of approximately 

80 to 100 ft below sea level

A close examination of the local geology of the 

study area indicates the presence of a contact between 

coarse-grained deltaic fore-set beds and the fine sand 

and silt bottom-set beds at about 70 ft below sea level 

(the boundary between layers 2 and 3) (fig. 2); this 

contact was not adequately simulated in the regional

model and, thus, the model had to be adjusted. To 

account for this contact between coarse- and finegrained sediments, the vertical conductance between 

model layers 2 and 3 was decreased from 0.002 to 

0.0002, and the horizontal hydraulic conductivity was 

increased in layers 1 and 2 from 200 to 350 ft/d, and 

decreased in layers 3 to 5 from 75 to 30 ft/d. These 

adjusted values are within the ranges of those reported 

by LeBlanc and others (1988), Barlow and Hess 

(1993), and Masterson and Barlow (1994), for the 

study area. As a result of these adjustments, most of 

the ground-water flow is confined to the upper 100 ft 

of saturated material.

292
7035' 7030'

7035' 

40

37' 

30"

40 

3T 

30"

41

37' 

30"

41 

32' 

30"

Figure 3. Simulated head contours, particle pathlines and known extent of Ashumet Valley plume in 1988 (a) original model res
A second simulation was then made using the 

adjusted model with the same particle- starting locations. The particle pathlines of the adjusted-model 

simulation (fig. 3b) coincide more closely with the 

extent of the observed plume. The vertical profile 

shown in figure 4b illustrates that the depth of the 

pathlines coincides with the depth of the contact 

between the coarse and fine sediments shown in figure 

2 and is consistent with the known vertical extent of 

the plume.

Despite significant adjustments made to the 

aquifer properties of the calibrated flow model, the 

error (root mean squared error/ maximum drop in head 

within the study area) in calculated heads in the 

adjusted model was 6 percent as compared to an error 

in head of 3 percent in the original flow model. In 

addition, calculated stream discharge rates in the 

adjusted model were similar to those of the original 

model. The two models produced very different particle pathlines, although both models would be considered calibrated on the basis of head distribution and 

stream-discharge rates.

SUMMARY AND CONCLUSIONS

Preliminary simulations using an existing threedimensional regional ground-water-flow model 

improve our understanding of the ground-water flow 

system near the MMR and will aid in the development 

of a fine-scaled MMR flow model. The effects of 

adjustments to model-calibrated aquifer properties on 

simulated particle pathlines illustrates how two 

ground-water-flow models with substantially different 

model input data sets can produce similar head distributions and stream-discharge rates, yet very different 

particle pathlines. This observation can be attributed 

to model calibration that is commonly based solely on 

water levels and stream discharge in the absence of 

data on ground-water velocities or extents of known 

plumes. Model calibration based simply on head and 

stream-discharge matching is insufficient for purposes 

of transport analysis and might require the incorporation of water-quality data to use ground-water-flow 

models as tools for predicting plume migration.

294
B

B1

B

4a

EXPLANATION

ACTIVE MODEL AREA

COLUMN 36

4b

B'

6000 FEET 

i

j ) HEAD-DEPENDENT FLUX BOUNDARY 

INACTIVE MODEL AREA

0 1000 METERS 

VERTICAL SCALE IS GREATLY EXAGGERATi O

 ) PARTICLE PATHLINE

Figure 4. Particle pathlines projected along column 36 of the model determined by (a) simulation of original model and (b) 

simulation of adjusted model. Line of section shown in figure 3.

295
REFERENCES

ABB Environmental Services Inc., 1992, Ground-Water 

Remediation Strategy Report; Installation Restoration 

Program: Portland, Maine, 115 p.

Barlow, P.M., and Hess, K.M., 1993, Simulated hydrologic 

responses of the Quashnet River stream-aquifer system to proposed ground-water withdrawals, Cape Cod, 

Massachusetts: U.S. Geological Survey WaterResources Investigations Report 93-4064, 53 p.

LeBlanc, D.R., Garabedian, S.P., Quadri, R.D., Morin, 

R.H., Teasdale, WE., and Paillet, F.L., 1988, Hydrogeologic controls on solute transport in a plume of 

sewage- contaminated ground water, in Ragone, S.E., 

ed., U.S. Geological Survey Program on Toxic 

Waste Ground-water Contamination Proceedings 

of second technical meeting, Cape Cod, Massachusetts, October 21-25, 1985: U.S. Geological Survey 

Open-File Report 86-481, p. B-7 to B-12.

Masterson, J.P., and Barlow, P.M., 1994, Effects of 

simulated ground-water pumping and recharge on 

ground-water flow in Cape Cod, Martha's Vineyard, 

and Nantucket Island basins, Massachusetts: U.S. 

Geological Survey Open-File Report 94-316, 78 p.

McDonald, M.G., and Harbaugh, A.W., 1988, A modular 

three-dimensional finite-difference ground-water 

flow model: U.S. Geological Survey Techniques of 

Water-Resources Investigations, book 6, chap. Al, 

586 p.

Pollock, D.W., 1989, Documentation of computer

programs to compute and display pathlines using 

results from the U.S. Geological Survey modular 

three-dimensional finitedifference ground-water flow model: U.S. Geological Survey Open-File Report 89-381, 188 p.

296
Overview of Research Activities on the Transport and

Fate of Chlorinated Solvents in Ground Water at

Picatinny Arsenal, New Jersey, 1991-93

By Thomas E. Imbrigiotta1 and Mary Martin 1

Abstract

The U.S. Geological Survey is conducting 

an interdisciplinary research study of ground-water 

contamination by chlorinated solvents, particularly 

trichloroethylene (TCE), at Picatinny Arsenal in 

north-central New Jersey. This paper summarizes 

the results of ongoing research studies investigating the processes of desorption, volatilization, and 

biotransformation and their effects on the fate and 

transport of TCE in the ground-water system.

Results of flow-through column experiments 

showed that contaminated aquifer sediments can 

act as a continuing source of TCE to the ground 

water. A one-dimensional model was developed to 

simulate the column desorption results. The model 

simulates an initial, fast-stage desorption by an 

equilibrium process and a second, slower-stage 

desorption by a kinetic mechanism. Concentrations of TCE sorbed to soil did not differ significantly whether the samples were air dried 

overnight prior to methanol extraction or whether 

the extraction was conducted on wet samples and 

the results were corrected for the TCE content of 

the soil moisture.

Results of a field experiment to study the 

dynamics of TCE volatilization in the unsaturated 

zone during infiltration indicated that an equilbrium distribution of TCE between soil gas and 

soil water was not achieved. A two-phase 

transport model of gas and aqueous phases was 

capable of simulating the field soil-gas and soilwater TCE concentrations after modification to

'U.S. Geological Survey, 810 Bear Tavern Rd., Suite 206, 

W. Trenton, NJ 08628

include a constant-flux term for desorption of TCE 

from soil to water.

Anaerobic biotransformation rates calculated on the basis of measured field TCE concentrations and estimated ground-water travel time 

between sites generally were greater than those 

previously measured in laboratory soil microcosm 

experiments.

A reactive two-dimensional multispecies 

transport model is being used to simulate desorption, volatilization, and microbial degradation of 

TCE along the central axis of the plume by using 

rates estimated from results of other studies at the 

site. The formation and transport of TCE degradation products cis-l,2-dichloroethylene and vinyl 

chloride also are simulated.

Aerobic cometabolic biotransformation of 

TCE and cis-1,2-dichloroethylene can be stimulated in soil microcosms constructed with soils 

from the unsaturated zone near Building 24 at the 

arsenal if the indigenous methanotrophic bacteria 

are supplied with appropriate amounts of oxygen, 

methane, and nutrients.

Preliminary results of a study to determine 

whether surfactants can enhance the removal of 

TCE from aquifer sediments during pump-andtreat remediation indicate that addition of the nonionic surfactant Triton-X 100 is effective in artificially increasing the rate of mass transfer of TCE 

from soil to the aqueous phase.

INTRODUCTION

The U.S. Geological Survey is conducting an 

interdisciplinary research study of ground-water 

contamination by chlorinated solvents and other

297
contaminants at Picatinny Arsenal in north-central 

New Jersey. The objectives of the study are to (1) 

identify and quantify the chemical, physical, and biological processes that control the movement and fate 

of these contaminants, particularly trichloroethylene 

(TCE), in the subsurface; (2) determine the relative 

importance of these processes in transporting TCE 

through the ground-water system; and (3) develop predictive models of contaminant transport.

From 1960 to 1981, the wastewater-treatment 

system in Building 24, which housed a metal-plating 

facility, discharged wastewater daily into two 8-ftdeep, sand-bottomed settling lagoons behind the 

building (fig. 1) (Benioff and others, 1990). The 

wastewater contained trace metals, other inorganic 

ions

used in plating solutions, and degreasing solvents 

(Imbrigiotta and Martin, 1991). From 1973 to 1985, 

solvent vapors from a degreasing unit were allowed to 

condense in an improperly installed overflow pipe and 

discharged into a 4-ft-deep dry well in front of Building 24. TCE was the degreasing solvent used from 

1960 to 1983. The infiltration of wastewater from the 

lagoons and chlorinated solvents from the dry well has 

created a plume of contaminated ground water downgradient from Building 24.

This paper (1) briefly describes the hydrogeology and current ground-water contamination at 

the Building 24 research site at Picatinny Arsenal, 

(2) presents and discusses a preliminary solutetransport mass balance for dissolved TCE in the 

shallow aquifer at the arsenal, and (3) summarizes the 

significant findings of the ongoing research studies at 

the Building 24 research site at the arsenal.

HYDROGEOLOGY AND GROUND-WATER 

CONTAMINATION

Picatinny Arsenal is located in a glaciated 

valley. The contamination plume at the Building 24 

site is in a 50- to 70-ft-thick unconfined aquifer 

consisting primarily of coarse to fine sand with some 

gravel and some discontinuous silt and clay layers. 

On the basis of results of aquifer-test analysis and the 

calibration of a multilayered ground-water-flow 

model, the horizontal hydraulic conductivity of the 

unconfined aquifer is estimated to be 50 to 360 ft/d 

and the estimated ratio of horizontal to vertical 

hydraulic conductivity is 100 to 1 (L.M. Voronin, U.S. 

Geological Survey, written commun., 1993).

Long-term water-table altitudes average about 

696 ft above sea level at Building 24, the contaminant 

source area, and about 686 ft above sea level at Green 

Pond Brook, the natural ground-water discharge point 

for the site. The general flow pattern in the unconfined 

aquifer is south-southeast from the edge of the glacial 

sediments to Green Pond Brook, with a slight downvalley component. Within the unconfined aquifer, 

flow generally is horizontal, with some downward 

flow near Building 24 and upward flow near Green 

Pond Brook. Estimated ground-water-flow velocities, 

based on calibrated flow-model hydraulic conductivities and measured head gradients, are about 1 to 3 ft/d 

in the plume area.

Ground-water contamination measured in the 

unconfined aquifer in 1987 and 1989 has been 

described previously by Sargent and others (1990) and 

Imbrigiotta and others (1991). Results of these studies 

showed that TCE was the most widespread organic 

contaminant in the system. Results of analyses of 

water samples collected from 53 wells in October and 

November 1991 confirmed that the areal extent of the 

TCE contaminant plume had changed little since the 

1987 synoptic sampling (fig. la). The plume extends 

1,640 ft from Building 24 to Green Pond Brook, 

where it is about 1,000 ft wide. The plume area in 

which TCE concentrations are greater than 10 (ig/L is 

estimated to cover 1.4 x 106 ft2 . The vertical distribution of TCE along the central axis of the plume 

(fig. Ib) indicates that the highest concentrations 

(> 10,000 jig/L) still are found near the base of the 

unconfined aquifer midway between Building 24 and 

Green Pond Brook. TCE concentrations greater than 

1,000 jig/L are present immediately downgradient 

from the source area. In 1991, TCE concentrations in 

water samples from most wells at the site are similar 

or slightly lower than those found previously.

The total estimated mass of dissolved TCE was 

calculated from results of six sets of synoptic waterquality analyses of samples collected during 1987-91. 

The estimated mass of dissolved TCE within the 

plume below the water table is about 1,000 kg. This 

mass is equal to about 660 L of pure TCE. The estimate of the mass of dissolved TCE within the plume 

appears to depend on the number of samples in which 

TCE concentrations exceeded 10,000 |j,g/L and the 

volume of ground water each of these samples is 

assumed to represent. As much as 60 to 70 percent of 

the total mass of dissolved TCE in the plume was 

estimated to be associated with wells in which TCE 

concentrations exceeded 10,000 jig/L.

298
A. Areal extent of trichloroethylene 

contaminant plume

EXPLANATION

p|j!j Area in which trichloroethylene concentration 

liiilil exceeds 10 micrograms per liter

10 - - LINE OF EQUAL TRICHLOROETHYLENE 

CONCENTRATION-Shows trichloroethylene 

concentration, in micrograms per liter. Dashed 

where approximate

Line of section 

Sampling site location

Bear Swamp Brook Dry we I 

FEET Se,,,,n9\B-24/9-B

600 400 200 0 200 400 600 800 1.0001.2001.4001.6001.8002.000 

DISTANCE FROM BUILDING 24 (B-24). IN FEET

B. Vertical distribution of trichloroethylene 

A , concentrations

EXPLANATION

 100 LINE OF EQUAL TRICHLOROETHYLENE CONCENTRATIONShows trichloroethylene concentration, 

in micrograms per liter. Dashed where 

approximate

A Well screen and trichloroethylene 

concentration, in micrograms per liter

NS Not sampled

< Less than

CAF-7 Location of well and local identifier

Figure 1. (A) Location of Building 24 study area at Picatinny Arsenal, New Jersey, and areal extent of trichloroethylene plume 

and (B) vertical distribution of trichloroethylene concentrations, October-November 1991. (Location of section A-A' is shown in 

figure 1 (A).

299
ESTIMATED MASS DISTRIBUTION OF 

TRICHLOROETHYLENE AND 

PRELIMINARY SOLUTE MASS BALANCE

TCE may be present at the Building 24 research 

site in several phases: dissolved in water, as a vapor in 

the soil gas, and sorbed onto solid surfaces or associated with biota. TCE also may be present as a dense 

nonaqueous-phase liquid (DNAPL). No estimate of 

the amount of DNAPL TCE at the site has been made, 

however. The amounts of TCE in the dissolved, 

sorbed, and vapor phases for a block of aquifer and 

unsaturated zone immediately downgradient from 

Building 24 were estimated on the basis of actual 

measurements of TCE concentrations in samples of 

all three phases (fig. 2). Most of the mass of TCE in 

the system near Building 24 is associated with the 

sediments in both the unsaturated and saturated zones.

A conceptual model of the physical, chemical, 

and biological processes that affect the transport and 

mass balance of TCE within the plume has been developed and presented previously by Imbrigiotta and 

Martin (1991). The physical processes of advection 

and dispersion in the saturated zone affect the movement of dissolved TCE and cause TCE to be removed 

from the system in the discharge to Green Pond Brook. 

TCE also is removed from the system by the biological process of anaerobic biotransformation (reductive 

dehalogenation) and the chemical processes of volatilization at the water table and sorption to saturatedzone sediments. Desorption of contaminated 

sediments can act as a source of TCE to the groundwater system. If TCE is present as a DNAPE, dissolution will result in an increase in dissolved TCE in the 

ground water.

Preliminary estimates of the fluxes of processes 

that affect the mass balance of TCE within the groundwater system at the Building 24 site are shown in 

figure 3. The estimated flux of TCE discharged to 

Green Pond Brook from the plume area, on the basis 

of measurements of TCE concentrations in ground 

water and base-flow discharge to the brook, is 1 to 

2 mg/s. The flux of TCE volatilized from the water 

table is estimated to be about 0.1 mg/s on the basis of 

measured soil-gas TCE gradients and estimates of the 

physical charateristics of the unsaturated zone. 

Although volatilization appears to be a minor factor 

affecting the mass balance of TCE in the ground-water 

system, Martin (this volume) states that results of a 

modeling sensitivity analysis show volatilization to be 

an important mechanism for removing solutes and

thereby affecting concentrations near the water table. 

Biotransformation probably is the mechanism by 

which most dissolved TCE leaves the ground-water 

system. First-order rate constants for TCE transformation ranging from less than 0.001 to 0.02/wk were 

estimated by Wilson and others (1991) on the basis of 

laboratory microcosm studies of soil from five sites 

within the plume area. By using the estimated mass of 

dissolved TCE, the flux of TCE lost from the plume 

through biotransformation is calculated to be in the 

range of 1 to 30 mg/s. Analogous first-order rate constants for TCE biotransformation calculated by Ehlke 

and others (1996) from field-measured TCE concentrations and time-of-travel data generally were higher 

than those measured in the laboratory experiments. 

Thus, the flux of TCE lost through biotransformation 

may actually be greater than that shown in figure 3.

Desorption of TCE from soils that have undergone long-term adsorption (years) and dissolution of 

DNAPL TCE probably arc the processes by which 

most TCE enters the ground-water system. Because 

no estimate of the amount of DNAPL has been made, 

no estimate of TCE dissolution can be made. Three 

first-order rate constants of TCE desorption from 

shallow aquifer sediments at the arsenal made by 

Roller and others (1996) ranged from 0.003 to 0.01 51 

wk. By assuming the estimated mass of TCE sorbed 

to the aquifer sediments to be three to four times the 

mass of TCE in the dissolved state, the 

estimated flux of TCE into the ground-water system 

through desorption is in the range of 15 to 85 mg/s. 

The rate constants were measured in laboratory flowthrough columns by using uncontaminated water as 

the influent fluid. Desorption rates in the field probably would be lower where ground water containing 

TCE is flowing past the desorbing sediments. This 

flux estimate is made by assuming that, over long 

periods (years), the short-term desorption rate (weeks 

and months) is equal to the short-term adsorption rate.

CURRENT RESEARCH ACTIVITIES

The ongoing studies at the Picatinny Arsenal 

site fall into five major areas of research: (1) chemical 

processes affecting transport of chlorinated solvents 

in the saturated zone, (2) transport of chlorinated solvents in the unsaturated zone, (3) biotransformation of 

chlorinated solvents in the saturated zone, (4) solutetransport modeling of chlorinated solvents in the

300
Building 24

10 feet 

10 feet

NOT TO SCALE

MASS DISTRIBUTION OF TRICHLOROETHYLENE

NEAR BUILDING 24

[kg, kilogram; <, less than; 

>, greater than; %, percent]

UNSATURATED ZONE Gas 0.001 kg 

Soil 3 kg 99.9%

SATURATED ZONE Water 1 kg 

Soil 4 kg

20% 

80%

Figure 2. Mass distribution of trichloroethylene in the saturated and unsaturated zones immediately downgradient from 

Building 24 at Picatinny Arsenal, New Jersey.

301
Building 24

VOLATILIZATION 

(0.1 mg/s)

Water table

ADVECTIVE 

TRANSPORT TO 

QREEN POND BROOK 

(1-2 mg/s)

Land surface

LONG-TERM 

DESORPTION 

(15-85 mg/s)

SHORT-TERM

ADSORPTION

(Not estimated)

SHORT-TERM

DESORPTION

(Not estimated)

ANAEROBIC

BIOTRANSFORMATION

(1-30 mg/s)

DISSOLUTION 

(Not estimated)

Estimated top of confining unit

NOT TO SCALE

GAINS

TRICHLOROETHYLENE MASS BALANCE COMPONENTS 

[mg/s, milligrams per second; <, less than]

LOSSES

LONG-TERM DESORPTION 

INFILTRATION 

DISSOLUTION 

SHORT-TERM DESORPTION

15-85 mg/s

< 0.1 mg/s

Not estimated

Not estimated

ANAEROBIC BIOTRANSFORMATION 1-30 mg/s 

ADVECTIVE TRANSPORT TO

GREEN POND BROOK 1-2 mg/s 

VOLATILIZATION 0 1 mq/s 

SHORT-TERM ADSORPTION Not estimated

Figure 3. Preliminary mass-balance estimates of fluxes of processes that effect the fate and transport of 

trichloroethylene in the ground-water system at Picatinny Arsenal, New Jersey.

saturated zone, and (5) remediation processes for 

chlorinated solvents in the saturated and unsaturated 

zones. A brief summary of the current research findings on each of these subjects is given in the following 

sections.

Chemical Processes Affecting Transport of 

Chlorinated Solvents in the Saturated Zone

Flow-through columns were constructed with 

sediments from four sites within the TCE plume to

determine the rate of TCE desorption in the saturated 

zone. Desorption appeared to occur at all sites in two 

stages an initial, rapid stage (days to weeks) during 

which 1 to 10 percent of the total sorbed mass of TCE 

was released, followed by a slow stage (months to 

years) during which the remaining 90 to 99 percent 

was desorbed. A column experiment with sediment 

artificially contaminated in the laboratory for only 5 

days showed the same two-stage desorption, but 65 

to 70 percent of the TCE mass was desorbed in the 

initial, rapid stage, and the remaining 30 to 35 percent

302
was desorbed in the slow stage. A one-dimensional 

model was developed to determine the desorption 

rates by simulating the concentrations measured in the 

desorption column experiments. Results of the model 

simulations compared well with the column data by 

simulating initial, fast-stage desorption as an equilibrium process and a second, slower-stage desorption as 

a kinetic process. Model-calculated long-term (slowstage) desorption rates of TCE from Picatinny Arsenal 

soils ranged from 0.5 x 10'8 to 2.5 x 10 ~ 8/s.

The effect of air drying TCE-contaminated soils 

prior to extraction with methanol to determine the concentrations of sorbed TCE was investigated. Aquifersediment samples collected at five locations within the 

TCE plume were split into two fractions; one fraction 

underwent extraction while wet and the other underwent extraction after being air-dried overnight. Concentrations of TCE in the wet soils were corrected for 

TCE in the soil moisture by subtracting the results of 

analyses of the ground-water samples. Comparison 

of the extraction results for air-dried and wet soils 

showed that, for four of the five samples, the concentration of sorbed TCE in the air-dried soils was not 

significantly different from that in the wet soils.

Transport of Chlorinated Solvents in the 

Unsaturated Zone

A study was conducted to quantify gas-water 

mass-transfer rates at the low flow rates of infiltrating 

water encountered in the unsaturated zone at the 

Building 24 site. This was done by conducting a field 

experiment under steady-state infiltration conditions 

and by using a mathematical model to simulate the 

results. A gas- and aqueous-phase transport model 

with desorption simulated as a constant-flux source of 

TCE at all depths was capable of simulating the field 

data. The gas-water mass-transfer rate constant used 

in the model was 4.5 x 10 ~6/hr. Equilibrium between 

the gas- and water-phase concentrations of TCE was 

not observed during infiltration at the field site. Masstransfer limitations between the soil-water and soilsolid phases also were observed during infiltration.

Biotransformation of Chlorinated Solvents in 

the Saturated Zone

Rates of anaerobic biotransfonnation of TCE 

and cis-1,2-dichloroethylene (cisDCE) were estimated by using field measurements from selected sites

along a flow path in the plume and time of travel for 

ground water between the sites. The first-order 

biotransformation rates for TCE calculated in this 

manner ranged from less than 0.001 to 0.08/wk, 

whereas the first-order biotransformation rates for cisDCE ranged from less than 0.001 to 0.03/wk. The 

field-calculated biotransformation rates for TCE generally were more rapid than the biotransformation 

rates measured previously in the laboratory soil microcosms (<0.01 to 0.02/wk) (Wilson and others, 1991). 

Field-calculated biotransformation rates for cisDCE 

generally were lower than the biotransformation rates 

measured in laboratory soil microcosms (O.01 to 

0.18/wk)(Ehlke and others, 1991). These results 

show that field TCE-concentration data and time-oftravel data can yield biotransformation-rate estimates 

for TCE and cisDCE that are usually the same order of 

magnitude as those measured in laboratory soilmicrocosm studies.

Solute-Transport Modeling of Chlorinated 

Solvents in the Saturated Zone

A modified version of the USGS SUTRA 

transport code (Voss, 1984) is being used to simulate 

areally variable desorption, volatilization at the water 

table, and microbial degradation of TCE. The transport of degradation products, cisDCE and vinyl 

chloride, also is simulated. The reactive multispecies 

solute-transport code was previously described by 

Martin (1991). The modified code simulates the transport and reaction of any number of species at the same 

time.

Sensitivity simulations with higher simulated 

ground-water velocities than those in the calibrated 

model generally resulted in increased simulated 

concentrations of TCE and decreased simulated concentrations of the degradation products. Sensitivity 

analysis on dispersivity showed that the simulated 

concentrations were too high or too low compared to 

measured concentrations when dispersivity rates were 

other than those used in the calibrated model. Results 

of sensitivity simulations run with various desorption 

and degradation rates generally showed that use of the 

laboratory estimates provided reasonable simulated 

concentrations.

303
Remediation Processes for Chlorinated 

Solvents in the Saturated and Unsaturated 

Zones

The feasibility of using aerobic cometabolic 

biotransformation of gas-phase TCE in the unsaturated 

zone as a remediation process was tested. In this process TCE is degraded as a consequence of stimulating 

methane degradation. Soil cores were collected near 

the contaminant source, where the soil-gas concentration of TCE was greatest (43 fag/L), and were used to 

construct soil microcosms. Results of the soilmicrocosm study showed that biotransformation of 

TCE was rapid (16 (^g/L)/d) in acclimated soil having 

a 1.2-percent methane headspace. Thus, this remediation process is feasible at the Picatinny Arsenal site 

and occurs at a much faster rate than anaerobic TCE 

biotransformation processes. Further studies to 

optimize the TCE, methane, oxygen, and nutrient 

concentrations are planned.

A study was begun to determine whether the 

addition of the nonionic surfactant Triton X-100 to 

ground water can artificially increase the rate of TCE 

mass transfer from aquifer sediments was begun. Soil 

samples from the field were used in laboratory experiments conducted with continuous-flow stirred tank 

reactors, with and without Triton X-100. Preliminary 

results indicate that the rate of desorption is increased 

by 15 to 20 percent by the presence of Triton X-100 in 

the aqueous phase. Two possible mechanisms that 

could be responsible for the increased desorption rate 

are proposed: (1) the addition of Triton X-100 above 

its critical micelle concentration increased the apparent water solubility of TCE and thus increased the 

concentration gradient between the sorbed and aqueous phases, or (2) the presence of Triton X-100 

increased the mass-transfer coefficient. Future experiments are planned to determine precisely the mechanism affecting the rate of desorption.

SUMMARY

The results of the ongoing interdisciplinary 

studies at the Picatinny Arsenal research site in northcentral New Jersey have been synthesized to yield 

preliminary estimates of the TCE mass distribution 

and the TCE mass balance (transport fluxes) in the 

ground-water system. Contaminated sediments are 

the primary repository for TCE in both the saturated 

and unsaturated zones. Desorption of TCE from contaminated sediments is a significant long-term source 

of TCE to the ground-water system. Anaerobic 

biotransformation of TCE apparently is the most 

important process for removal of TCE from the 

ground-water system. These results can be used to 

guide future investigations of other sites contaminated 

with chlorinated solvents. In addition, some of the 

results of the remediation studies may help in the 

eventual cleanup of such sites.

Plans for future work at Picatinny Arsenal 

include (1) determination of the presence or absence 

of DNAPL TCE at the site and its relative importance 

as a source of TCE compared to desorption; (2) application of the modified solute-transport model to test 

hypotheses of plume formation, plume aging, and 

the expected effectiveness of different remediation 

processes in cleaning up the plume; (3) direct 

measurement of the TCE flux volatilizing through 

the unsaturated zone; (4) field testing of surfactants 

as a method to enhance desorption of TCE from 

contaminated sediments; (5) application of aerobic 

cometabolic biotransformation of TCE in the field to 

test its effectiveness as a remediation process at pilot 

scale; and (6) application and evaluation of other 

remediation technologies for chlorinated solvents at 

the field scale.

REFERENCES

BeniofT, P.A., Bhattacharyya, M.H., Biang, C, Chiu, S.Y., 

Miller, S., Patton, T., Pearl, D., Yonk, A., and Yuen, 

C.R., 1990, Remedial investigation concept plan for 

Picatinny Arsenal, Volume 2: Descriptions of and 

sampling plans for remedial investigation sites: 

Argonne, 111., Argonne National Laboratory, 

Environmental Assessment and Information Sciences 

Division, p. 22-1 -22-24.

Ehlke, T.A., Imbrigiotta, T.E., Wilson, B.H., and Wilson, 

J.T., 1991, Biotransformation of cis-1,2-dichloroethylene in aquifer material from Picatinny Arsenal, 

Morris County, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 689-697.

Ehlke, T.A., Wilson, B.H., Wilson, J.T., and Imbrigiotta, 

T.E., 1996, In situ biotransformation of trichloroethylene and cis-l,2-dichloroethylene at Picatinny 

Arsenal, New Jersey, in Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the

304
technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological WaterResources Investigations Report 94-4015.

Imbrigiotta, T.E., Ehlke, T.A., and Martin, Mary, 1991, 

Chemical evidence of processes affecting the fate and 

transport of chlorinated solvents in ground water at 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 681-688.

Imbrigiotta, T.E., and Martin, Mary, 1991, Overview of 

research activities on the movement and fate of chlorinated solvents in ground water at Picatinny Arsenal, 

New Jersey, in Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 673-680.

Koller, David, Imbrigiotta, T.E., Bachr, A.L., and Smith, 

J.A., 1996, Desorption of trichloroethylcne from 

aquifer sediments at Picatinny Arsenal, New Jersey, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Water-Resources Investigations 

Report 94-4015.

Martin, Mary, 1991, Simulation of reactive multispecies 

transport in two-dimensional ground-water flow 

systems, in Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology

Program-Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 698-703.

Martin, Mary, 1996, Simulation of transport, desorption, 

volatilization, and microbial degradation of trichloro 

ethylene in ground water at Picatinny Arsenal, New 

Jersey, in Morganwalp, D.W., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Sargent, B.P., Fusillo, TV., Storck, D.A., and Smith, J.A., 

1990, Ground-water contamination in the area of 

Building 24, Picatinny Arsenal, New Jersey: U.S. 

Geological Survey Water-Resources Investigations 

Report 90-4057, 94 p.

Voss, C.I., 1984, SUTRA: A finite-element simulation 

model for saturated-unsaturated, fluid-dcnsitydependent ground-water flow with energy transport or 

chemically-reactive single-species solute transport: 

U.S. Geological Survey Water-Resources Investigations Report 84-4369, 409 p.

Wilson, B.H., Ehlke, T.A., Imbrigiotta, T.E., and Wilson, 

J.T., 1991, Reductive dechlorination of trichloroethylene in anoxic aquifer material from Picatinny 

Arsenal, New Jersey, in Mallard, G.E., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceed ings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 704-707.

305
306
Simulation of Transport, Desorption, Volatilization, and

Microbial Degradation of Trichloroethylene in Ground

Water at Picatinny Arsenal, New Jersey

By Mary Martin1

Abstract

The fate of dissolved trichloroethylene in 

ground water at Picatinny Arsenal, New Jersey, is 

being analyzed using a reactive multispecies 

transport model of a two-dimensional vertical 

section along the central axis of the plume. The 

model simulates the desorption, volatilization at 

the water table, and microbial degradation of 

trichloroethylene. The formation and transport of 

the degradation products cis-1 -2-dichloroethylene 

and vinyl chloride also are simulated. Results of 

this modeling analysis demonstrate the 

applicability and usefulness of the multispecies 

two-dimensional transport code to contaminantfate and -transport studies.

INTRODUCTION

The U.S. Geological Survey (USGS) is 

conducting an interdisciplinary research study of the 

contamination of ground water by chlorinated solvents 

and other contaminants at Picatinny Arsenal, New 

Jersey. As part of the study, the fate and transport of 

dissolved trichloroethylene (TCE) in ground water at 

the arsenal are being analyzed with a reactive 

multispecies transport model. The transport 

simulations are being used to test hypotheses 

concerning the fate of TCE in the ground-water-flow 

system by using sensitivity simulations.

TCE entered the ground-water-flow system 

from a metal-plating facility and wastewater-settling

U.S. Geological Survey, 810 Bear Tavern Rd., West Trenton, 

NJ08619

lagoons at Building 24 at the arsenal during 1960-85 

(fig. 1). TCE concentrations measured during 1985-93 

were used to define a 1.4 x 106-ft2 plume area 

extending 1,630 ft from Building 24 to Green Pond 

Brook in a thickness of about 60 ft of unconfined 

glacial sediments. The amount of dissolved TCE in 

the ground water within the plume area is estimated to 

be about 1,000 kg (about 2,200 Ib) on the basis of 

these measurements. This mass is equal to about 175 

gallons of liquid TCE. The chemistry of the ground 

water, measured TCE concentrations, and the 

chemical processes affecting the fate of TCE within 

the plume are discussed by Imbrigiotta and others 

(1988 and 1991). The model simulates transport from 

Building 24 to Green Pond Brook, the natural 

discharge area, along the approximate flow lines 

within the unconfined aquifer near the central axis of 

the plume.

A modified version of the USGS SUTRA 

transport code (Voss, 1984) is being used to simulate 

the desorption, volatilization at the water table, and 

microbial degradation of TCE. The formation and 

transport of degradation products, cis-1-2- 

dichloroethylene (cisDCE) and vinyl chloride (VC), 

also are simulated. The reactive multispecies solutetransport code is described generally in Martin (1991). 

The modified code retains all the simulation 

capabilities of the SUTRA code and also simulates the 

transport and reaction of any number of species. The 

mathematics that describe the chemical reactions must 

be programmed by the user.

This paper presents the results of sensitivity 

simulations used in testing hypotheses concerning the 

fate of TCE in ground water at Picatinny Arsenal. 

Sensitivity simulations were designed to compare 

simulated solute concentrations from the calibration 

simulation to concentrations from simulations in

307
41'

40'

A. Areal extent of trichloroethylene 

contaminant plume

EXPLANATION

mm Area in which trichloroethylene concentration 

raisi exceeds 10 miorograms per liter

_ 10 LINE OF EQUAL TRICHLOROETHYLENE 

CONCENTRATION-Shows trichloroethylene 

concentration, in micrograms per liter. Dashed 

where approximate

A-A' Line of section

 Sampling site location

74*

Bear Swamp Brook

620

600 400 200 0 200 400 600 800 1.0001.2001.4001.6001,8002.000 

DISTANCE FROM BUILDING 24 (B-24). IN FEET

B. Vertical distribution of trichloroethylene 

concentrations

EXPLANATION

 100-- LINE OF EQUAL TRICHLOROETHYLENE CONCENTRATION-- 

Shows trichloroethylene concentration, 

in micrograms per liter. Dashed where 

approximate

A Well screen and trichloroethylene 

concentration, in micrograms per liter

NS Not sampled

< Less than

CAF-7 Location of well and local identifier

Figure 1. (A) Location of Building 24 study area at Picatinny Arsenal, New Jersey, and areal extent of trichloroethylene 

plume and (B) vertical distribution of trichloroethylene concentrations, October-November 1991. (Location of section is shown 

in fig. 1A.)

308
which different rates were used for the various massbalance components. Concentrations from the calibration and sensitivity simulations are compared. The 

applicability and usefulness of the multispecies twodimensional transport code to analyze contaminant 

fate and transport is demonstrated.

MODEL DESIGN AND CALIBRATION

Average steady-state flow conditions within the 

95,000-ft , 1-ft-wide cross-sectional area are 

simulated by using a grid with 14,329 nodes. Element 

areas range in size from 0.1 to 20 ft2 . Specified-flux 

boundary conditions are used to represent groundwater recharge and flows across the horizontal and 

vertical boundaries of the cross-sectional area (fig. 2). 

Simulated inflow at the flux boundaries is about 

7.5xlO~4 ft3/s and simulated outflows are about 

2.5xlO~4 ft3/s. The magnitudes of the simulated 

boundary fluxes were calculated by using simulated 

flows from a three-dimensional ground-water flow 

model of the valleywide flow system at Picatinny 

Arsenal (Voronin, 1991). A constant-head node is 

used to represent the head in Green Pond Brook where 

it crosses the line of section. Simulated outflow at 

Green Pond Brook is about 5xlO'4 ft3/s. The resulting

Ground-water flow path

altitude of the simulated water table is 696 ft above sea 

level near Building 24 and 686 ft at Green Pond 

Brook.

Solute transport is simulated for 5 years in 50 

time steps. The final simulated concentrations 

represent essentially steady-state plume conditions 

after 1985. Constant-concentration nodes and 

desorption within the plume area are solute sources to 

the simulated system. Simulations were made with 

constant-concentration TCE sources of 110 mg/L 

(one-tenth of the TCE saturation concentration) at the 

water table beneath Building 24 and at the base of the 

unconfined aquifer 170 ft downgradient from Building 

24. During 1987-91, TCE concentrations in ground 

water within the plume have consistently been greatest 

at these locations (greater than 1,000 fig/L near 

Building 24 and greater than 14,000 jig/L near the 

base of the unconfined aquifer). The constantconcentration solute sources are assumed to represent 

high rates of desorption or dissolution of TCE near the 

settling lagoons, at the overflow dry well, and near the 

base of the unconfined aquifer 770 ft downgradient 

from Building 24. Nonaqueous-phase flow of TCE 

from Building 24 to the base of the unconfined aquifer 

is hypothesized as the cause of high concentrations of 

TCE at the base of the unconfined aquifer (Imbrigiotta 

and others, 1991).

TRICHLOROETHYLENE MASS-BALANCE COMPONENTS

Gains

Desorption 

Dissolution

Losses

Volatilization

Flow to Green Pond Brook

Microbial degradation

EXPLANATION

Area in which desorption and mfcrobial 

degradation of trichloroethylene occur

Flow boundary; arrow shows direction 

of flow

Constant-concentration source 

Constant-head boundary

NOT TO SCALE

Figure 2. Schematic drawing showing conceptual solute-transport model and simulated boundary conditions.

309
Both volatilization and microbial degradation 

are simulated with first-order decay rates that are a 

function of the simulated concentration of TCE. 

Microbial degradation was simulated throughout the 

plume area, but volatilization was simulated only at 

the nodes representing the water table. Desorption 

was simulated by using a first-order production term 

and was a function of the estimated amount of TCE 

sorbed to sediments within the plume. The rates at 

which these processes were simulated and the methods 

used to determine the rates are discussed in the 

following section on sensitivity analysis.

The flow system was calibrated by using heads 

and head gradients measured synoptically in 13 to 21 

monitor wells nine times during 1987-93. Generally, 

heads are simulated to within 1 ft of the 1993 waterlevel measurements, which are most representative of 

average flow conditions. The solute-transport system 

was calibrated by matching simulated TCE, cisDCE, 

and VC concentrations to those measured synoptically 

in samples from 15 to 47 monitor wells six times 

during 1987-91. TCE, cisDCE, and VC 

concentrations generally are simulated to within an 

order of magnitude of average concentrations 

measured in micrograms per liter in water from each 

well. Simulated TCE, cisDCE, and VC concentrations 

from the calibrated model, with simulated desorption, 

volatilization, and microbial degradation, are shown in 

figure 3.

A precise calibration was not attempted because 

of the limitations of using a two-dimensional 

simulation along an estimated flow path and of 

estimating flow and solute-transport characteristics 

along the section. The model was calibrated to 

represent a realistic flow field and approximate 

concentrations within the plume. This type of 

calibration is consistent with the purpose of the 

transport simulations that is, to test hypotheses 

concerning the fate of TCE in the ground-water flow 

system. Relatively small-scale variations in the flow 

and solute-transport characteristics are not expected to 

be represented accurately in the model. Also, this 

modeling approach does not provide a unique estimate 

of the magnitude of the mass-balance components; 

however, the overall characterization of the flow and 

transport systems, as well as the testing of hypotheses 

and the comparison of factors affecting the fate of 

TCE, is reasonable and is the most detailed and 

comprehensive possible within the limitations of the 

model design and data availability.

SENSITIVITY ANALYSIS

A sensitivity simulation in which TCE is 

transported as a conservative solute results in 

simulated TCE concentrations that are much higher 

than those simulated by using the calibrated model. 

The simulation results show desorption, volatilization, 

and microbial degradation to be important massbalance components affecting TCE concentrations. 

Simulated TCE concentrations within the same 

hydrogeologic system represented in the calibrated 

model, but with no simulated desorption, 

volatilization, or microbial degradation, are shown in 

figure 4.

Horizontal hydraulic conductivities in the 

calibrated model are 150 to 320 ft/d for the sandy parts 

of the aquifer and 50 ft/d for the fine-grained parts of 

the aquifer. The ratio of simulated horizontal to 

vertical hydraulic conductivity is 100 to 1 throughout 

the section except near Green Pond Brook, where the 

ratio is 10 to 1. The porosity of all the sediments is 

0.3. Sensitivity analysis conducted by using higher 

horizontal hydraulic conductivities than those in the 

calibrated model caused simulated concentrations to 

increase to unreasonable values over most of the 

plume area. Decreasing either the anisotropy 

throughout the section or the hydraulic conductivity of 

the fine-grained layers resulted in irregularities in the 

plume configuration with relatively low concentrations 

of TCE and relatively high concentrations of the 

degradation products in the fine-grained layers. 

Simulated flow velocities generally were 1 to 3 ft/d at 

the simulated location of the monitor wells. Lower 

simulated porosities or higher simulated hydraulic 

conductivies caused simulated ground-water velocities 

to increase, which in turn generally caused simulated 

TCE concentrations to- increase and simulated 

concentrations of the degradation products to 

decrease.

Simulated lateral dispersivity is 1.0 ft and 

simulated transverse dispersivity is 0.1 ft. Results of 

sensitivity analysis conducted by using dispersivity 

values that varied one-half of an order of magnitude 

from those in the calibrated model indicated that 

lateral and transverse dispersivity values greater than 

1.0 ft and 0.1 ft, respectively, cause the simulated 

plume orientation to be more vertical and simulated 

concentrations of TCE and the degradation products to 

increase to unreasonable values over most of the 

plume area. Rates of lateral and transverse 

dispersivity less than those used in the calibrated

310
FEET 

710 r

700 -

Bear Swamp Brook

(A) Trichloroethylene

600 400 200 200 400 600 800 1.000 1.200 1 400 1.600 1 800 2,000

Bear Swamp Brook

600 400 200 200 400 600 800 1.0001.2001.4001.6001.8002000

600 400 200 0 200 400 600 800 1.0001.200 1.4001.6001.8002.000 

DISTANCE FROM BUILDING 24 (B-24), IN FEET

(B) cis-l-2-Dichloroethylene

(C) Vinyl chloride

EXPLANATION

. 10 - - LINE OF EQUAL TRICHLOROETHYLENE CONCENTRATIONShows trichloroethylene concentration, 

in micrograms per liter. Dashed where 

approximate

fl Location of well screen

Figure 3. Simulated concentrations from calibrated model with simulated desorption, volatilization, and microbial 

degradation: (A) trichloroethylene, (B) cis-1-2-dichloroethylene, and (C) vinyl chloride. (Location of section A-A' shown in 

fig. 1.)

311
660

650

640

630

I.OOO / I!

A'

FEET

710

680

670

660

650

640

630

EXPLANATION

-100 LINE OF EQUAL TRICHLOROETHYLENE CONCENTRATIONShows trichloroethylene concentration, 

in micrograms per liter. Dashed where 

approximate

fl Location of well screen

600 400 200 0 200 400 600 800 1.0001.2001.4001.6001.8002.000 

DISTANCE FROM BUILDING 24 (B-24), IN FEET

Figure 4. Simulated trichloroethylene concentrations with no desorption, volatilization, or microbial degradation. (Location of 

section A-A' shown in fig. 1.)

model cause the simulated plume orientation to be 

more horizontal and simulated concentrations in the 

central part of the plume to decrease to unreasonable 

values.

Volatilization at the water table was simulated at 

a rate of 8.6x10 "6/s on the basis of estimates of the 

physical characteristics of the unsaturated zone. Use 

of higher simulated rates of volatilization in the 

sensitivity simulations than those used it the calibrated 

model resulted in simulated concentrations that more 

closely match measured concentrations at the 

locations of monitor wells near the water table. 

Because of the model's limitations, however, estimates 

of volatilization based on model calibration and 

sensitivity are not considered to be more accurate than 

estimates based on physical characteristics. Results of 

the sensitivity analysis showed that volatilization is an 

important mechanism for removing solutes and 

thereby affecting concentrations near the water table.

Desorption, biological degradation of TCE and 

the formation of cisDCE, and the degradation of 

cisDCE and the formation of VC all were simulated at 

a rate of 0.005/week. Because the degradation of VC 

was not simulated, simulated concentrations are higher 

than measured concentrations. Desorption rates were 

estimated on the basis of results of laboratory 

desorption measurements made after 20 days (Roller, 

1992) and degradation rates were estimated on the

basis of laboratory degradation measurements (Wilson 

and others, 1991). Although quantification of the rates 

of desorption and degradation in the plume area is 

difficult, the proceses probably control the major gains 

and losses of TCE. Results of sensitivity simulations 

conducted with various desorption and degradation 

rates generally show that use of the laboratory 

estimates resulted in reasonable simulated 

concentrations.

CONCLUSIONS

Although the calibrated model does not provide 

a unique estimate of the magnitudes of the various 

mass-balance components of the plume of TCEcontaminated ground water at the Building 24 site, the 

use of sensitivity simulations to test hypotheses 

concerning the fate of TCE in ground water is useful. 

Sensitivity simulations in which simulated groundwater velocities were higher than those in the 

calibrated model generally resulted in increased 

simulated concentrations of TCE and decreased 

simulated concentrations of the degradation products. 

Results of sensitivity analysis on dispersivity showed 

that the simulated concentrations were unreasonable 

when dispersivity rates differed from those used in the 

calibrated model by one-half order of magnitude. 

Although volatilization is not a major mass-balance

312
component, this process was shown to be an important 

mechanism in removing solutes and thereby affecting 

solute concentrations near the water table. Results of 

sensitivity simulations with various desorption and 

degradation rates generally showed that use of the 

laboratory estimates resulted in reasonable simulated 

concentrations.

The ability of the model to simulate 

concentrations of TCE and its microbial degradation 

products, cisDCE and VC, as well as the response of 

the simulated concentrations in the sensitivity 

simulations has provided useful insights into 

determining the fate of TCE in ground water at 

Picatinny Arsenal and analyzing the applicability of 

estimated physical and chemical characteristics. This 

modeling analysis also demonstrates the applicability 

and usefulness of the multispecies two-dimensional 

transport code to contaminant-fate and -transport 

studies.

REFERENCES

Imbrigiotta, T.E., Martin, Mary, Sargent, B.P., and Voronin, 

L.M., 1988, Preliminary results of a study of the 

chemistry of ground water at the Building 24 research 

site, Picatinny Arsenal, New Jersey, in Mallard, G.E., 

and Ragone, S.E., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

technical meeting, Phoenix, Arizona, September 26- 

30, 1988: U.S. Geological Survey Water-Resources 

Investigations Report 88-4220, p. 351-359.

Imbrigiotta, T.E., Ehlke, T.A., and Martin, Mary, 1991, 

Chemical evidence of processes affecting the fate and 

transport of chlorinated solvents in ground water at 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 681-688.

Roller, David, 1992, Desorption of trichloroethylenc from 

contaminated sediment at Picatinny Arsenal, New 

Jersey: Boulder, Colorado, University of Colorado, 

unpublished master's thesis, 61 p.

Martin, Mary, 1991, Simulation of reactive multispecies 

transport in two-dimensional ground-water flow 

systems, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 698-703.

Voronin, L.M., 1991, Simulation of ground-water flow at 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 713-720.

Voss, C.I., 1984, SUTRA: A finite-element simulation 

model for saturated-unsaturated, fluid-densitydependent ground-water flow with energy transport or 

chemically-reactive single-species solute transport: 

U.S. Geological Survey Water-Resources 

Investigations Report 84-4369, 409 p.

Wilson, B.H., Ehlke, T.A., Imbrigiotta, T.E., and Wilson, 

J.T., 1991, Reductive dechlorination of 

trichloroethylene in anoxic aquifer material from 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 704-707.

313
314
Simulation of Trichloroethylene Volatilization in the

Unsaturated Zone during a Field Infiltration Experiment

at Picatinny Arsenal, New Jersey

By H. Jean Cho1 , James A. Smith2 , and Peter R. Jaffe3

ABSTRACT

A field experiment was conducted at 

Picatinny Arsenal, New Jersey, to study the 

dynamics of trichloroethylene as it volatilizes 

from contaminated ground water to land surface 

during infiltration. It was found that the gas- and 

water-phase concentrations of TCE were not in 

equilibrium in the unsaturated zone during 

infiltration. A mathematical model was 

developed to simulate the volatilization of TCE in 

the unsaturated zone. Gas water mass-transferrate constants were calibrated to the data. The 

mass-transfer limitations of a volatile organic 

compound between the gas and liquid phases 

were described explicitly in the model. The 

water-infiltration rate was 0.34 centimeters per 

hour, and TCE desorption from the soil matrix 

was significant. Slow desorption of TCE from the 

soil matrix was incorporated in the mathematical 

model.

INTRODUCTION

A field experiment was conducted at Picatinny 

Arsenal in Morris County, New Jersey, to study the 

transport of trichlorethylene (TCE) in the unsaturated 

zone during infiltration. In 1960, TCE was the 

primary solvent used for metal degreasing at Picatinny

'Department of Civil and Environmental Engineering, 

Vanderbilt University, Box 1658-B, Nashville, TN 37235

Department of Civil Engineering, University of Virginia, 

Thornton Hall, Charlottesville, VA 22903--2442

3 Department of Civil Engineering and Operations Research, 

Princeton University, Princeton, NJ 08544

Arsenal. The wastewater from the degreasing 

operations was discharged into sand-lined lagoons and 

into an overflow dry well adjacent to Building 24 

(fig. 1). As a result, a plume of TCE-contaminated 

ground water has been formed in the unconfined 

aquifer that now extends from Building 24 to Green 

Pond Brook, the ground-water discharge point (Smith 

and others, 1992).

This research was conducted to study the 

distribution of TCE between the aqueous and gaseous 

phases in an unsaturated soil system and to quantify 

the rate of exchange of TCE between the two phases. 

The primary focus of the research is the movement of 

a volatile organic compound (VOC) which is 

dissolved in ground water and which diffuses upward 

to land surface. Specifically, the research is aimed at 

evaluating the effect of infiltrating water on the 

volatilization process. During infiltration, the VOC 

will tend to dissolve in the percolating water, so that 

mass transfer of the volatilized VOC will occur 

between the soil-air and soil-water phases.

The main objective of this research was to 

quantify gas water mass-transfer rates at the low flow 

rates of infiltrating water encountered in the 

unsaturated zone. This was done by means of a field 

experiment conducted under steady-state infiltration 

conditions and use of a mathematical model of twophase VOC transport.

METHODS

Mathematical Model

A model was developed to simulate the 

transport of a VOC in the unsaturated zone during

315
infiltration. The governing equations at steady state 

are

(i)

(2)

where cg and cw are the concentrations of the solute 

in the soil gas and soil water, respectively; 6^ and 

6VV are the volumetric soil-gas and soil-water 

fractions, respectively; qw is the infiltration rate; DG< 

and Dj^are the molecular diffusion coefficients of 

the solute in soil air and soil water, respectively; H 

is Henry's constant; ATgw is the gas water masstransfer-rate constant; and S is a constant source 

term describing TCE mass transfer from the soil 

matrix to the soil water.

O Thermisters 

O Neutron Probe

O O O

EXPLANATION Green Pond Brook

10 m-mm Line of equal trichloroethylene

concentration, in micrograms per 0 400 FEET 

liter. Dashed where approximate. | ' ^ L-j '

O 6-8 Borehole and identifier. 0 100 METERS

Figure 1. Map of Picatinny field site and distribution of trichloroethylene in ground water.

316
Field Apparatus

Lysimeters for sampling soil water in the 

unsaturated zone were constructed from porous 

ceramic cups. Soil-gas samplers were constructed 

from stainless-steel tubing and fine-mesh stainlesssteel screening (Cho and others, 1993).

The lysimeters for the field infiltration 

experiment were installed in March 1991 (fig. 2). Six

bentonite 

Soil Gas Probe

bentonite 

Soil Water Probe

bentonite

Soil Gas Probe

bentonite 

Soil Water Probe

bentonite 

Soil Gas Probe

bentonite 

Soil Water Probe

bentonite 

Soil Gas Probe

bentonite 

Soil Water Probe

Shallow 

Ground-water Probe

land surface

0.6 meters

0.9 meters

1.2 meters

1.5 meters

1.8 meters

2.1 meters

2.4 meters

2.7 meters

NOT TO SCALE

3.6 meters

Figure 2. Configuration of boreholes 6-6,6-7, 6-8, 8-6, 

and 8-8.

boreholes were drilled to a depth of 4 m using a 5.1 - 

cm-outer-diameter drive point attached to a 4.8-cmdiameter-outer pipe. The locations of these six 

boreholes are shown in figure 1. The water table at the 

site was approximately 3.5 m below land surface. A 

probe installed at a 3.6-m depth was used to sample 

shallow ground water. Above the water table, four 

lysimeters with 2.7-m, 2.1-m, 1.5-m, and 0.9-m 

lengths of tubing and four soil gas probes, with 2.4-m, 

1.8-m, 1.2-m, and 0.6-m lengths of tubing, were 

installed. Each probe was surrounded by a mixture of 

sieved (no. 10 sieve) native soil and commercial fine 

sand (95 percent finer than 0.59 mm diameter). A 

bentonite layer was emplaced between probes to 

prevent preferential flow along the borehole. 

Additional boreholes was drilled for the emplacement 

of thermisters and a neutron probe (Cho and others, 

1993) (fig. 1).

Unsaturated-zone water was collected using a 

5-mL gastight syringe with a Mininert valve, as 

described by Smith and others (1992). When 5 mL of 

water had been collected, the Mininert valve was 

closed, a liquid field surrogate was injected into the 

sample, and the syringe was placed on ice in a cooler 

for transport to the laboratory. Soil-gas samples were 

collected in 125-mL gas sampling bulbs by a 

procedure described by Smith and others (1990). The 

samples were transported at room temperature in a 

cooler to the laboratory.

In the laboratory, soil-water samples were 

analyzed using a gas chromatograph with a purge-andtrap concentrator, a capillary column, and an 

electrolytic conductivity detector. Soil-gas samples 

were analyzed using a gas chromatograph with a 

purge-and-trap concentrator, a packed column and an 

electrolytic conductivity detector (Cho and others, 

1993).

Results and Discussion

To maintain steady infiltration rates during the 

field experiment, a sprinkler was turned on at the site 

on June 5, 1991, and remained on until July 1, 1991. 

The infiltration rate remained relatively constant at 

approximately 0.34 cm/h, with a standard deviation of 

0.15 cm/h. The moisture-content profile was 

determined with a neutron probe and gravimetric soil 

analyses. At the Picatinny field site, moisture contents 

stabilized after the first day of infiltration (Cho, 1992).

317
Although steady-state TCE concentrations were 

measured in gas and water samples from boreholes 6-7 

and 6-8 on June 11,12,13,15, and 17, only those from 

borehole 6-8 are discussed here. TCE concentrations 

from for borehole 6-7 were similar to those from 

borehole 6-8. One sample was collected from each 

gas probe and two from each lysimeter on each day of 

sampling. Mean TCE concentrations and standard 

deviations as a function of depth are shown in table 1.

Table 1 . Mean concentrations of TCE in soil gas and soil 

water at borehole 6-8

[m, meters; ug/L, micrograms per liter]

Gas Probes

Lysimeters

Depth 

(m)

0.6

1. 2

1.8

2.4

.9

1. 5

2. 1

2.7

Mean 

concentration 

(H/L)

0.003

.022

.051

.113

.39

.43

.55

.68

Standard 

deviation

0.005

.009

.005

.043

.20

.17

.23

.23

Henry's constants were calculated as a function 

of depth using temperature measurements and the 

temperature Henry's constant relation of Gossett 

(1987). The measured gas-phase concentrations were 

divided by Henry's constant to compute the waterphase concentrations expected under equilibrium 

conditions. Gas-phase TCE concentrations 

normalized with Henry's constant and observed waterphase TCE concentrations are plotted in figure 3, 

along with one standard deviation error bars. The 

dotted lines in figure 3 are linear regression lines 

relating both the normalized TCE concentration in soil 

gas and the TCE concentration in soil water to depth. 

Under equilibrium conditions, the normalized TCE 

concentrations in soil gas (triangular symbols) and the 

measured TCE concentrations in soil water (square 

symbols) would be coincident. The data in figure 3 

show that the soil-water-phase TCE concentrations 

were higher than those predicted with the equilibrium 

assumption, suggesting that the soil at Picatinny 

Arsenal was acting as a source of TCE to the 

percolating soil water.

Laboratory Desorption Experiments

In order to test this hypothesis, a series of 

column desorption experiments was conducted in the 

laboratory on soil cores collected at the outer 

perimeter of the site of the infiltration experiment. 

Significant quantities of TCE were observed in the 

column effluent when clean deionized water was

BELOW DEPTH LAND SURFACE, METERS IN 

-30025 -0015 -500 i ti

 ,

\ \ Cg/H

\ * Cw

\ 1 1

^ \ 

1 ' ' 1

\ ' \ '

H \ 

\ ^ 1

 . H\ \

t \h * 1 

\ \

\i \ 

 . \

0.0 0.5 1.0 1.5 

TCE CONCENTRATION, IN MICROGRAMS PER LITER

Figure 3. Variation of soil water and soil gas TCE 

concentrations with depth in the unsatu rated zone, Borehole 

6-8. Mean and 1 standard deviation confidence intervals of 

TCE concentration in the soil-water phase, Cw, shown as 

squares. Mean and 1 standard deviation confidence 

intervals of TCE concentration in the soil-gas phase 

normalized by Henry's constant, Cg/H, shown as triangles.

pumped into the soil (Cho, 1992). Fluxes of TCE from 

the soil matrix to the soil water were calculated using 

the equation

S = ', (3)

where Cw is the concentration of TCE in soil water 

which has traveled a distance ^through the soil. 

Using equation 3 and concentration data from the soil 

desorption experiments, a constant S of 1 .4 mg/(m «h) 

was computed for the Picatinny soil.

Numerical Simulations of Field Data

Numerical simulations of the field experiment 

were conducted to calibrate the steady state model of 

equations 1 and 2. Input parameters and boundary 

conditions are shown in table 2.

Table 2. Input parameters and boundary conditions for 

numerical simulation.

[°C, degrees Celsius; ug/L, micrograms per liter]___________

Parameter Value

Element size 0.01 cm

Soil Temperature 20°C

Henry's constant 0.32

Gas concentration of TCE at upper boundary 0

Water concentration of TCE at upper boundary 0

Gas concentration of TCE at lower boundary 36 (ig/L

Water concentration of TCE at lower boundary 8 (ig/L

318
In an initial simulation, the source term of TCE 

from the soil matrix, 5, was set at 1.4 mg/(m3 «h) as 

determined from desorption experiments. Under these 

conditions, the model significantly overpredicts 

concentrations of TCE in the soil water. A likely 

explanation is that the source term, 5, obtained from 

desorption experiments on highly disturbed cores 

could not accurately predict desorption from 

undisturbed field soil. It was concluded that the data 

collected from laboratory desorption experiments 

could not be used to simulate the field experiments. 

Therefore, S was added to the calibration parameters. 

The simulations were conducted to calibrate two 

parameters: S to account for processes within the soilwater phase and K^ to account for interaction 

between the soil-water and soil-gas phases, which 

were not in equilibrium. The best-fit values of Kg^, 

and S are shown in table 3. The model output for 

borehole 6-8 is shown in figure 4. In all three plots, 

the estimated soil-moisture content is shown by a 

dotted line. In this simulation, both (a) the

Table 3. Parameters for numerical simulation of field data 

for borehole 6-8

[h , per hour; ng/(m *h), micrograms per cubic meter per hour]_____

Parameter Best fit 

value

Standard

deviation

of estimate

Gas-water mass transfer-rate

constant, K^,» h" 1

TCE flux from soil matrix to soil 

water, S1, ng/(m3-h)

4.5x1 0-6

1.6

l.OxKT5

0.4

observed concentrations of TCE in soil water and (b) 

the observed concentrations of TCE in soil gas were 

fairly well matched by the model (solid line). The 

third plot in figure 4 shows the degree of disequilibrium, defined as the ratio of the concentration of TCE 

in the soil gas to the concentration of TCE in the soil 

water divided by Henry's constant. At equilibrium, the 

degree-of-disequilibrium is one. The dashed line in 

the degree of disequilibrium plot denotes equilibrium 

conditions. Throughout the unsaturated zone, the 

degree of disequilibrium was less than one, indicating

SOIL MOISTURE CONTENT 

0.0 0.10 0.20 0.30

SOIL MOISTURE CONTENT 

0.0 0.10 0.20 0.30

SOIL MOISTURE CONTENT 

0.0 0.10 0.20 0.30

o -

CO

DC 

LLJ In 8

o

2

ul

8ul

rr

CO

Q. 

HI 

Q

Moisture Content

,Measured Concentration

O 

un

§ .

Borehole 6-8 

K_gw = 4.5e-6 1/h 

S = 1.6 mg/(m*3*h)

o . in 

o 

o .

o m .

o

LO 

CM

0.0 0.4 0.8 0.0 0.05 0.10 0.15 o.o 0.5 1.0 1.5

TCE CONCENTRATION IN WATER 

IN MICROGRAMS PER LITER

TCE CONCENTRATION IN GAS, 

IN MICROGRAMS PER LITER

DEGREE OF DISEQUILIBRIUM, 

(CG/CW)/H

(a) (b) (c)

Figure 4. Measured and predicted trichlorethylene concentrations and degree of disequilibrium for Borehole 6-8.

319
that the concentration of TCE in the soil gas was less 

than the concentration predicted based on equilibrium 

with the aqueous phase.

Nevertheless, the low Kgw values obtained show that 

equilibrium conditions were not met during the 

experiments.

Summary

A field experiment was conducted at Picatinny 

Arsenal, where TCE has contaminated an unconfined 

sand and gravel aquifer. A desorption experiment on 

field soils showed that a considerable amount of TCE 

has sorbed to the soil matrix. Therefore, there were 

two sources of TCE to unsaturated soil gas and soil 

water: TCE from the contaminated ground water and 

TCE sorbed onto unsaturated-zone soils. As a first 

approximation, the source of TCE from the soil matrix 

to the soil water was treated as a constant-flux term for 

soil water at all depths. Using this term, a two-phase 

transport model, originally developed to treat only the 

gas and aqueous phases of a VOC, was capable of 

simulating the field data, with a best fit Kgw of 

approximately 10~6 h~'.

The equilibrium distribution of TCE between 

the soil-water and soil-gas phases was not achieved 

during infiltration experiments at the Picatinny site. 

Mass-transfer limitations between the soil-water and 

soil-solid phases also were observed during. Because 

of the importance of desorption, a precise value of Kgw 

was difficult to obtain in the field, and the calibrated 

value was not significantly different from zero.

REFERENCES

Cho, H.J., 1992, The Volatilization of Organic Compounds 

in Unsaturated Porous Media during Infiltration: 

Princeton, Princeton University, unpublished Ph.D. 

thesis, 192 p.

Cho, H.J., Jaffe, PR., and Smith, J.A., 1993, Simulating the 

volatilization of solvents in unsaturated soils during 

laboratory and field infiltration experiments: Water 

Resources Research, v. 29, p. 3329-3342.

Gossett, J.M., 1987, Measurement of Henry's law constants 

for C | and 2 chlorinated hydrocarbons: 

Environmental Science and Technology, v. 21, p. 202- 

208.

Smith, J.A., Chiou, C.T., Kammer, J.A., and Kile, D.E., 

1990, Effect of soil moisture on the sorption of 

trichloroethene vapor to vadose-zonc soil at Picatinny 

Arsenal, New Jersey: Environmental Science and 

Technology, v. 24, p. 676 683.

Smith, J.A., Cho, H.J., Jaffe, PR., MacLeod, C.L., and 

Koehnlein, S.A., 1992, Sampling unsaturated-zone 

water for trichloroethene at Picatinny Arsenal, New 

Jersey: Journal of Environmental Quality, v. 21, 

p. 264-271.

320
Cometabolic Biotransformation of Trichloroethylene and

cis-1,2-Dichloroethylene in Unsaturated-Zone

Soil at Picatinny Arsenal, New Jersey

By Theodore A. Ehlke1 and Thomas E. Imbrigiotta1

Abstract

This study determined the feasibility of in 

situ Cometabolic remediation of gas-phase chlorinated ethenes at Picatinny Arsenal, New Jersey. 

Cometabolic biotransformation of trichloroethylene (TCE) and cis-1,2-dichloroethylene (cisDCE) was studied in a laboratory experiment with 

unsaturated-zone soil cores collected within a 

contaminant plume near the source. Concentrations of TCE and cis-DCE in soil, soil gas, and 

ground water within the plume also were quantified. At the highest concentrations studied, TCE 

(3.2 jaM) mixed with cis-DCE (4.3 jaM) was 

degraded most rapidly (0.05 jamol TCE/L/d, 0.09 

(amol cis-DCE/L/d) in acclimated soils with a 

3-percent methane headspace. Degradation of 

TCE at lesser concentrations (1.3 jaM) as the only 

chloroethene present in fertilized, acclimated soil 

was much faster (0.12 jamol TCE/L/d) in the presence of 1.2 percent methane. Most of the unsaturated-zone TCE (>99.9 percent) near the 

contaminant source was sorbed to soil. Results of 

soil-gas analyses indicate that the concentration 

of unsaturated-zone TCE near the contaminant 

source was highest (0.32 jaM) in a shallow clay 

layer, and decreased with depth to the water table. 

This indicates that the original contamination was 

mostly from condensed TCE which drained from 

a degreasing tank to a nearby dry well. The concentration of TCE in unsaturated-zone soil elsewhere throughout the plume increased with depth, 

indicating that most unsaturated-zone TCE

'U.S. Geological Survey, 810 Bear Tavern Road, Suite 206, 

West Trenton, New Jersey 08628

contamination resulted from volatilization losses 

of TCE from contaminated shallow ground water.

INTRODUCTION

Cometabolism is a non-energy-yielding 

chemical reaction that occurs fortuitously as a result of 

microbial metabolic processes (Brock and others, 

1984). Cometabolic bioremediation of soil 

contaminated with chlorinated ethenes is a promising 

technology that offers the potential for rapid 

remediation. Cometabolic biotransformation of 

trichloroethylene (TCE) and cis-1,2-dichloroethylene 

(cis-DCE), originally described by Wilson and Wilson 

(1985), typically involves biostimulation of a mixed 

population of methanotrophs (methane-oxidizing 

bacteria) to produce methane-monooxygenase enzyme 

(MMO). This enzyme is synthesized by soil 

microorganisms to catalyze the reaction of methane to 

methanol, but can also epoxidize chlorinated ethenes 

such as TCE, cis-DCE, and vinyl chloride (VC). 

TCE-epoxide is an unstable compound that rapidly 

reacts spontaneously to form harmless compounds, 

such as glycolate and acetate (Roberts and others, 

1990).

Picatinny Arsenal, a weapons-development 

facility in northern New Jersey, is the site of an 

ongoing study by the U.S. Geological Survey, in 

cooperation with the U.S. Environmental Protection 

Agency, of the fate and transport of TCE in an 

unconfined aquifer. Previous studies at Picatinny 

Arsenal (Smith and others, 1990; Ehlke, 1988) 

indicated the presence of elevated concentrations of 

TCE in unsaturated-zone soil near building 24 as well 

as the possibility of TCE Cometabolic activity. A 

ground-water-treatment process that came on line in 

1992 produces TCE-contaminated gases that require 

treatment. Development of a cometabolic-based

321
y

Bar Stamp Brook

Building 24

Overflow pit D 

E-2-V*

0 »-E-V

Third Av«nu«

C.

o

b

Building 31

0 50 FEET 

0 5 10 15 METERS

Di««»l fuel 

lank lit*

EXPLANATION 

^ 9- A OBSERVATION WELL AND LOCAL IDENTIFIER

O »-A-V SOIL-CORE AND SOIL-GAS SAMPLING SITE 

 E-2-V SOIL-GAS SAMPLING SITE

40'

39'

75*

Figure 1. Location of soil-core sampling sites, soil-gas sampling sites, and ground-water monitoring wells in the vicinity of 

building 24, Picatinny Arsenal, New Jersey.

treatment system could be used either to treat TCEcontaminated soil gas near building 24 in situ or to 

process gases containing TCE ex situ.

The major objective of this study was to 

determine the TCE cometabolic activity of the 

indiginous microflora in the unsaturated zone at 

Picatinny Arsenal and to study the effect of a series of

amendments on the rate of biotransformation of TCE 

and cis-DCE. Other study objectives include the 

quantification of volatile-organic-compound (VOC) 

contaminant levels in soil gas in the unsaturated zone 

and concentrations of VOC's sorbed to soil. This 

report describes TCE and cis-DCE cometabolic 

activity of the indiginous soil microflora at Picatinny

322
Arsenal and methods developed to quantify the mass 

of unsaturated-zone chlorinated ethenes.

STUDY METHODS

Soil cores were collected at two locations near 

building 24 (fig. 1) by using a weight-driven 6.98-cm 

inside-diameter by 60.9-cm-long split-spoon coring 

device. The apparatus was autoclaved at 121°C for 1 

hour prior to transport to the field and was rinsed 

onsite with sterile, deionized water between 

successive core samplings. Continuous cores were 

collected from a depth of 0.6 m to the water table 

(about 3 m). Soil samples were transferred to 1-qt 

sterile mason jars, placed on ice, and transported to the 

laboratory for further processing. After arrival at the 

laboratory, soil samples were refrigerated at 4 °C prior 

to analysis for physical properties and sorbed VOC 

content, as well as for cometabolic activity.

Multilevel soil-gas sampling probes were 

installed at two locations near building 24 by using 

one existing soil-gas probe nest previously installed 

near the overflow pit (dry well) as a model. The probes 

were constructed from 0.63-cm outside-diameter 

stainless-steel tubing slotted at the lower end. The 

probes were installed by augering a 10-cm-diameter 

borehole to slightly below the static water level, then 

installing probes at different depths typically depths 

of 2.74 m (in water), 2.28 m, 1.37 m, and 0.61 m. The 

borehole was sealed between successive probes with 

0.12m of bentonite pellets and with 0.15 m of sand 

above and below the slot openings.

Soil-gas probes were sampled periodically by 

connecting a short length of tubing to Swagelok 

fittings on each probe and removing a 1-L or smaller 

soil-gas sample, which was captured on standard gaschromatograph (GC) traps for subsequent laboratory 

gas-chromatographic analysis.

Laboratory assay of GC traps containing soilgas samples was done by installing a trap containing a 

soil-gas sample on a Tekmar 4000 purge-and-trap 

thermal desorber. After testing the trap for gas leaks, 

GC analysis proceeded by addition of 10 jaL of 

toluene gas in 3 mL of helium to the purge tube as an 

internal standard, then purging the impinged contents 

of the GC trap onto the column of a Varian 3600 GC 

for quantification of volatile hydrocarbons by standard 

methods (Wershaw and others, 1987).

Ground-water samples were collected from the 

lowermost soil-gas probe nest by peristaltic pump with

a short section (0.3 m) of silicone pump tubing and 

were stored in amber 40-mL glass vials for subsequent 

laboratory assay by standard methods (Wershaw and 

others, 1987). Other water samples from deeper 

monitoring wells were collected, transported, stored, 

and analyzed as previously described (Imbrigiotta and 

others, 1991).

A laboratory study was conducted on 

unsaturated-zone soil cores to determine TCE and cisDCE cometabolic activity. The <2-mm size fraction 

from the 0.6- to 3-m depth was composited and 

transferred to sterile 100-mL serum vials for 

determination of TCE and cis-DCE biotransformation 

activity. The procedure was to transfer 25 g of wellmixed soil to a series of serum vials. Triplicates were 

used for all treatments. The serum vials were 

amended with 1 mL of methane or 170 |aL of a sterile 

nutrient (fertilizer) solution (1 g/L each of KH2 PC>4, 

(NH4 )2 SO4 , and KNO3) or a methanotroph 

inoculum to stimulate methanotroph activity. 

Acclimation of the soil microorganisms to 1-percent 

methane in the headspace of individual serum vials 

took 1 to 2 months to reach maximum methanotrophic 

activity. Peak methanotrophic activity was determined 

as the point at which methane consumption per unit of 

time no longer increased. Following acclimation, 

serum vials were spiked with 1 mL of a concentrated 

ethene-gas standard, giving an initial headspace 

concentration of either 1.3 |aM TCE or 3.2 |aM TCE 

and 4.3 |aM cis-DCE. By using a gas-tight syringe, 

600 to 3,000 jaL of methane was introduced into 

certain serum vials, giving a headspace concentration 

of 0.6 to 3 percent volume-in-volume (v/v) methane. 

Following introduction of methane and ethenes, 1-mL 

headspace samples were initially and periodically 

assayed for methane and ethene concentrations by 

means of gas chromatography. The results were used 

to determine the TCE or cis-DCE cometabolic activity 

of the soil.

MICROBIOLOGY

Soil from site 9A (fig. 2) was studied for 

biotransformation activity at an initial TCE 

concentration of 185 |ag/L (1.4 |aM) in 0.6 percent, 1.2 

percent, and 3 percent v/v methane headspace 

concentrations. The 1.2-percent methane amendment 

was studied with and without fertilizer. Results (fig. 2) 

indicate that the greatest TCE removal after 8 days 

occurred in soil having a 1.2-percent methane

323
200

EXPLANATION

-* 0.8 PERCENT METHANE 

" 1.2 PERCENT METHANE

- 3 PERCENT METHANE

-»- FERTILIZED. 1.2 PERCENT METHANE 

"» CONTROL

4 5 

DAY

1.53

1.37

1.22

1.07

0.91

0.76

0.61

0.45

0.30

0.15

Figure 2. Biotransformation of gas-phase trichloroethylene in soil samples from core 9A near building 24, Picatinny Arsenal, 

New Jersey.

EXPLANATION 

0.6 PERCENT METHANE 

1.2 PERCENT METHANE 

+"+ 3 PERCENT METHANE

0.6 PERCENT METHANE, FERTILIZED 

CONTROL

Figure 3. Biotransformation of gas-phase trichloroethylene in soil samples from core 9E near building 24, Picatinny Arsenal, 

New Jersey.

324
headspace concentration with fertilizer amendment 

(0.12 jamol TCE/L/d removal). Lesser but significant 

TCE removal also occurred in 0.6 percent, 1.2 percent 

(no fertilizer), and 3 percent methane (0.10, 0.07, and 

0.06 jamol TCE/L/d removal, respectively). No 

significant losses of TCE occurred in unacclimated 

controls over the same period.

Soil from site 9E (fig. 3) also significantly 

removed TCE and cis-DCE but less rapidly than soil 

from site 9A, because the initial ethene concentrations 

studied were higher (nominally 3.2 jjM TCE, 4.3 jjM 

cis-DCE) and because TCE and cis-DCE were present 

as cosubstrates. The results indicate that both TCE 

and cis-DCE were significantly removed relative to 

controls over a 28-day period. The greatest TCE 

removal (0.05 jamol TCE/L/d) occurred in soil 

amended with 3 percent methane. The least TCE 

removal (0.037 jamol TCE/L/d) occurred in soil 

amended with 0.6 percent methane. Cis-DCE likely 

inhibited TCE biotransformation as a result of 

competition for the MMO enzyme, as previously 

reported by Palumbo and others (1991).

Most of the TCE losses in dead controls (0.02 

jamol TCE/L/d) are attributed to sorption. Slight 

biotransformation activity in unacclimated (live) 

controls (0.03 jamol TCE/L/d, data not shown) is most

likely the result of low-level methanotrophic activity, 

as evidenced by slight production of methane (which 

stimulates methanotrophs) over the 28-day duration of 

the experiment. This apparent low-level 

biotransformation was used to estimate the natural, in 

situ cometabolic activity at the site. The amendment 

with a fertilizer solution slightly increased the rate of 

TCE biotransformation (0.044 jamol TCE/L/d) 

compared to that in soil amended with 0.6 percent 

methane (0.037 jamol TCE/L/d); the increase probably 

would have been greater at a higher methane 

concentration. Amendment with a methanotrophic 

inoculum gave much the same result as fertilization 

with nitrogen and phosphorus.

Cis-DCE was removed much more rapidly than 

was TCE over 28 days in all treatments containing 

both compounds (fig. 4) as a result of competitive 

inhibition for MMO, as has been observed previously 

(Wackett and others, 1989; Semprini and others, 1990; 

and Palumbo and others, 1991). The removal of cisDCE was least in soil amended with 0.6 percent 

methane (0.056-jamol cis-DCE/L/d removal). 

Removal of cis-DCE ranged from 0.08 jamol cisDCE/L/d to 0.09 jamol/L/d in 1.2 percent and 3 

percent methane, in soil amended with 0.6 percent 

methane and fertilizer or in soil amended with 0.6 

percent methane, fertilizer, and a methanotrophic

EXPLANATION 

0.6 PERCENT METHANE 

1.2 PERCENT METHANE 

3 PERCENT METHANE 

FERTILIZED, 0.6 PERCENT METHANE 

CONTROL

5.15

4.12

3.61

3.09

2.58

2.06

1.55

Figure 4. Biotransformation of gas-phase cis-1,2-dichloroethylene in soil samples from core 9E near building 24, Picatinny 

Arsenal, New Jersey.

325
inoculum. Losses of cis-DCE in dead controls, 

presumably to sorption, were much less than TCE 

losses (0.006 ^mol cis-DCE/L/d and 0.02 ^mol 

TCE/L/d, respectively).

SOIL-GAS VOLATILE-ORGANICCOMPOUND CONCENTRATIONS

Soil-gas TCE concentrations were highest (0.32 

|^M TCE, or 42.8 (ag/L) at a depth of 0.6 m near a 

former dry well adjacent to building 24 (table 1). A 

clay layer that is present from 0.6 to 1.2 m is evidently 

highly contaminated with VOC's and may restrict 

upward diffusive transport of TCE in un saturated-zone 

soil gas. Soil-gas TCE concentrations decreased with 

depth near building 24 (0.06 |uM TCE, 7.7 jug/L at a 

depth of 2.3 m), indicating that leakage of condensed 

solvents to the dry well adjacent to building 24 

probably was a major contaminant source. Cis-DCE 

concentrations in soil gas near building 24 also were 

highest (0.005 |uM cis-DCE, 0.5 (ag/L) at shallow 

depths, indicating that significant reductive 

dehalogenation of TCE occurs in unsaturated-zone 

soil near building 24.

Concentrations of TCE and cis-DCE in soil gas 

decreased with distance from the overflow pit near 

building 24. Generally, the highest TCE and cis-DCE 

concentrations at sites 9A and 9E, and elsewhere in the 

plume, were found closest to the water table (table 1).

Table 1. Concentration of trichloroethylene and cis-1,2- 

dichloroethylene in sil gas near building 24, Picatinny 

Arsenal, New Jersey

[TCE, trichloroethylene; cis-DCE, cis-l,2-dichloroethylene; m, meters; 

fiM, micromoles; <, less than; detection level=O.OOI |.iM; d/mo/yr, 

day/month/year]

Sampling TCE cis-DCE 

depth concentration concentration 

(m)

Location Date 

(d/mo/yr)

9E-V 

9E-V 

9E-V 

9E-V

8/27/92 

8/27/92 

12/14/92 

12/14/92

1.38

2.16

1.38

2.16

0.05 

.10 

.03 

.07

E2-V

E2-V

E2-V

1-2-V

E2-V

8/27/92

8/27/92

8/27/92

1 2/14/92

12/14/92

0.61

1.38

2.28

1.38

1.83

.32

.31

.06

.03

.05

9A-V 

9A-V 

9A-V

8/27/92 

8/27/92 

12/14/92

1.38

2.08

2.08

.02 

.02 

.02

0.005

.001

<.OOI

<.OOI

.005

.003

<.OOI

<OOI

<.OOI

<.OOI

<.OOI

.003

This indicates that unsaturated-zone soil 

contamination, except in the immediate vicinity of the 

overflow pit near building 24, is likely the result of 

diffusive losses from contaminated shallow ground 

water. For example, the highest concentrations of 

TCE at site 9E were 0. 10 (aM (13.3 |ug/L) at a depth of 

2.16 m and 0.05 |uM (6.6 |ug/L) at a depth of 1.38 in.

VOLATILE ORGANIC COMPOUNDS 

SORBED TO UNSATURATED-ZONE SOIL

Concentrations of sorbed TCE in unsaturatedzone soil near building 24 range from 0.21 |ug/g to 2.0 

(ig/g (table 2). The highest concentration of sorbed 

TCE (2.0 Hg/g) was found near a former dry well 

(overflow pit) adjacent to building 24 in a clay layer in 

a depth interval of 0.91 to 1 m. Concentrations of 

sorbed TCE in the unsaturated zone decreased with 

depth near building 24; this finding is consistent with 

soil-gas results at the same site.

Cis-DCE was also sorbed to unsaturated-zone 

soil near building 24, as a result of reductive 

biotransformation of TCE in the unsaturated zone, but 

at a far lower concentration than TCE (0.10 to 0.05 

Hg/g). Concentrations of cis-DCE were highest at 

shallow depths (0.61 to 1.22 m); this result is 

consistent with results found for soil gas.

The mass of TCE distributed between the soil 

gas and soil in the unsaturated zone was estimated. 

The results indicate that a 32-m square by 3-m deep 

volume of unsaturated-zone soil near building 24 

contains about 1 to 2 g TCE in the gas phase and about 

2.6 to 3 kg TCE or greater than 99.9 percent sorbed to 

unsaturated-zone soil. This conclusion is described in 

detail by Roller (1992).

Table 2. Concentration of trichloroethylene and cis-1 ,2- 

dichloroethylene sorbed to unsaturated-zone soil at Picatinny 

Arsenal, New Jersey

[TCE, trichloroethylene; cis-DCE, cis-l,2-dichloroethylene; m, meters; 

, micrograms per gram; nd, not determined; detection level=O.OI |.ig/g]

Location

9A

9A

9A

9A

Dry well

Sampling 

interval 

(m)

0.61-1.22

1.22-1.83

1.83-2.44

2.44-3.05

.91-1.00

TCE 

concentration

frig/g)

1. 27

.5!

.35

.2!

2.0

cis-DCE 

concentration

(M9/g)

O.IO

.10

.07

.05

nd

326
CONCLUSIONS

Cometabolic biotransformation of TCE and cisDCE occurs slowly in situ in unsaturated-zone soils at 

Picatinny Arsenal and can be increased by acclimation 

to methane concentrations ranging from 0.6 percent v/v 

to 3 percent v/v. Fertilizing the soil increased 

cometabolic biotransformation activity of TCE and cisDCE.

The fastest TCE removal (0.12 jamol TCE/L/d) 

occurred in the presence of 1.2 percent methane, with 

fertilizer amendment, at an initial TCE concentration of 

1.3 jaM. TCE removal was much slower (0.05 jamol 

TCE/L/d) when cis-DCE also was present, as a result of 

competition for the MMO enzyme as well as the 

concentration of ethenes studied. Bio-transformation of 

cis-DCE occurred much more rapidly than TCE 

removal when both ethenes were present (0.09 jjmol 

cis-DCE/L/d and 0.05 [imol TCE/L/d in 3 percent 

methane, respectively).

The methods developed for quantification of 

volatile organic compounds in soil gas worked well and 

could be used effectively elsewhere to measure organic 

compounds that can be trapped on standard GC traps. 

Concentrations of TCE in soil gas below an overflow pit 

near building 24 ranged from 0.03 to 0.32 \iM (3.9 to 

42.8 ng/L) and were greatest in a clay layer near land 

surface. Elsewhere throughout the plume, soil-gas TCE 

concentrations ranged from 0.02 to 0.10 ^iM (2.6 to 

13.3 ^g/L) and generally increased with depth to the 

water table. By far the greatest mass of unsaturatedzone TCE and cis-DCE (more than 99.9 percent) is 

sorbed to soil from which the ethenes would have to be 

desorbed prior to cometabolic remediation.

REFERENCES CITED

Brock, T.D., Smith, D.W., and Madigan, M.T., 1984, 

Biology of microorganisms (4th ed.): Englewood 

Cliffs, New Jersey, Prentice-Hall, 265 p.

Ehlke, T.A., 1988, Microbiological transformation of 

trichloroethylene in soil at Picatinny Arsenal, New 

Jersey, m Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Phoenix, Arizona, September 26-30, 1988: U.S. 

Geological Survey Water-Resources Investigations 

Report 88-4220, p. 371-376.

Imbrigiotta, T.E., Ehlke, T.A., and Martin, Mary, 1991, 

Chemical evidence of processes affecting the fate and 

transport of chlorinated solvents in ground water at 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 681-688.

Koller, David, 1992, Desorption of trichloroethylene from 

contaminated sediment at Picatinny Arsenal, New 

Jersey: Boulder, Colorado, University of Colorado, 

unpublished master's thesis, 61 p.

Palumbo, A.V., Eng, W., Boerman, PA., Strandberg, G.W., 

Donaldson, T.L., and Herbes, S.E., 1991, Effects of 

diverse organic contaminants on trichloroethylene 

degradation by methanotrophic bacteria and methaneutilizing consortia, in Hinchee, R.E., and Olfenbuttel, 

R.F., eds., On-site bioreclamation processes for 

xenobiotic and hydrocarbon treatment: Boston, 

Butterworth-Heinemann, p. 77-91.

Roberts, P.V., Hopkins, G.D., Mackay, D.M., and Semprini, 

Lewis, 1990, A field evaluation of in situ biodegradation of chlorinated ethenes, Part 1, Methodology and 

field site characterization: Ground Water, v. 28, no. 4, 

p. 591-604.

Semprini, Lewis, Roberts, P.V., Hopkins, G.D., and 

McCarty, PL., 1990, A field exaluation of in situ 

biodegradation of chlorinated ethenes, Part 2, Results 

of biostimulation and biotransformation experiments: 

Ground Water, v. 28, no. 5, p. 715-727.

Smith, J.A., Chiou, C.T., Kammer, J.A., and Kile, D.E., 

1990, Effect of soil moisture on the sorption of 

trichloroethylene vapor to vadose-zone soil at 

Picatinny Arsenal, New Jersey: Environmental 

Science and Technology, v. 24, no. 5, p. 676-682.

Wackett, L.P, Brusseau, G.A., Householder, S.R., and 

Hanson, R.S., 1989, Survey of microbial oxygenases, 

trichloroethylene degradation b propane-oxidizing 

bacteria: Applied and Environmental Microbiology, 

v. 55, no. 11, p. 2960-2964.

Wershaw, R.L., Fishman, M.J., Grabbe, R.R., and Lowe, 

L.E., 1987, Methods for the determination of organic 

substances in water and fluvial sediments: U.S. 

Geological Survey Techniques of Water-Resources 

Investigations, book 5, chap. A3, 80 p.

Wilson, J.T., and Wilson, B.H., 1985, Biotransformation of 

trichloroethylene in soil: Applied and Environmental 

Microbiology, v. 49, no. 1, p. 242-243.

327
328
Desorption of Trichloroethylene From Aquifer Sediments 

at Picatinny Arsenal, New Jersey

By David Koller1 , Thomas E. Imbrigiotta1 , Arthur L Baehr1 , and James A. Smith2

Abstract

Desorption of trichloroethylene (TCE) from 

contaminated aquifer sediments is postulated to 

be a continuing source of this contaminant to 

ground water downgradient from the site of a 

former metal-plating/degreasing operation at 

Picatinny Arsenal in north-central New Jersey. 

Flow-through columns were constructed using 

sediments from four sites within the TCE plume 

to determine whether TCE desorption was occurring and, if so, at what rate. Results of the column 

experiments with contaminated sediments indicate that TCE desorption is occurring at all sites 

tested. Desorption in these columns appeared to 

occur in two stages an initial, rapid stage (days to 

weeks) during which 1 to 10 percent of the total 

sorbed mass of TCE is released, followed by a 

slow stage (months to years) during which the 

remaining 90 to 99 percent is desorbed. Results 

of a column experiment using sediment artificially contaminated in the laboratory for only 

5 days showed the same two-stage desorption, but 

65 to 70 percent of the sorbed TCE was desorbed 

in the initial, rapid stage, and the remaining 30 to 

35 percent was desorbed in the slow stage.

A one-dimensional model was 

developed to determine the desorption rates by 

simulating the desorption measurements from the 

column experiments. The model simulates the 

initial, rapid-stage desorption as an equilibrium

'U.S. Geological Survey, 810 Bear Tavern Road, Suite 206, 

West Trenton, NJ 08628

2 Department of Civil Engineering and Applied Mechanics, 

University of Virginia, Charlottesville, VA 22903-2442

process, and simulates the second, slower stage 

desorption as a kinetic process. Results of model 

simulations compared well with results of the 

column experiments. Long-term (slow-stage) rate 

constants for TCE desorption from Picatinny 

Arsenal soils calculated by using the model 

ranged from 0.5 x 10"8 to 2.5 x 10 "8 per second.

INTRODUCTION

From 1973 to 1983, the improperly placed overflow pipe of a metal degreaser released pure-phase 

TCE into a dry well located in the unsaturated zone in 

front of Building 24 at Picatinny Arsenal in northcentral New Jersey (fig. 1). Leakage of the TCE from 

the dry well into the underlying unconfined aquifer 

helped to create a ground-water-contamination plume 

approximately 500 m long by 250 to 300 m wide at 

Green Pond Brook, the discharge point of the aquifer 

(Imbrigiotta and Martin, 1991). Although input of 

TCE to the ground-water system stopped in 1983, concentrations of TCE in water from wells immediately 

downgradient from the dry well continue to range 

from 1,000 to 1,500 U£/L, even after 10 years. This 

observation implies that a continuing source of TCE 

exists in the Building 24 area.

One possible explanation is that slow TCE 

desorption from contaminated aquifer sediments is 

acting as a persistent source of TCE to the ground 

water. Solutes may diffuse into intraparticle 

micropores or dead-end pores of soil particles and 

become physically inaccessible to the bulk ground 

water, thereby hindering release of the solutes. Other 

researchers have noted that the release rates of many 

low-molecular-weight organic compounds are typically lower than the uptake rates when both are determined in laboratory studies (Szecsody and Bales, 1989;

329
40

39* 0 20 MILES _ 

20 KILOMETERS

EXPLANATION

:^x:;:;:;:^x;x;>: Shaded area between lines represents the area where 

t^MtZ. trichloroethylene concentrations are greater than 10 

micrograms per liter

 10 LINE OF EQUAL CONCENTRATION OF

TRICHLOROETHYLENE-Concentrationin 

micrograms per liter. Dashed where approximately 

located

Site 5 Soil-coring site, sampled well, and identifiers

41-9

Figure 1. Location of Building 24 study area at Picatinny Arsenal, New Jersey; areal extent of trichloroethylene plume, 

October-November 1991; and locations of soil-coring sites and wells sampled.

330
Miller and Weber, 1984). Di Cesare and Smith (1993) 

suggest that the slow desorption of organic solutes 

from soil organic matter may be due to a chemical 

weathering process in which the arrangement of functional groups of soil humic and fulvic acids is changed 

by variations in temperature, pH, and ionic strength. 

The increased tortuosity of the path the solute must 

travel to be released from the soil organic matter 

reduces the release rate. In either situation, the 

assumption of an instantaneous equilibrium between 

the trapped or adsorbed organic solutes and the bulk 

ground water is not valid. Additionally, this effect 

may increase with exposure time of TCE to aquifer 

sediments (Pavlostathis and Jaglal, 1991; Di Cesare 

and Smith, 1993). Many studies of desorption rates 

have been conducted with artificially contaminated 

sediments and may not accurately reflect long-term 

desorption rates.

For the purposes of this study, TCE trapped in 

dead-end or intraparticle pore spaces and TCE partitioned into the soil organic matter are referred to 

together as "sorbed TCE." Consequently, by this 

definition "desorbed TCE" includes both TCE that 

partitions out of soil organic matter and TCE that diffuses from the dead-end or intraparticle pore spaces.

This report describes the results of research to 

(1) determine whether desorption of TCE from aquifer 

sediment is a continuing source of TCE to the ground 

water at Picatinny Arsenal, and (2) estimate the rate of 

TCE release from field-contaminated sediments. Use 

of flow-through columns with field-contaminated sediments and artificially contaminated sediments to 

investigate the desorption phenomenon is described, 

and a one-dimensional model that simulates the 

column results and determines the release rate 

constants is presented.

MATERIALS AND METHODS

Samples of ground water and aquifer sediment 

were obtained from each of the four sites shown in 

figure 1. At each site, soil cores were collected within 

3 m of an observation well. Soil core samples were 

collected from 1.5 m above to 1.5 m below the 

screened interval of the observation well. Groundwater samples were collected from the observation 

wells shortly before or after the cores were collected.

Soil samples were collected by driving a 7.6-cm 

by 61-cm split-spoon coring device to the desired 

depth with a rig-mounted slam bar. The sample was

transferred immediately to clean 1-L wide-mouth 

screw-cap glass bottles. The bottles were filled to 

minimize headspace, sealed, and stored at 4 °C.

Ground-water samples were collected from 

the observation well at each coring site according to 

standard U.S. Geological Survey well-sampling protocols (Wood, 1976; Gibs and Imbrigiotta, 1990). All 

wells were purged with a 4.4-cm-outside-diameter 

stainless-steel gear submersible pump. Volatile 

organic compound (VOC) samples were collected in 

40-mL septum-lined screw-capped glass vials and 

stored at 4 °C.

In the laboratory, three sets of duplicate flowthrough column experiments were conducted using 

sediments from core sites 1, 3, and 5 (fig. 1) as 

described in detail by Koller (1992). Each column 

was constructed by transferring a known mass of fieldcontaminated aquifer sediment to each of the 2.5-cminside-diameter by 100-cm-long glass columns. 

Ground water from the well adjacent to the core site 

was purged with nitrogen gas to remove VOC's, then 

used as the feed water for the columns. Sodium azide 

(200 mg/L) was added to the influent water to inhibit 

biodegradation of TCE in the columns during the 

experiments (Pavlostathis and Jaglal, 1991). A lowflow-rate peristaltic pump was used to maintain a 

steady flow of 0.15 to 0.20 mL/min of the feed water 

through the columns. Effluent from the columns was 

sampled several times per day during the first week of 

an experiment. The sampling frequency was 

decreased to once daily in the second and third weeks 

and further decreased to once every 2 or 3 days for the 

remainder of the experiment. The mass of the watersaturated columns was measured at the end of each 

experiment to determine the porosity and pore volume 

of the sediment column.

A fourth column experiment was conducted 

with sediment that was artificially contaminated with 

TCE in the laboratory. To prepare this sample, sediment from site 4 was baked at 95 °C for 3 days to 

remove all sorbed TCE. A portion of the baked sediment was methanol-extracted and checked for complete removal of TCE. The baked sediment was then 

recontaminated by mixing it with a 40-mg/L aqueous 

TCE solution for 5 days. The sediment was then 

separated from the solution and allowed to air dry 

overnight in a fume hood. Subsamples were analyzed 

to determine the sorbed TCE concentration. Most of 

the artificially contaminated sediment was transferred 

to a set of flow-through columns, and an experiment

331
identical to those done with the field-contaminated 

sediment was conducted.

Tracer tests with bromide solution were conducted to determine whether preferred channels of 

flow existed in the columns. These tests also allowed 

estimation of dispersion coefficients for use in the 

model.

Sediment TCE concentrations were determined 

by using the methanol- extraction procedure described 

in Koller and others (1996). Concentrations of dissolved TCE in the column effluent samples, in the 

extract solutions, and in ground-water samples were 

determined by means of a purge-and-trap/gas chromatographic method described in Kammer and Gibs 

(1989) and Slator and Ho (1989). The sediment 

organic-carbon content and particle-size distribution 

also were measured (Wershaw and others, 1987; Guy, 

1969).

DESCRIPTION OF MODEL

A physically based model was developed to test 

the hypothesis that a fraction of the sorbed TCE is not 

in equilibrium with or is "less accessible" to the bulk 

ground water flowing through the sediment. The fraction of TCE that is in equilibrium is described by the 

dimensionless equilibrium storage coefficient s, which 

is defined as follows:

s = (0 + Kd p(l-0)), (1)

where 0 is porosity (unitless), p is bulk sediment density (g/cm3 ), and Kd is the ratio of sorbed TCE (in 

grams of TCE per gram of sediment) to the aqueousphase concentration of TCE (in grams of TCE per 

cubic centimeter of aqueous phase) at equilibrium. 

Therefore, the units of Kd are cubic centimeters per 

gram. The mass-balance equation describing the 

behavior of the two fractions includes dispersion, 

advection, and a source term that controls the release 

of TCE stored in the sediment that is not in equilibrium with the aqueous phase:

rate of TCE mass transfer from sediment not in 

equilibrium with the aqueous phase (in (g/cm3 )/s). 

The boundary condition for simulating the column 

influent is as follows:

c=Cjatx=0, (3)

where Cj is the influent concentration (in g/cm3 ). 

The other boundary condition is obtained by:

limc = C

where C0 is the initial column effluent concentration

(in g/cm ), which gives the initial condition:

= C0 )att = O. (5)

The release rate R is assumed to be proportional to 

M~the concentration of TCE sorbed to the sediment 

but not in equilibrium with the aqueous phase (in

g/cm3 ):

R =

dt = kM,

where k is the coefficient determining the release 

rate (in s' 1 ). Therefore,

R = kMoe-kt ,

where M0 is the initial total concentration of TCE 

sorbed to the sediment (in g/cm3 ), and M0= M 

at t=0. Equation (2) subject to the boundary and 

initial conditions has the following solution, found 

by Laplace transform:

_n , (ax + pt) c=C0 + w e v *

where w = W|+w2 + w3 , a = -b/2a, P = -b /4a, 

and

(?)

- e P '/, (0 (9)

dc d c , , dc ^7 = a + b at 2 dx nR, (2)

where c is the aqueous concentration of TCE (in 

g/cm3 ), t is time (in s), x is the length of the soil 

column (in cm), a is D/s, b is -q/s, D is the dispersion 

coefficient (in cm2/s), q is flow rate (in cm/s), and R is

where C = Cj - C0, and

KMx

a ) Jnci

t . . . «a'/ F ,

I2 (t)-e 73 (i

(10)

332
O = J

-x 1 exp - + Br

3/2

/ f-*2 ' e*p -7 .

 di , (12)

3/2 £/T , and (13)

Sediment Characteristics

The results of physical and chemical analyses of 

the sediment and ground-water samples are shown in 

table 2. These results show that all four soil samples 

were primarily sand with small amounts of silt and 

clay. In addition, the organic-carbon content of all 

four samples was low (<0.38 percent). The maximum 

concentration of TCE in ground water was three 

orders of magnitude less than the aqueous solubility 

limit for TCE; therefore, dissolution of pure-phase 

TCE was unlikely to be a significant alternative source 

of TCE to the ground water.

= J

/ * _ ^Yf^l exPv i .^..^_ m * ^"-4- j. /I f"Y a 2 TT

3/2 (14)

The solution presented above is similar to that presented by Cameron and Klute (1977) who developed a 

combined equilibrium-kinetic model and applied it to 

adsorption of pesticides, nutrients, and metals in soils. 

Each column simulation requires the data-input 

parameters listed in table 1. All input parameters for a 

given column simulation are measured directly in the 

experiment or are calculated from measured parameters. The model fits two parameters to simulated data: 

the fraction of total sorbed TCE in the less accessible 

phase, M/M0, and the constant, k, which describes the 

rate of release of the less accessible TCE.

DESORPTION OF TRICHLOROETHYLENE

Tracer Tests

Data collected during bromide-ion tracer tests 

showed good agreement between the bromide breakthrough times and the predicted breakthrough times 

based on the gravimetrically determined pore 

volumes. These findings confirmed that no preferred 

flow channels existed in the columns. Results of the 

tracer test also provided estimates of the dispersion 

coefficient for the flow through the columns for use in 

simulations of TCE desorption.

Desorption of Trichloroethylene from FieldContaminated Sediment

Results of the three column desorption experiments conducted with field-contaminated sediments 

are shown in figure 2. Concentrations of TCE remaining on the sediment at any time were normalized by 

dividing by the initial concentration of TCE sorbed to 

the sediment. Thus, all curves start at 1.0 and decrease 

with time as desorption occurs.

In column experiment 1, conducted with the 

most contaminated sediment collected close to the 

source (table 1), only 20 percent of the initial sorbed 

TCE was desorbed after 98 days (98 pore volumes). 

In column experiment 2, conducted with the moderately contaminated sediment collected farther from the 

source, 34 percent was desorbed after 110 days (190 

pore volumes). In column experiment 3, conducted 

with the least contaminated sediment collected farthest 

from the source, 20 percent of the initial sorbed TCE 

was removed after 78 days (85 pore volumes). 

Expressed on the basis of percent removal per 100 

pore volumes of flow through the columns, removal of 

initial sorbed TCE concentrations for columns 1, 2, 

and 3 was 20, 18, and 24 percent, respectively. Thus, 

desorption in the columns from field-contaminated 

sediments from the three locations in the plume tested 

was occurring slowly and at very similar rates. The 

desorption rate did not seem to depend on the initial 

sorbed concentration of TCE on the sediment.

Desorption in these columns appeared to occur 

in two stages an initial, rapid stage (days to weeks) 

during which 1 to 10 percent of the total sorbed mass 

of TCE was released, followed by a slow stage 

(months to years) during which the remaining 90 to 99 

percent was desorbed. A possible explanation for this

333
Table 1. Summary of column-experiment model-input and -output parameters

[u£/g, micrograms per gram; g, grams; cm2 , square centimeters; cm, centimeters; g/cm\ grams per 

cubic centimeters; mL/d, milliliters per day; mL, milliliters; ug/L, micrograms per liter; s, seconds; 

Field, field-contaminated sediment; Artificial, artificially contaminated sediment; TCE, trichloroethylene; <, less than]

Column-experiment number and 

______________________________sediment contamination

Column experiment ________1______2_____3 ____4 parameter ~~~~ ~

Field Field Field Artificial

Model-input parameter:

Initial sorbed TCE 

concentration (ug/g)

Sediment mass (g) 

Cross-sectional area (cm2) 

Length of soil column (cm)

Initial sorbed TCE concentration (g/cm3)

Sediment porosity 

Sediment density (g/cm ) 

Flow rate (mL/d)

Initial TCE concentration 

in effluent (ug/L)

Initial TCE concentration 

in influent (ug/L)

Pore volume (mL) 

Model-output parameter:

Less accessible fraction of 

sorbed TCE concentration

Release-rate constant (s"')

(Cj)

1.7

(L)

(M0)

(0)

(r)

(q)

(C0)

713

5.07

100

2.4x1 0"6

.45

2.6

230

20

230

0.78

594

5.07 

76.5 

.2xlO'6

.39

2.5

260

30

150

0.48

688

5.07 

89.7 

.OxIO'6

.50

3.0

245

12

226

0.17

651 

5.07 

74.6 

.2x10'7

.50

1.7

245

100

228

(M/M0) .99 .90 .90 .33 

(k) 2.5x10'8 I.OxlO'8 0.5x10'8 5.0x10'8

Table 2. Field data from core sediments and ground water collected for 

column experiments

[%, percent; ug/L, micrograms per liter; TCE, trichloroethylene; <, less than; well and core-site 

locations shown in fig. I ]

Core- or ground-watersample characteristic

Core site

Fraction organic carbon (%)

Sand (%)

Silt + clay (%)

Well sampled

TCE concentration in ground

water (ug/L)

Column-experiment number

1

1

38

92

8

9-H

3,500

2

3

.17

100

<l

9-E

1,300

3

5

22

94

6

41-9

1,900

4

4

.03

99

1

92-5

220

334
(Q FRACTION OF INITIAL CONCENTRATION OF SORBED TRICHLOR

REMAINING ON SEDIMENTS

CO 

CO

en
is that the more accessible TCE sorbed to the surface 

of the sediments is desorbed in the initial, rapid stage, 

whereas the less accessible TCE that is trapped in 

dead-end pore spaces or weathered organic matter is 

desorbed in the second, slower phase.

Desorption of Trichloroethylene from 

Artificially Contaminated Sediments

Results of the fourth column desorption experiment conducted with artificially contaminated sediments also are shown in figure 2. Desorption of TCE 

from sediments exposed to contamination for only 5 

days proceeded much more quickly than desorption 

from the field-contaminated sediments. In column 

experiment 4,67 percent of the initial sorbed TCE was 

desorbed after only 54 days (58 pore volumes). All 

TCE theoretically is desorbed by the time 100 pore 

volumes flow through this column. Apparently, the 

elapsed time was insufficient for the TCE to penetrate 

as many dead-end pore spaces or as much of the 

organic-material matrix as it did in the field-contaminated sediments. A two-stage desorption curve was 

observed again; approximately 65 to 70 percent of the 

sorbed TCE was desorbed in the first, rapid stage, and 

the remaining 30 to 35 percent was desorbed in the 

second, slow stage. These findings indicate that TCEdesorption studies conducted with artificially contaminated sediments likely will yield overestimates of field 

desorption rates.

Simulation of Desorption

Desorption was simulated by using the inputparameter values listed in table 1 for each of the four 

column experiments. Model-simulated desorption 

curves were generated and plotted as the solid lines in 

figure 2. The model simulates a rapid release of the 

equilibrium-sorbed TCE followed by a relatively slow 

release of the less accessible TCE. The simulated 

desorption curves agree fairly well with the measured 

desorption values over most of each experiment. 

The model output parameters, M/M0 and k, also are 

given for each of the experiments in table 1. The large 

fractions of less accessible TCE remaining on the 

sediments at the beginning of the slow desorption 

stage (90 to 99 percent of initial sorbed TCE) match 

the estimates made from the measured desorption 

plots. These results support the hypothesis that 

removal of sorbed TCE is kinetically limited, and that

equilibrium-based models will fail to predict the 

removal rates because most of the TCE is inaccessible 

to the water. Mass removal rates for field-contaminated sediments are limited by the low release-rateconstant values controlling the desorption of a large 

fraction of unavailable sorbed TCE. In contrast, only 

33 percent of the sorbed TCE in the artificially contaminated sediment is less accessible, and the removal 

rate is higher than that for field-contaminated samples. 

The long-term desorption rate constants calculated 

with the model ranged from 0.5 x 10"8 to 2.5 x 10"8 per 

second.

This model overestimates the removal rate at the 

end of the desorption experiments, as seen in figure 2. 

In an updated model the release-rate constant, k, 

would be a function of the decreasing concentration of 

TCE on the sediment. The release rate R would be 

expressed as follows:

R = k(M(t)) x M(t). ( 15 >

This change in the expression for R would 

improve the accuracy of the simulation of the gradual 

transition between the two desorption stages over that 

achieved in this study.

SUMMARY AND CONCLUSIONS

Results of laboratory column experiments using 

aquifer sediments from Picatinny Arsenal in northcentral New Jersey indicate that desorption of TCE is 

occurring and acts as a continuing source of TCE to 

the plume. Desorption in the columns appeared to 

occur in two stages an initial, rapid stage (days to 

weeks) during which 1 to 10 percent of the total 

sorbed mass of TCE is released, followed by a slow 

stage (months to years) during which the remaining 90 

to 99 percent is desorbed. Desorption of TCE from 

field-contaminated sediments is slow compared to desorption of TCE from artificially contaminated sediments. A one-dimensional model was developed to 

simulate the column desorption data with an initial, 

rapid-stage desorption as an equilibrium process and a 

second, slow-stage desorption as a kinetic process. 

Results of model simulations agreed well with measured column desorption values over time. These 

results support the hypothesis that most sorbed TCE in 

the system is in a less accessible fraction (trapped 

either in dead-end and intraparticle pore spaces or in 

weathered organic matter) that is not in equilibrium 

with the ground water and undergoes kinetically 

limited desorption.

336
REFERENCES

Camcron, D.A., and Klutc, A., 1977, Convcctive-dispcrsivc 

solute transport with a combined equilibrium and 

kinetic adsorption model: Water Resources Research, 

v. 13, p. 183-188.

Di Ccsarc, David, and Smith, J.A., 1994, Effects of

surfactants on dcsorption rate of nonionic organic compounds from soil to water: Reviews of Environmental 

Contamination and Toxicology, v. 134, p. 1-29.

Gibs, Jacob, and Imbrigiotta, T.E., 1990, Well-purging 

criteria for sampling purgcablc organic compounds: 

Ground Water, v. 28, no. 1, p. 68-78.

Guy, H.P., 1969, Laboratory theory and methods for sediment analysis: U.S. Geological Survey Techniques of 

Water Resources Investigations, book 5, chap. Cl, 

58 p.

Imbrigiotta, T.E., and Martin, Mary, 1991, Overview of 

research activities on the movement and fate of chlorinated solvents in ground water at Picatinny Arsenal, 

New Jersey, in Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Montercy, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 673-680.

Rammer, J.A., and Gibs, Jacob, 1989, An analytical technique for screening purgeablc volatile organic compounds in water: U.S. Geological Survey Open-File 

Report 89-53, 25 p.

Roller, David, 1992, Dcsorption of trichlorocthylene from 

contaminated sediment at Picatinny Arsenal, New 

Jersey: Boulder, Colorado, University of Colorado, 

Department of Geological Sciences, unpublished 

Master's thesis, 61 p.

Roller, David, Imbrigiotta, T.E., and Smith, J.A., 1996,

Effect of air drying on solvent extraction of trichloroethylcnc-contaminated soils from Picatinny Arsenal, 

New Jersey, in Morganwalp, D.W., and Aronson, D. A., 

cds., U.S. Geological Survey Toxic Substances 

Hydrology Program--Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 

20-24, 1993: U.S. Geological Water-Resources Investigations Report 94-4015.

Miller, C.T., and Weber, W.J., Jr., 1984, Modeling organic 

contaminant partitioning in ground-water systems: 

Ground Water, v. 22, no. 5, p. 584-591.

Pavlostathis, S.G., and Jaglal, Kcndrick, 1991, Dcsorptive 

behavior of trichlorocthylene in contaminated soil: 

Environmental Science and Technology, v. 25, no. 2, 

p. 274-279.

Slator, R.W., and Ho, J.S., 1989, Method 502.2, revision 2, 

in Methods for the determination of organic compounds in drinking water: U.S. Environmental 

Protection Agency Environmental Monitoring and 

Systems Laboratory, EPA-600/4-88/039, Cincinnati, 

Ohio, December 1988, p. 31-62.

Szccsody, J.E., and Bales, R.C., 1989, Sorption kinetics of 

low-molccular-wcight hydrophobic organic compound on surface-modified silica: Journal of Contaminant Hydrology, v. 4, p. 181-203.

Wcrshaw, R.L., Fishman, M.J., Grabbc, R.R., and Lowc, 

L.E., 1987, Methods for determination of organic 

substances in water and fluvial sediments: U.S. 

Geological Survey Techniques of Water Resources 

Investigations, book 5, chap. A3, 80 p.

Wood, W.W., 1976, Guidelines for collection and field analysis of ground-water samples for selected unstable 

constituents: U.S. Geological Survey Techniques of 

Water Resources Investigations, book 1, chap. D2, 

24 p.

337
338
Effect of Air Drying on Solvent Extraction of

Trichloroethylene-Contaminated Soil,

Picatinny Arsenal, New Jersey

By David Koller1 , Thomas E. Imbrigiotta1 , and James A. Smith2

Abstract

The effect of air drying trichloroethylene 

(TCE)-contaminated soils prior to extraction with 

methanol to determine the concentrations of 

sorbed TCE was investigated. Saturated-aquifer 

soil and ground-water samples were collected at 

five locations within a TCE plume in ground 

water at Picatinny Arsenal, New Jersey. Each soil 

sample was split into two fractions: one fraction 

underwent extraction while wet and the other 

underwent extraction after being air dried 

overnight. Concentrations of TCE in the wet soils 

were corrected for TCE in the soil moisture by 

subtracting the TCE concentrations determined in 

the ground-water samples. Comparison of the 

extraction results for air-dried and wet soils 

showed that, for four of the five samples, the 

concentration of sorbed TCE in the air-dried 

fraction was higher than or not significantly 

different from that in the wet fraction. In addition, 

the precision of the extraction results for air-dried 

soils was not significantly different from the 

extraction results for wet soils.

INTRODUCTION

Study of the fate and transport of volatile 

organic compounds (VOC's) in ground water and the 

unsaturated zone requires analytical methods to 

determine the mass of contaminants sorbed to soil.

'U.S. Geological Survey, West Trenton, NJ 

Department of Civil Engineering and Applied Mechnaics, 

University of Virginia, Charlottesville, VA

Present standard techniques for analyzing soil 

containing VOC's involve concentration of the 

contaminants by extracting them into an organic 

solvent or by gas-purging them from the soil samples 

onto a solid adsorbent (Sawhney and others, 1988; 

U.S. Environmental Protection Agency, 1986). The 

VOC's in the resulting extracts can then be separated 

and quantified with a gas chromatograph and a 

suitable detector. One important limitation of these 

analytical procedures is that neither procedure can 

accurately quantify the mass of VOC's sorbed only to 

the soil. Because the above-described concentration 

procedures are routinely applied to the soil sample 

without drying, they yield the combined masses of 

VOC's sorbed to the soil and VOC's dissolved in the 

water in the saturated soil sample. For soil samples 

with low organic-carbon content that are exposed to 

ground water containing high concentrations of 

VOC's, the mass of VOC's in the soil water may be of 

comparable magnitude to the mass sorbed to the soil. 

For these cases, analysis of the whole-soil sample-- 

that is, soil and water will result in an overestimation 

of the mass of VOC's sorbed to the soil.

Recently, several researchers have reported that 

VOC's sorbed to natural soil that has been exposed to 

contamination for long periods are resistant to 

desorption into both water and air (Steinberg and 

others, 1987; Sawhney and others, 1988; Smith and 

others, 1990; Koller, 1992). If these findings are 

correct, it is possible that certain soil samples could be 

air-dried prior to extraction and analysis. Air drying of 

the soil would effectively remove the VOC's in the soil 

water by volatilization and leave only the fraction of 

the VOC's that is truly sorbed to the soil sample. 

Although repartitioning of volatilizing VOC's back 

onto the soil is possible, this process is minimized if 

the air drying is performed in a setting where the

339
vapors are constantly being removed, such as a fume 

hood. Direct determination of the concentration of 

VOC's sorbed to air-dried soils would avoid potential 

errors involved in analyzing wet soils and groundwater samples and correcting for the TCE content of 

the soil moisture. Air drying also would simplify the 

handling of soil samples, because precautionary 

measures to prevent the volatilization of the VOC's 

would not be necessary. This paper describes the 

results of a study to determine whether overnight air 

drying of contaminated soil significantly reduces the 

concentration of sorbed TCE.

DESCRIPTION OF THE FIELD SITE

Picatinny Arsenal, in north-central New Jersey, 

has been the site of an interdisciplinary research study 

of the fate and transport of chlorinated solvents in 

ground water by the U.S. Geological Survey (USGS) 

since 1986. Ground-water contamination by TCE 

downgradient from Building 24, site of a former metal 

plating/degreasing operation, is well-documented 

(Imbrigiotta and Martin, 1991; Imbrigiotta and others, 

1991; Sargent and others, 1990). From 1973-83, TCE 

vapors from a degreasing unit condensed in an 

overflow pipe and discharged to a dry well in front of 

the building (fig. 1). Discharge of TCE from the dry 

well into the aquifer has created a plume of groundwater contamination that extends 500 m downgradient 

from Building 24 to Green Pond Brook, the point of 

ground-water discharge from the system. 

Contamination from Building 24 has remained 

primarily in the 15- to 18 m-thick unconfined sand and 

gravel aquifer. The hydrogeology and environmental 

processes affecting the fate and transport of 

chlorinated solvents in the ground water at this site 

have been discussed in detail by Imbrigiotta and 

Martin (1991), Ehlke and others (1991), Martin 

(1991), Wilson and others (1991), and Smith and 

Others (1990).

MATERIALS AND METHODS

Samples of ground water and aquifer material 

were collected from the TCE-contaminated part of the 

aquifer at each of the five sites shown in figure 1. At 

each site, soil cores were collected within 3 m of an 

observation well. Soil-core samples were collected 

from 1.5 m above to 1.5m below the screened interval 

of the observation well. Ground-water samples were

collected from the observation wells shortly before or 

after the cores were collected.

Soil samples were collected by driving a 7.6-cm 

by 61-cm split-spoon coring device to the desired 

depth with a rig-mounted slam bar. The sample was 

transferred quickly to clean 1-L wide-mouth screwcap glass bottles. The bottles were filled to minimize 

headspace, sealed tightly, and stored at 4 °C until 

analysis.

Ground-water samples were collected from the 

observation well at each coring site according to 

standard USGS well-sampling protocols (Wood, 1976; 

Gibs and Imbrigiotta, 1991). All wells were purged 

with a 4.4-cm o.d. stainless-steel, gear submersible 

pump. VOC samples were collected in 40-mL septumlined screw-capped glass vials and stored at 4 °C until 

analysis.

In the laboratory, soil samples were removed 

from the sealed bottles and immediately passed 

through a 2-mm sieve. Sieving homogenized the 

sample and resulted in a more uniform soil for 

extraction. The wet, sieved soil sample was then 

divided into four subsamples. The first three 

subsamples were analyzed for organic-carbon content, 

moisture content, and particle-size distribution 

(Wershaw and others, 1987; Fishman and Friedman, 

1985; Guy, 1969). The fourth subsample was used to 

quantify the concentration of TCE sorbed to the soil. 

One-half of this wet subsample was extracted and 

analyzed immediately by using the procedure 

described below, whereas the other half was air-dried 

for 24 hr at 23 °C in a fume hood and then subjected to 

the same extraction and analytical procedure.

TCE was extracted from all soil samples by a 

slight modification of the procedure described in 

Sawhney and others (1988) and Smith and others 

(1990). From 10 to 15 g of soil were measured into 

preweighed 50-mL glass serum vials. 

Chromatography-grade methanol (approximately 

50 mL) was weighed into to each vial until the 

headspace comprised less than 5 percent of the vial 

volume. Vials were capped with aluminum-foil liners, 

Teflon-lined septa, and aluminum crimp caps. The 

vials were shaken thoroughly by hand and placed in an 

oven at 75 °C. The small headspace in each vial 

accommodated expansion of the methanol during 

heating. After 24 hours, each vial was shaken again 

and returned to the oven. After a total of 48 hours, the 

vials were removed and allowed to cool. Each vial 

contained a settled soil layer and a clear methanol 

layer.

340
40

39 SMILES .

6 ' z'c KILOMETERS

EXPLANATION

^^^\m Shaded area between lines represents the area where 

W-'-M^M trichloroethylene concentrations are greater than 10 

micrograms per liter

 10 LINE OF EQUAL CONCENTRATION OF

TRICHLOROETHYLENE-Concentrationin 

micrograms per liter. Dashed where approximately 

located

Site 5 Soil-coring site, sampled well, and identifiers

^ I «f

Figure 1. Location of Building 24 study area at Picatinny Arsenal, New Jersey; areal extent of trichloroethylene plume, 

October-November 1991, and locations of soil-coring sites and wells sampled.

341
Subsamples of the methanol supernatant were 

analyzed for TCE by using a purge-and-trap/gaschromatography (PT/GC) method. The gas 

chromatograph used was equipped with a wide-bore 

capillary column and a photoionization detector and a 

Hall electolytic conductivity detector in series. This 

analytical procedure is a modification of that described 

by Kammer and Gibs (1989) and Slator and Ho 

(1989). The chromatographic data were collected and 

processed by using a personal computer with datareduction software. Standards in the range of TCE 

concentrations in the samples were run each day. 

Ground-water samples were analyzed by the same 

PT/GC method.

The detection limit for TCE in water or extract 

samples was 0.2 jag/L. On the basis of the maximum 

dilution factor, the average volume of methanol used 

in each extraction, and the average mass of soil 

extracted, the detection limit for TCE sorbed to the 

soil was about 15 fag/kg.

Quality-control procedures to ensure reliable 

and accurate results included the following: (1) 

Selected soil samples were subjected to a second 48- 

hour extraction to detemine whether all the TCE had 

been removed during the first extraction. (2) For each 

set of extractions, two to four vials containing 

methanol with known TCE concentrations underwent 

the same extraction procedure to determine whether 

losses due to sorption to the glass vial or volatilization 

through the cap occurred. (3) All soil samples were 

extracted at least in triplicate; several soil samples 

were extracted up to 10 times. (4) Replicate analyses 

of individual extract solutions and ground-water 

samples were performed frequently.

The concentrations of TCE sorbed on air-dried 

soils were calculated on the basis of the mass of TCE 

determined in the methanol extract and the dry weight 

of the soil extracted. The mass of TCE detemined in 

the methanol extract for wet soils was corrected 

(decreased) by the mass of TCE contained in the soil 

moisture (based on the results of the analysis of the 

ground-water samples). Depending on the sample, this 

correction ranged from 3 to 53 percent of the 

methanol-extracted TCE mass. In addition, the weight 

of the wet soil was corrected by subtracting the mass 

attributable to soil moisture.

RESULTS OF SOIL-EXTRACTION 

ANALYSES

The results of physical and chemical analyses of 

the ground-water and soil samples (other than the 

extraction analyses) are shown in table 1. All five soil 

samples were primarily sand with small amounts of 

silt and clay. In addition, the organic-carbon content of 

all five samples was low (<0.38 percent). Soilmoisture contents ranged from 12 to 24 percent. TCE 

concentrations in ground water differed by two orders 

of magnitude among the five sites.

Results of sequential extractions on three 

different soil samples showed that 95 to 99 percent of 

the TCE present in the soil samples was recovered 

during the first 48-hour extraction and little more TCE 

was recovered by extracting for an additional 48 

hours. All subsequent soil extractions were, therefore, 

48-hour extractions.

The results of the extraction of methanol 

samples spiked with known TCE concentrations 

showed that an average of 95 percent of the spiked 

amount was recovered. This finding indicates that the 

amount of TCE lost by sorption to the glass vial or 

volatilization through the cap was insignificant and did 

not affect the accuracy of the procedure. In addition, 

methanol blanks and soil samples containing no TCE 

that underwent the extraction process showed no 

detectable TCE concentrations when analyzed, 

indicating that no cross-contamination occurred during 

the extraction and analysis procedures.

The results of the extractions of the air-dried 

and wet soils are shown in table 2. In four of the five 

samples tested, the concentrations of TCE extracted 

from the air-dried soils were essentially equal to or 

greater than the concentrations of TCE extracted from 

the wet soils. A two-tailed t test was applied to the five 

sets of analysis results to determine whether the means 

were significantly different at the 95-percent 

confidence level (Zar, 1974). The t-test results, shown 

in the last column of table 2, confirm that no 

significant difference in mean sorbed TCE 

concentrations exists between air-dried and wet soils 

from sites 1, 2, and 4.

At site 5, the concentration of TCE on the airdried soils was significantly higher than that on the 

wet soils, possibly because some of the aqueous TCE 

may have volatilized before the soil could be sealed in 

the extraction vessel. Therefore, the apparent total 

concentration of TCE in the extraction vessel would 

have been lower than the actual value. Consequently,

342
the apparent concentration of sorbed TCE after 

subtracting the correction for the actual concentration 

of TCE in the soil water (1,900 fig/L) also would have 

been lower than the actual value. Concentrations of 

TCE sorbed on wet soil samples from highly 

contaminated sites may be overadjusted downward as 

a result of this phenomenon.

Overadjustment does not occur for soil samples 

from sites where TCE concentrations in ground water 

are low because the TCE contribution from the soil 

water is negligible. At site 2, for example, the 

correction for the low TCE concentration in ground 

water makes almost no difference in the calculated 

concentration of sorbed TCE (3 percent), because the

Table 1. Physical and chemical characteristics of aquifer soils and ground water at five 

sites at Picatinny Arsenal, New Jersey

[m, meter; %, percent; (ag/L, micrograms per liter; Ohs.-well name, observation-well name; TCH cone., trichloroethylene concentration of ground water at site; <, less than]

Aquifer soils

Site

1

2

3

4

5

Soilsample 

depth 

(m)

3.0 - 7.6

2.7 - 8.8

2.7 - 7.3

6.4-11.0

3.4 - 7.9

Soilmoisture 

content

(%)

24

15

14

12

14

Soil 

particle size 

distribution

(%)

Sand 92

Silt + clay 8

Sand 100

Silt + clay <1

Sand 96

Silt + clay 4

Sand 94

Silt + clay 6

Sand 99

Silt + clay 1

Organiccarbon 

content

(%)

0.38

.16

.17

.03

.22

Ground water

Obs.- 

well 

name

9-H

9-A

9-E

92-5

41-9

Depth of 

interval

4.6-6.1

3.0-7.3

4.3-5.8

7.9-9.4

4.9-6.4

TCE 

cone.

(M9/L)

3,500

70

1,300

220

1,900

Table 2. Concentrations of sorbed trichloroethylene on airdried and wet soils sampled at Picatinny Arsenal, New 

Jersey

[Sorbed TCE cone., sorbed trichloroethylene concentration; ng/kg, 

micrograms per kilogram; Std. dev., standard deviation; CV, coefficient of 

variation; %, percent; Sig. diff. @ p=0.05, statistically significant 

difference between means at the 95-percent confidence level; NSD, no 

significant difference; SD, significant difference]

Site

1

2

3

4

5

Soilmoisture 

condition

air-dried

wet

air-dried

wet

air-dried

wet

air-dried

wet

air-dried

wet

Number 

of 

replicate 

samples

4

4

10

10

9

8

7

5

9

8

Sorbed 

TCE 

cone, 

(ng/kg)

2,600

2,600

420

410

700

900

25

27

460

330

Std. 

dev.

900

400

29

95

150

60

6

9

41

110

CV(%)

35

15

7

23

22

7

22

34

9

34

Sig. 

diff. 

@ 

p=0.05

NSD

NSD

SD

NSD

SD

concentration of sorbed TCE is much greater than that 

of dissolved TCE in the wet sample.

At site 3, the concentration of TCE on the airdried soils was significantly lower than that on the wet 

soils. Air drying resulted in a loss of approximately 20 

percent of the TCE concentration on the soil at this 

site. It is not known why the results for this sample 

were different from those for the other four samples, 

because all were treated identically in the extraction 

and analysis process. It is possible that the sample split 

for the extractions was not homogeneous.

The precision of the air-dried-sample analyses 

was better than the precision of the wet-sample 

analyses for three of the five soil samples tested. The 

overall coefficient of variation (standard 

deviation/mean) of the five air-dried-sample analyses 

was 19 percent and of the five wet-sample analyses 

was 23 percent. A t test comparing these means 

showed no significant difference between the 

coefficients of variation for air-dried samples and wet 

samples at the 95-percent confidence level.

343
One possible reason that TCE does not desorb 

when field contaminated soil samples are air-dried is 

that the TCE may be physically trapped in dead-end 

pore spaces or in the intraparticle pore spaces of the 

aquifer materials. In this case, the TCE may not be 

able to diffuse readily to the surface of the soil to 

volatilize.

Another possible explanation is that the TCE is 

held on the soil by a nonequilibrium, kinetically 

limited desorption mechanism. Koller (1992) 

measured field TCE soil/water partition coefficients 

(Koc's) at Picatinny Arsenal and found a strong 

disequilibrium in favor of the soils that is, the 

concentrations of TCE on the soils are much higher 

than would be predicted if an equilibrium with the 

ground-water TCE concentrations is assumed. Smith 

and others (1990) previously found an analogous 

disequilibrium for TCE in the unsaturated zone, with 

much higher concentrations of TCE on the soils than 

would be predicted from the soil-gas TCE 

concentrations. In addition, Koller (1992) determined, 

by means of column experiments with aquifer 

sediments from Picatinny Arsenal, that TCE desorbs 

much more slowly from soils that have been 

contaminated for a long time (years) than from similar 

soils that have been artificially contaminated for a 

short time (days). The kinetic limitation may result 

from a weathering process that changes the 

arrangement or composition of the soil organic matter 

over long periods. Di Cesare and Smith (1993) have 

proposed that functional groups of soil humic and 

fulvic acids may reconfigure and rearrange over time 

in response to changes in temperature, pH, and ionic 

strength of the ground water. When desorption begins, 

the rate of solute diffusion from the soil organic matter 

may be much less than the original rate of diffusion 

into the soil because the tortuosity of the path the 

solute must travel has increased greatly over the 

period of contamination. This hindered desorption of 

TCE from the organic phase may help prevent the loss 

of TCE by volatilization during air drying.

SUMMARY AND CONCLUSIONS

The data presented here show that overnight air 

drying of aquifer soils that have been contaminated for 

long periods does not result in a significant loss of 

TCE from most samples. The overall precision of the 

extraction results was not significantly different for 

air-dried sediments than for wet sediments. Because

there appears to be little difference in the accuracy or 

precision of air-dried and wet-soil extractions, and 

because air drying soils simplifies sample handling in 

the field and laboratory, air drying of TCEcontaminated soil samples at Picatinny Arsenal prior 

to extraction is deemed acceptable and advantageous. 

In addition, analysis of air-dried soil samples avoids 

potential errors analyzing wet-soil samples, which 

must be corrected for the TCE content of the soil 

water. Though caution must always be used in 

applying the findings from one site to another, the 

results of these analyses of samples from Picatinny 

Arsenal probably have transfer value to many other 

sites at which the geology and duration of TCE 

contamination are similar.

The extraction procedure described here yields 

accurate and reproducible results for soil with TCE 

concentrations greater than 15 (ig/kg. This technique 

has the advantage of requiring no special 

instrumentation other than the PT/GC setup ordinarily 

used for the analysis of VOC's in water.

REFERENCES

Di Cesare, David, and Smith, J.A., 1993, Effects of

surfactants on the desorption rate of nonionic organic 

compounds from soil to water: Reviews of 

Environmental Contamination and Toxicology, v. 134, 

p. 1-29.

Ehlke, T.A., Imbrigiotta, T.E., Wilson, B.H., and Wilson, 

J.T., 1991, Biotransformation of cis-1,2- 

dichloroethylene in aquifer material from Picatinny 

Arsenal, Morris County, New Jersey, in Mallard, G.E., 

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 689-697.

Fishman, M.J. and Friedman, L.C., 1985, Methods for 

determination of inorganic substances in water and 

fluvial sediments: U.S. Geological Survey Techniques 

of Water Resources Investigations, book 5, chap. Al, 

709 p.

Gibs, Jacob, and Imbrigiotta, T.E., 1990, Well-purging 

criteria for sampling purgeable organic compounds: 

Ground Water, v. 28, no. 1, p. 68-78.

Guy, H.P., 1969, Laboratory theory and methods for

sediment analysis: U.S. Geological Survey Techniques 

of Water Resources Investigations, book 5, chap. Cl, 

58 p.

344
Imbrigiotta, T.E., and Martin, Mary, 1991, Overview of 

research activities on the movement and fate of 

chlorinated solvents in ground water at Picatinny 

Arsenal, New Jersey, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 673-680.

Imbrigiotta, T.E., Ehlke, T.A., and Martin, Mary, 1991, 

Chemical evidence of processes affecting the fate and 

transport of chlorinated solvents in ground water at 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 681-688.

Kammer, J.A., and Gibs, Jacob, 1989, An analytical 

technique for screening purgeable volatile organic 

compounds in water: U.S. Geological Survey OpenFile Report 89-53, 25 p.

Koller, David, 1992, Desorption of trichloroethylene from 

contaminated sediment at Picatinny Arsenal, New 

Jersey: Boulder, Colorado, University of Colorado, 

Department of Geological Sciences, unpublished 

master's thesis, 61 p.

Martin, Mary, 1991, Simulation of reactive multispecies 

transport in two-dimensional ground-water-flow 

systems, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program- Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 698-703.

Sargent, B.P., Fusillo, T.V., Storck, D.A., and Smith, J.A., 

1990, Assessment of contamination of ground water 

and surface water in the area of Building 24, Picatinny 

Arsenal, New Jersey, 1986-87: U.S. Geological 

Survey Water-Resources Investigations Report 

90-4057, 120 p.

Sawhney, B.L., Pignatello, J.J., and Steinberg, S.M., 1988, 

Determination of 1,2-dibromoethane (EDB) in field 

soils: Implications for volatile organic compounds: 

Journal of Environmental Quality, v. 17, no. 149-152.

Slator, R.W., and Ho, J.S., 1989, Method 502.2, revision 2, 

in Methods for the determination of organic 

compounds in drinking water: U.S. Environmental 

Protection Agency Environmental Monitoring and 

Systems Laboratory, EPA- 600/4-88/039, Cincinnati, 

Ohio, December 1988, p. 31-62.

Smith, J.A., Chiou, C.T., Kammer, J.A., and Kile, D.E., 

1990, Effect of soil moisture on the sorption of 

trichloroethylene vapor to vadose-zone soil at 

Picatinny Arsenal, New Jersey: Environmental 

Science and Technology, v. 24, no. 5, p. 676-683.

Steinberg, S.M., Pignatello, J.J., and Sawhney, B.L., 1987, 

Persistence of 1,2-dibromoethane in soils: Entrapment 

in intraparticle micropores: Environmental Science and 

Technology, v. 21, no. 12, p. 1201-1208.

U.S. Environmental Protection Agency, 1986, Test methods 

for evaluating solid waste (SW-846), Method 8240, 3d 

ed.: U.S. Environmental Protection Agency Office of 

Solid Waste and Emergency Response.

Wershaw, R.L., Fishman, M.J., Grabbe, R.R., and Lowe, 

L.E., 1987, Methods for determination of organic 

substances in water and fluvial sediments: U.S. 

Geological Survey Techniques of Water Resources 

Investigations, book 5, chap. A3, 80 p.

Wilson, B.H., Ehlke, T.A., Imbrigiotta, T.E., and Wilson, 

J.T., 1991, Reductive dechlorination of 

trichloroethylene in anoxic aquifer material from 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91- 4034, p. 704-707.

Wood, W.W., 1976, Guidelines for collection and field

analysis of ground-water samples for selected unstable 

constituents: U.S. Geological Survey Techniques of 

Water Resources Investigations, book 1, chap. D2, 

24 p.

Zar, J.H., 1974, Biostatistical analysis: Englewood Cliffs, 

N.J., Prentice- Hall, Inc., 620 p.

345
346
In Situ Biotransformation of Trichloroethylene and 

cis-1,2-dichloroethylene at Picatinny Arsenal,

New Jersey

By Theodore A. Ehlke1 , Barbara H. Wilson2 , John T. Wilson3 , 

and Thomas E. Imbrigiotta1

Abstract

Biotransformation of trichloroethylene (TCE) 

and cis-1,2-dichloroethylene (cis-DCE) was 

estimated within a part of the unconfined aquifer at 

Picatinny Arsenal, New Jersey, that is contaminated with chlorinated solvents. The first-order 

in situ biotransformation rate for TCE was estimated to range from about 0.001 to 0.05 per week. 

The in situ removal rates generally are more rapid 

than the biotransformation rates that were 

previously measured in the laboratory in soil 

cores collected at the arsenal.

In situ biotransformation was estimated to 

occur more slowly for cis-DCE than TCE. The 

first-order biotransformation rate for cis-DCE 

ranged from about 0.01 to 0.03 per week. Longterm concentrations of TCE and cis-DCE in ground 

water at one location indicate that TCE concentration probably was not the major determining factor 

limiting the transformation of TCE to cis-DCE in 

ground water.

INTRODUCTION

Picatinny Arsenal, a weapons-development facility 

in northern New Jersey, is the site of an ongoing U.S. 

Geological Survey (USGS) study of the fate and transport of chlorinated solvents in an unconfined aquifer. 

The primary sources of TCE to the unconfined aquifer 

near building 24 were two unlined wastewater lagoons 

behind the building and an overflow pit in front of the 

building that received TCE waste from a vapor

'U.S. Geological Survey, West Trenton, N.J.

2 Dynamac Corporation, Robert S. Kerr Environmental 

Research Laboratory, Ada, Okla.

'U.S. Environmental Protection Agency, Robert S. KenEnvironmental Research Laboratory, Ada, Okla.

degreaserfrom 1960 to 1983 (Sargent and others, 1990). 

The TCE plume in the unconfined aquifer extends 

about 540 m from building 24 to Green Pond Brook, 

the local ground-water discharge area (fig. 1). Ground 

water in the aquifer contained an estimated 970 kg of 

TCE in 1991 (Mary Martin, U.S. Geological Survey, 

oral commun., 1993).

The major mechanisms for removal of chlorinated 

ethenes from ground water in the unconfined aquifer 

are reductive dehalogenation by soil microorganisms, 

volatilization from contaminated ground water to the 

unsaturated zone, sorption to aquifer material, and 

advective transport toward Green Pond Brook. 

Estimates of volatilization, sorption, and advectivetransport at the arsenal are described elsewhere (Smith 

and others, 1990; Martin, 1991; Voronin, 1991; and 

Roller, 1992) and were estimated to be significantly 

less than losses of ethenes from ground water as a result 

of reductive dehalogenation. Cometabolic biotransformation of TCE and cis-DCE, although probably 

important in the unsaturated zone, is likely a minor 

removal mechanism in the saturated zone because 

concentrations of oxygen and methane are low there 

(Ehlke and Imbrigiotta, 1996).

Reductive dechlorination (dehalogenation) is the 

process by which a chlorine atom is replaced by a 

hydrogen atom under anaerobic conditions, forming 

cis-DCE. Cis-DCE is similarly dechlorinated, forming 

vinyl chloride (VC), a potent carcinogen. The electrontransfer mechanism producing reductive dehalogenation in situ has not been well described, but has been 

shown elsewhere to be largely microbially mediated 

(Freedman and Gossett, 1989; Sewell and Gibson, 

1991).

Reductive dechlorination of TCE and cis-DCE at 

Picatinny Arsenal has been described in a series of 

reports (Wilson, 1989; Ehlke and others, 1991; Wilson 

and others, 1991). Rates of dechlorination of TCE in 

the laboratory with soil cores collected from several 

sites within the unconfined aquifer at the arsenal

347
41*

40*

39*

75' 74'

20 MILES _ 

0 ' 20 KILOMETERS

EXPLANATION

,.:x::.v:v:;:; : AREA WITHIN WHICH TRICHLOROETHYLENE CONCENTRATIONS 

':W®& ARE GREATER THAN 10 MICROGRAMS PER LITER

 8B GROUND-WATER MONITORING WELL AND SITE IDENTIFIER 

 8 82-3 GROUND-WATER AND SOIL-CORE COLLECTION AND SITE IDENTIFIER

Figure 1. Location of building 24 study area at Picatinny Arsenal, N.J., and areal extent of trichloroethylene 

plume, October-November 1991.

ranged from O.008 to 0.02 per week (Wilson and 

others, 1991). Results of field studies at the arsenal 

indicate that rates of reductive dechlorination of TCE 

and cis-DCE (as estimated from appearance of end 

products and disappearance of reactants in ground 

water) vary spatially in the unconfined aquifer

(Imbrigiotta and others, 1991). This report describes 

estimated in situ reductive-dechlorination rates for 

TCE and cis-DCE at different sites within the unconfined aquifer, variation in the biotransformation rate 

over time, and differences between laboratory and field 

estimates of dechlorination rates.

348
METHODS

In situ removal rates of TCE and cis-DCE were 

estimated from concentrations of TCE and cis-DCE in 

samples of ground water from a series of wells along 

the path of ground-water flow within the contaminant 

plume (fig. 1) by using estimates of advective transport 

of TCE and cis-DCE (Martin, 1991; Voronin, 1991). 

Gains and losses of TCE and cis-DCE in ground water 

due to nonbiological (physical) mechanisms, such as 

sorption, desorption, dispersion, and volatilization, 

were considered to be small in comparison to those due 

to dechlorination. Because physical losses of TCE and 

cis-DCE from ground water, particularly at the boundaries of the contaminant plume, may not be insignificant everywhere, the in situ rates of TCE and cis-DCE 

dechlorination (computed as removal rates, but implied 

to be largely biologically mediated) reported in this 

paper may be overestimated. Desorption of TCE from 

contaminated soils, primarily near building 24, has 

been the major mechanism by which TCE is released to 

ground water at Picatinny Arsenal since its use there 

ceased in 1983. Desorption and sorption mechanisms 

were assumed to be constant over the time interval for 

advective flow of TCE between successive downgradient wells. Results of a previous study (Koller, 1992) 

indicate that this assumption probably is reasonable. 

The presence of nonaqueous-phase liquid (NAPL) 

TCE in the aquifer near the former overflow pit has not 

been verified in soil-core samples, but is indicated by 

modeling results (Mary Martin, U.S. Geological Survey, oral commun., 1993). If NAPL TCE were present 

at Picatinny Arsenal, the rate of solution of TCE probably would be relatively constant and probably would 

not affect biotransformation (removal) computations. 

Losses of TCE by dispersion and volatilization mechanisms were considered to be insignificant at the scale of 

the plume for the purposes of removal-rate computation. TCE flux to and from the unsaturated zone is relatively constant and is estimated to be minor (Cho and 

others, 1993). The time-of-travel for ground water 

between building 24 and Green Pond Brook is about 

2.1 years, at an average velocity of 0.7 m/d (Martin, 

1991).

The rate of microbial removal of TCE and cis-DCE 

between downgradient wells within the plume was estimated by means of a first-order equation that includes 

concentrations of TCE and cis-DCE in ground-water 

samples collected at various times and the estimated 

advective-transport rate. The computation incorporates 

the assumption that a particle of water intercepted by

an upgradient well is transported advectively to the 

downgradient well, and that the possible nonbiological 

losses of TCE and cis-DCE described previously are 

insignificant. The rate of removal was computed as 

follows:

(0

Tw

where

Kd is the estimated in situ removal rate (per week); 

C, is the upgradient concentration of chlorinated ethene; 

C2 is the downgradient concentration of chlorinated ethene 

at a later time (date when upgradient well was 

sampled, plus Tw); 

Tw is the advective transport time, in weeks, between

up- and down-gradient wells; and 

L is the natural logarithm.

Change in the rate of removal of TCE at one well 

(9A; fig. 2) over the time interval from 1982 to 1992 

was estimated by using geochemical data collected by 

non-USGS personnel during 1982-87 and by USGS 

personnel during 1987-92. The former results are listed 

in Sargent and others (1990).

ESTIMATE OF THE RATE OF IN SITU 

REMOVAL OF TRICHLOROETHYLENE 

AND CIS-1,2-DICHLOROETHYLENE

Removal of TCE and cis-DCE was estimated along 

a transect parallel to ground-water flow within the contaminated part of the unconfined aquifer (fig. I) by 

using a first-order reaction (eq. I). Most transport of 

chlorinated ethenes in ground water near the contaminant source occurs in the shallow flow system (3.5- 

11 m depth). Results of particle-tracking analysis 

indicate that concentrations of ethenes in ground water 

at wells 9B, 9E, CAF-2, and 92-5 (fig. 1) provide a 

reasonably good estimate of local ground-water quality 

in the shallow flow regime (Voronin, 1991). Advective 

flow between these wells varies from about 75 days 

(wells 9B-9E) to 180 days (wells 9E-CAF-2). Advective flow in the deeper part of the unconfined aquifer 

(depth greater than 12 m) is an important transport 

mechanism for ethenes in ground water near Green 

Pond Brook. For additional details of flow in the 

unconfined aquifer, see Voronin (1991).

Removal of ethenes in the deeper part of the 

unconfined aquifer was estimated for that part of the 

aquifer between wells 92-3 and 112-7 (fig. 1). This 

location corresponds to ground-water flow 15 to 16.8m 

deep in permeable layer 2, as described by Voronin

349
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CONCENTRATION, IN MICROGRAMS PER LITER

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(1991). Ground-water flow in this system is slightly 

slower than flow at shallower depths (about 0.5 m/d 

between wells 92-3 and 112-7 compared to 0.7 m/d in 

the shallower part of the unconfined aquifer between 

wells CAF-2 and 92-5). The increase in contact time 

per unit of distance increases the removal rate of 

ethenes in the deeper aquifer.

The in situ removal of TCE in shallow ground 

water near the contamination source (between wells 9B 

and 9E) ranges from <0.001 to 0.08 per week (median 

0.02 per week) on the basis of concentrations of 

ethenes in ground water collected at different times 

corresponding to the estimated ground-water-flow time 

between these wells. Most of the removal is attributed 

to anaerobic biotransformation processes for the 

reasons previously stated.

Removal of TCE from ground water in the shallow 

flow system between wells 9E and CAF-2 is more rapid 

than removal nearer the contamination source (median 

0.05 per week, range 0.02-0.08 per week). This difference is attributed to increased biological activity associated with biotransformation of fuel oil. A plume of 

degradation products of fuel oil near building 31 partly 

overlaps the TCE plume in this part of the study area 

(fig. 1). A leaking diesel-fuel tank was removed from 

near building 31 in 1986. This tank was a likely source 

of hydrocarbons, which could be metabolized by methanogens to form methane and other products, such as 

fatty acids and reduced nonvolatile organic compounds 

that have been reported to be indirectly associated with 

in situ biotransformation of chlorinated ethenes. Major 

and others (1991) reported that in situ anaerobic 

biotransformation of organic solvents (largely methanol) enhanced the dechlorination of tetrachloroethylene (PCE) and TCE.

Results of analyses of ground-water samples 

collected at a well upgradient from the fuel-oil spill (well 

13-2) and two wells downgradient from the spill (wells 

34-1 and 111-1) (table 1) indicate greater-than-background concentrations of methane, ammonia, dissolved 

organic carbon, and certain minor elements (such as iron 

and manganese) in wells 34-1 and 111-1 compared to 

concentrations of these constituents at the uncontaminated well (13-2). The redox potential and concentrations of dissolved oxygen and nitrate were lower in wells 

34-1 and 111-1 than in well 13-2. In situ biotransformation of TCE and cis-DCE in the unconfined aquifer near

wells 34-1 and 111-1 probably is rapid, as indicated 

by the relatively high ratio of the concentration of products of TCE biotransformation (cis-DCE and VC) to 

the reactant concentration. For example, the ratio of 

TCE:cis-DCE concentrations at well 34-1 was 1.5:1, and 

the ratio of TCE:VC concentrations at well 111-1 was 

2:1 (table 1). By comparison, the TCE cis-DCE concentration ratio at well 92-3, which is near a part of the 

unconfined aquifer in which biotransformation of TCE 

is slow, ranged from 90:1 to 140:1 during the same 

period.

Removal of TCE in the part of the unconfined 

aquifer between wells CAF-2 and 92-5 is very slow, 

ranging from O.001 to 0.001 per week (median 0.001 

per week). The low concentrations of cis-DCE and VC 

observed at the downgradient well (92-5) (from < 1 ng/L 

to 10 ug/L cis-DCE, and <1 ug/L VC) indicate that 

biotransformation of TCE and cis-DCE there probably 

is significantly slower than biotransformation closer to 

the contamination source. Analyses of soil samples 

collected at location CAF-2 during a previous study 

indicated very slow biotransformation of TCE (O.01 

per week) (Wilson and others, 1991).

Removal of TCE from ground water in a deeper 

part of the unconfined aquifer (12-16.8 m deep) was 

estimated from concentrations of TCE and cis-DCE 

in ground water in wells 92-3 and 112-7. In general, 

results indicate that removal of TCE is fairly rapid

Table 1. Results of water-quality analyses of ground-water 

samples from Picatinny Arsenal, New Jersey 

[mg/L, milligrams per liter; N, nitrogen; ug/L, micrograms per liter; 

<, less than]

Constituent or 

characteristic

Redox potential, millivolts

Oxygen, dissolved, mg/L

Ammonia, dissolved as N (mg/L)

Nitrate, dissolved as N (mg/L)

Iron, dissolved, ng/L

Manganese, dissolved, ug/L

Organic carbon, dissolved, mg/L

Methane, ng/L

Trichloroethylene, ng/L

cis- 1 ,2-Dichloroethylene, ug/L

Vinyl chloride, ng/L

Well identifier

13-2

209

4.3

.01

1.9

10

10

.7

1**

4

<l

<!

34-1

1 60

.4

.02

2

20

42

1. 2

77*

1 5

l()

<l

111-1

34

.2

.45

<.l

2,300

2IO

4.2

325

16

<l

8

*Samples were collected in December 1989, June 1989. 

**Samples were collected in April 1990.

351
(median 0.025 per week) in this part of the aquifer. 

The presence of modest concentrations of daughter 

products of TCE biotransformation at well 112-7 

(22-120 ug/L cis-DCE and 3-4 ug/L VC) indicate 

that biotransformation of TCE probably is much 

more rapid at the 14 to 15.7 m depth of well 112-7 

than at shallower depths in this area. In comparison, 

the concentration of cis-DCE at well 112-8 (same 

location as well 112-7 but screened at 4.8 to 6.4 m 

instead of 14.0 to 15.7 m) ranged from <1 to 7 ug/L 

over the same period.

Cis-DCE removal occurred more slowly in the 

unconfined aquifer than did removal of TCE. Removal 

of cis-DCE in a shallow part of the aquifer between 

wells 9A and 111-1 ranged from 0.01 to 0.03 per week. 

Removal of cis-DCE in the deeper part of the aquifer 

between wells 92-3 and 112-7 was estimated to range 

from about 0.001 to 0.01 per week. This finding is consistent with results reported elsewhere, which indicate 

that the rate of cis-DCE biotransformation to VC is 

slower than the rate of production of cis-DCE from 

TCE (Barrio-Lage and others, 1986). Similar rates for 

the biotransformation of cis-DCE (0.01 -0.06 per week) 

were reported by Barrio-Lage and others (1986) for 

Florida Everglades muck. The observed concentrations 

of TCE, cis-DCE, and VC in ground water at Picatinny 

Arsenal also strongly indicate that the in situ removal 

rate of cis-DCE (largely by biotransformation) is less 

than that for TCE.

VARIATION IN BIOTRANSFORMATION 

RATE WITH TIME

Variation in the rate of biotransformation of TCE to 

cis-DCE during 1982-92 was estimated from results of 

analyses of ground-water samples collected at well 9A. 

Time-series plots of the reactant (TCE) and the intermediate product of biotransformation (cis-DCE) are 

shown in figure 2. The plot indicates that, although the 

concentration of TCE in ground water decreased from 

about 300 ^ig/L in 1982 to 66 ug/L in 1992, the concentration of cis-DCE in ground water increased slightly 

from about 43 ug/L in 1982 to 79 ug/L in 1992. The 

biotransformation of TCE to VC is shown in equation 

2. If the overall biotransformation rate described by the 

following series of reactions was limited by the concentration of the first reactant (TCE), the formation of 

cis-DCE would be expected to decrease as the concentration of TCE decreased

H

TCE

However, the rate of formation of cis-DCE during 

1982-92 was nearly constant. The reaction rates of 

TCE to cis-DCE and from cis-DCE to VC are more 

likely to depend on the concentration of other constituents (electron donors), that have not yet been identified 

at Picatinny Arsenal. These results also provide field 

evidence that the in situ biotransformation of cis-DCE 

to VC occurs more slowly than the formation of cisDCE from TCE. If the biotransformation of cis-DCE to 

VC occurred more rapidly than the formation of 

cis-DCE, the in situ pooling of cis-DCE at certain 

points in the unconfined aquifer would not be observed. 

DiStefano and others (1992) have hypothesized that 

electrons transferred during methanogenesis may be 

involved in reductive dechlorination of PCE and TCE.

COMPARISON OF FIELD-AMD 

LABORATORY-MEASURED 

BIOTRANSFORMATION RATES

Anaerobic biotransformation rates of TCE and 

cis-DCE, which were estimated previously in a series 

of laboratory batch experiments (Ehlke and others, 

1991; Wilson and others, 1991), generally indicated 

similar or lower biotransformation rates than were estimated from field results (table 2). The laboratory estimate of the rate of TCE biotransformation ranged from 

<0.01 to 0.02 per week within the contaminant plume, 

compared to field estimates of the rate of TCE biotransformation at Picatinny Arsenal, which range from 

0.001 to 0.05 per week. The reasons for these 

differences are unclear at this time but may reflect 

experimental conditions in the laboratory, including the 

inability to reproduce in situ concentrations of reduced 

electron donors, which are required for the reductive 

dehalogenation process to proceed. Also, observed in 

situ removal of TCE and cis-DCE includes losses from 

dispersion, volatilization, and advective transport-- 

processes that were not included in laboratory removal 

estimates.

352
Table 2. Laboratory and field estimates of trichloroethylene 

biotransformation in the unconfined aquifer at Picatinny Arsenal, 

New ersey

[m/d, meters per day; GW, ground water; <, less than. Laboratory data 

from Wilson and others, 1991]

Location and 

sample type

9Uto9F-,GW 

9U, soil

9[i to CAF-2, GW 

CAF-2, soil

CAF-2 to 92-5, GW 

92-5, soil

92-3 to II2-7.GW

1 1 2-7, soil 

4 1 -9, soil

Biotransformation rate, 

per week

<0.0()l 0.08, median 0.02 (field) 

0.02 (laboratory)

0.02 0.08, median 0.05 (field) 

<O.OI (laboratory)

<0.()0 1 0.00 1 , median 0.00 1 ( field) 

0.0 1 2 (laboratory)

0.025 (field)

0.008 (laboratory) 

0.004 (laboratory)

Ground-water 

velocity 

between 

locations 

shown in fig. 1 

(m/d)

0.3

.6

.7

.5

SUMMARY AND CONCLUSION

Rates of biotransformation of TCE and cis-DCE 

within the contaminated part of the unconfined aquifer 

at Picatinny Arsenal, N.J., were estimated. Rate constants for anaerobic biotransformation of TCE in situ 

were estimated to range from about 0.001 to 0.05 per 

week in different parts of the unconfined aquifer. In situ 

biotransformation of TCE generally is estimated to 

occur more rapidly than was previously measured in 

laboratory studies with soil cores from the arsenal. The 

rate of in situ biotransformation of cis-DCE is slower 

than the rate of biotransformation of TCE, ranging 

from about 0.01 to 0.03 per week. These estimates are 

within the range of reported rates of biotransformation 

of TCE and cis-DCE in soil at other locations. Concentrations of TCE and cis-DCE in ground water during 

1982-92 indicate that the TCE concentration is not a 

limiting factor in determining the rate of biotransformation of TCE to cis-DCE.

REFERENCES

Barrio-Lage, Gladys, Parsons, F.Z., Nassar, R.S., and 

Lorcnzo, P. A., 1986, Sequential dehalogenation of 

chlorinated ethenes: Environmental Science and 

Technology, v. 20, p. 96-99.

Cho, H.J., JafTe, P.R., and Smith, J.A., 1993, Simulating the 

volatilization of solvents in unsaturated soils during 

laboratory and field infiltration experiments: Water 

Resources Research, v. 29, no. 10, p. 3329-3342.

DiStcfano, T.D., Gossctt, J.M, and Zindcr, S.H., 1992, 

Hydrogen as an electron donor for dechlorination of 

tetrachloroethcne by an anaerobic mixed culture: 

Applied and Environmental Microbiology, v. 58, no. I, 

p. 3622-3629.

Ehlkc, T.A., Imbrigiotta, T.E., Wilson, B.H., and Wilson, 

J.T., 1991, Biotransformation of cis-l,2-dichloroethylene in aquifer material from Picatinny Arsenal, Morris 

County, New Jersey, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey Calif., March H-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 689-697.

Ehlke, T.A., and Imbrigiotta, T.E., 1996, Comctabolic 

biotransformation of trichloroethylene and cis-1,2- 

dichlorocthylene in unsaturated-zone soil at Picatinny 

Arsenal, New Jersey, in Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Freedman, D.L., and Gossett, J.M., 1989, Biological reductive dechlorination of tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions: 

Applied and Environmental Microbiology, v. 55, no. 9, 

p. 2144-2151.

Imbrigiotta, T.E., Ehlke, T.A., and Martin, Mary, 1991, 

Chemical evidence of processes affecting the fate and 

transport of chlorinated solvents in ground water at 

Picatinny Arsenal, New Jersey in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, Calif., March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 681-688.

Koller, David, 1992, Desorption of trichloroethylene from 

contaminated sediment at Picatinny Arsenal, New 

Jersey: Boulder, University of Colorado, unpublished 

Master's thesis, 61 p.

Major, D.W., Hodgins, E.W., and Butler, B.J., 1991, Field 

and laboratory evidence of in situ biotransformation of 

tctrachloroethene to ethcne and ethane at a chemical 

transfer facility in North Toronto, in Hinchce, R.E., and 

Olfenbuttel, R.F., eds., On-site bioreclamation processes for xcnobiotic and hydrocarbon treatment: 

Boston, Butterworth-Heinemann, p. 147-171.

Martin, Mary, 1991, Simulation of reactive multispecies 

transport in two-dimensional ground-water-flow 

systems, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey Calif, March 11-15, 1991: U.S. Geological

353
Survey Water-Resources Investigations Report 

91-4034, p. 698-703.

Sargent, B.P., Fusillo, T.V., Storck, D.A., and Smith, J.A., 

1990, Ground-water contamination in the area of 

Building 24, Picatinny Arsenal, New Jersey: U.S. Geological Survey Water-Resources Investigations Report 

90-4057, 94 p.

Sewell, G.W., and Gibson, S.A., 1991, Stimulation of the 

reductive dechlorination of tetrachloroethene in 

anaerobic aquifer microcosms by the addition of 

toluene: Environmental Science and Technology, v. 25, 

no. 5, p. 982-984.

Smith, J.A., Chiou, C.T., Kammer, J.A., and Kile, D.E., 

1990, Effect of soil moisture on the sorption of trichloroethylene vapor to vadose-zone soil at Picatinny 

Arsenal, New Jersey: Environmental Science and 

Technology, v. 24, no. 5, p. 676-682.

Voronin, Lois, 1991, Simulation of ground-water flow at 

Picatinny Arsenal, New Jersey, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the

technical meeting, Monterey, Calif., March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 713-720.

Wilson, B.H., 1989, Biotransformation of chlorinated 

hydrocarbons and alkylbenzenes in aquifer material 

from Picatinny Arsenal, New Jersey, in Mallard, G.E., 

and Ragone, S.E., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Phoenix, Ariz., September 26-30, 

1988: U.S. Geological Survey Water-Resources 

Investigations Report 88-4220, p. 389-394.

Wilson, B.H., Ehlke, T.A., Imbrigiotta, T.E., and Wilson, 

J.T., 1991, Reductive dechlorination of trichloroethylenc in anoxic aquifer material from Picatinny Arsenal, 

New Jersey, in Mallard, G.E., and Aronson, D. A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, Calif., March 11-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 

91-4034, p. 704-707.

354
Surfactant-Enhanced Remediation of Ground Water at 

Picatinny Arsenal, New Jersey

By James J. Deitsch 1 and James A. Smith 1

Abstract

The efficiency of pump-and-treat remedial 

systems is limited by the slow, kinetic mass 

transfer of organic contaminants from soil to 

water. The scope of this research is to determine 

whether the addition of the nonionic surfactant, 

Triton X-100, to the aqueous phase can artificially 

increase the rate of trichloroethene (TCE) mass 

transfer from soil to water at Picatinny Arsenal. 

The rate of TCE mass transfer (desorption) from 

soil to water is modeled as the product of a masstransfer coefficient and a concentration gradient. 

Laboratory experiments were conducted using 

continuous-flow stirred tank reactors (CFSTR's), 

with and without Triton X-100, to study the effect 

of the surfactant on the rate of desorption of TCE 

from soil to water. Preliminary results indicate 

that the rate of desorption is increased by the 

presence of Triton X-100 in the aqueous phase. 

Two possible mechanisms are responsible for the 

increased desorption rate: First, the addition of 

Triton X-100 above its critical micelle 

concentration can increase the apparent water 

solubility of TCE and thus increase the 

concentration gradient between the sorbed and 

aqueous phases. Second, Triton X-100 can 

increase the mass-transfer coefficient. Future 

experiments are planned to determine the precise 

mechanism affecting the rate of desorption.

'Department of Civil Engineering and Applied Mechanics 

University of Virginia, Charlottesville, Virginia 22903-2442

INTRODUCTION

Picatinny Arsenal is located in a northeastsouthwest trending valley in Morris County, New 

Jersey. During 1960-81, Building 24 (fig.l) was the 

site of metal-plating and degreasing operations that 

used trichloroethene (TCE) as the degreasing solvent. 

Wastewater from these operations was discharged into 

an overflow dry well and two unlined lagoons adjacent 

to Building 24. As a result of these activities, the 

unconfined, sand-and-gravel aquifer beneath the site is 

contaminated with TCE. The plume of ground-water 

contamination is indicated in figure 1. Details of the 

site's hydrogeology and history of ground-water 

contamination have been given by Imbrigiotta and 

others (1996).

Remediation of TCE-contaminated ground 

water at Picatinny Arsenal was begun in September 

1992 with the implementation of a pump-and-treat 

system. The system employs five withdrawal wells 

which extract ground water from the subsurface at a 

combined rate of approximately 150 gal/min. TCE is 

removed from the extracted water by two air-stripping 

towers and an activated carbon filter. The water is then 

returned to Green Pond Brook. The locations of the 

withdrawal wells are shown in figure 1.

Pump-and-treat remediation is a commonly 

used technology for the remediation of contaminated 

ground-water systems. However, the overall 

effectiveness of the technology is limited by the slow, 

kinetic mass transfer of organic contaminants from 

soil to water (Smith and others, 1991; Pavlostathis and 

Mathavan, 1992; Di Cesare and Smith, 1993). The 

removal of contaminated ground water and subsequent 

replacement with upgradient water induces a 

contaminant-concentration gradient between the 

sorbed and aqueous phases. The concentration 

gradient causes the sorbed organic contaminants to 

desorb slowly from the soil organic matter and

355
X*

EXPLANATION

: Tricbloroetbene concentrations greater than 10 

1 micrograms per liter. October-November 1991 0

Pump-and-treal system withdrawal well Q 

Soil sampling location

400 FEET

100 METERS

Figure 1. Map of study area showing building 24, region of 

trichloroethene concentations greater than 10 micrograms 

per liter, remedial withdrawal wells, and the soil sampling 

location.

contaminate the inflowing ground water. Therefore, 

pump-and-treat remedial systems may have to operate 

for tens of years before a large fraction of the total 

subsurface contamination is removed.

This paper presents preliminary data that will be 

used to determine whether the rate of TCE mass 

transfer from soil to water at Picatinny Arsenal can be 

artificially increased by the addition of a nonionic 

surfactant, Triton X-100, to the aqueous phase.

MATERIALS AND METHODS

The soil sample used in this study was collected 

from Picatinny Arsenal from a composite depth of 3 to 

6 ft at the location indicated in figure 1. The organiccarbon content of the sample is 31 percent. 500 fiCi of 

[ I4C] TCE (specific activity equal to 6.2 mCi/mmol) 

was obtained and mixed with nonradioactive TCE to 

yield a net volume of 6 mL pure liquid. The resultant 

chemical and radiochemical purity of the radioisotope 

was greater than 98 percent.

In preparation for the experiment, the soil was 

air dried in a fumehood for 24 hours and then heated 

for 24 hours at 105°C. These precautions were taken to 

reduce the presence of TCE on the soil. The trace 

amounts of TCE remaining on the soil do not affect the 

results of the experiment. After heating, the soil was 

passed through a #10 sieve (2-mm opening) to insure 

that the soil subsamples would be homogeneous.

The kinetic desorption of TCE from soil to 

water was studied using a continuous-flow stirred tank 

reactor (CFSTR) that consisted of a glass column with 

an inside diameter of 1.0 cm and a length of 15 cm. 

The nominal volume was approximately 12 cm3 . A 

Teflon O-ring shield provided a leak-free seal and a 

chemically inert surface. The bed supports were also 

Teflon and had 20 |im openings. All surfaces inside 

the CFSTR were either glass or Teflon. Each CFSTR 

contained approximately 5 g of soil.

The inflow and outflow junctions of each 

CFSTR were prepared identically. 5-cm lengths of 

Teflon tubing were connected to the inflow and 

outflow ports of the CFSTR. The thread adapters, nuts, 

and ferrules formed an airtight seal. The opposite ends 

of the Teflon tubes were connected to separate twoway Teflon valves. A l-m length of silicone rubber 

tubing was attached to the inflow valve. The outflow 

valve was connected to a 30-cm length of Teflon 

tubing. After all the hydraulic junctions of the system 

were sealed, 10 mL of distilled water was poured into 

the CFSTR and the end fitting was closed.

The CFSTR was fastened to a rotary shaker. The 

inflow tubing was run through a peristaltic pump, and 

the end of the tube was immersed in a reservoir of 

distilled water containing 200 mg/L sodium azide. The 

addition of sodium azide eliminates any 

microbiological activity in the CFSTR (Pavlostathis 

and Mathavan, 1992). The CFSTR was then shaken at 

150 revolutions per minute (rpm) continuously to 

insure a completely mixed system. Distilled water was 

pumped through the CFSTR at a rate of 12 mL/hr for 

3 days. This procedure was followed for three reasons: 

First, air bubbles in the CFSTR were removed, thus 

limiting the potential for volatilization after TCE is 

injected. Second, small amounts of soil exit the 

CFSTR during early flow periods. After the third day, 

little or no additional soil exits the CFSTR. Thus, the 

mass of soil in the CFSTR remains constant after 

3 days, allowing the exact mass to be calculated. 

Third, the 3-day flow period insures that the system is

356
flowing at a constant rate before injecting the TCE. A 

total of six CFSTR's were prepared in this manner.

After 3 days, the pumping was discontinued and 

the valves closed. The thread adapters at the effluent 

ports were removed. A 25 j^L syringe was used to inject 

a known volume of the [ 14C]-labeled TCE solution into 

each of the six CFSTR's. The thread adapters were then 

replaced. The CFSTR's were continuously shaken at 

150 rpm for 1 week. At the end of the equilibration 

period, the CFSTR's were prepared for sampling.

The CFSTR's were separated into two groups of 

three. A solution of distilled water and 200 mg/L 

sodium azide was the reservoir for group one. The 

second group used a solution of distilled water, 

200 mg/L sodium azide, and 300 mg/L Triton X-100 

surfactant. The critical micelle concentration (CMC) 

of Triton X-100 is 140 mg/L (Kile and Chiou, 1989). 

The two solutions were pumped through the columns 

at flow rates of approximately 2 mL/hr. (One column 

flowed at 6 mL/hr because of an equipment problem).

Approximately 0.5 mL of effluent from the 

CFSTR's was collected periodically in 7-mL 

scintillation vials. The exact amount of effluent 

captured was determined gravimetrically. The 

radioactivity of the samples was measured with a 

Packard Tri-Carb 1900CA liquid-scintillation 

analyzer, and the corresponding TCE concentrations 

were calculated with a standard curve relating 

concentration to counts per minute. During periods 

when no samples were collected, the effluent was 

collected in glass tubes to quantify the flow rate . The 

sampling was conducted for 7 days.

A CFSTR was prepared with no soil to verify 

that no loss of TCE occurred during the sampling 

period, and to verify that the system was completely 

mixed. A known volume of [C 14]-labeled TCE was 

injected and allowed to equilibrate for 1 week. The 

CFSTR was then sampled continuously for 8 hours 

using the described procedure.

BACKGROUND THEORY

The governing equation for the conservation of 

mass for the aqueous phase in the CFSTR, assuming a 

completely mixed system, is as follows:

i/-~i

V = CHC C^+rV 

w dt in a '

(1)

where Vw is the volume of water in the reactor (in 

mL), Q is the flow rate through the CFSTR (in mL/hr),

Cjn is the inflow solute concentration (in mg/L), C is 

the aqueous phase concentration throughout the reactor (in mg/ L), ra is the reaction term (in mg/Lhours ), 

and dC/dt is the time rate of change of the solute concentration. In this study, TCE is the solute being 

modeled and desorption is assumed to be the only 

reaction component.

The desorption of TCE is represented by a firstorder kinetic mass-transfer relation. This relation can 

be expressed mathematically as

dC

-s =-K(C-KdC), (2)

where Cs is the sorbed solute concentration (in mg/g), 

K is the mass-transfer coefficient (in hr" 1 ), Kd is the 

partition coefficient (L/g), and dCs/dt is the mass of 

solute desorbed per gram of soil per hour (in 

mg/gxhours). The rate of desorption is directly proportional to K and the chemical potential gradient 

between the sorbed and aqueous phases.

Equation 2 can be multiplied by Ms, the mass of 

soil in the CFSTR (g), and divided by Vw. The 

resulting equation is substituted into equation 1 to 

account for the desorption of TCE. When Cjn is zero, 

as in the experiments reported in this paper, the 

following equation results:

dt V

(3)

where all terms have been described previously. Equation 3 can be solved analytically when desorption is 

not present, yielding

C = (4)

where C0 is the initial concentration (in mg/L). 

Equation 4 quantifies the effluent TCE concentration 

from the CFSTR with no soil. The effluent concentrations obtained from this experiment are compared to 

concentrations generated by equation 4 in the Results 

section of this paper.

A fourth-order Runge-Kutta method was used to 

numerically solve equation 3 with desorption present. 

The only unknown in the system of equations is K. By 

finding a functional form of K to fit the experimental 

data, the value of K was quantified for each CFSTR. 

The resulting values of K were compared to determine 

whether the presence of the surfactant affected the rate 

of desorption.

357
 Analytical solution

x Laboratory measurement

200 250 300 

TIME, IN MINUTES

Figure 2. Trichloroethene concentration curves obtained 

from laboratory measurements and the analytical solution of 

equation 4.

RESULTS

The decline of aqueous TCE concentrations 

over a period of 8 hr for the CFSTR with no soil 

present is shown in figure 2. According to equation 4, 

the decrease of aqueous TCE concentrations follows 

an exponential decline if the flow rate and volume are 

constant. Both conditions were met. The TCE 

concentrations observed by sampling the CFSTR 

matched the TCE concentrations generated by the 

analytical solution. In fact, an exponential regression 

analysis of the data provided a correlation coefficient, 

r, of-0.9799. Based on the strong correlation, two 

important conclusions are drawn. First, the assumption 

of a completely mixed reactor is justified by the 

experimental results. Second, because the observed 

TCE concentrations do not lie below the curve 

produced by equation 4, the sampling procedure 

appears to retain all of the TCE in the sample effluent.

The numerically fitted aqueous TCE 

concentration curves for the six CFSTR's with soil are 

shown in figure 3. The independent variable is the 

number of pore volumes that have flowed through the 

CFSTR's. The number of pore volumes is defined as 

the overall volume of flow divided by the volume of 

water in the reactor, Vw. In this report, Vw is 10 mL. 

The use of pore volumes facilitates the comparison of 

CFSTR's with different flow rates. The accuracy of the 

numerical solutions in describing the data of two

20 25 30 35 40 45 

PORE VOLUMES

Figure 3. Numerically generated trichloroethene 

concentration curves for the six continuous-flow stirred tank 

reactors with soil.

CFSTR's is shown in figure 4. The numerical solutions 

applied to the remaining four CFSTR's also closely 

matched the laboratory data. The TCE concentrations 

shown in figure 3 rapidly decline in all six CFSTR's 

during the first five pore volumes. However, the rate 

of decline of TCE concentrations in the CFSTR's with 

Triton X-100 decreases after five pore volumes. In 

contrast, the rapid rate of decline in the three CFSTR's 

without Triton X-100 continues until 10 pore volumes. 

After 10 pore volumes, the rate of decline of the TCE 

concentrations is slightly higher for Triton X-100 

CFSTR's. This continues until 40 pore volumes, where 

the six curves become essentially identical.

The relation shown by the TCE concentration 

curves in figure 3 is directly related to the rate of 

desorption in the six CFSTR's. According to 

equation 3, the rate of decrease of the aqueous TCE 

concentrations is a function of dilution and the rate of 

desorption; dilution increases the rate of decrease, 

whereas, desorption decreases the rate of decrease. 

Because dilution is standardized for all six CFSTR's 

by using the number of pore volumes as the 

independent variable, the deviations in figure 3 are 

caused by differences in desorption rates. Because the 

TCE concentrations are nearly identical for the first 

five pore volumes, the rates of desorption in the six 

CFSTR's are similar in this time interval. However, 

after five pore volumes, the rate of desorption in the

358
250

-.200

150

100

1 

uj

I

t 50

o Without Triton X-100: Numerical solution

e With Triton X-100: Numerical solution

 Without Triton X-100: Laboratory measurement

+ With Triton X-IOfc Laboratory measurement

10 15 20 25 

PORE VOLUMES

30 35

Figure 4. Numerical solutions fitted to match laboratory 

data obtained from two continuous-flow stirred tank 

reactors.

CFSTR's with Triton X-100 is greater than in the 

CFSTR's without Triton X-100. This accounts for the 

higher observed TCE concentrations in the CFSTR's 

with Triton X-100 than in the CFSTR's without Triton 

X-100. Triton X-100 increases the rate of TCE 

desorption from the soil to water relative to results of 

experiments without Triton X-100, as shown in 

figure 3.

The curves generated in figure 3 use a timevarying mass-transfer coefficient. The experimental 

data could not be described adequately with a constant 

mass-transfer coefficient. A scheme using two 

different constant values also was investigated but 

proved to be unsuccessful. The results of this study 

indicate that the rate of desorption varies with time.

Finally, the mass fraction of TCE removed from 

the soil phase as a function of pore volumes of flow for 

the six CFSTR's is shown in fkunx- 5. Using 35 pore 

volumes as a reference point, the mass fractions of 

TCE desorbed from the soil by the CFSTR's with 

Triton X-100 are 0.64, 0.72, and 0.72. In comparison, 

the fraction of TCE desorbed from the soil by the 

CFSTR's without Triton X-100 after 35 pore volumes 

of flow are 0.41, 0.54, and 0.58. On average, the 

CFSTR's with Triton X-100 removed 18.33 percent 

more TCE after 35 pore volumes than did the CFSTR's 

without Triton X-100. It should be noted that even

0.8

0.7

0.6

0.4

g 0.3

S.

2 0.2

90.1

With Triton X-100

Without Triton X-100

10 15 20 25 30 

PORE VOLUMES

35

Figure 5. Mass fraction of trichloroethene removed from 

the soil as a function of pore volumes of flow for the six 

continuous-flow stirred tank reactors.

though the equilibration time of this experiment was 

just 1 week, only 51 percent of the TCE on average 

was desorbed from the soil in the CFSTR's without 

Triton X-100 after 35 pore volumes. The desorption of 

TCE from soil to water is kinetically limited even for 

short equilibration periods.

DISCUSSION

Two possible mechanisms are responsible for 

the higher desorption rate in CFSTR's containing 

Triton X-100 relative to the rate in CFSTR's without 

Triton X-100. The first mechanism is well documented 

(Kile and Chiou, 1989; Kile and Chiou, 1990); the 

addition of a surfactant above its CMC can increase a 

solute's apparent water solubility. By increasing a 

solute's apparent water solubility, the sorption 

coefficient (Kd) of the compound is decreased. 

Analysis of equation 2 shows that, as Kd decreases, the 

concentration gradient between the sorbed and 

aqueous phases increases, therefore increasing the rate 

of desorption.

Second, the addition of Triton X-100 to the 

CFSTR may increase the mass-transfer coefficient, K. 

The addition of Triton X-100 to the water decreases 

surface tension. A decrease in surface tension allows 

the water to fill and swell the organic polymer matrix

359
to a greater extent than under normal surface tensions. 

By swelling the organic matrix, it is possible that 

diffusive paths which previously were not passable are 

now open as pathways for the TCE to desorb.

In the analysis of the previous results, Kd was 

assumed to remain constant in both systems. The 

differences in desorption rate were assumed to be the 

result of an alteration of the mass-transfer coefficient. 

At this point, the effect of Triton X-100 on the Kj of 

TCE has not been quantified. However, future 

experiments are planned to investigate the effects of 

the surfactant on Kd . Once the relation between Kj 

and surfactant concentration is determined, the 

mechanism responsible for the increased desorption 

rate can be identified.

Other planned experiments include 

determination of the effect of residence time on the 

desorption of TCE using Triton X-100. The 1-week 

equilibration period for this study clearly shows a 

kinetic effect on the rate of desorption. It is expected 

that equilibration times longer than 1 week will show 

stronger kinetic effects. By performing experiments 

with equilibration times longer than 1-week, the 

effectiveness of Triton X-100 and other surfactants on 

long-term contaminated soils can be evaluated.

SUMMARY

Preliminary experimental results indicate that 

the addition of the nonionic surfactant Triton X-100 is 

effective in artificially increasing the rate of TCE mass 

transfer from the soil to the aqueous phase. Two 

mechanisms are potentially responsible: First, the 

mass-transfer coefficient could have been increased by 

the addition of Triton X-100. Second, the

concentration gradient could have been increased by a 

reduction of Kj because of the presence of Triton 

X-100. The precise mechanism for this increased 

desorption rate is still under study.

REFERENCES

Di Ccsarc, D., and Smith, J.A., 1993, Effects of surfactants 

on the desorption rate of nonionic organic compounds 

from soil to water : Reviews of Environmental 

Contamination and Toxicology, v. 134, p. 1-29.

Imbrigiotta, T.E., and Martin, Mary, 1996, Overview of 

research at Picatinny Arsenal, New Jersey , 1991- 

1993: in Morganwalp, D.W., and Aronson, D.A., 

eds.,U.S. Geological Survey Toxics Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 20- 

24, 1993, Water Resources Investigations Report 94- 

4015.

Kile, D.E., and Chiou, C.T., 1989, Water solubility

enhancements of DOT and trichlorobenzene by some 

surfactants below and above the critical micelle 

concentration: Environmental Science and 

Technology, v. 23, no. 7, p. 832-838.

Kile, D.E., and Chiou, C.T., 1990, Effect of some petroleum 

sulfonate surfactants on the apparent water solubility 

of organic compounds: Environmental Science and 

Technology, v. 24, no. 2, p. 205 -208.

Pavlostathis, S.G., and Mathavan, G.N., 1992, Desorption 

kinetics of selected volatile organic compounds from 

field contaminated soil: Environmental Science and 

Technology, v. 26, no. 3, p. 532-538.

Smith, J.A., Tuck, D.M., Jaffe, P.R., and Mueller, R.T., 

1991, Effects of surfactants on the mobility of 

nonpolar organic contaminants in porous media: in 

Organic Substances and Sediments in Water, Volume I, 

R. Baker, ed., (Chelsea , Mich.) Lewis Publishers, 

p. 201-230.

360
Sampling Borehole Flow to Quantify Aquifer 

Cross-Contamination by Volatile Organic Compounds

By Ronald A. Sloto1

Abstract

A combination of borehole geophysical 

methods, measurements of vertical borehole flow, and 

analyses of borehole-fluid samples were used to assess 

the extent of aquifer cross-contamination in the Stockton Formation in Hatboro, Pennsylvania. The Stockton 

consists of interbedded nonmarine sandstone and siltstone-mudstone. Most industrial, public-supply, and 

monitoring wells drilled into the Stockton Formation 

are constructed as open holes with short casings and 

are open to multiple water-bearing zones. Caliper, 

fluid-resistivity, fluid-temperature, natural-gamma, 

and single-point-resistance logs were run in 19 boreholes 149 to 656 feet deep to locate water-bearing 

fractures and determine zones of vertical boreholefluid movement. The direction and rate of vertical 

borehole-fluid movement was measured by injecting a 

slug of high-conductance fluid at a specific depth in 

the borehole and monitoring the movement of the slug 

with the fluid-resistivity tool. After intervals of borehole flow were determined, samples of moving fluid 

were extracted from nine boreholes at a rate less than 

that of the measured borehole flow and analyzed for 

volatile organic compounds. An estimated 80.9 kilograms per year of volatile organic compounds were 

moving downward through the sampled boreholes 

from the contaminated, upper part of the aquifer to the 

lower part, which is tapped by public-supply wells. 

Trichloroethylene accounts for 94 percent and 

1,1,1-trichloroethane accounts for 3 percent of the 

compounds.

U.S. Geological Survey, Malvern, Pa.

INTRODUCTION

Many public supply, industrial, and monitoring 

wells in the United States are completed as open holes 

that obtain water from several formations or from several water-bearing zones in a single formation. The 

advantage to this construction practice is that a much 

greater yield can be obtained than if a well is open to a 

single formation or water-bearing zone. The disadvantage is that these boreholes, which connect several 

aquifers or water-bearing zones, commonly short circuit the ground-water-flow system and act as conduits 

for the transport of contaminants. The existence and 

magnitude of this problem is largely undetermined.

This paper presents some of the results of a 

study done by the U.S. Geological Survey (USGS) in 

cooperation with the U.S. Environmental Protection 

Agency (USEPA) to identify and assess cross-contamination by abandoned multiaquifer boreholes. A combination of borehole geophysical methods, 

measurements of vertical borehole flow, and analyses 

of borehole-fluid samples were used to assess the 

extent of aquifer cross-contamination in the Stockton 

Formation in Hatboro, Pa. The purpose of this paper is 

to show how those techniques were used to identify 

and determine the magnitude of cross-contamination.

Location and Description of Study Area

The study area is part of Hatboro Borough in 

Montgomery County, Pa. (fig. 1). It is centered around 

an older industrial area typical of many other communities in southeastern Pennsylvania. Much of the 

industry dates back to or prior to the World War II era. 

Industrial and public-supply wells drilled into the 

Stockton Formation were constructed as open holes 

with short casings and are open to multiple waterbearing zones. Downward fluid movement in

361
75°07' 75°05'

40°11'

40°10'

3000 4,000 5,000 FEET

1,000 METERS

CONTOUR INTERVAL 10 FEET 

NATIONAL GEODETIC DATUM OF 1929

Base from U.S. Geological Survey Hatboro 1:24,000,1983

EXPLANATION

 1236 BOREHOLE AND IDENTIFICATION NUMBER (Prefix MG

omitted from number) 

O 1221 MONITORING WELL CLUSTER AND IDENTIFICATION

NUMBER (Prefix MG omitted from number)

Figure 1. Location of boreholes and monitoring-well clusters.

362
boreholes in the Stockton Formation was identified 

using borehole geophysical methods in the township 

north of Hatboro (Sloto and Davis, 1983).

The study area includes the Raymark site, which 

has been designated as an National Priorities List (NPL) 

site by the USEPA (fig. 1). The Raymark NPL site has 

been the location of a metal fabrication shop since 1948. 

Solvent containing trichloroethylene (TCE) was used in 

the manufacturing process to clean and degrease metal 

parts. Over several decades of operation, TCE was 

introduced into the environment at the location of a solvent storage tank, a metal degreaser, and four disposal 

lagoons. Other industrial facilities in the vicinity, which 

included manufacturers of kitchen ranges, toys, television picture tubes, and metal parts, also used solvents in 

their manufacturing processes.

Geological Setting

Hatboro is underlain by the Upper Triassic 

Stockton Formation. The Stockton is the basal unit in 

the Triassic-Jurassic Newark basin, which contains 

16,000 to 20,000 ft of nonmarine sedimentary rocks of 

the Newark Supergroup. Following deposition, the 

sediments were tilted, faulted, uplifted, and eroded. 

The Stockton Formation is about 6,000 ft thick and 

includes alluvial fan deposits, fluvial and lacustrine 

sandstones, and fluvial and near-shore lacustrine mudstones and siltstones (Turner-Peterson and Smoot, 

1985). Thick, poorly defined upward fining cycles 

possibly were deposited by large, perennial, meandering rivers. One of the characteristics of the Stockton is 

thick-bedded to locally massive arkosic sandstones.

In the Hatboro area, the Stockton Formation 

strikes approximately N. 65° E. and dips approximately 9° NW. The rocks are chiefly interbedded 

arkosic sandstone and siltstone-mudstone. Grain size 

changes both laterally and vertically. The rocks that 

underlie the study area belong to the middle arkose 

member (Rima and others, 1962), but boreholes may 

penetrate the underlying lower arkose member.

Hydrological Setting

The rocks of the Stockton Formation form a complex, heterogeneous, multiaquifer system. This aquifer 

system is comprised of a series of gently dipping sandstone and siltstone-mudstone beds that form lithologic 

units, a few inches to tens of feet thick; each lithologic 

unit has different hydraulic properties. Secondary permeability is dominant. In general, the sandstone beds

are the principle water-bearing units, but some of the 

finer-grained beds may contain water-bearing zones. 

Because of the softness and fine grain size of the siltstone-mudstone beds, water-bearing openings tend to 

clog. In addition, the soft siltstone-mudstone beds 

deform without breaking under stress. Harder sandstone 

beds tend to develop joints and are more permeable.

All of the lithologic units are hydraulically connected. Ground water in units at or near the land surface 

is under unconfined conditions; ground water in the 

deeper units may be confined or semiconfined. Most 

deep boreholes penetrate several water-bearing zones 

and are multiaquifer boreholes. Each water-bearing 

zone usually has a different hydraulic head. The hydraulic head in a deep borehole is the composite head of the 

several water-bearing zones it penetrates. In boreholes 

penetrating several water-bearing zones, borehole fluid 

will flow in the direction of lower head. Water-level 

measurements made at the MG-1222 borehole cluster 

on August 5, 1991, showed a water level 36.48 ft below 

land surface for the open interval 33 to 63 ft below land 

surface, a water level 49.91 ft below land surface for the 

open interval 118 to 138 ft, and a water level 55.85 ft 

below land surface for the open interval 185 to 195 ft. 

Water moves downward through the aquifer in response 

to this downward head gradient, which is caused in 

part by the pumping of deep public supply wells.

MEASUREMENT OF BOREHOLE FLOW

Caliper, fluid-resistivity, fluid-temperature, 

natural-gamma, and single-point-resistance logs were 

run in 19 boreholes (fig. 1) 149 to 656 ft deep to 

interpret lithostratigraphy, locate water-bearing fractures, and determine zones of vertical borehole-fluid 

movement. A complete suite of geophysical logs from 

MG-1222, the deepest borehole in a three-borehole 

cluster, is shown on figure 2. Construction of a threedimensional lithostratigraphic model for the study area 

using caliper, natural-gamma, and single-pointresistance logs was described by Sloto and others (1992).

Correlation of caliper logs with single-pointresistance, fluid-resistance, and fluid-temperature logs 

was used to identify fractures and water-producing and 

water-receiving zones. Caliper logs were used to identify fractures and possible water-bearing openings. 

Not all fractures shown on the caliper log of borehole 

MG-1222 (fig. 2), for example, are water-bearing fractures. Single-point-resistance logs were sometimes 

helpful for identification of water-bearing zones.

363
HOLE DIAMETER 

IN INCHES

NATURAL-GAMMA

COUNTS PER

SECOND

SINGLE-POINT

RESISTANCE

IN OHMS

6 8 1000 4000 5000 6000 7000 8000

0

£ 75 -

TEMPERATURE

IN DEGREES

CELCJUS

FLUID

RESISTIVITY 

IN OHM-METERS

13 13.5 H 14.5 15 5000 5500 6000 6500 7000 

0 i I i i , 0

Figure 2. Caliper, natural-gamma, single-point resistance, fluid-temperature, and fluid-resistivity logs from borehole MG-1222.

364
Fluid-resistivity and fluid-temperature logs were 

used to identify water-producing and water-receiving 

zones and to determine intervals of vertical borehole 

flow. Water-producing and water-receiving zones were 

identified by sharp changes in resistivity. Intervals of 

borehole flow were identified by a low resistivity gradient between water-producing and water-receiving 

zones. The fluid resistivity log from borehole MG1222 (fig. 2) shows sharp changes in fluid resistivity at 

83, 128, and 189 feet below land surface, indicating 

water-producing or water-receiving zones. Fluid-temperature logs from wells with no borehole flow generally show a decrease in fluid temperature with depth in 

the upper part of the borehole and an increase in fluid 

temperature with depth as a function of the geothermal 

gradient in the lower part of the borehole. Intervals of 

vertical borehole flow were identified by little or no 

temperature gradient. Intervals of nearly constant temperature in borehole MG-1222 (fig. 2) at depths of 50 

to 80, 80 to 130, and 130 to 192 ft below land surface 

indicate intervals of borehole flow.

Upon completion of geophysical logging, the 

suite of logs was evaluated to choose zones of potential borehole flow. The direction and rate of vertical 

borehole-fluid movement was then determined by 

injecting a slug of high-conductance fluid at a specific 

depth in the borehole and monitoring the movement of 

the slug with the fluid-resistivity tool. This is the 

brine-tracing method described by Patten and Bennett 

(1962). The lower limit of flow measurement is about 

0.5 gal/min in a 6-in.-diameter borehole. Borehole 

flow was calculated by:

INCREASING RESISTIVITYQ = 7.481 (1)

where Q is borehole flow, in gallons per minute;

V is the rate of vertical borehole fluid movement, in feet per minute; and 

r is the radius of the borehole, in feet.

For borehole MG-1222, slugs of high-conductance fluid were injected at 85, 95, 120, and 175 ft 

below land surface. Downward flow at the rate of 2.3, 

2.6, 2.8, and 3.4 gal/min, respectively, was measured. 

Figure 3 shows the movement of the slug of high-conductance fluid injected 175 ft below land surface. The 

slug moved downward at 2.3 ft/min, which corresponds to a flow rate of 3.4 gal/min.

High-conductance-fluid slugs were injected at 

different depths in 18 boreholes to determine the 

direction and rate of vertical borehole flow. As many

w 180

01234 

TIME SINCE INJECTION, IN MINUTES

Figure 3. Movement of a high-conductance fluid slug 

injected 175 feet below land surface in borehole MG-1222.

as eight slugs were injected in a single borehole. Fluid 

movement at rates up to 14 gal/min was measured 

(table 1). The predominant direction of fluid movement was downward. Downward borehole flow was 

measured in 15 boreholes. Generally, fluid moves 

downward from fractures in both the siltstone-mudstone beds and sandstone beds in the upper part of the 

aquifer, which is unconfined, to fractures in the sandstone beds in the lower part of the aquifer, which is 

semiconfined or confined. Both upward and downward flow was measured in borehole MG-220; flow 

was downward in the upper part of the borehole and 

upward in the lower part of the borehole. No measurable flow was detected in boreholes MG-487 and MG488, which are about 50 ft apart. The direction and rate 

of borehole flow was not determinable for borehole 

MG-212 because of the large diameter (14 in.), rough 

borehole wall, and possible horizontal flow through 

the formation.

SAMPLING BOREHOLE FLOW

After intervals of borehole flow were determined, a sample of the moving fluid was extracted 

with a nitrogen-driven bladder pump at a rate less 

than that of the measured borehole flow. Samples 

were collected from nine boreholes (MG-220, 951, 

1236, 1237, 1238, 1239, 1240, 1241, and 1242; see 

fig. 1) and analyzed for volatile organic compounds 

(VOCs) by the USGS National Water-Quality 

Laboratory in Arvada, Colo. None of the boreholes 

sampled were on the Raymark NPL site. For boreholes with downward flow, these samples represent 

water moving from the shallow to the deeper part of 

the aquifer.

365
Table 1. Borehole-flow measurements

Borehole identification number

MG-212

MG-220

MG-295

MG-487

MG-488

MG-951

MG-1II4

MG-1221

MG-1222

MG-1223

MG-1235

MG-1236

MG-1237

MG-1238

MG-1239

MG-1240

MG-1241

MG-1242

MG-1283

Logged 

depth 

(feet)

244

470

656

510

186

323

149

195

200

196

143

122

169

200

166

159

174

178

300

Purpose for which 

borehole 

was drilled

Public supply

Public supply

Industrial supply

Industrial supply

Industrial supply

Public supply

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Monitoring

Direction of 

borehole flow

Not determinable

Down and up

Down

None

None

Down

Down

Down

Down

Not measured

Down

Down

Down

Down

Down

Down

Down

Down

Down

Maximum measured flow 

(gallons per minute)

Not deterrninable

4.1 (Down) 

.8 (Up)

3.0

0

0

2.0

5.7

4.3

3.4

Not measured

<.5

14

1.6

.9

.7

.7

1.5

.7

6.1

VOCs detected in borehole flow are summarized in table 2. Fluid from all boreholes sampled contained detectable concentrations of VOCs. Of 37 

VOCs analyzed for, 19 (51 percent) were detected. 

Trichloroethylene (TCE) and tetrachloroethylene 

(PCE) were detected in samples from all boreholes. 

Chloroform, 1,1,1-trichloroethane (TCA), and trans1,2-dichloroethylene (DCE) were detected in eight of 

nine boreholes sampled. The source of chloroform in 

ground water may be leaking water-distribution and 

sewer lines in the study area. Some compounds, such 

as trans-DCE, 1,1 -dichloroethane (DCA), and vinyl 

chloride (VC) were never used in the study area and

are degradation products of other VOCs. Under anaerobic conditions, PCE successively degrades by reductive dechlorination to TCE, cis-DCE or trans-DCE, 

1,1 -dichloroethene, and VC (Parsons and others, 1984; 

Vogel and McCarty, 1985; Freeman and Gossett, 

1989). TCA anaerobically degrades to DCA. Concentrations of TCE, TCA, and 1,1-dichloroethylene in 

borehole-fluid samples as great as 5,800 fig/L, 1,400 

(4,g/L, and 260 p,g/L, respectively, indicate that some of 

the water moving downward in the aquifer through 

open boreholes is highly contaminated and that open 

boreholes may contribute substantially to groundwater contamination.

366
Table 2. Volatile organic compounds detected in borehole-flow samples

[Concentrations given in micrograms per liter]

Compound

Benzene

Bromoform

Carbon tetrachloride

Chlorodibromomethane

Chloroform

Dichlorobromomethane

1,1-Dichloroethane

1,2-Dichloroethane

1 , 1 -Dichloroethylene

trans- and cis-l,2-Dichloroethylene

1 ,2-Dichloropropane

Methylene chloride

Tetrachloroethylene

Toluene

1,1,1-Trichloroethane

1 , 1 ,2-Trichloroethane

Trichloroethylene

Trichlorofluoromethane

Vinyl chloride

Number of boreholes in 

which compound was 

detected

1

1

4

1

8

2

6

2

6

8

">

3

9

1

8

2

9

1

2

Maximum 

concentration 

detected

0.5

.4

25

.2

15

.4

14

5.6

260

72

60

.4

57

.2

1 ,400

.4

5,800

.9

1.4

CROSS-CONTAMINATION BY VOLATILE 

ORGANIC COMPOUNDS

The rate of borehole-fluid movement is known; 

therefore, concentrations of VOCs in borehole-fluid 

samples can be used to estimate the downward mass 

flux and the extent of aquifer cross-contamination. The 

mass of a VOC moving downward in the aquifer 

through an open borehole annually was estimate based 

on the measured borehole-flow rate and results of 

chemical analysis of the fluid by:

M= 1.99 QC, (2)

where M is the quantity of a compound moving down 

a borehole annually, in grams per year;

Q is the measured borehole flow, in gallons per 

minute; and

C is the concentration of the compound in the 

borehole fluid, in micrograms per liter.

Because most of the wells were sampled once 

and concentrations of VOCs vary temporally, the mass 

flux estimates should be considered an approximation. 

The data can be used to provide an estimate of the 

magnitude of cross-contamination. An estimated 

80.9 kg/yr of VOCs (table 3) were moving downward 

through the nine open boreholes sampled from the 

contaminated, upper part of the aquifer to the lower 

part. Trichloroethylene accounts for 94 percent and 

1,1,1-trichloroethane accounts for 3 percent of the 

compounds.

367
Table 3. Estimated quantity of volatile organic compounds moving downward in the Stockton 

Formation through nine sampled boreholes

Compound Grams per year

Carbon tetrachloride

1.1-Dichloroethane

1.2-Dichloroethane 

Chloroform

1.1-Dichloroethylene

trans- and cis-l,2-Dichloroethylene

1.2-Dichloropropane 

Tetrachloroethylene

1.1.1-Trichloroethane

1.1.2-Trichloroethane

Trichloroethylene

Total

566

203

8.2

35.7

727

246

449

449

2,049

1

75,861

80,851

CONCLUSIONS

A combination of borehole geophysical methods, 

measurements of vertical borehole flow, and analyses of 

borehole-fluid samples provided effective methods to 

identify and assess an aquifer cross-contamination 

problem. Borehole geophysical methods were used to 

identify zones of fluid movement. Borehole-flow 

measurements provided data on the direction and rate of 

borehole-fluid movement. Sampling and analysis of 

moving borehole fluid provided concentrations of 

VOCs. An estimated 80.9 kg/yr of VOCs (table 3) were 

moving downward through the nine open boreholes 

sampled from the contaminated, upper part of the aquifer to the lower part, which is tapped by public-supply 

wells. This magnitude of this flow indicates that crosscontamination caused through open boreholes is a serious problem affecting ground-water supplies.

In response to a verbal presentation of study 

results, the USEPA Superfund Removal Branch 

plugged or reconstructed five of the sampled boreholes 

plus four other boreholes. The remaining four sampled 

boreholes were reconstructed by the USGS so that they 

are now open only to a single water-bearing zone. Wells 

in the six monitoring clusters on the Raymark NPL site 

were constructed or reconstructed by the USEPA so that 

they are open only to a single water-bearing zone. The 

aquifer cross-contamination problem caused by known 

boreholes in the study area has been eliminated.

REFERENCES

Freedman, D.L., and Gossett, J.M., 1989, Biological reductive dechlorination of tetrachloroethylene and trichloroethylene to ethylene under methanogenic 

conditions: Applied and Environmental Microbiology, 

v. 55, p. 2144-2151.

Parsons, Frances, Lange, G.B., and Rice, Ramona, 1984, 

Biotransformation of chlorinated organic solvents in 

static microcosms: Environmental Toxicology and 

Chemistry, v. 4, p. 739-742.

Patten, E.P., Jr. and Bennett, G.D., 1962, Methods of flow 

measurement in well bores: U.S. Geological Survey 

Water-Supply Paper 1544-C, 28 p.

Rima, D.R, Meisler, Harold, and Longwill, Stanley, 1962, 

Geology and hydrology of the Stockton Formation in 

southeastern Pennsylvania: Pennsylvania Geologic 

Survey, 4th sen, Water Resources Report 14, 111 p.

Sloto, R.A., and Davis, O.K., 1983, Effect of urbanization on 

the water resources of Warminster Township, Bucks 

County, Pennsylvania: U.S. Geological Survey WaterResources Investigations Report 82-4020, 72 p.

Sloto, R.A., Macchiaroli, Paola, and Towle, M.T., 1992, Identification of a multiaquifer ground-water cross-contamination problem in the Stockton Formation by use of 

borehole geophysical methods, Hatboro, Pennsylvania: 

Proceedings, Symposium on the Application of Geophysics to Engineering and Environmental Problems, 

Oakbrook, Illinois, April 26-29, 1992, Society of 

Engineering and Mineral Exploration Geophysicists, 

p. 21-35.

Turner-Peterson, C.E. and Smoot, J.P., 1985, New thoughts 

on facies relationships in the Triassic Stockton and 

Lockatong Formations, Pennsylvania and New Jersey: 

U.S. Geological Survey Circular 946, p. 10-17.

Vogel, T.M., and McCarty, PL., 1985, Biotransformation of 

tetrachloroethylene to trichloroethylene, dichloroethylene, vinyl chloride, and carbon dioxide under 

methanogenic conditions: Applied and Environmental 

Microbiology, v. 49, p. 1080-1083.

368
A Sample-Freezing Drive Shoe for a Wireline-Piston

Core Sampler

By Fred Murphy 1 andW.N. Herkelrath 1

Abstract

The loss of fluids and sample during retrieval 

of cores of noncohesive sediments results in incorrect measures of fluid distributions and a less-thanideal measure of the stratigraphic position of the 

sample. In order to reduce these errors, a hollow 

drive shoe was developed that allows freezing in 

place the lowest 75 millimeters of a 48-millimeterdiameter, 1.5-meter-long sediment sample taken 

using a commercial wireline piston-core sampler. 

The end of the core is frozen by piping liquid 

carbon dioxide at ambient temperature through a 

steel tube from a bottle at the land surface to the 

drive shoe where it evaporates and expands, cooling the interior surface of the shoe to about -70 °C. 

The freezing process takes about 10 minutes. This 

device was used to collect samples for a study of oilwater-air distributions in unconsolidated sediments 

at the site of an oil spill near Bemidji, Minnesota. 

Before freezing was employed, samples of sandy 

sediments from near the water table sometimes 

flowed out of the core barrel as the sampler was 

withdrawn. Freezing the bottom of the core 

ensured the retention of all material that entered 

the core barrel and lessened the redistribution of 

fluids within the core.

INTRODUCTION

Because saturated, noncemented, well sorted sands 

have high permeability and low cohesion, collection of 

core samples without losing sediment or pore fluids as 

the sampler is removed from its position after driving 

is difficult. Various devices have been used to reduce 

sample loss, including pistons or remotely operated 

valves that maintain a vacuum above the sample

'U.S. Geological Survey, Menlo Park. Calif.

(Munch and Killey, 1985; Zapico and others, 1987), 

inflatable bladders that expand into and seal the lower 

end of a core barrel (McElwee and others, 1991), movable steel fingers that are driven into the sediment just 

below the core liner (McElwee and others, 1991), and 

sample-freezing jackets (Durnford and others, 1991). 

Munch and Killey (1985) and Zapico and others (1987) 

incorporated a left-hand-threaded coupling between 

the drill string and the core barrel to allow the core 

barrel to be detached from the drill string and retrieved 

rapidly by means of a wireline.

The hollow drive shoe described in this paper (fig. 1) 

enables freezing in place of the lowest 75 mm of a 

48-mm-diameter, 1.5-m-long core after the core sampler has been driven. The hollow drive shoe replaces 

the standard drive shoe of the Solinst wireline-piston 

core sampler (Zapico and others, 1987). Interstitial ice 

in the frozen sediment prevents fluid from flowing 

through it, and friction between the ice-sediment plug 

and the plastic core liner prevents the otherwise freeflowing sand from leaving the core barrel liner. The 

purpose of this paper is to describe the design, fabrication, and field use of the freezing sampler.

BACKGROUND

In 1979, a broken high-pressure pipeline spewed 

about 2x 106 L of crude oil over several thousand square 

meters of sandy glacial outwash deposits near Bemidji, 

Minn. (fig. 2). The impetus for developing the samplefreezing drive shoe was the difficulty we experienced 

in recovering cores of saturated and nearly saturated 

sands at the Bemidji site by means of a simple wirelinepiston core sampler. The core samples are needed to 

measure the distributions of oil, air, and water within 

the sediments in connection with modeling the movement of the oil at the site (Hess and others, 1992; Essaid 

and others, 1993).

369
left-hand thread

core barrel cap

core barrel

polycarbonate 

plastic liner

hard-faced

cutting

shoe

wireline

slotted liner

C02 feed tube

C02 feed tube 

attachment

freezing jacket

piston

Figure 1. A cross-sectional view of the sample-freezing drive shoe attached to the 

Solinst core barrel. The scale is broken-the sampler is about 2-m long and 75-mm in 

diameter.

370
97° 90°

49° - --

44°

c^_c; site of oil spill

. = 

Bemidji

(Minnesota

o

 "~7

160 KILOMETERS V

100 MILES

r1 - 

I V

Figure 2. Location of the oil spill near Bemidji, Minn.

DESCRIPTION OF THE DEVICE

The sample-freezing drive shoe (fig. 3) is a triplewalled steel cylinder that threads onto the Solinst core 

barrel in place of the standard, solid, hardened-steel 

drive shoe. The device is assembled from brazed and 

welded mild-steel parts except for the cutting edge of 

the drive shoe, which is faced with a hard al loy for wear 

resistance.

In order to freeze the bottom end of a core sample, 

liquid carbon dioxide is sprayed into the lower end of 

the inner chamber. The liquid evaporates, depositing 

solid carbon dioxide (dry ice) at about -70 °C in the 

distribution channels cut in the inner wall of the freezing chamber (fig. 3). As the dry ice sublimes, and the 

liquid evaporates, the resulting carbon dioxide gas 

circulates upward through the channels, exhausts 

though holes in the upper end of the interior wall, 

moves through the annulus between the plastic core 

liner and the core barrel, passes from the exterior of the 

liner to its interior through slots in the upper end of the 

liner, and leaves the sampler through the cable hole in 

the core barrel cap.

Flow of carbon dioxide is restricted near the entry 

of the feed tube to the freezing chamber by a 1.5-mmdiameter wire that occupies most of the 1.9-mm bore of 

a 10-cm length of stainless-steel tubing. This restriction ensures that the carbon dioxide remains a liquid

under a pressure of about 6000 kPa within the feed tube. 

Carbon dioxide vaporization and resulting cooling only 

occurs downstream from the restriction within the 

freezing chamber; this cools the sample efficiently and 

prevents ice from binding the feed tube to the sediment.

USE OF THE SAMPLE-FREEZING 

DRIVE SHOE

Use of the freezing sampler in the field is shown 

in figure 4. Carbon dioxide is piped from a bottle 

equipped with a dip tube, through a short length of 

flexible tubing, and then through segments of 1/4-inch 

outside-diameter stainless-steel tubing to the drive 

shoe. The upper pieces of tubing are interchangeable 

1.52-m-long segments, but the two lowest sections of 

tubing are special. The tube that attaches to the drive 

shoe extends without a break from the drive shoe to 

above the core barrel cap; this tube includes a thickwalled, threaded lower section about 10-cm long that 

fits a mating part of the drive shoe. A filter attached to 

the upper end of the tube prevents particles from entering the drive shoe and plugging the flow-restricting 

narrow tube in the shoe. The next segment of tubing 

above the drive shoe has a U-shaped fold that is 

refolded to fit snugly against the drive rod used to 

emplace the sampler. This double bend makes the feed 

tube assembly slightly elastic and so prevents breakage 

of the tube joints when the feed tube scrapes or catches 

on the auger during driving.

In use, the core barrel is assembled with a 1.75-mlong piece of polycarbonate liner fitted between shoulders cut in the drive shoe and in the core barrel cap. The 

piston is passed through the liner and set in the nose of 

the drive shoe, and the core barrel cap is threaded onto 

the sampler. The core sampler is placed in the bore of 

the auger and lowered to the bottom of the hole. The 

1.52-m-long segments of carbon-dioxide feed tubing 

are attached to the sampler as it is lowered. Once the 

sampler is in place at the bottom of the bore, the wireline is pulled taut, and the sampler is hammered 1.52 m 

into the ground.

After the sampler has been driven, the carbon dioxide 

bottle is attached to the feed tube through a short length of 

flexible tubing. The tank valve is opened, and a 1/8-inch 

ball valve is opened to allow liquid carbon dioxide to flow 

to the drive shoe. A hissing or bubbling sound made by the 

escaping gas can usually be heard coming from inside the 

auger. When the drive shoe becomes plugged with solid 

carbon dioxide, the noise stops or becomes intermittent, 

and the valve is closed. Closure of the valve prevents 

possible damage to the drive shoe from the pressure of 

liquid carbon dioxide and reduces unnecessary chilling of

371
(B)

O O

threaded upper 

attachment

C02 feed tube 

attachment

exhaust ports

distribution 

channels

inner sleeve

outer sleeve

cutting shoe

Figure 3. (A) An exploded view of the sample-freezing drive shoe. (B and C) Two 

transparent views of the assembled device, shown rotated 90 degrees with respect to 

each other.

372
flexible tubing

liquid C02 

drive rod

feed tube 

auger

bent tube and 

filter

feed tube 

and core 

barrel

sample-freezing 

drive shoe

Figure 4. A partly transparent view of the emplaced sample-freezing drive shoe as it 

might appear during the freezing operation.

373
the exterior of the drive shoe. After a minute or so, the 

sputtering sound will settle to a steady hiss, or, if the shoe 

has been completely plugged, the hissing will resume. 

The valve can again be opened until the shoe again plugs 

with dry ice. This process is repeated until about 10 minutes have elapsed. The lowest 75 mm of the sample will 

then be frozen.

Although the drive shoe is triple walled and the 

exterior near the freezing jacket is insulated by an air 

gap, ice does adhere to the exterior of the shoe. Sometimes the use of the drill rig's hydraulic cylinders was 

necessary to break the sampler loose from the bore. 

Once the sampler is loose it can be recovered on the 

drill string by means of a winch. The sample end 

remains frozen for several minutes. To date this device 

has been used to collect about 20 cores. The device was 

successfully used to sample at depths more than 2 m 

below the water table.

DISCUSSION

Use of the sample-freezing drive shoe takes more 

time and effort than using a simple wireline-piston core 

sampler. When the core end is frozen, the use of a piston 

is not necessary to retain sediment in the core barrel. 

However, the piston does prevent loose material and 

fluid, particularly oil floating on water in the bore, from 

contaminating the sample. Judging from the preservation of laminae and sharp interfaces in the cores that 

have been retrieved, samples recovered using the sample-freezing drive shoe appear to be no more disturbed 

than those collected by means of the standard drive 

shoe with the Solinst sampler.

Poor sample recovery can have several causes; for 

example, a cobble could jam in and plug the drive shoe 

opening, the material through which the barrel is being 

driven could be compressed or forced aside by the 

drive shoe and not enter the core barrel, or the end of 

the core barrel liner could fold into the interior of the 

drive shoe and prevent sediment from entering. The 

sample-freezing drive shoe ensures only that all 

material that enters a core barrel is retained there and 

lessens the redistribution of fluids in the sample.

The cryogenic sampler described by Durnford and 

others (1991) is similar to the device described here in 

that both samplers use bottled carbon dioxide to freeze 

core samples. That device is described as being capable 

of retrieving 32-mm-diameter, 75-mm-long cores over 

a depth range of 1.2 to 1.8 m, does not employ a piston, 

and is not adaptable to a commercially available core 

sampling tool. The sampler described in this note can 

be used to obtain larger cores (47 mm diameter and 

1.5 m long) from depths of at least 8 m, but it is

374

more complex design than the sampler described by 

Durnford and others (1991) and has a larger outside 

diameter. The sample-freezing drive shoe described 

here requires an auger with an inside diameter of at 

least 90 mm, whereas the sampler described by Durnford could be passed through an auger with a smaller 

bore.

SUMMARY AND CONCLUSIONS

A hollow drive shoe was developed that is compatible with a commercially available core-sampling tool 

that freezes the lowest 75-mm of a core after the sampler is emplaced. In-place freezing of the end of a core 

of noncohesive sediment prevents the loss of any material that enters the barrel and lessens the redistribution 

of fluids as the sample is retrieved and during subsequent handling. The time needed to collect a sample is 

somewhat greater when the sample-freezing drive shoe 

is used, because time is needed to assemble and disassemble the carbon dioxide feed tubes, and about 10 

minutes is needed to freeze the core end. The samplefreezing drive shoe offers a means to enhance recovery 

of samples of noncohesive materials and the fluids they 

contain from the shallow saturated zone.

REFERENCES

Durnford, D., Brookman, J., Billica, J., and Milligan, J., 

1991, LNAPL distribution in a cohesionless soil: A 

field investigation and cryogenic sampler: Ground 

Water Monitoring Review, v. 11, no. 3, p. 115-122.

Essaid, H.I., Herkelrath, W.N., and Hess, K.M., 1993, Simulation of fluid distributions observed at a crude-oil 

spill site incorporating hysteresis, oil entrapment and 

spatial variability of hydraulic properties: Water 

Resources Research, v. 29, no. 6, p. 1753-1770.

Hess, K.M., Herkelrath, W.N., and Essaid, H.I., 1992, 

Determination of subsurface fluid contents at a crudeoil spill site: Journal of Contaminant Hydrology, v. 10, 

p. 75-96.

McElwee, C.D., Butler, J.J., and Healy, J.M., 1991, A new 

sampling system for obtaining relatively undisturbed 

samples of unconsolidated coarse sand and gravel: 

Ground Water Monitoring Review, v. 11, no. 3. p. 

182-191.

Munch, J.H., and Killey, R.W.D., 1985, Equipment and 

methodology for sampling and testing cohesionless 

sediments: Ground Water Monitoring Review, v. 5, 

no. 1, p. 38-42.

Zapico, M.M., Vales, S., and Cherry, J.A., 1987, A wireline 

piston core barrel for sampling cohesionless sand and 

gravel below the water table: Ground Water Monitoring 

Review, v. 7, no. 3, p. 74-82.
Determination of Surface Area of Soil Components by

Sorption of Nitrogen and Ethylene Glycol Monoethyl

Ether Vapors and Evaluation of Adsorption, Partition and

Cation Solvation Processes

By Gary T. Chiou 1 and David W. Rutherford

Abstract

The vapor uptake isotherms of nitrogen and 

ethylene glycol monoethyl ether (EGME) were 

determined for soil and soil components. The cross 

sectional area of EGME was determined to be 40.0 

x 10"20 irT by comparing adsorption isotherms on 

reference surface area standards. For samples 

where EGME does not penetrate into the solid 

(Ottawa Sand, hematite, aluminum oxide, synthetic 

hydrous iron oxide, and kaolinite), the EGME and 

nitrogen isotherms show consistent surface uptake 

and the EGME isotherm can be reasonably estimated from the nitrogen isotherm by multiplying 

the nitrogen uptake by a factor of 1.3 to account for 

different molecular area and molecular weight. For 

other samples where the EGME can penetrate into 

the solid (peat, Ca-montmorillonite, and illite), the 

total uptake can be resolved to surface adsorption, 

partition into soil organic matter, and/or cation solvation. For peat, the nitrogen uptake is very small, 

resulting in a surface area of 1.3 square meters per 

gram; uptake of EGME on peat is large and linear, 

which is indicative of large bulk solubility of 

EGME in soil organic matter (190 milligrams per 

gram). For Ca-montmorillonite, the resulting 

cation solvation isotherm shows a stoichiometric 

interaction of about 4 EGME molecules per 

exchangeable Ca ion. Exchange of organic cations

'U.S. Geological Survey. Box 25046. MS 408. Denver Federal Center. Denver. CO 80225

for metal ions in montmorillonite sharply decreases 

the cation solvation effect of EGME. This cation 

exchange may result in a higher surface area of the 

clay due to opening of interlayer surfaces.

INTRODUCTION

Because of various molecular components and 

lattice structures in such solids as soils, minerals, and 

clays, these materials may exhibit multiple interactions 

in the sorption of specific chemical species. While 

experimental sorption data reflect the total uptake from 

all contributing mechanisms, it is of practical interest to 

determine the relative contributions to the observed 

uptake from all of these interactions. In principle, the 

uptake of a chemical by a solid can result either from its 

adsorption on the solid's intrinsic (external) surfaces or 

from its penetration into the solid's internal matrix by 

processes other than surface adsorption, such as cation 

solvation in expanding clays or the dissolution into a 

portion of the solid matrix. To separate these contributions, one begins by determining the uptake capacity by 

one (or more than one) of the mechanisms using 

unequivocal analytical methods and subtracting it from 

the total capacity and repeating this process in a stepwise manner.

For a chemically inert vapor showing no 

specific reaction or penetration effect, its uptake by a 

solid results mainly from physical adsorption on the 

solid's surfaces. In such cases, the adsorption ranging 

from submonolayer to multilayer capacity can be well 

characterized by the Brunauer-Emmett-Teller (BET) 

model (Brunauer and others, 1938). In systems where 

vapors penetrate into the solid, the observed sorption

375
capacity would exceed the amount predicted by the 

BET equation. It is, therefore, prudent to analyze the 

vapor sorption by the BET equation for the vapor/solid 

pair of interest in comparison with the capacity of a 

reference inert gas (such as nitrogen) on the same solid 

to account for the extent of uptake attributable to 

chemical penetration.

This paper describes the uptake of a nonpolar 

vapor (nitrogen) and a polar vapor (ethylene glycol 

monoethyl ether) (EGME) on soil organic matter 

(peat), kaolinite, Ca-montmorillonite, and two organoclays, and the analysis of individual effects of surface 

adsorption, bulk solubility, and cation solvation pertinent to the system. In soil science, it has been 

assumed that the uptake of polar EGME measures the 

"total surface" of a solid, while the uptake of nonpolar 

nitrogen measures the external surface of the solid 

(Dyal and Hendricks, 1950; Bower and Gschwend, 

1952; Martin, 1955; Sor and Kemper, 1959); the difference has been described as "internal surface." It is 

arguable that the term "internal surface" is appropriate, since the difference in uptake caused by penetration of a polar solvent (or vapor) can be more properly 

attributed either to the cation solvation effect (as in 

expanding clays) (Tiller and Smith, 1990; McNeal, 

1964; Dowdy and Mortland, 1967) or to solubility (as 

in soil organic matter) (Chiou and others, 1990; Chiou 

and others, 1988; Rutherford and others, 1992).

Experimental

Two surface-area reference standard samples 

of aluminum oxide were used; surface areas were 

109 m2/g (reference standard A) and of 29.9 m2/g 

(reference standard B). An additional aluminum oxide 

sample (nonactivated reagent grade) had a purity of 

99.99 percent. The Ottawa sand is a standard sand 

with a 20 to 30 mesh particle size. The sample was 

washed with 10 percent HC1 and treated with F^C^ to 

remove any residual iron oxides and organic matter. 

The hematite sample was highly pure with a total Fe 

content of 70.28 percent on a dry-weight basis; X-ray 

diffraction data showed the sample to be a well crystallized material. The synthetic hydrous iron oxide 

was prepared in the laboratory by hydrolyzing 0.5 M 

aqueous ferric nitrate with 10 percent NaOH at pH 8.0 

to 8.5. The sample has a total Fe content of 62.4 percent. X-ray diffraction data showed it to be an essentially amorphous iron oxide with low goethite content. 

The Ca-montmorillonite sample had a cation exchange

capacity of 120 meq/100 g. The fifthian illite sample 

contained greater than 90 percent illite (with minor 

amounts of quartz and muscovite), had a cation 

exchange capacity of about 25 meq/100 q (with K as 

the major cation), and an organic carbon content of 

about 0.8 percent. The illite sample was treated with 

H2O2 to reduce the organic content before its use in 

sorption experiments. The treated sample contains 

about 0.4 percent organic carbon. The peat sample is a 

reference peat of the International Humic Substances 

Society, collected from the Everglades, FL. Reagent 

grade tetramethylammonium (TMA) bromide and 

tetraethylammonium (TEA) bromide were used. 

Woodburn soil is a silt loam from Oregon and its dryweight composition is 1.9 percent organic matter, 

68 percent silt, 21 percent clay, and 9 percent sand. 

X-ray diffraction showed that fine-grained mica (illite) 

and kaolinite are the dominant clays.

The TMA and TEA exchanged montmorillonites 

were prepared by suspending 10 g of Ca-montmorillonite in 800 mL of water. Approximately 3 mol of the 

organic cation per mol of cation exchange capacity 

(CEC) was added and mixed for 24 hours. The suspension was washed on a filter membrane until a bromide 

test with AgNO3 was negative. Organic carbon analysis showed that the TMA-montmorillonite was 87 percent exchanged with TMA and the TEAmontmorillonite 73 percent exchanged with TEA.

The apparatus and procedure used for the determination of vapor uptake by soils and minerals has 

been described previously (Chiou and others, 1988). 

Briefly, the experiment consisted of equilibrating the 

vapor in a sorption chamber containing an electronic 

microbalance from which a test solid sample (sorbent) 

was suspended in a small glass cup. The sorbent sample was heated at 100 °C for 8 to 10 hours inside the 

chamber and then cooled to room temperature for 6 to 

8 hours under a vacuum of 10"6 torr to remove moisture and to determine the "dry sample weight." Typically, 100 to 500 mg of sample was used in sorption 

experiments. The test liquid (EGME) was purified by 

vacuum distillation to remove residual air and then 

introduced into the sorption chamber to equilibrate. A 

change in sample weight resulting from vapor uptake 

was recorded as an electrical signal from the balance. 

The partial pressure of EGME at the point of vaporsample equilibrium was recorded by a pressure gauge. 

The saturation vapor pressure of EGME at room temperature (about 5 mm Hg at 24 °C) was determined by

376
monitoring the pressure of saturated EGME vapor 

prior to the sorption experiment.

The vapor uptake of nitrogen gas by the sample 

was determined at liquid nitrogen temperature using a 

surface area analyzer. The test samples were prepared 

by heating at 100 °C under a flow of pure nitrogen gas 

for 24 hours. The surface areas of samples were calculated by use of the linear BET plot of the nitrogen 

isotherms at relative pressures of 0.05 to 0.30 to obtain 

the monolayer adsorption capacities, together with the 

cross sectional area of the nitrogen molecule.

RESULTS AND DISCUSSION

For surface adsorption, according to the Bet 

theory (Brunauer and others, 1938), the adsorption 

capacity at the beginning of the linear range following 

the initial sharp rise corresponds roughly to the completion of monolayer adsorption (Qm); multilayer adsorption is signified by increasing slope of the isotherm as 

the relative pressure approaches 1. The Qm value of the 

adsorbate and the proper molecular area (a^) of the 

adsorbate are then used to calculate the surface area 

of the solid (adsorbent). The linear BET plot allows the 

value of Qm of an adsorbate on a given solid to be 

calculated, as follows:

C-l

Q(PVP-l) CQmpo CQ, (1)

where Q is the amount of vapor adsorbed at given 

P/P°, P is the equilibrium partial pressure of the adsorbate, P° is the saturation vapor pressure of the adsorbate, and C is a constant related to the net heat of 

monolayer adsorption, which affects the sharpness of 

the curvature. A plot of 1/Q(P°/P-1) against P/P° 

yields a straight line with a slope of (C-l)/CQm and an 

intercept of l/CQm , from which the Qm and C values 

are calculated. The linearity of the BET plot is usually 

observed at P/P° = 0.05 - 0.30 in systems involving 

physical adsorption (Brunauer and others, 1938; Dyal 

and Hendricks, 1950; Bower and Gschwend, 1952). 

The cross sectional area of 16.2 x 10"20 in 2 for the 

nitrogen molecule is well established from the literature (Brunauer and others, 1938; Dyal and Hendricks, 

1950; Bower and Gschwend, 1952). To determine the 

molecular area of EGME, we determined the EGME 

monolayer capacities of two alumina reference samples (29.9 and 109 m2/g, respectively) as 11.3 and

40.4 mg/g from the BET plot (eq. 1) of the respective 

EGME isotherms; the EGME corresponding cross 

sectional areas are 39.6 x 10"20 and 40.4 x 10'20 m2, 

giving an average of 40.0 x 10 m.

The BET model describes vapor adsorption on 

surfaces that are external to the solid. This adsorption 

may include penetration into pores but not into the solid 

matrix. In cases where penetration of the solid matrix 

occurs along with surface adsorption, the system may 

exhibit an "apparent monolayer capacity" giving a 

higher "apparent surface area." Therefore, a comparison of the "apparent monolayer capacity" from a vapor 

(such as EGME) that could potentially penetrate into a 

solid sample with the surface area of the solid characterized by an inert vapor (such as N2 ) provides a measure 

of the uptake by penetration of the vapor.

The isotherms for nitrogen and EGME uptake 

by kaolinite are shown in figure 1. These isotherms 

are typical of a type II isotherm, which is characteristic of the transition from submonolayer to multilayer 

surface adsorption. The surface area calculated from 

the BET equation is 21.0 m2/g using the nitrogen isotherm and 21.1 m2/g using the EGME isotherm. This 

shows the consistency of the BET method in determining the surface uptake of vapors, as long as the vapors 

do not penetrate the solid matrix. The equivalence of 

nitrogen and EGME isotherms for other nonswelling 

mineral samples (sand, hematite, aluminum oxide, and 

synthetic iron oxide) is also demonstrated and is 

shown in table 1.

S 70

<

O

-! 40

£ 30

20

10

 Peai-EGME

>linite-EGME

-Nitrogen

t Scale)

0.1 0.2 0.3 0.4 0.5 0.6 0.7 

RELATIVE PRESSURE, P/P°

0.8 0.9

Figure 1. Isotherms of ethylene glycol monoethyl ether 

vapor at room temperature and nitrogen vapor at 77 °K for 

kaolinite and Florida peat.

377
Table 1. Qm (N2) monolayer capacities, BET-(N2) surface areas, Qm (EGME)eq equivalent monolayer capacities, and 

Qm (EGME)ap apparent monolayer capacities of selected soils and minerals

[The Qm(N2) values are from N2 adsorption isotherms and the BET equation; Qm (Ni) values are used to determine the surface areas of samples; the 

Qm (EGME)eq are values equivalent to Qm(N 2 ) for the same surface areas of samples; Q(EGME)ap values are from the EGME sorption isotherms and the 

BET equation, assuming that no penetration or specific interation occurs. The ratio of Qni(EGME)ap to Qm (EGME)eq , when significantly greater than 1, 

expresses the extent of EGME penetration or specific interaction.]

Sample

Ottawa Sand

Hematite

Aluminum Oxide

Synthetic Hydrous Iron Oxide

Kaolinite

Peat

Ca-Montmorillonite

Illite

TMA-Montmorillonite

TEA-Montmorillonite

Woodburn soil

Qm <N2 ) 

(mg/g)

0.032

.14

.87

53.4

6.03

.36

21.8

19.3

52.1

19.5

3.22

BET-(N2 ) Area 

(m2/g)

0.11

.50

3.03

186

21.0

1.26

75.9

67.2

185

69.4

11.2

(mg/g)

0.042

.18

1.13

69.4

7.84

.47

28.3

25.1

67.7

25.4

4.19

(mg/g)

0.035

.18

1.02

61.2

7.90

...

215

38.0

93.8

45.1

13.4

Qm (EGME)ap/ 

Qm (EGME)cq

0.83

1.00

.90

.88

1.01

7.60

1.51

1.38

1.77

3.20

The isotherms for nitrogen and EGME on peat 

are shown in figure 1. The nitrogen isotherm is type 

II, whereas the EGME isotherm shows a large and 

linear uptake that is characteristic of bulk EGME solubility in peat (Chiou and others, 1990; Chiou and 

others, 1988; Rutherford and others, 1992). The 

nitrogen isotherm gives a surface area of 1.5 m2/g. 

Because the EGME isotherm on peat is linear (rather 

than type II), the BET calculation does not apply; 

obviously, the small EGME adsorption is obscured by 

the much larger partition effect with peat. The limiting solubility of EGME in peat at P/Po = 1, normalized for the organic content of peat, is 190 mg/g.

As shown above, in the absence of vapor penetration, nitrogen and EGME isotherms give equivalent 

surface coverage; therefore, the nitrogen isotherm can 

be converted to an equivalent EGME isotherm for surface adsorption by multiplying the nitrogen adsorption 

capacity by a factor of 1.3, to account for differences 

in molecular weights and molecular cross sectional 

areas. This procedure was applied to vapor sorption 

by Ca-montmorillonite, fifthian illite, and two organoclays to account for surface adsorption on (preexisting) external surfaces (from nitrogen adsorption data).

The amount of EGME penetration was then determined from the difference of the total EGME 

isotherm and the calculated EGME isotherm for surface adsorption and is plotted as the EGME cation 

solvation isotherm. The original EGME and nitrogen 

isotherms and the EGME cation solvation isotherm for 

Ca-montmorillonite are shown in figure 2.

0.3 0.4 0.5 0.6 0.7 

RELATR-E PRESSURE, P/P°

Figure 2. Isotherms of ethylene glycol monoethyl ether 

vapor at room temperature and nitrogen vapor at 77 °K on 

Ca-montmorillonite.

378
The equivalent EGME monolayer capacity 

[Qm (EGME)eq] calculated from the nitrogen monolayer capacity and apparent EGME monolayer capacity [Qm (EGME)ap] calculated from the EGME 

uptake isotherm using the BET equation for all samples are summarized in table 1. The Qm(EGME)ap for 

Ca-Montmorillonite is 7.6 times higher than the 

Qm (EGME)eq . The ratio of Qm (EGME)ap to 

Qm (EGME)^ is 1.5 illite, 1.4 for TMA-montmorillonite, and 1.8 for TEA-montmorillonite. The large ratio 

for Ca-montmorillonite is attributed to the large cation 

solvation effect of EGME relative to its surface 

adsorption. The EGME cation solvation isotherm in 

figure 2 shows rapid uptake of EGME at relative pressures below 0.1 and complete solvation above this 

point as indicated by the relatively flat isotherm. 

Based on CEC of 120 meq/lOOg, this would indicate a 

stoichiometric relation of about 4 EGME molecules 

per exchangeable Ca ion. The smaller ratio for illite is 

in keeping with its lower CEC (15.5 meq/lOOg) and 

the lower cation solvation of EGME with K relative to 

Ca (0.9 - 1.4 EGME molecules per exchangeable K 

ion). Similarly, substitution of TMA or TEA for Ca in 

the montmorillonite decreases the cation solvation 

effect, as these organic cations are presumably less 

effectively solvated by EGME than the Ca ion. In case 

of TMA-montmorillonite, a great increase in surface 

area is noted over that of Ca-montmorillonite, indicating that the interlayer opened up by TMA ions is relatively accessible to nitrogen and EGME. This is not 

observed for TEA-montmorillonite since the surface 

area remains comparable with that of Ca-montmorillonite, indicating that the bigger TEA ions occupy most 

of the interlayer space.

As shown, the uptake of polar vapors (as 

EGME) by dry natural samples may be expected to 

occur by multiple mechanisms, including surface 

adsorption on external surfaces, partition (dissolution) 

into the sample's organic matter (or organic impurities), and solvation of the interlayer cations in expanding clays. The significance of each interaction 

depends on the composition and morphology of the 

solid samples as well as on the polarity of vapors. In 

general, as the polarity of the vapors decreases, both 

the partition effect with organic matter and the cation 

solvation effect decrease. The expected decrease in 

partition follows from the observation that the solubility of vapors or liquids in relatively polar soil organic 

matter increases sharply with the polarity of the vapors 

or liquids (Chiou and others, 1990; Rutherford and

others, 1992). The expected decrease in cation solvation has been demonstrated from published data showing the consistency of the nitrogen and nonpolar 

organic vapor uptake by expanding clays and the 

excess uptake of polar compounds, as demonstrated in 

this study.

A general analytical procedure to determine the 

relative contributions of surface adsorption, partition, 

and cation solvation for sorption of a vapor on a natural solid sample containing clays, other inorganic 

matter, and organic matter would involve (1) determination of the sorption isotherms of the vapor of interest and of nitrogen on the sample (2) determination of 

the sample's organic and ash contents (3) subtraction 

of the portion of the non-penetrating vapor isotherm 

(calculated from the nitrogen isotherm using a suitable 

conversion factor) from the total isotherm, and 

(4) subtraction from the resulting curve of the linear 

isotherm calculated for the organic content of the 

sample. The last step requires either the experimental 

or the estimated vapor sorption (partition) isotherm on 

soil organic matter. The result from the above procedure would give an isotherm that can be used to estimate the cation solvation contribution. This procedure 

has been applied to Woodburn soil as shown in figure 3.

SUMMARY

The differences in sorption capacities of polar 

EGME and nonpolar, inert N2 on various dry solids 

illustrate the various interactions that control the vapor 

uptake. The extent of physical adsorption on a solid's

5 30

25

20

 TotaJ Uptake

O Cation Solvation

A Surface Adsorption

D Partition in SOM

0.1 0.2 0.3 0.4 0.5 0.6

RELATIVE PRESSURE, P/P°

0.7 0.8

Figure 3. Analysis of Woodburn soil showing total uptake of 

ethylene glycol monoethyl ether vapor and contributions by 

surface adsorption, partition into soil organic matter (SOM) 

and cation solvation.

379
surfaces can be accounted for by BET equation using 

the nitrogen adsorption data and a constant adjusting 

factor. The excess over physical adsorption on the 

solid's surfaces is attributed to either vapor solubility 

in SOM or cation solvation with clay components. 

The procedure used to separate these effects for 

EGME sorption can be used to separate the different 

sorptive contributions of other organic vapors on a 

natural solid. Thus, the sorption of nonpolar vapors 

(such as aliphatic and aromatic hydrocarbons) on dry 

soils and minerals should consist primarily of surface 

adsorption, because nonpolar vapors have low solubility in polar SOM and cannot effectively solvate clay 

cations. This result has been verified by the consistency of the total uptake of various low-polarity 

vapors on soils in comparison with the nitrogen 

adsorption data in an earlier study. The sorption 

capacities of polar liquids or vapors on dry natural 

solids may be expected to be greatly enhanced over 

their surface adsorption capacities if these samples 

contain significant amounts of organic matter and/or 

solvating clays because of the concurring solubility in 

SOM and/or the cation solvation effect.

REFERENCES

Bower, C.A., and Gschwend, F.B., 1952, Ethylcne glycol 

retention by soils as a measure of surface area and 

interlayer swelling: Soil Science Society of America 

Proceedings, v. 17, p. 342-345.

Brunauer, S., Emmett, P.H., and Teller, E., 1938, Adsorption 

of gases in multimolecular layers: Journal of 

American Chemical Society, v. 60, p. 309-319.

Chiou, C.T., Dilc, D.E., and Malcolm, R.L., 1988, Sorption 

of vapors of some liquids on soil humic acid and its 

relation to partition of organic compounds in soil 

organic matter: Environmental Science and Technology, v. 22, p. 298-303.

Chiou, C.T., Lee, J.F., and Boyd, S.A., 1990, The surface 

area of soil organic matter: Environmental Science 

and Technology, v. 24, p. 1164-1166.

Chiou, C.T., Lee, J.F., and Boyd, S.A., 1992, Reply to 

comment on "The surface area of soil organic matter": Environmental Science and Technology, v. 26, 

p. 404-406.

Dowdy, R.H., and Mortland, M.M., 1967, Alcohol-water 

interactions on montmorillonitc surfaces. 1 Ethanol: 

Clays and Clay Minerals, v. 15, p. 259-271.

Dyal, R.S., and Hendricks, S.B., 1950, Total surface of 

clays in polar liquids as a character index: Soil 

Science , v. 69, p. 421-432.

Martin, R.T., 1955, Ethylenc glycol retention by clays: 

Soil Science Society of America Proceedings, v. 19, 

p. 160-164.

McNeal, B.L., 1964, Effect of exchangeable cations on 

glycol retention by clay minerals: Soil Science, v. 97, 

p. 96-102.

Rutherford, D.W., Chiou, C.T., and Kile, D.E., 1992. 

Influence of soil organic matter composition on the 

partition of organic compounds: Environmental 

Science and Technology, v. 26, p. 336-340.

Sor, K., and Kemper, W.D., 1959, Estimation of hydrateable 

surface area of soils and clays from the amount of 

absorption and retention of ethylene glycol: Soil 

Science Society of America Proceedings, v. 23, 

p. 105-110.

Tiller, K.G., and Smith, L.H., 1990, Limitations of EGME 

retention to estimate the surface area of soils: 

Australian Journal of Soil Research, v. 28, p. 1-26.

380
Isolation of Lipophilic Organic Contaminants

along the Upper Mississippi River Using a

Semipermeable Membrane Device

By Geoffrey S. Ellis 1 and Colleen E. Rostad 1

Abstract

The detection of many environmental contaminants in,water is often hindered by their transient 

nature and low concentration. However, these 

compounds can represent a significant environmental hazard to biota and humans through bioconcentration. Traditional methods of analyzing 

aquatic biota are complicated by variations in 

species, sex, age, reproductive stage, location, 

behavioral patterns, metabolism, and water conditions. A new sampling device consisting of a 

tubular semipermeable membrane of low-density 

polyethylene filled with synthetic fish fat, triolien, 

which mimics the mechanism of bioconcentration, 

may resolve some of these problems. To evaluate 

the effectiveness of these sampling devices, they 

were deployed at nine sites along the upper Mississippi River; caged fish at three of these sites were 

used for comparison purposes. The devices and 

caged fish were collected over an eight-week 

period, and indigenous fish were collected at the 

end of the study. Compounds were extracted from 

the triolien by back dialysis into an appropriate 

solvent. Fish tissues were ground and extracted 

wet into a solvent and the lipid removed. All samples were analyzed by gas chromatography/mass 

spectrometry (GC/MS). Target compounds were 

pesticides, herbicides, and polychlorinated biphenyls (PCBs). Comparisons of compound concentrations in the fish samples and sampling devices 

will be used to determine the effectiveness of 

this technique as a monitoring tool for bioconcentration.

U.S. Geological Survey, Arvada, Colo.

INTRODUCTION

The uptake and concentration of lipophilic organic 

contaminants by aquatic organisms is a significant 

pathway for toxic compounds in the environment. In 

conjunction with food chain effects, this constitutes 

bioaccumulation which magnifies concentrations of 

chemicals in fatty tissues of organisms in successively 

higher trophic levels (Neely, 1980). This process has 

the potential to increase the toxicological significance 

of all lipophilic contaminants in the environment. 

Although the bioavailability of toxins in the environment is of great importance for human health and 

environmental quality concerns, it is often difficult to 

assess. The transient nature of direct chemical discharges and low concentrations of many nonpointsource contaminants in the water column frequently 

preclude compound detection and quantification due to 

limitations of sampling and analytical technology.

Historically, bioavailability has been described in 

terms of equilibrium partitioning models, which 

attempt to express this process mathematically in terms 

of two compartments and the flux rates into and out of 

the compartments. Aquatic organisms are viewed as 

lipid pools separated from an aqueous medium by a 

biomembrane. Flux rates are assumed to be determined 

by passive diffusion of contaminants across the membrane. Although these models generally correct for differences in percent lipid between individual organisms, 

they assume that lipid-water partitioning is independent of known variations in lipid composition and content between species, and age and gender differences 

within a species. These models often rely on empirically derived partition coefficients between an organic 

solvent and water (usually octanol-water systems 

expressed as KoW) however, these constants may not 

be sufficiently representative of actual bioconcentration factors (Huckins and others, 1990, Opperhuizen 

and others, 1988).

381
Several other factors complicate the application of 

equilibrium partitioning models to complex environmental systems. The area in which organisms are found 

may range from a static location for immotile organisms to thousands of kilometers for migratory species; 

this areal variability can confound efforts to identify 

sources and pathways for the compounds of interest. 

Many absorbed contaminants are metabolized by 

organisms to other compounds which also may be 

toxic, and some or all of these metabolic products may 

be excreted on a relatively short time scale. Bioavailability studies typically require large numbers of biological samples that are very time consuming and labor 

intensive to collect and analyze. Consequently, the 

specific processes involved in bioaccumulation have 

been completely determined for only relatively simple 

environmental systems.

By the late 1970's, these problems were well documented and efforts were being made to address them. 

In 1980, Bryne and Aylott (1980) received a British 

patent for a passive sampling device based on liquidliquid partitioning between a nonpolar solvent and 

water separated by a semipermeable membrane. 

Sodergren (1987) was the first to use a hydrophilic 

dialysis membrane filled with hexane. These devices 

have many advantages over conventional biomonitoring. Absorbed compounds are not metabolized and 

variations among different species with respect to gender, age, lipid content, and so forth are eliminated, and 

the location of the sampling site is fixed.

Although these methods represent distinct advantages over previous methods, several problems are 

associated with them. Because the process of diffusion 

across a semipermeable membrane is essentially 

passive through small pores in the membrane, lowmolecular-weight compounds may pass through the 

membrane unimpeded; whereas comparatively large 

molecules are excluded. The Solvents used inside these 

semipermeable membrane devices (SPMDs) are low 

molecular weight and, therefore, tend to diffuse out of 

the sampling device by osmotic pressure. Over long 

periods of time, these solvent losses can be significant. 

Huckins and others (1990) found the measured concentration factors for the compounds studied determined 

by use of the solvent-filled devices were one to two 

orders of magnitude smaller than the bioconcentration 

factors reported in the literature. This may indicate that 

organic solvents are not valid surrogates for biologic 

lipids, and(or) that the rate of solvent diffusion out of 

the sampling device is great enough to prohibit the

system from reaching equilibrium over the sampling 

time period. Additionally, hydrophilic dialysis membranes have been found to be susceptible to degradation by organic solvents and biota.

Researchers at the U.S. Fish and Wildlife Service 

(FWS) National Fisheries Contaminant Research 

Center have developed a passive SPMD that addresses 

some of the problems of earlier designs. Nonporous, 

low-density polyethylene (or polypropylene) flat tubing is used as a semipermeable membrane. Although 

this material is referred to as nonporous, it does, in fact, 

contain small cavities and transient holes of about 5 to 

10 Am diameter. This hydrophobic material is more 

resistant to degradation by organic solvents and biota 

than the hydrophilic dialysis membranes used in other 

samplers, although biofouling on the surface occurs. 

The replacement of the solvent with lipid makes these 

samplers more analogous to real fish than earlier 

designs. This is exemplified by the fact that empirically 

determined lipid/water partition coefficients, Klw 

(or lipid concentration factors) are closer to actual 

bioconcentration factors than published partition coefficients (KoW) for most organic contaminants. The 

larger molecular size of lipid molecules compared to 

solvent molecules makes the membrane relatively 

impermeable to the lipid.

The efficacy of the lipid-filled SPMD for concentrating organic contaminants has been tested extensively in laboratory flow-through diluter studies. A few 

small scale field studies of this technique also have 

been made. This study is the first large-scale field 

investigation of the application of lipid-filled SPMDs 

to monitoring bioavailability of organic contaminants 

in the environment. This paper briefly describes the 

SPMD technique and the progress made in this study.

STUDY DESIGN

This study involves the deployment of lipid-filled 

SPMDs, caged fish, and the collection of indigenous 

fish in the upper Mississippi River. All field work was 

carried out by FWS personnel, and was timed to coincide with the U.S. Geological Survey (USGS) Upper 

Mississippi River study. SPMDs and caged fish were 

deployed at three sites and samples were collected after 

28 days. Indigenous fish were collected 57 days after 

deployment of the SPMDs.

Initial preparation of SPMD and fish samples was 

done by FWS personnel. SPMDs were back-dialyzed 

into hexane and concentrated to 1 mL by rotoevapora382
tion. Fish were dissected in the field to remove their 

livers and then frozen until ready for further processing. Whole-fish samples were homogenized by FWS 

personnel, frozen, and sent to the USGS laboratory in 

Arvada, Colo. Fish-tissue samples were prepared for 

analysis by extraction with methylene chloride, lipid 

removal by gel permeation chromatography (GPC), 

concentration in Kuderna-Danish glassware, and dry 

nitrogen evaporation. Analysis of sample extracts was 

performed by gas chromatography/negative chemical 

ionization/mass spectroscopy (GC/NCI/MS).

STUDY OBJECTIVES

This study has several objectives, each of which is 

specific to the various organizations involved. The primary objective for FWS researchers is to evaluate the 

efficacy of lipid-filled SPMDs for monitoring the bioavailability of organic contaminants under field conditions. The USGS is interested in assessing the 

bioavailability of various lipophilic organic contaminants in the Upper Mississippi River. Additionally, the 

USGS is investigating the partitioning of organic 

contaminants among SPMDs, fish, suspended silt, 

colloids, bed sediments, and the aqueous phase.

We hypothesize that, although certain organic 

compounds will be found in the aqueous phase and in 

the SPMDs, only their metabolites will be found in the 

fish. The degree to which SPMDs indicate the bioavailability of these compounds will be assessed quantitatively by comparing lipid concentration factors 

calculated from SPMD and aqueous concentrations to 

published bioconcentration factors.

PROJECT STATUS

Field sampling has been completed by FWS 

personnel. SPMD samples have been back-dialyzed, 

concentrated, and sent to the USGS laboratory. Dialysates have been further concentrated and analyzed by 

GC/NCI/MS. All SPMD data have been tabulated. The 

presence of the following compounds has been confinned at quantifiable levels in some or all of the 

samples from the SPMDs:

pentachlorobenzene,

hexachlorobenzene,

pentachloroanisole,

dacthal (DCPA),

trifluralin,

lindane,

heptachlorepoxide,

cis- and trans-chlordane (mostly his-),

cis- and trans-nonachlor (mostly trans), and

tetra-, penta-, hexa-, hepta-, octa-, and nonachlorobiphenyls.

All fish samples were collected, homogenized, 

frozen, and sent to the USGS laboratory by FWS personnel. A fish-tissue extraction protocol has been 

developed. Initial experiments on the column-extraction procedure showed that post-extraction tissue contained negligible residues of the target compounds. 

GPC experiments showed good chromatographic separation of lipid and target compounds. With the exception of three problematic compounds (DDD, DDE, and 

DDT), recoveries for the standard compounds quantitated based on isodrin ranged from 60 to 167 percent; 

the average recovery was 99 percent. The fish-tissue 

samples were extracted and have been analyzed by 

GC/NCI/MS. The fish tissue data are currently being 

correlated for comparison with the SPMD data.

REFERENCES

Brync, B.A., and Aylott, R.I., 1980. Concentrator for 

removing organic materials from aqueous systems: 

British Patent No. 1566253.

Huckins, J.N., Tubcrgcn, M.W., and Manuwecra. O.K., 

1990, Semipermcablc membrane devices containing 

model lipid: A new approach to monitoring the bioavailability of lipophilic contaminants and estimating 

their bioconcentration potential: Chcmosphcrc v. 20, 

no. 5, p. 533-552.

Necly. W.B., 1980. Chemicals in the environment: Distribution, transport, fate, and analysis: N.Y.. Marcel Dckker, 

Inc.. 245 p.

Opperhuizcn. A., Scrnc. P., and Van dcr Stecn. J.M.D. 1988, 

Thermodynamics offish/water and octan-1-ol/watcr 

partitioning of some chlorinated benzenes: Environmental Science and Technology, v. 22. p. 286-292.

Sodcrgrcn. A., 1987, Solvent filled dialysis membranes simulate uptake of pollutants by aquatic organisms:. Environmental Science and Technology, v. 21. p. 855-859.

383
384
Measurement of Concentrations of Alachlor and Its

Ethanesulfonic Acid Metabolite in Water by Solid-Phase

Extraction and Enzyme-Linked Immunosorbent Assay

By Diana S. Aga1 , E. Michael Thurman1 , and Michael L. Pomes 1

Abstract

Solid-phase extraction (SPE) and enzymelinked immunosorbent assay (ELISA) were 

combined for the trace analysis of 2-chloro-2',6'- 

diethyl-N-(methoxymethyl)acetanilide(alachlor) 

and2-[(2,6-diethylphenyl)(methoxymethyl)amino- 

]-2-oxoethanesulfonic acid (ESA, a major soil 

metabolite of alachlor). The anti-alachlor 

antibody reacted not only with alachlor but also 

with ESA, which produced false-positive 

immunoassay detections of alachlor in natural 

water samples. The combination of SPE with 

ELISA effectively separates and determines both 

parent and metabolite compounds; thus, the same 

antibody may be used for two ELISA methods. 

Alachlor and ESA were isolated from water by 

SPE on CIS resin and were eluted sequentially 

with ethyl acetate and methanol. The separation 

involves the difference in solubility of the ionic 

sulfonic acid in ethyl acetate and methanol. The 

SPE-ELISA method has a detection limit of 

0.01 jag/L (micrograms per liter) for alachlor and 

0.05 fig/L for ESA with a precision of 

± 10 percent. Analyses of surface-water and 

ground-water samples were confirmed by highperformance liquid chromatography with 

photodiode-array detection and gas 

chromatography/mass spectrometry. Results 

showed widespread presence of ESA in surface 

water and ground water of the midwestern United

U.S. Geological Survey, Lawrence, KS

States in concentrations of 0.10 to greater than 

10 Mg/L.

INTRODUCTION

Solid-phase extraction (SPE) and enzymelinked immunosorbent assay (ELISA) are two 

analytical techniques that recently have found wide 

application in environmental chemistry. For example, 

ELISA has been applied to the analysis of herbicide 

residues in soil and water (Bushway and others, 1988; 

Vanderlaan and others, 1988; Schlaeppi and others, 

1989; Feng and others, 1990a, 1990b; Thurman and 

others, 1990; Goh and others, 1991; Van-Emon and 

Lopez-Avila, 1992). ELISA is a sensitive, fast, and 

cost-effective technique that can be conducted both in 

the laboratory and in the field. However, most of the 

applications of ELISA are limited to the screening of 

samples in a complex environmental matrix, which 

may lead to false-positive detections of the target 

analytes (Feng and others, 1990a, 1990b; Baker and 

others, 1993). Less commonly, false-negative 

detections are observed when the detection limit of the 

assay is reached. Furthermore, the immunoassay 

performance can be affected substantially not only by 

cross-reacting compounds, but also by matrix 

components that interfere with the assay detection 

system and the antibody-antigen interactions.

SPE is used extensively as a clean-up procedure 

for clinical and environmental samples (Morris and 

Ruthann, 1988; Tippins, 1988). In contrast to ELISA, 

which can detect known substances, SPE can separate 

unknown substances by mechanism similar to those of 

high-pressure liquid chromatography (Morris and 

Ruthann, 1988). The complimentary features of these 

two techniques may be combined for a selective and

385
sensitive analytical method that is analogous to a 

chromatographic separation, where the SPE cartridge 

is the column and ELISA is the detector. With 

properly designed SPE procedures, the reliability of 

the immunoassay may be greatly enhanced by 

separating closely related compounds and reducing 

false-positive detections. Moreover, because SPE is 

also a preconcentration technique, the sensitivity of 

the ELISA may be increased by several orders of 

magnitude. Because SPE is easily automated for 

water analysis (Castellani and others, 1990; Meyer and 

others, 1993 ), the reproducibility and speed of 

immunoassay is not compromised. In addition, cross 

reactivity, which commonly is viewed as a major 

drawback for quantitative immunoassay, can be a 

positive aspect of ELISA by using the same antibody 

to quantitate a cross-reacting compound after the 

compounds have been separated by SPE.

Alachlor, a widely used herbicide in the 

midwestem United States, and its relatively persistent 

and very mobile ethanesulfonic acid metabolite, ESA 

(Baker and others, 1993), are important target 

compounds for the development of a sensitive assay 

that combines SPE and ELISA. The high frequency of 

false-positive detections observed in the ELISA 

screening kits for alachlor have been attributed to the 

significant cross reactivity of the ESA metabolite 

towards the anti-alachlor antibody (Macomber and 

others, 1992; Baker and others, 1993). The specific 

objectives of the study reported herein were to (1) 

understand the principles involved in separating parent 

herbicides and ionic metabolites by SPE, (2) couple 

SPE and ELISA for analysis of alachlor and ESA, (3) 

apply the method on samples of surface water and 

ground water and compare the results with gas 

chromatography/mass spectrometry (GC/MS) and 

high-performance liquid chromatography with 

photodiode-array detection (HPLC/PDA).

METHODS

Solid-Phase Extraction Procedure

The SPE procedure was automated with a 

Millipore Workstation2 . The SPE cartridges

The use of brand names in this paper is for identification purposes only and docs not constitute endorsement by the U.S. Geological Survey.

containing 360 mg of 40-mm C ]X -bonded silica were 

washed sequentially with 2 mL of methanol, 6 mL of 

ethyl acetate, 2 mL of methanol, and 2 mL of distilled 

water. One hundred (100) mL of sample were passed 

through the cartridge at a flow rate of 10 mL/min. The 

cartridge was eluted with 3 mL of ethyl acetate, 

followed by a transfer step to remove the ethyl acetate 

(top layer) containing the alachlor from the residual 

water (bottom layer) in the eluate. Then, the cartridge 

was eluted with methanol to remove the ESA, which 

was collected in a separate test tube. Both ethylacetate and methanol extracts were evaporated to 

dryness under nitrogen at 45 °C (degrees Celsius). 

The ethyl-acetate extracts were reconstituted with 1 

mL of 20 percent methanol-water mixture, and the 

methanol extracts were reconstituted with 5 mL of 

water for analysis by ELISA (fig. 1).

The preparation of samples for GC/MS and 

HPLC analysis was performed in a similar SPE 

procedure. The ethyl-acetate extracts were spiked 

with phenanthrene-d| 0, which was used as the external 

standard for quantitation, and the sample volume was 

reduced to about 100 \j.L for the analysis of alachlor by 

GC/MS. The methanol extracts for HPLC analysis 

were evaporated to dryness and redissolved in 100 ^L 

of 10 mM phosphate buffer/methanol (20/80) mixture.

Enzyme-Linked Immunosorbent Assay 

Procedure

ELISA kits from two different manufacturers 

were used to analyze each solid-phase eluate. One 

type of alachlor ELISA kit had antibodies coated on 

the wells of the microtiter plate, and the other type of 

alachlor ELISA kit had antibodies covalently attached 

to magnetic particles. Each extracted sample was 

analyzed in duplicate. For the microtiter plate-based 

ELISA kit, 80 fiL of sample and 80 fiL of haptenenzyme conjugate were mixed in the well, and the 

mixture was incubated at 30 °C in an orbital shaker 

(200 revolutions per minute). After 1 hour, the plate 

was rinsed five times with deionized water, and excess 

water was removed. One hundred sixty (160) mL of 

color reagent (1:1 mixture of hydrogen peroxide and 

3,3',5,5'-tetramethylbenzidine or TMB) were added, 

and the color developed for 30 minutes. Then, the stop 

solution (40 jaL, 2 M sulfuric acid) was added. Optical 

densities were read on a microplate reader using 

computer software . The alachlor concentration in the 

samples were calculated by interpolation from four

386
rn

Alachlor and sulfonic acid metabolite 

(100-milliliter sample)

i

SPEC) 8 cartridge

Effluent to waste Elute with 3 milliliiers of 

ethyl acetate 

(contains alachlor)

Elute with 2 milliliters of methanol 

(contains ethanesulfonic acid 

and oxo-acetic acid metabolites)

(a) Analysis by ELISA: evaporate 

to dryness, then reconstitutewiih 

1 milliliter of 20-percent 

methanol-water mixture.

(b) Analysis by GC/MS: spike with 

phenanthrene-dioand evaporate to 

100 microliters.

(a) Analysis by ELISA: evaporate to 

dryness, then reconstitute with 5 

milliliters of distilled water.

(b) Analysis by HPLC: spike with 

metribuzin and evaporate to dryness, 

then reconstitute with 100 microliters 

of 20/80 buffer/methanol mixture .

Figure 1. Flow chart for illustrating analytical method using solid-phase extraction (SPE) and enzyme-linked 

immunosorbent assay {ELISA), gas chromatography/mass spectrometry (GC/MS), or high-performance liquid 

chromatography (HPLC).

sets of alachlor standards at concentrations of 5.0, 

0.10, and 0 (ag/L in 20-percent methanol-water 

mixture.

l.O,

For the magnetic particle-based ELISA kits, 

250 ^L of sample were mixed with 250 [iL of haptenenzyme conjugate and 500 [iL of paramagnetic 

particles. This mixture was incubated for 30 minutes 

at room temperature, and a strong magnetic field was 

applied. The excess reagent was removed. Then, 500 

[iL of color reagent were added, and the mixture was 

incubated for 20 minutes at room temperature (25 °C), 

followed by addition of the stop solution (500 |iL). 

The optical densities were read using a photometric 

analyzer, and concentrations were calculated from 

logarithmically transformed data with the same set of 

alachlor standards used for the microtiter plate-based 

ELISA. ESA analysis was performed using the 

alachlor magnetic particle-based ELISA that was 

calibrated with ESA standards at concentrations of 20, 

5.0, 1.0, and 0 [ig/L.

Analysis by Gas Chromatogaphy/Mass 

Spectrometry

Operating conditions for the GC/MS analysis of 

alachlor were identical to those described by Thurman 

and others (1990). The detector was operated in a 

selected-ion-monitoring (SIM) mode, and 

confirmation was based on the presence of the 

molecular-ion peak, two confirming ions (with area 

counts of ±20 percent) and a retention time match of 

±0.2 percent relative to phenanthrene-d 10. The ratio of 

the base-peak ion of alachlor and the 188-amu (atomic 

mass unit) ion of phenanthrene-d 10 were used to 

construct the calibration curve. A 12-m fused-silica 

capillary column of methylsilicone with a film 

thickness of 0.33 fxm and 0.2 mm inside diameter was 

used for separation. Helium was used as the carrier 

gas with a flow rate of I mL/min and a head pressure 

of 35 kPa. The samples were injected in the spfitless 

mode by an auto-injector. The column temperature 

was held at 50 °C for I minute, then increased at 6 °C

387
per minute to 250 °C and held at this temperature for 

10 minutes. Injector temperature was 280 °C.

Analysis by High-Performance Liquid 

Chromatography

The HPLC analysis of the methanol extracts for 

the confirmation of ESA was performed in liquid 

chromatograph with a photodiode-array (PDA) 

detector. The HPLC was equipped with a 2.1 x 100 

mm narrow bore, reversed-phase column, packed with 

5 mm of octadecylsilyl. The mobile phase consisted 

of 38.5-percent HPLC-grade methanol and 61.5- 

percent 10 mM dibasic sodium phosphate (pH 7 

buffer). The flow rate of the mobile phase was 1.2 

mm/min. The sample injection volume was 90 uL. 

ESA was monitored at a wavelength of 200 nm with a 

4-nm bandwidth. The reference wavelength was set at 

450 nm with a 80-nm band width. To confirm the 

identity of the ESA, the ultraviolet spectra were 

scanned from 190 to 400 nm and the spectra matched 

to standard spectra in an automated library search. A 

fixed quantity of metribuzin (500 ng) was spiked into 

each methanol extract to serve as the internal standard 

for quantitation. A detailed discussion on the HPLC 

method is presented by Pomes and others (this 

volume).

RESULTS AND DISCUSSION

Acetanilide Cross-reactivity and Enzymelinked Immunosorbent Assay Development

The cross reactivity of the antibodies in the two 

alachlor ELISA kits with other acetanilides and 

alachlor metabolites was determined (fig. 2). Cross 

reactivities were expressed as IC50, which is the 

concentration of the compound that causes 50-percent 

inhibition of absorbance, and as the least detectable 

dose (LDD), which is the concentration at 90-percent 

inhibition of absorbance. Both ELISA kits were most 

sensitive for alachlor and had a LDD of 0.07 ug/L and 

0.06 |*g/L, for the magnetic particle-based and 

microtiter plate-based ELISA, respectively. The 

magnetic particle-based ELISA had an IC50 of 0.89 

fig/L, whereas the microtiter plate-based ELISA had 

an IC50 of 0.75 ug/L for alachlor. The compound with 

the highest cross reactivity was ESA, which had an 

IC5() of 5.4 [ig/L and a LDD of 0.19 ug/L for the

magnetic particle-based ELISA, and an IC^o of 1.7 

ug/L and a LDD of 0.12 ug/L for the microtiter platebased ELISA. The oxo-acetic acid (OXA) metabolite 

of alachlor had an IC50 of 335 fig/L and a LDD of 21 

ug/L (magnetic particle-based ELISA) and an IC50 of 

66 fig/L and a LDD of 2.7 |ig/L (microtiter plate-based 

ELISA).

The cross reactivity with ESA has caused 

problems in previous water-quality surveys for 

alachlor by ELISA because of false-positive 

detections of alachlor (Macomber and others, 1992; 

Baker and others, 1993). Furthermore, it has been 

reported that the ELISA for alachlor does not correlate 

well with GC/MS and gives more than 10-percent 

false-positive detection near the detection limit of the 

assay, from 0.10 to 0.20 jig/L (Feng and others, 1990a, 

1990b). This poor correlation has been attributed to 

the presence of high ESA concentrations as confirmed 

by liquid chromatography and mass spectrometry 

(Macomber and others, 1992). ESA and OXA are two 

major soil metabolites of alachlor (Sharp, 1988); 

however, only ESA appears to have enough significant 

cross reactivity to cause interference with ELISA 

(fig. 2).

Other soil metabolites of alachlor, such as 

chlorodiethylacetanilide and hydroxydiethylacetanilide, did not show any cross reactivity up to a 

concentration of 1,000 M-g/L. Metolachlor, another 

major corn herbicide, showed small cross reactivity, 

with an IC50 of 109 and 27 fig/L for magnetic particlebased and microtiter plate-based ELISA, respectively. 

This cross-reactivity pattern suggests that the binding 

of alachlor towards the antibody is greatly affected by 

the presence of the methoxymethyl side chain. These 

results are not surprising because the antibodies 

against alachlor were generated by the use of an 

alachlor-protein conjugate, formed through the 

chlorine-bearing carbon of alachlor via a thioether 

bond (Feng and others, 1990a, 1990b). The thioether 

may explain the cross reactivity of ESA, which has 

nearly the same structure as the hapten-protein 

conjugate used for antibody generation, whereas the 

OXA does not show any significant cross reactivity 

despite of the similar structure of the molecule.

Separation by Solid-Phase Extraction

The separation of alachlor and ESA was 

achieved by SPE using a C )8 resin. Both compounds 

adsorbed quantitatively from water onto the C | g

388
. COCH2CI

Alachlor

ICso= 0.89,0.75 

LDD= 0.07 ,0.06

CH3OCH2v . COCH2SO3H

H3CH2C CH2CH3

Ethanesulfonic acid metabolite

1050=5.4,1.7 

LDD= 0.19, 0.12

CH3OCH2,CH(CH3^ . COCH2CI

Metolachlor

ICso= 109,27 

LDD= 3.0,1.2

CH3OCH2x , COCOOH

H3CH2C CH2CH3

Oxoacetic acid metabolite

ICso= 335 ,66 

LDD= 21 , 2.7

H3CH2C

H . . COCH2CI 

N

CH2CH3

Chlorodiethylacetanilide

No cross reactivity at less than 

1,000 micrograms per liter

H3CH2C

H . . COCH2OH 

N

CH2CH3

Hydroxydiethylacetanilide

No cross reactivity at less than 

1,000 micrograms per liter

Figure 2. Cross-reactivity of antibodies with chloroacetanilide herbicides and alachlor metabolites expressed as the 

concentration required for 50-percent inhibition (IC50), in micrograms per liter, and in least detectable dose (LDD) at 

90-percent inhibition. The first value is for the magnetic particle-based ELISA and the second value is for the 

microtiter plate-based ELISA.

cartridge (table 1). Alachlor has a large capacity on 

the C | 8 resin; more than 7 L of sample could pass 

through the cartridge before a 10-percent 

breakthrough was observed. In contrast, ESA has 

considerably less capacity; a 10-percent breakthrough 

and a 100-percent breakthrough were observed after 

175 mL and 750 mL of sample, respectively, had 

passed through the cartridge (table 1). This difference 

in capacity is caused by the higher water solubility of 

ESA compared to that of alachlor. Likewise, OX A has 

a relatively small capacity on the C\% resin, with a 10- 

percent breakthrough beginning at 150 mL and a 100- 

percent breakthrough observed after 750 mL. A more 

soluble herbicide, 2,4-dichlorophenoxyacetic acid 

(2,4-D), also was examined. It was found that 2,4-D 

had a very small capacity for sorption, with a 10- 

percent breakthrough observed after only 10 mL of

sample had passed through the cartridge and a 100- 

percent breakthrough observed after 90 mL (table 1). 

These results suggest that the isolation of ionic 

compounds on C 18 resin is possible, but this isolation

Table 1. Breakthrough capacity of alachlor and other ionic 

compounds on C18 cartridges

[capacity in mL; > , greater than. , not measured]

Compound

alachlor

ethanesulfonic acid metabolite (ESA)

oxo-acetic acid metabolite (OXA)

2,4-dichlorophenoxyacetic acid (2.4-D)

10-percent 

breakthrough

>7,00()

1 75

ISO

l()

100 percent 

breakthrough

750

750

90

389
technique is quite sensitive to the aqueous solubility of 

the analyte. Thurman and others (1978) reported that 

ionic character was critical in the sorption of organic 

acids onto a macroporous acrylic resin and was a 

function of chain length. The same concept applies to 

the isolation of ionic metabolites on C ]8 resin because 

both isolations involve hydrophobic interactions.

The separation of ESA and alachlor occurs in 

the elution step. Alachlor is eluted with ethyl acetate 

without removal of ESA (table 2). ESA then is eluted 

with methanol; elution efficiency is greater than or 

equal to 90 percent. HPLC data showed that OXA 

also elutes with methanol (Pomes and others, this 

volume), but because OXA has almost no cross 

reactivity with the ELISA kits, it does not interfere 

with the analysis of ESA by ELISA.

Combining Solid-Phase Extraction and 

Enzyme-linked Immunosorbent Assay

The SPE-ELISA method for ESA is shown in 

figure 1. A 100-mL sample was passed through the 

C1 8 cartridge, and the alachlor and ESA were 

quantitatively sorbed to the resin. The alachlor was 

eluted with ethyl acetate, which, in turn, was analyzed 

by ELISA using the magnetic particle-based alachlor 

assay and then verified by GC/MS. The ESA was 

eluted with methanol, then methanol was evaporated 

to dryness, and the ESA was reconstituted with water. 

This mixture was analyzed by ELISA using the 

magnetic particle-based alachlor assay but calibrated 

with ESA standards and then verified by HPLC. It 

should be noted that solvent exchange of the organic 

solvent to an aqueous medium was performed prior to 

ELISA.

The method worked well with fortified water 

samples (table 2). Alachlor recovery from the ethylacetate extract was 100 percent and from the methanol 

extract was 0 percent, whereas 90 to 99 percent of the 

ESA was recovered from the methanol extract and 

none was recovered from the ethyl-acetate extract. 

Thus, the separation technique allowed for good 

quantitation of ESA by immunoassay. The calibration 

curve for ESA was linear for concentrations of 1.0 to 

20 jig/L. Because SPE preconcentrates the sample, 

the detection limit can be as low as 0.05 ug/L when a 

100-mL sample is used for SPE and the dried extract is 

diluted to 5 mL water.

Table 2. Recovery of alachlor and ethanesulfonic acid 

(ESA) metabolite from a C18 resin by sequential elution with 

ethyl acetate and methanol by SPE

, micrograms per liter

Concentrations in fortified 

distilled water

Percentage

« i Li of alachlor 

. recovered

 Percentage

, COA Of coA

A. 0 Mg/L alachlor + 0.20 M g/L ESA

B. 0.050 Mg/L alachlor + 0.10ng/L ESA

C. 0. lOng/L alachlor + 0.050 ng/L ESA

D. 0.025 Mg/L alachlor + 0.025 ng/L ESA

E. 0.20 ng/L alachlor + 0.0 ng/L ESA

0

100

97

105

98

92

97

99

90

0

The automated workstation combined with SPE 

performs as a liquid chromatograph with a fraction 

collector by pumping the sample through the cartridge, 

separating the compounds by sequential elution, and 

collecting the fractions for analysis by immunoassay, 

GC/MS, or HPLC. The automated SPE procedure is 

capable of reproducing the separation within ±10 

percent.

The ability of SPE to separate alachlor and ESA 

was demonstrated on ground-water samples from 

various locations. The presence of both alachlor and 

ESA in the ground-water samples resulted in an 

overestimation of alachlor when water samples were 

analyzed directly by ELISA. The discrepancy 

between the amount of alachlor measured by GC/MS 

and that measured by ELISA was proportional to the 

amount of ESA in the sample. However, when the 

water samples were passed through SPE and the ethylacetate eluate was analyzed for alachlor by ELISA, the 

concentration of alachlor agreed with that obtained by 

GC/MS. In addition, analysis of the methanol eluates 

by ELISA showed high concentrations of ESA 

whenever there was a discrepancy between alachlor 

concentrations measured by ELISA and GC/MS. This 

indicates further that ESA, which caused the false 

alachlor concentration in ELISA, was recovered in the 

methanol extract and not in the ethyl-acetate extract. 

These results suggest that the SPE-ELISA method is 

quite effective for the separation and analysis of 

alachlor and ESA.

The correlation of HPLC and SPE-ELISA for 

22 surface-water samples from reservoirs in the 

midwestern United States is shown in figure 3A. The 

correlation coefficient (r) was 0.93, and the slope of 

0.69 demonstrates that the two methods correlate quite 

well and that HPLC concentrations were generally 

smaller. The smaller ESA concentrations determined

390
z 3- 

< E

2 w SC PI

i

il

4.0

y = 0.044 -i- 0.69x 

r = 0.93

2.0 4 0 6.0 10.0

4.0 6 0 8.0

CONCENTRATION OF ETHANESirLFONlC ACID METABOLITE 

DETERMINED BY ENZYME-LINKED IMMUNOSORBENT ASSAY, 

IN MICROGRAMS PER LITER

Figure 3. Correlation of concentrations of ethanesulfonic 

acid metabolite in (A) 22 surface-water samples and (B) 20 

ground-water samples as determined by high-performance 

liquid chromatography and enzyme-linked immunosorbent 

assay.

by HPLC analysis may be attributed to high 

concentrations of humic and fulvic acids that were 

coextracted in the methanol fraction. ESA 

concentrations of less than 0.10 ng/L are difficult to 

measure using HPLC because natural organic 

substances that are recovered in the methanol fraction 

produce a large background signal in the HPLC. 

However, ground-water samples were comparatively 

easy to analyze by HPLC because of the relatively 

small quantities of humic and fulvic acids in the 

samples, which produce clean extracts. Figure 3B 

shows the high correlation between ESA 

concentrations determined by ELISA and by 

HPLC (r= 0.97). The ESA concentration in water 

samples from the various reservoirs ranged from less 

than 0.10 (ig/L to more than 10 (ig/L. Further studies 

are underway to understand the fate and transport of 

ESA in the soil and in the aquatic environment.

SUMMARY

The coupling of SPE and ELISA results in a 

reproducible immunoassay method for the trace 

analysis of alachlor and its major soil metabolite, 

ESA, in ground water and surface water. The results 

are comparable to both GC/MS and HPLC results, for

alachlor and ESA, respectively. The procedure uses 

SPE to isolate and concentrate the herbicide and 

metabolite and uses sequential elution with ethyl 

acetate and methanol to separate the two compounds. 

The method is sensitive with a quantitation limit of 

0.01 fig/L for alachlor and 0.05 (ig/L for ESA, and a 

precision of ±10 percent.

REFERENCES

Baker, D.B., Bushway, R.J., Adams, S.A.. and Macomber, 

C.S., 1993, Immunoassay screens for alachlor in rural 

wells-false positives and an alachlor soil metabolite: 

Environmental Science and Technology, v. 27, p. 562- 

564.

Bushway, R.J., Perkins. B., Savage, S.A., Lckousi. S.L.. and 

Ferguson, B.S., 1988, Determination of atrazinc 

residues in water and soil by enzyme immunoassay: 

Bulletin of Environmental Contamination and 

Toxicology, v. 42, p. 899-904.

Castellam, W.J., Vanlente. F.. and Chou. D.. 1990,

Applications of robotics in clinical laboratory: Journal 

of Automatic Chemistry, v. 12, p. 141-144.

Feng, P.C.. Wratten, S.J., Logusch, E.W.. Horton. S.R.. and 

Sharp, C.R., 1990a. Immunoassay detection method 

for alachlor-Application to analysis of environmental 

water samples: American Chemical Society (ACS) 

Symposium Series, v. 442, p. 180-192.

 1990b, Development of an enzyme-linked

immunosorbent assay for alachlor and its application 

to the analysis of environmental water samples: 

Journal of Agricultural and Food Chemistry, v. 38. 

p. 159-163.

Goh, K.S., Hernandez, J., Powell, S.J., Garretson, C,

Troiano, J., Ray, M., and Grccnc. C.D., 1991, Enzymcimmunoassay for the determination of atrazinc residue 

in soil: Bulletin of Environmental Contamination and 

Toxicology, v. 46, p. 30-36.

Macomber, Carol, Bushway. R.J., Perkins. L.B., Baker. 

D.B, Fan, T.S., and Ferguson, B.S., 1992, 

Determination of the ethancsulfonatc metabolite of 

alachlor in water by high-performance liquid 

chromatography: Journal of Agricultural Food 

Chemistry, v. 40, p. 1450-1452.

Meyer, M.T., Mills. M.S., and Thurman, E.M., 1993,

Automated solid-phase extraction of herbicides from 

water for gas chromatographic-mass spcctrometric 

analysis: Journal of Chromatography, v. 629. p. 55-59.

Morris, Z., and Ruthann. K., 1988, Solid phase extraction 

for sample preparation: Phillipsburg, New Jersey. J.T. 

Baker Inc., p. 7.

391
Schlaeppi, J.M., Fory, W., and Ramsteiner, K., 1989,

Hydroxyatrazine and atrazine determination in soil and 

water by enzyme-linked immunosorbent assay using 

specific monoclonal antibodies: Journal of Agricultural 

Food Chemistry, v. 37, p. 1532-1538.

Sharp, D.B., 1988, Alachlor, in Hearney, P.C., and 

Kaufman, D.D., eds., Herbicides Chemistry, 

degradation and mode of action: New York, Dekker, 

p. 301-333.

Thurman, E.M., Malcolm, R.L.. and Aiken, G.R., 1978, 

Prediction of capacity factors for aqueous organic 

solutes adsorbed on a porous acrylic resin: Analytical 

Chemistry, v. 50, p. 775-779.

Thurman, E.M., Meyer. M.T., Pomes, M.L., Perry. C.A., 

and Schwab, A.P.. 1990, Enzyme-linked 

immunosorbent assay compared with gas 

chromatography/mass spectromctry for the 

determination of triazine herbicides in water: 

Analytical Chemistry, v. 62, p. 2043-2048.

Tippins, B., 1988, Solid-phase extraction fundamentals: 

Nature, v. 334, p. 273-274.

Vanderlaan, Martin. Watkins, Bruce. and Stanker. Larry. 

1988, Environmental monitoring by immunoassay: 

Environmental Science and Technology, v. 22, p. 247- 

254.

Van Emon, J.M., and Lopez-Avila, Viorica, 1992,

Immunochemical methods for environmental analysis: 

Analytical Chemistry, v. 62. p. 79A-88A.

392
Isocratic Separation of Alachlor Ethanesulfonic Acid,

Alachlor Oxoacetic Acid, and Hydroxyatrazine by

Reversed-Phase Liquid Chromatography

By Michael L. Pomes 1 , Douglas F. Holub1 , Diana S. Aga, 1 and EM. Thurman 1

Abstract

The polar nature of two alachlor metabolites, alachlor ethanesulfonic acid and alachlor 

oxoacetic acid, makes their detection by gas 

chromatography/mass spectrometry impossible 

without derivatization. Enzyme-linked immunosorbent-assay techniques cannot distinguish 

alachlor metabolites from alachlor; thus, 

reversed-phase high-performance liquid chromatography with photodiode-array detection is 

required for the separation and spectral detection 

of these analytes. Use of an isocratic, methanol/10 

millimolar sodium phosphate dibasic buffer mixture as the mobile phase with reversed-phase 

(C-18) high-performance liquid chromatography 

allows for the separation and quantification of 

alachlor ethanesulfonic acid, alachlor oxoacetic 

acid, and hydroxyatrazine at concentrations 

greater than or equal to 0.10 micrograms per liter 

in 100-milliliter water samples. The buffer in the 

methanol mixture provides cations that ion pair 

with the alachlor metabolite anions to decrease 

polarity and promote nonpolar interactions 

between the analytes and the reversed-phase 

column. Enhanced chromatographic resolution 

results from injecting samples in a matrix that 

contains less methanol than the mobile phase. The 

disparity between the methanol content of the 

sample matrix and the mobile phase focuses the 

analytes at the head of the chromatographic 

column to produce sharper peaks because the 

analytes are less soluble in the sample matrix than

U.S. Geological Survey, Lawrence, Ks.

in the mobile phase. Results by reversed-phase 

high-performance liquid chromatography show 

that one can detect alachlor ethanesulfonic acid in 

surface-water, along with lesser concentrations of 

alachlor oxoacetic acid and hydroxyatrazine.

INTRODUCTION

While developing an enzyme-linked immunosorbent assay (ELISA) for alachlor, Feng and others 

(1990) reported that their technique produced false 

positive results. The ELISA tested positive for 

alachlor, but the presence of alachlor could not be confirmed by gas chromatography/mass spectrometry 

(GC/MS). Recently, one of the alachlor metabolites, 

2-[(2,6-diethylphenyl)(methoxymethyl)amino-]-2- 

oxoethanesulfonic acid (ESA), analyzed by Feng and 

others (1990), has been identified as the source of 

many alachlor ELISA false positives detected in 

ground-water samples (Macomber and others. 1992; 

Baker and others, 1993).

ESA exists as an anion in solution and is too 

polar to be detected by GC/MS techniques without 

derivitization. ESA can be extracted from water samples using solid-phase extraction (SPE) cartridges, but 

the polarity of ESA requires elution of the cartridges 

with methanol, a polar solvent (Macomber and others, 

19^2; Aga and Thurman, 1994). Methanol also is 

used in the mobile phase for the analysis of ESA by 

high-performance liquid chromatography (HPLC) 

(Macomber and others, 1992). This paper addresses 

the HPLC separation of ESA oxoacetic acid [(2,6- 

diethylphenyl)-(methoxymethyl)amino]oxoacetic 

acid] (another soil metabolite of alachlor) (Feng and 

others, 1990), and hydroxyatrazine in surface-water 

samples. Like the alachlor metabolites, hydroxyatrazine is relatively polar, can not be analyzed by GC/MS 

techniques without derivatization, and elutes from the

393
SPE cartridges with methanol (D.S. Aga, U.S. 

Geological Survey, oral commun., 1993).

PREVIOUS WORK

ESA and Oxoacetic Acid

Macomber and others (1992) reported reversedphase chromatography techniques for the analysis of 

ESA and oxoacetic acid using a C-18 column. Before 

analysis, 1 mL sodium phosphate dibasic buffer was 

added to the 1 -mL methanol extract containing the 

analytes and injected on the HPLC with a 50-|iL 

injection. The mobile phase used to separate ESA 

consisted of a 50/50 (v/v: volume to volume) mixture 

of methanol and 50 mM (millimolar) sodium phosphate dibasic (Na2HPO4) pH 7.0 buffer. The buffer 

allowed sodium (Na) cations to pair with the analyte 

anions to produce neutral compounds that interact 

with the C-18 column. ESA and oxoacetic acid were 

detected with a UV detector set at wavelength of 

205 nm.

Hydroxyatrazine

Various chromatographic techniques have been 

used for the detection and quantification of 

hydroxyatrazine. Rustum and others (1990), Qain and 

others (1991), and Wang and others (1992) have used 

reversed-phase chromatography with C-18 columns to 

determine concentrations of hydroxyatrazine in soil 

extracts. Vermeulen and others (1982) used a 

reversed-phase C-8 column for analyzing atrazine and 

selected metabolites because they found some of the 

metabolites too polar to be retained on C-18 columns.

Mobile phases used to analyze hydroxyatrazine 

ranged from gradient elutions to combinations of acetonitrile and buffers or water at different pH's. Wang 

and others (1992) used a 50/50 acetonitrile/water 

mobile phase containing 3.18 mM hydrochloric acid 

(HC1) to give a pH of 2.5. Rustum and others (1990) 

used a gradient elution consisting of 0.01 M potassium 

phosphate dibasic (K^HPC^) (pH 2.0 stabilized with 

phosphoric acid) and acetonitrile (5 percent for 0 to 

10 minutes, 20 percent for 10 to 15 minutes, 30 percent for 15 to 20 minutes, and 70 percent for 20 to 35 

minutes). Qain and others (1991) also used a 50/50 

acetonitrile/water mixture but added 0.020 M n-heptane sulfonic acid (pH of 2.8) and phosphoric acid to

perform ion-pair reverse-phase liquid chromatography. Vermeulen and others (1982) used a mobile phase 

consisting of 40/60 methanol/water with 50 mM 

ammonium acetate dissolved in each component to 

produce a solution pH of 7.4.

Detection strategies used to identify 

hydroxyatrazine range from single wavelength detection to use of photodiode array (PDA) detectors. PDA 

detectors consist of a ultraviolet/visible lamp (UV/ 

VIS), diffraction grating, detector cell, and a siliconchip array with slots dedicated to individual wavelengths (Harris, 1991). Radiant energy striking a 

particular slot causes that part of the chip to partially 

discharge. During recharging, a computer monitors 

the amount of energy required to recharge the entire 

array. Regions of the spectrum where maximum 

absorbance occurs correspond to the parts of the diode 

array that needed the most recharging. Rustum and 

others (1990) used a wavelength of 230 nm for the 

detection of hydroxyatrazine with deethylatrazine, 

deisopropylatrazine, deethylhydroxyatrazine, and 

deisopropylhydroxyatrazine. Qain and others (1990) 

used a PDA detector that scanned the absorption spectrum of each peak between 190 and 300 nm. They 

reported the maximum ultraviolet absorbance of 

hydroxyatrazine to be 215 nm. Vermeulen and others 

(1982) used a wavelength of 220 nm for the detection 

of hydroxyatrazine with deethylatrazine, deisopropylatrazine, deethylhydroxyatrazine, and deisopropylhydroxyatrazine. Finally, Wang and others (1992) used a 

wavelength of 254 nm for the detection and quantification of atrazine and hydroxyatrazine.

METHODS

Solvents and Standards

Solvents used during the course of this study 

included ethyl acetate and HPLC grade methanol. 

Alachlor ethanesulfonic acid, alachlor oxoacetic acid 

and hydroxyatrazine were prepared as 1-mg/mL 

(milligram per milliter) stock solutions in HPLC grade 

methanol and stored in a freezer.

Apparatus and Instrumentation

Syringes used for the preparation of standards 

and the spiking of appropriate volumes of water 

included 10-u,L and 100-u,L gas-tight syringes.

394
Automated solid-phase extraction of the analytes from 

water was accomplished with the automated workstation using C-18 SPE cartridges. Volume reduction 

of the eluted herbicide samples was accomplished 

with nitrogen evaporation. Analysis and quantitation 

of the herbicide metabolites in this study was completed with a liquid chromatograph and controller. A 

3-|am particle size C-18 column, 100 mm x 4.6 mm, 

was used in the HPLC for separation. The autosampler on the HPLC used a 250-jaL syringe, and 

detection and spectral analysis of analytes used a 

PDA. The PDA detector allowed the simultaneous 

monitoring of eight separate wavelengths and the 

scanning of spectra between 190 and 600 nm.

Extraction and Analysis

The analytes under consideration for this study 

were extracted from 100-mL sample volumes using 

C-18 SPE cartridges and the automated workstation. 

Details of the extraction process are discussed in Aga 

and Thurman (1994). The final step of the process 

consisted of separate elutions of ethyl acetate and 

methanol. The cartridge was eluted with 3.0 mL ethyl 

acetate to remove the less polar compounds such as 

atrazine and alachlor. This fraction then was prepared 

for GC/MS analysis. The SPE cartridge was eluted 

with 5.0 mL of methanol into a centrifuge tube. 

Because of their polarity, hydroxyatrazine and the 

alachlor metabolites remained on the cartridge during 

the ethyl acetate elution but were removed by the more 

polar methanol. The 5.0-mL methanol extract was 

spiked with metribuzin (4-amino-6-(l,l-dimethylethyl)-3-(methylthio)-1,2,4-triazin-5(4H)-one), the 

internal standard, then evaporated to dryness with a 

nitrogen stream at 45 °C. One hundred microliters of a 

20/80 methanol/10 mM Na2HPO4 buffer mixture was 

added to the centrifuge tube, and the tube was vortexed. The sample then was transferred to a crimpcapped microvial for analysis on the HPLC.

The following operating conditions were used to 

detect and quantify the analytes of interest in this 

study. The mobile phase consisted of a 31.5/68.5 

methanol/10 mM Na2HPO4 buffer mixture. The 

mobile phase, heated to a temperature of 60 °C, flowed 

through the standard bore C-18 column at 1.20 mL/min. 

The detector wavelength was set at 200 nm with a 4-nm 

bandwidth; the reference wavelength was set to 450 

nm with a 80-nm bandwidth. Finally, an injection 

volume of 90 |aL was used to introduce samples to the

HPLC. Use of the 90-|aL injection in conjunction with 

a 100-mL sample volume allowed the separation and 

detection of alachlor ethanesulfonic acid, alachlor 

oxoacetic acid, and hydroxyatrazine at concentrations 

greater than or equal to 0.10 ng/L using 100-mL 

water samples.

The last step in the HPLC analysis involved the 

identification and quantitation of the herbicide metabolite analytes in the water sample. Two options in the 

HPLC operating software were used for the identification. The first, a peak purity check, measured the 

spectrum on the upslope of the peak, top of peak, and 

downslope of the peak. If the peak consisted of coeluting analytes, different, nonmatching spectra would 

be collected. If a peak consisted of one analyte, the 

upslope, top, and downslope spectra would all match; 

therefore, the peak was considered pure. The second 

option, an automated library search, matched analyte 

spectra against standards stored in the library. Analyte 

areas and peak heights were quantified relative to 

metribuzin.

DISCUSSION

The UV spectra and the chemical structures of 

the analytes of interest in this study are illustrated in 

figure 1. Of the three compounds and internal standard, both ESA and oxoacetic acid had maximum UV 

absorbances of 196 nm. Because ESA and oxoacetic 

acid had similar UV spectra, a wavelength of 200 nm 

was used as the detector wavelength as opposed to the 

205-nm wavelength used by Macomber and others 

(1992). The choice of 200 nm, which is less than the 

UV cutoff of 205 nm for methanol hence necessitated 

zeroing the detector with a 100-percent methanol 

mobile phase before any analysis was completed. 

Hydroxyatrazine and metribuzin had UV maxima 

greater than 200 nm, and the wavelengths corresponding to their maximum absorbances could be monitoredwith a diode-array detector. However, only the 

200-nm wavelength was monitored because the peak 

purity and automated library search routines did not 

function properly if more than one wavelength was 

monitored.

Other changes made to the operating conditions 

originally used by Macomber and others (1992) arose 

from addition of oxoacetic acid as an analyte. In chromatographs completed at ambient temperatures, 

oxoacetic acid standards had two peaks rather than one 

(fig. 2A). Having two peaks in a standard rather than

395
ALACHLOR ETHANESULFONIC ACID 

UV MAXIMUM=196 nanometers

ALACHLOR OXOACETIC ACID 

UV MAXIMUM=196 nanometers

HYDROXYATRAZINE 

UV MAXIMUM=216 nanometers

METRIBUZIN 

UV MAXIMUM=212 nanometers

ALACHLOR ETHANESULFONIC ACID

ALACHLOR OXOACETIC ACID 

OH

CH, ~

CH3

<2 CH3

HYDROXYATRAZINE

SCH3

METRIBUZIN

200 250 300 350 

WAVELENGTH, IN NANOMETERS

Figure 1. Ultraviolet (UV) maxima and chemical structures of alachlor ethanesulfonic acid, alachlor oxocetic acid, 

hydroxyatrazine, and metribuzin.

one may indicate a contaminant in the standard. Heating 

the mobile phase caused the two peaks to blend until 

essentially one peak occurred at temperatures greater 

than 50 °C (fig. 2B). The presence of a contaminant in 

the oxoacetic acid standard was discounted by the fact

that a pure peak was indicated by the peak purity test at 

temperatures greater than 50 °C. Apparently, some equilibrium must exist for oxoacetic acid between the two 

isomers so that two peaks occurred at ambient temperatures, and one peak occurred at higher temperatures.

396
220

200

180

160

140

120

100

80

CO

2 60

Q 

CL 

< 

Q

1

CO

LU 

O

i

CC

O

CO

m

40

20

0

240

220

200

180

160

140

120

100

80

60

40

20

0

ALACHLOR OXOACETIC ACID 

48 nanograms per microliter 

27 degrees Celsius

1

ALACHLOR OXOACETIC ACID 

- 48 nanograms per microliter 

55 degrees Celsius

B

2 3 

TIME, IN MINUTES

Figure 2. Chromatographs of oxocetic acid at (A) 27 degrees Celsius showing two peaks and (B) 55 degrees Celsius 

showing a single peak. Oxoacetic acid standard mixed in sample matrix consisting of 50/50 mixture of methanol and 20 

millimolars sodium phosphate dibasic pH 7.0 buffer. Mobile phase consists of a 44/56 mixture of methanol and 20 millimolars 

sodium phosphate dibasic pH 7.0 buffer flowing at a rate of 1.2 milliliters per minute through the column.

397
A rotational isomer of oxoacetic acid may be responsible for the second peak observed at ambient temperatures (Drew Kline, Monsanto, oral commun., 1993). 

Because of oxoacetic acid rotational isomers, the 

mobile phase was heated to 60 °C. Although one 

oxoacetic acid peak was evident at 60 °C, the peak had 

a width exceeding 1 minute when measured along the 

baseline. Additional work was undertaken to decrease 

peak width. The oxoacetic acid peak width decreased 

and peak height increased when the concentration of 

NaoHPC^ was decreased in the mobile phase and 

sample matrix. Specifically, peak height nearly doubled when the concentration of Na 2HPO4 was 

decreased from 20 to 10 mM. Use of the heated 

mobile phase and decreased buffer concentration 

enhanced the ability to detect oxoacetic acid.

Two aspects involving the width of chromatographic peaks became apparent in deciding the appropriate conditions for the analysis of ESA, oxoacetic 

acid, and hydroxyatrazine. Peak width could be 

broadened noticeably if the methanol part of the 

sample matrix exceeded that of the mobile phase. 

Alternately, peak width could be narrowed, and peak 

height noticeably increased if the methanol part of the 

mobile phase exceeded that of the sample matrix. In 

the latter case, increased chromatographic resolution 

resulted from injecting samples in a matrix that contained proportionally less methanol than did the 

mobile phase. The disparity between the methanol 

content of the sample matrix and the mobile phase 

focused the analytes at the head of the chromatographic column to produce sharper peaks because the 

analytes were less soluble in the sample matrix than in 

the mobile phase. The chromatographic conditions 

used for the set of analytes under discussion used a 

sample matrix consisting of 20 percent methanol and a 

mobile phase consisting of 38.5 percent methanol.

The order of elution of the analytes in a 25 ng/fiL 

standard mixture is illustrated in figure 3A. Oxoacetic 

acid has an average retention time of 2.81 minutes; 

hydroxyatrazine, 3.14 minutes, ESA, 3.40 minutes; 

and metribuzin, 4.71 minutes. Results by re versed - 

phase high-performance liquid chromatography show 

that one can detect alachlor ethanesulfonic acid in surface water, along with lesser concentrations of alachlor 

oxoacetic acid and hydroxyatrazine. Chromatographs 

of surface-water samples with positive detections of 

oxoacetic acid, ESA, and hydroxyatrazine are illustrated in figures 3B and 3C. Additionally, the

chromatograms of the surface-water samples share 

large peaks that elute during the early part of the analysis. These large peaks probably are due to the humic 

material present in surface water. Humic material 

accumulates on the SPE cartridges and is partially 

removed from the cartridge with methanol (Mills and 

Thurman, 1992) to the extent that many of the sample 

extracts are yellow-colored prior to injection on the 

HPLC. The presence of humic material in these 

ction of oxoacetic acid, hydroxyatrazine, and ESA by 

causing elevated baselines. More work needs to be 

done to minimize the effect of humic material using 

reversed-phase HPLC.

SUMMARY

Hydroxyatrazine was isolated by solid-phase 

extraction and methanol elution of two polar alachlor 

metabolites, oxoacetic acid and ESA. The polarity of 

these metabolites requires that the buffer used in the 

mobile phase for HPLC analysis must contain cations 

to pair with the anionic alachlor metabolites. Using a 

wavelength of 200 nm (closer to the UV maxima for 

oxoacetic acid and ESA) and an injection volume of 

90 fiL, analyte concentrations greater than or equal to 

0.10 fig/L were detected and quantified from 100-mL 

water samples. Chromatography of oxoacetic acid 

was improved by heating the mobile phase to 60 °C to 

assure that only one oxoacetic acid peak was detected, 

and decreasing the concentration of the buffer to 10 

mM greatly increased peak height. Additional chromatographic resolution was achieved by injecting a 

sample matrix containing proportionally less methanol 

than the mobile phase. The disparity between the 

methanol content of the sample matrix and the mobile 

phase focuses the analytes at the head of the chromatographic column to produce sharper peaks because the 

analytes are less soluble in the sample matrix than in 

the mobile phase. Results by reversed-phase HPLC 

show that one can detect alachlor ethanesulfonic acid 

in surface water, along with lesser concentrations of 

alachlor oxoacetic acid and hydroxyatrazine. However, the presence of humic material in surface-water 

samples, isolated by solid-phase extraction and eluted 

into the extract with methanol, poses a potential interference to the method.

398
140

120

100

80

60

40

20

o ai

CO O.UUU fz*"^

0 2,500

CC

Q

< 2,000

CO

- 1,500

LJJ 

O

Z

CD 1 - 000

CC

o

CO

9 500

<L

_J_J

5 0

3,000

2,500

2,000

1,500

1,000

500

n

i i i

-

-

-

jl

LU

Z

cr > 

\ >

\ x

r V V g> J . ^-^~^L_

B

-

-

-

z

N

CD

CC t 

LU

5

yv ,

.

-

-

" /H

a

o

o

Q Z

§ §

O D

ui § 

^ <

1 H | «

\i § 2 11 _) T

1 A 5 i

/ ^^ < ^

J , ;^^" -^-i A^_

C

-

-

-

z

N

CD

CC

LU:E

345 

TIME, IN MINUTES

Figure 3. Chromatographs of (A) 25 nanograms per microliter standard mixture of oxoacetic acid hydroxyatrazine, ESA, and 

metribuzin, 10-microliter injection, (B) surface-water sample showing positive identifications of oxoacetic acid and ESA, and 

(C) surface-water sample showing positive identification of hydroxyatrazine.

399
REFERENCES

Aga, D.S., Thurman, E.M., and Pomes, M.L., 1994, Determination of alachlor and its sulfonic acid metabolite in 

water by solid-phase extraction and enzyme-linked 

immunosorbent assay: Analytical Chemistry, 

v. 66, p. 1495-1499.

Baker, D.B., Bushway. R.J., Adams. S.A., and Macomber. 

Carol, 1993, Immunoassay screens for alachlor in rural 

wells False positives and an alachlor soil metabolite: 

Environmental Science and Technology, v. 27. 

p. 562-564.

Feng, P.C.C., Wratten, S.J.. Horton, S.R., Sharp, C.R., and 

Logusch, E.W., 1990, Development of an enzymelinked immunosorbent assay for alachlor and its application to the analysis of environmental water samples: 

Journal of Agricultural and Food Chemistry, v. 38, 

p. 159-163.

Harris, D.C.. 1991, Quantitative chemical analysis: New 

York, W.H. Freeman and Company, 782 p.

Macomber, Carol, and others, 1992. Determination of the 

ethane sulfonate metabolite of alachlor in water by 

high-performance liquid chromatography: Journal of 

Agricultural and Food Chemistry, v. 40, p. 1450-1452.

Mills, M.S., and Thurman, E.M., 1992, Mixed-mode isolation of triazine metabolites from soil and aquifer sediments using automated solid-phase extraction: 

Analytical Chemistry, v. 64, p. 1985-1990.

Qain, Wenheng, Schultz, N.A., Stuart, J.D., Hogan, J.C., Jr.. 

and Mason, A.S., 1991, Determination of atrazinc and 

hydroxyatrazine residues in soil by ion-pair, revcrsedphase high performance liquid chromatography: Journal of Liquid Chromatography: v. 14. p. 1327-1392.

Rustum, A.M., Ash, Steve, Saxcna Adcsh, and Balu, K... 

1990. Rcverscd-phase high-performance liquid 

chromatographic method for the determination of soilbound [ 14C]atrazinc and its radiolabeled metabolites in 

a soil metabolism study: Journal of Chromatography. 

v. 514, p. 209-218.

Vermeulen, N.M.J., and others. 1982, Separation of atrazinc 

and some of its degradation products by highperformance liquid chromatography: Journal of 

Chromatography, v. 240, p. 247-253.

Wang, Zhendl, Gamble, D.S., and Langford, C.H., 1992, 

Interaction of atrazine with Laurcntian soil: Environmental Science and Technology, v. 26, p. 560-565.

400
Use of 15N to Trace Movement of Nitrogen Fertilizer at a

Field Plot

By Dale W. Blevins 1 , Donald H. Wilkison 1 , Steven R. Silva2, and Brian P. Kelly 1

Abstract

A simplified, lower-cost technique for 

preparation of water samples for nitrogen isotope 

(5 N) analysis of nitrate (NO3 ) is being tested. 

Anion exchange resins were used to concentrate 

nitrate in the field and samples were prepared for 

5 15N - NO3 (ratio of 15N to 14N in NO3 , relative 

to atmospheric N, expressed in per mil) analysis 

using a modification of the Dumas combustion 

method. Initial test results of the 5 l5N - NO3 technique indicate that a precision of 0.05 per mil can 

be achieved. More than 99 percent of nitrate in 

solution is sorbed onto the resins. Hydrochloric 

acid was used to elute 95 to 100 percent of the 

nitrate from the resins. Chloride and organic 

molecules can interfere with 5 I5N - NO3 values, 

but not at concentrations detected at the field plot.

The 15N-labeled fertilizer was applied to a 

0.1 -acre field plot in May 1992 to determine the 

quantity of nitrogen fertilizer reaching ground 

water. Twenty-two percent of the labeled fertilizer 

was removed from the plot in harvested grain. 

Soil-gas analyses indicated no significant denitrification of the applied fertilizer. Labeled fertilizer 

accounted for 34 percent of the nitrate in runoff in 

the first storm after application. Small quantities 

of labeled fertilizer were detected 1.5 feet below 

ground surface 1 month after application, but 

evapotranspiration prevented further downward 

movement until November 1992. Labeled fertilizer was detected in shallow wells within 6 days 

of the first recharge event in November 1992.

U.S. Geological Survey, Independence. MO 

2 U.S. Geological Survey, Mcnlo Park, CA

The 5 15N - NO3 values in deep wells increased 

1 month later. Labeled fertilizer accounted for as 

much as 0.74 milligram per liter of the nitrate in 

ground water.

INTRODUCTION

Water from about 20 percent of farmstead wells 

in Missouri contains concentrations of nitrate (NO3 ) 

larger than the Missouri drinking-water standard of 

10 mg/L as nitrogen [(N); Mesko and Garlson, 1988; 

Sievers and Fulhage, 1990; Ziegler and others, 1993a; 

Ziegler and others, 1993b]. Although N fertilizers are 

a suspected NO3 source, other sources, such as animal 

waste and septic systems, have commonly been implicated. This study was designed to determine the input 

of NO} fertilizer to NO3 concentrations in ground 

water. The nitrogen isotope composition (5 I5N), in 

combination with ground-water chemistry and hydrologic monitoring, was used to investigate the movement of NO3 through the hydrologic system from 

fertilizer applied to a field plot.

In order to perform a large number of 8 I5N 

analyses for this study, a new technique for the collection and preparation of NO3 from natural waters for 

determination of 5 I5N - NO3 (ratio of I5N to I4N in 

NO3 , relative to atmospheric N, expressed in per mil) 

analyses was developed in collaboration with the 

Stable Isotope Tracers Project of the National 

Research Program of the U.S. Geological Survey, 

Menlo Park, California. The technique combines the 

use of anion-exchange resins to concentrate NO3 in 

the field and the Dumas combustion method (Fiedler 

and Proksch, 975) of preparation as modified by 

Kendall and Grim (1990). The combination of these 

methods facilitates the collection, concentration, storage, and shipment of NO3 samples, and decreases the 

time and expense involved in laboratory preparation.

401
The purpose of this paper is to present preliminary 

results of a study of nitrogen isotope-tracer movement 

beneath a field plot and a test of the sample-preparation technique.

USE OF ANION EXCHANGE RESINS FOR 

NITROGEN ISOTOPE ANALYSIS

Field activities for the resin technique involved 

the concentration of dissolved NO3 onto anionexchange resins. Filtered-sample solutions were 

dripped through resin-filled columns at a uniform flow 

rate. Sample solutions were analyzed to confirm that 

more than 99 percent of NO3 was removed from the 

sample.

The NO3 can be removed from the resins with 

an elutant that either contains anions with a stronger 

affinity for the resin or a much higher concentration of 

an ion with lower affinity. After some experimentation, hydrochloric acid (HCI) was chosen as the 

elutant (table 1). This allowed for the concentration of 

NO3 [initially in a 250 to 1,000 mL sample] into 

15 mL of desorbate. Nitrate analyses of the desorbate 

indicated that at least 95 percent of the NO3 was 

removed from the resins and preliminary isotope data 

indicated that nearly 100 percent of the NO3 was 

recovered from the resins. Initial experiments, using a 

solution of known 5 I5N - NO3 , indicate that a precision of about 0.05 per mil is possible. Also, the concentrated desorbate facilitates sample storage, 

shipment, and preparation for the mass spectrometer 

by decreasing sample volume. It also eliminates the 

need for mercuric chloride (HgCI 2 ) or sulfuric acid 

(H2SO4 ) preservatives.

The possible effects of interfering anions commonly present in natural waters are being investigated. 

Sample solutions of a potassium nitrate (KNO3 ) standard with a known 5 N - NO3 were prepared. These 

standards were spiked with differing concentrations of

Table 1. Summary of elutant solutions, percent recovery of nitrate, and associated problems

[mL, milliliters; <, less than; M, molar, CaCl2, calcium chloride; >, greater than; SrQ2> strontium chloride; KCI, potassium chloride; H2SO4, sulfuric 

acid; HCI, hydrochloric acid]

Elutant 

solution

Percent recovery 

of nitrate in 

desorbate

Associated 

problems

Saturated potassium 

citrate

Saturated calcium 

citrate

1.25MCaCl2 

5.0MCaCl2

2.6 M SrCl2 

4.0 M KCI

0.4 M H2SO4 

1.0MHC1 

3.0 M HCI

<25

<25

25 

>95

>95 

>95

50-80 

>95

Poor recovery of nitrate; reaction 

with Vycor glass tubes causes 

combustion tube failure

Poor recovery of nitrate

Poor recovery of nitrate

Hydrous CaCl2 salt cumbersome 

to handle and dry

Hydrous SrCl2 salt cumbersome 

to handle and dry

Reaction with Vycor glass causes 

combustion tube failure

Poor recovery of nitrate 

Poor recovery of nitrate

Sample must be neutralized 

before drying

402
chloride (CI) and dissolved organic carbon (DOC). 

Initial results indicated large concentrations of CI and 

DOC can prevent some of the NO3 from being 

adsorbed, but not at concentrations detected in water 

sampled at the field study site.

Final laboratory preparation for isotopic analysis involved neutralizing and freeze-drying the desorbate, and combusting the dried salt by the Dumas 

technique (Fiedler and Proksch, 1975). As modified by 

Kendall and Grim (1990), the Dumas technique disassociates NO3 salts at high temperature in a closed tube 

in the presence of copper (Cu), copper oxide (CuO), 

and calcium oxide (CaO). The Cu and CuO convert 

various N species to nitrogen gas (N2 ), while the CaO 

reacts with residual water (H2O) and carbon dioxide 

(CO2). The combustion method, in comparison to the 

traditional method outlined by Bremner (1965) and 

Bremner and Edwards (1965), is much faster (thereby 

decreasing laboratory costs) and eliminates many 

possible sources of contamination and isotopic 

fractionation.

MOVEMENT OF 15N AT THE FIELD PLOT

A tracer plot was installed at the Missouri 

Management Systems Evaluation Area near Centralia, 

Missouri (fig. I). The plot covers 0.1 acre and has 

17 in. of silty loam topsoil overlying about 24 in. of 

low- permeability claypan. The claypan is underlain 

by about 55 ft of loess and glacial till that comprise the 

surficial aquifer.

Plot instrumentation includes a rainfall collector, a continuous-recording flume, an automatic runoff 

sampler, interflow collectors, 2 neutron-probe access 

tubes, 8 gravity lysimeters, 8 suction lysimeters, 

14 monitoring wells, and 6 tensiometers (fig. 1). The 

plot is bordered by a 6-in.-high berm and a subsurfacewater barrier to prevent ofTsite runoff and soil water 

from moving onto the plot. Conventional tillage was 

used on corn during 1991 and 1992, and fertilizer was 

applied at a rate of 83 Ibs per acre in 1991. Background concentrations of various N species and 

' 5N - NO3 were determined from May 28, 1991, 

through May 12, 1992. The dominant species of N in 

soil and ground water was NO3 , however, organic N 

concentrations were substantial in runoff. On May 13, 

1992, 3,800 g of NO3 fertilizer with a S I5N value of 

3,080 per mil (spiked fertilizer) was applied as a liquid 

and tilled into the top 4 in. of soil just before planting. 

This enrichment was chosen so that even small contributions of spiked fertilizer could be detected in surface 

and ground water but low enough to prevent analytical 

problems. The cost of the enriched fertilizer was 

$6,000.

A hydrologic budget was calculated for the plot 

from May 28, 1991, through May 15, 1992 (Kelly and 

Blevins, 1993). About 53 percent of the 30.3 in. of 

rainfall was lost to evapotranspiration, 36 percent 

recharged ground water, 10 percent was lost to runoff, 

and 1 percent was lost as interflow. The 5 I5N - NO3 

values in three samples of rainfall were 0.2, 0.5, and 

1.2 per mil. Data from the National Atmospheric Deposition Program (1992) collected at a site about 50 mi 

south of the plot indicate that annual loadings of N in 

rainfall on the plot average about 180 g per year or 

about 4.8 percent of the N applied as fertilizer.

The 5 I5N - NO3 values in two prespike storms 

were 11.4 and 12.2 per mil (table 2). Background total 

15N in four topsoil samples ranged from 7.7 to 9.1 per 

mil and averaged 8.5 per mil. Therefore, the background 8 15N in runoff was about 10 per mil. The first 

runoff event following spike application had one of 

the larger NO3 concentrations (2.1 mg/L) and the largest I5N enrichment from the spike. The 8 I5N - NO3 

value in the runoff sample was 1,065 per mil, and the 

spike was applied as NO3 . Therefore, 34 percent of 

the NO3 in the storm-composited runoff sample was 

from the recently applied spiked fertilizer. Nitrogen in 

harvested grain accounted for 22.2 percent of the 

spiked fertilizer, leaving 65.1 Ibs per acre of the spiked 

N available for leaching, runoff, and denitrification.

Denitrification in clay-rich soils and aquifers 

can cause a significant loss of N. However, analyses of 

soil-gas samples collected after application of the 

spiked N fertilizer indicate that no significant denitrification occurred. The N2 to argon (Ar) ratios were 

nearly identical to those in the atmosphere and 8 I5N 

values of soil gas were nearly identical to those in the 

atmosphere (table 3). Also, concentrations of 

dissolved oxygen in ground water ranged from 1.9 to 

10.1 mg/L. Near zero concentrations of dissolved 

oxygen are expected if denitrification is substantial. 

Therefore, denitrification is not present in the saturated zone.

The spiked fertilizer was detected in the 

shallowest suction lysimeters (1.5 ft deep) on June 1 

(125 per mil); however, it was not detected in soil 

moisture at depths of 3.0 and 4.0 ft during the growing 

season (from May through October). Values of 8 15N - 

NO3 in suction lysimeter samples collected in November and December ranged from 516 to 835 per mil.

403
O

Gage house, 

automatic sampler, 

and rainfall collector

Q

Tile drains intersect / 

water at the top ' 

of the claypan soil

Location of study area

<

 

 

 A 

ow1 oM1 0E2

Owo r>EO

ow2 o"2 sE1

 A 

» i / » \

t 1 i * i

k' \ ! / '' \ \ \ 1 1 \ »t

\\\ '/ Gravity lysimeters \» 

\M // intercept preferential flow \\

v *! * *

Subsurface '

flrrps«f chflmh*»r5

O

0 10 20 3C

I i i 1 I i ' L

0 5

~)

O

^v

Bermand

subsurface

water barrier

J

JFEET

10 METERS

EXPLANATION

O STILLING WELL AND STAGE 

^-^ RECORDER

() INTERFLOW COLLECTOR 

A NEUTRON PROBE ACCESS TUBE 

 GRAVITY LYSIMETER

 TILE DRAIN

 SUCTION LYSIMETER

O MONITORING WELL

^E1 MONITORING WELL WITH 

(§ CONTINUOUS WATER-LEVEL 

RECORDER AND IDENTIFIER

 TENSIOMETER

Figure 1. Instrumentation at study plot.

404
s15 Table 2. Analyses of 8 5N in samples collected at the field plot near Centralia, Missouri

[--; no data]

Source

(«g. 1)

WellEO

Well El

Well E2

Well Ml

WellM2

Well WO

WellWl

WellW2

Well 

depth 

(in feet)

11.8

15.0

25.0

15.0

23.2

11.8

19.0

28.2

Date 

collected

5-11-92

8-10-92

11-18-92

12-16-92

6-28-92

9-17-91

12-03-91

5-11-92

8-10-92

11-18-92

12-16-92

6-28-91

5-12-92

8-11-92

11-19-92

12-16-92

5-11-92

8-10-92

11-19-92

12-16-92

5-11-92

5-11-92

8-10-92

11-18-92

12-16-92

6-28-91

9-17-91

5-11-92

8-10-92

11-19-92

12-16-92

5-11-92

8-10-92

11-19-92

12-17-92

S1SN 

(total 

nitrogen, 

per mil)

~

-

__

-

-

_

-

-

_

~

~

_

-

--

-

-

~

81SN - NO3 

(dissolved nitrate, 

per mil)

10.3

9.8

42.3

32.7

13.5

13.1

13.7

13.2

13.4

14.7

33.6

14.5

15.5

15.9

16.2

16.2

11.1

11.6

11.3

56.5

14.0

9.9

13.4

14.7

33.6

11.0

10.9

11.3

11.2

11.1

11.3

15.1

15.1

15.1

15.0

405
Table 2. Analyses of 615N in samples collected at the field plot near Centralia, Missouri--"Continued"

Source

(fig. 1)

Storm runoff

(storm-composite

samples)

Topsoil at EO

Topsoil at WO

Topsoil

Topsoil

Grain

Grain

Stover1

Date 

collected

12-17-92

3-29-92

7-02-92

7-10-92

7-11-92

7-19-92

11-20-92

11-21-92

11-22-92

12-14-92

1-03-93

1-06-92

1-07-92

11-21-91

11-21-91

10-15-91

10-14-92

10-14-92

815N 

(total 

nitrogen, 

per mil)

..

--

-

9.1

8.5

8.7

7.7

7.0

648

670

815N - NO3 

(dissolved nitrate, 

per mil)

12.2

11.4

1,065

1,117

970

210

300

285

65

362

380

-

_

~

The part of com above the ground, except for grain.

Table 3. Soil-gas analyses

[-- no data]

Depth of 

sample 

collection 

(In feet)

1.5

1.5

3.0

3.0

4.0

4.0

atmospheric gas

Molar nitrogen 

to 

argon ratio

6-24-92

84.0

84.6

84.0

83.6

2-22-93

84.9

84.7

83.7

84.8

84.7

83.6

815N 

relative to 

atmospheric nitrogen 

6-24-92 2-22-93

40.28

-

-.07

-.15

.00

40.13

+.18

+.40

+.13

-

+.09

.00

406
The gravity lysimeters, constructed to intercept 

ground-water recharge, did not yield water between 

the application of spiked fertilizer and November 12, 

1992. The 6 15N - NO3 in the first gravity-lysimeter 

samples collected in November had enrichment 

that ranged from 297 to 1,145 per mil. Values of 

5 I5N - NO3 in gravity lysimeter samples collected 

from December 1, 1992, through January 4,1993, 

ranged from 665 to 1,050 per mil; that is, about 

30 percent of the NO3 in gravity lysimeter samples 

came from the spiked N fertilizer.

Background 5 15N - NO3 in 15 ground-water 

samples collected from 8 wells from June 27, 1991, 

through May 12, 1992, ranged from 9.9 to 15.5 per 

mil, with an average of 11.8 per mil. Background 

concentrations of NO3 ranged from 4.0 to 45.5 mg/L, 

averaged 20.8 mg/L, and accounted for more than 

95 percent by weight of all dissolved N in ground 

water. The spike was applied in May 1992 and no 

enil-eH fertjlirer was detected in ground water by 

August 10-11. The constant decline of water levels 

indicates minimal recharge during this time (fig. 2). 

This minimal recharge is caused by the large quantity 

of evapotranspiration that typically occurs in agricultural crops within this region. In fact, spiked fertilizer 

probably did not reach ground water until substantial 

ground-water recharge from fall rain began after 

November 12 (fig. 2). About 0.24 mg/L of spiked 

fertilizer was detected in the two shallowest wells, WO 

and EO, as soon as 6 days after the rain began. Twentyeight days later, after continued recharge, spiked N 

fertilizer in well EO decreased slightly, but increased to 

0.74 mg/L in well WO. Small quantities of spiked NO3 

were detected in samples from the intermediate depth 

wells El and Ml (table 2). However, the total contribution of N fertilizer to ground-water NO3 cannot be 

determined for several years until the carry-over of 

NO3 in the soil and glacial till from year to year can be 

determined.

SUMMARY AND CONCLUSIONS

A simplified procedure for 15N analysis of NO3 

in natural waters involves concentration of the NO3 on 

anion exchange resins in the field and the modified 

Dumas combustion method. This procedure decreases 

laboratory preparation time and analysis cost. It also 

eliminates the need for toxic preservatives and 

decreases storage and shipment requirements. Precision of the method is currently about 0.05 per mil. 

High concentrations of dissolved CI and DOC can 

affect the accuracy of the method, but not at concentrations detected at the study site.

The N fertilizer applied to corn in the spring 

typically is not leached to ground water until after the 

fall rain, which occurs after the harvest. At this time, 

evapotranspiration is low and soils become saturated. 

About 22 percent of the N fertilizer applied was 

removed from the grain, leaving about 65 Ibs per acre 

available for leaching, washoff, and denitrification. 

Soil-gas analyses and substantial quantities of 

dissolved oxygen in shallow ground water indicate 

denitrification was not a significant cause of loss of N 

in fertilizer.

Although small quantities of NO3 reached 

ground water as soon as fall recharge occurred, the 

total contribution of N fertilizer to ground water cannot be determined until samples have been collected 

over several years to determine the carry-over of NO3 

in the soil and glacial till from one year to the next. 

One month after recharge began, less than 1 mg/L of 

the NO3 in ground water was derived from spiked 

fertilizer. However, as of March 1993, enrichment of 

5 I5N - NO3 is still increasing in most wells.

With substantially decreased costs of 15N 

analysis, use of 15N may become feasible for many 

more field studies. However, the labeled fertilizer used 

in this study cost $6,000 for 0.1 acre and tracer studies 

over large areas are not likely unless substantial 

decreases in the cost of fertilizer enriched with N are 

achieved.

407
O

cc

LU 

I 

Q.

LU

5

LU

O 

CC

LU

CC

160

140

120

100

80

60

40

20

0

i i i i i i i i i i i

- D SHALLOW WELLS 

WO AND EO

_ + INTERMEDIATE WELLS 

W1.M1, ANDE1

A DEEP WELLS 

W2, M2, AND E2

-

Mean 8 15 N enrichment 

relative to atmospheric nitrogen 

before spike application 

(11.8 per mil) \

~~4 * ± ^ i

I 1 1 I I I III!

1 1 1 1 1 1 1

Spiked Qfertilizer 

applied

/ '

D "

+"

D 

ffi.

h A $ f

9 H ^ ^* 

i i i i i i i

J J A S O N D 

1991

JFMAMJJASOND 

1992

ESTIMATED 

WATER LEVEL

MEASURED 

WATER LEVEL

J J A S O N D 

1991

JFMAMJJASOND 

1992

Figure 2. 5 I5N enrichment relative to atmospheric nitrogen, ground-water 

level in well E1, and precipitation May 28, 1991, through December 31,1992.

408
REFERENCES

Brcmner, J.M., 1965, Isotope-ratio analysis of nitrogen in 

nitrogen-15 tracer investigations, in Black, C.A., ed.. 

Methods of soil analysis-Part 2, Chemical and Microbiological Properties: Madison, Wisconsin, American 

Society of Agronomy, Madison, Wisconsin, p. 1,256- 

1,286.

Bremncr, J.M., and Edwards, A.P., 1965, Determination and 

isotope-ratio analysis of different forms of nitrogen in 

soils-I. Apparatus and procedure for distillation and 

determination of ammonium: Soil Science Society of 

America Proceedings, v. 29, p. 504-507.

Fiedler, R., and Proksch, 1975, The determination of nitrogen by emission and mass spcctrometry in biochemical analysis A review: Analytica Chimica Acta. 

v. 78, no. l,p. 1-62.

Kelly, B.P., and Blevins, D.W., in press, Hydrology and 

nitrogen distribution in claypan soil and glacial till 

near Centralia, Missouri-May 1991 to May 1992, in 

Proceedings of Soil and Water Conservation Society, 

Minneapolis, Minn.

Kendall, Carol, and Grim, E., 1990, Combustion tube 

method for measurement of nitrogen isotope ratios 

using calcium oxide for total removal of carbon 

dioxide and water: Analytical Chemistry, v. 62, no. 5, 

p. 526-529.

National Atmospheric Deposition Program, 1992, NADP/ 

NTN annual data summary precipitation chemistry in 

the United States, 1991: Fort Collins, Colorado State 

University, Natural Resource Ecology Laboratory, 

475 p.

Mesko, T.O., and Carlson, G.M., 1988, Occurrence of pesticides, nitrate, volatile organic compounds, and trace 

elements in ground water and streams, southeastern 

Missouri, 1986-87: U.S. Geological Survey Open-File 

Report 88-495, 73 p.

Sievers, D.M., and Fulhage, C.D., 1990, Survey of rural 

wells in Missouri - Region I sampling: Naltional Water 

Well Association Cluster of Conferences, Agricultural 

Impacts on Groundwater Quality, Feb. 20- 21, 1990, 

16 p.

Ziegler, A.C., Wallace, W.C., Blevins, D.W., and Maley, 

R.D., 1993a, Occurrence of pesticides, nitrite plus 

nitrate, arsenic, and iron in water from two reaches of 

the Missouri River alluvium, northwestern MissouriJuly 1988 and June-July 1989: U.S. Geological Survey 

Open-File Report 93-101, 30 p.

Ziegler, A.C., Wilkison, D.H., Blevins, D.W., Maley. R.D., 

1993b, Occurrence of herbicides and nitrate in rural 

domestic wells, west-central Missouri, July 1990, 

in Geological Society of America Abstracts with 

Programs, North Central Section, v. 25, no. 3, 

March 1993, p. 92.

409
410
Occurrence, Deposition, and Long Range Transport of

Herbicides in Precipitation in the Midwestern and

Northeastern United States

By Donald A. Goolsby1 , E. Michael Thurman2 , Michael L. Pomes2 , 

Michael Meyer2 , and William A. Battaglin 1

Abstract

Herbicides were detected in precipitation 

throughout the midwestern and northeastern 

United States during late spring and summer of 

1990 and 1991. During May and June atrazine, 

alachlor, or both were detected in 60 to 75 percent 

of weekly accumulations of precipitation 

collected at 81 sampling sites in 23 States. 

Atrazine or alachlor were detected in at least one 

sample from all States, including samples from 

remote wilderness areas such as Isle Royale in 

northern Lake Superior. During the remainder of 

the year, herbicides generally were detected in 

weekly samples at less than about 20 percent of 

the sites. Atrazine was the most frequently 

detected herbicide, followed by alachlor, 

desethylatrazine, and metolachlor. Herbicide 

concentrations and frequency of detection were 

much higher in the Midwest than elsewhere. 

Precipitation-weighted herbicide concentrations 

for mid-April through mid-July typically were 0.2 

to 0.4 |ig/L (micrograms per liter) in the Midwest 

and weighted concentrations as large as 0.9 j^g/L 

were present at a few sites. Concentrations of 1 to 

3 |ig/L were measured in a few individual 

samples representing small amounts of 

precipitation. Deposition rates for both atrazine 

and alachlor ranged from more than 200 

(jag/m2)/yr (micrograms per square meter per 

year) at a few sites in the Midwest to less than

'U.S. Geological Survey, Lakewood, CO 

U.S. Geological Survey. Lawrence, KS

10 (jag/m2 )/yr in the Northeast. These amounts 

represent less than 1 percent of the atrazine and 

alachlor applied annually to crops. The 

geographic pattern of herbicide deposition 

provides evidence for long-range atmospheric 

transport.

INTRODUCTION

Current agricultural practices in the United 

States require extensive use of herbicides for 

economical production of corn, soybeans, sorghum, 

and other row crops. Data compiled by Gianessi and 

Puffer (1991) show that about two-thirds of the 

285 million kg of herbicides used annually in the 

United States (Aspelin and others, 1992) are used in 

crop production in the Midwestern United States, and 

principally in the Mississippi River drainage basin. As 

a consequence, the Midwest is the major source area 

for transport of herbicides into streams and ground 

water in the midcontinent as well as to the atmosphere. 

Most herbicides are relatively soluble but significant 

amounts (a few percent of the amounts applied) can be 

transported into streams (Baker and Richards, 1989; 

Goolsby and others 1991; Thurman and others, 1992) 

and ground water (Hallberg, 1989; Holden and others, 

1992; Kolpin and others, 1994). Previous research has 

also shown that herbicides and other organic 

compounds can be transported into the atmosphere by 

various processes. In the atmosphere, these 

compounds can be dispersed by air currents and 

redeposited on the land surface, lakes, and streams by 

rainfall, snow, and dry deposition, often at 

considerable distances from their origin. For example, 

the transport of chlorinated insecticides and

411
polychlorinated biphenyls (PCB's) into the Great 

Lakes is believed to be largely from the atmospheric 

sources (Eisenreich and others, 1981). Herbicides 

have been reported in fog (Glotfelty and others, 1987) 

and rainfall in the midwestern United States (Capel 

and others, 1991; Nations and Hallberg, 1992; 

Richards and others, 1987; Williams and others 1992), 

eastern United States (Glotfelty and others, 1990; Wu, 

1981), and in Europe (Buser, 1990; Trevisan, 1993). In 

spite of these studies, relatively little is known on a 

regional or multistate scale about deposition patterns 

of herbicides in precipitation and the magnitude of 

herbicide deposition in precipitation in comparison to 

quantities applied to cropland or to surface-runoff 

losses.

To learn more about herbicides in precipitation, 

a study was conducted during 1990 and 1991 to 

determine the occurrence and distribution of 

herbicides in precipitation (rain and snow) for a large 

part of the United States, including much of the 

Mississippi River basin where most of the herbicides 

are used. The study was conducted as part of the Toxic 

Substances Hydrology Program and was mass 

oriented in that it attempted to determine the mass 

deposited in precipitation over nearly one-quarter of 

the United States for the two most heavily used 

herbicides in the country, atrazine and alachlor 

(Gianessi and Puffer, 1991). Specific objectives of the 

study were to (1) determine the occurrence and 

temporal distribution of several major herbicides in 

precipitation, (2) estimate the amounts of atrazine and 

alachlor deposited in precipitation annually in 

individual States and over a large part of the 

conterminous United States, (3) relate annual 

deposition of atrazine and alachlor to amounts applied 

annually, and (4) compare annual herbicide deposition 

in precipitation within the Mississippi River basin to 

the estimated annual amount transported out of the 

basin in streamflow. A summary of results from this 

study is presented in this paper.

DESCRIPTION OF STUDY AREA AND 

DATA COLLECTION METHODS

The study area for this investigation 

encompassed 26 States (fig. 1). It includes States in the 

upper Midwest where the use of herbicides, such as 

atrazine and alachlor, is most intense, and States 

eastward to the Atlantic Ocean and northward to the 

Canadian border that potentially may receive

atmospheric deposition of herbicides applied in the 

Midwest. Precipitation samples were obtained from 

the study area through the cooperation and assistance 

of the National Atmospheric Deposition Program / 

National Trends Network (NADP/NTN), Fort Collins, 

Colo., and the Illinois State Water Survey, Champaign, 

111. The National Atmospheric Deposition Program 

(1992) operates a network of about 200 sites 

nationwide to monitor the status and trends in acid 

deposition. Weekly accumulations of precipitation 

(rain and snow) were obtained at 81 NADP/NTN 

sampling sites located in 23 States within the 26-State 

study area (fig. 1). In addition, five NADP/NTN sites 

in high-elevation areas along the Rocky Mountains 

(fig. 1) and in Alaska (not shown on fig. 1) were 

sampled to provide background data on herbicides in 

precipitation at sites far from the study area. A 

complete description of each sampling site is given in 

a directory of precipitation monitoring sites (National 

Atmospheric Deposition Program, 1986). Collection 

of samples began in March 1990 and continued 

through September 1991 -a period of 19 months.

Each NADP/NTN monitoring site is equipped 

with an Aerochem Metrics Wet/Dry collector to 

collect precipitation and a separate Belfort 5-780 

Universal Rain Gage to make an independent 

measurement of precipitation amount. The wet/dry 

collectors are equipped with two 13-liter polyethylene 

buckets that alternately collect wet or dry deposition. 

The "wet side" bucket is covered with a tight fitting lid 

until precipitation begins, at which time a heated 

sensor opens the lid to allow collection of wet 

deposition. When precipitation ceases, the sensor 

closes the lid until the next precipitation event. 

Precipitation is accumulated in this manner for a 

period of 1 week. Each Tuesday, regardless of whether 

precipitation has occurred the previous week, a site 

operator removes the "wet side" bucket, seals it with a 

polyethylene lid, and sends it to the Central Analytical 

Laboratory (CAL) operated by the Illinois State Water 

Survey in Champaign, III. for analysis. For the 86 

NADP/NTN sites used in this study, a subsample was 

obtained from the polyethylene buckets upon receipt 

at the CAL. The subsample, consisting of 10 to 125 

mL (milliliters), depending on the amount of 

precipitation collected, was transferred to a 125-mL 

baked glass bottle, chilled, and sent to the U.S. 

Geological Survey's laboratory in Lawrence, Kans., 

for herbicide analysis. Buckets from sites where no 

precipitation occurred were leached for 24 hours with 

50 mL of distilled water, which then was used as a 

system blank. Subsamples of these blanks were 

provided by the CAL for herbicide analysis. Only

412
L______

EXPLANATION

 *»NADP/NTN SAMPLING 

SITE AND CAUCODE

Background 

Area /

Study Area

500 MILES

500 KILOMETERS

Figure 1. Location of study area, background area, and National Atmospheric Deposition Program/National Trends Network 

(NADP/NTN) sampling sites.

herbicides deposited in precipitation (rain and snow) 

were analyzed in this study. No samples of dry 

deposition were analyzed.

All precipitation samples were analyzed by 

enzyme-linked immunosorbent assay (ELISA) for 

both alachlor and atrazine. The reporting limits for 

ELISA were 0.1 u,g/L for atrazine and 0.2 u,g/L for 

alachlor. Samples in which herbicides were detected 

and that had a sample volume of at least 60 mL were 

confirmed by gas chromatography/mass spectrometry 

(GC/MS). Analysis of samples with less than 60 mL 

were not confirmed. About 10 percent of all samples 

that were found to contain no herbicides by ELISA 

also were analyzed by GC/MS. These samples were 

chosen randomly from all negative samples. An 

exception to the above procedure occurred during the 

first few months of the study when 20 percent of all

samples with 60 mL or more of water were randomly 

selected for GC/MS analysis regardless of whether 

herbicides were detected by ELISA. Of about 6,000 

samples analyzed by ELISA during the study, more 

than 2,200 were confirmed by GC/MS analysis. 

Included in the 6,000 samples were about 1,200 

samples analyzed by ELISA and 400 samples 

analyzed by GC/MS that were used for qualityassurance (QA) purposes.

Samples for ELISA analysis were analyzed in 

duplicate (Pomes and others, 1991) using microtiter 

plates. Concentrations were determined with a 

microtiter plate reader and calibration curves were 

developed from four standard solutions of known 

atrazine or alachlor concentration that ranged from 0 

to 5 f.ig/L. Samples were analyzed in duplicate and the 

results averaged. Standards were analyzed in triplicate

413
or quadruplicate. GC/MS samples were analyzed by 

an automated procedure (Thurman and others, 1 990) 

following solid-phase extraction of samples on C 18 

cartridges. Samples were analyzed for 1 1 parent 

herbicides: alachlor, ametryn, atrazine, cyanazine, 

metolachlor, metribuzin, prometon, prometryn, 

propazine, simazine, and terbutryn, and two 

metabolites of atrazine desethylatrazine and 

deisopropylatrazine.

Several experiments conducted on the 

NADP/NTN plastic collection buckets over periods of 

3 weeks confirmed that no significant sorption, 

degradation, or other losses of the major herbicides 

studied occurred in the buckets. However, the buckets 

contributed a compound with a GC/MS peak that 

interfered with the identification of 

deisopropylatrazine in some of the samples.

Linear-regression equations, given below, were 

developed from more than 2,200 samples analyzed for 

atrazine and alachlor by both ELISA and GC/MS. 

Because of slight changes in methodology and 

differing batches of ELISA kits, regression results 

were improved by developing separate equations for 

samples collected in 1990 and 1991.

1990: atrazine = -0.01 +0.70 ELISA. R2 =

0.94; standard error of estimate = 0.09 ng/L.

1991: atrazine = 0.00 + 0.80 ELISA. R2 =

0.78; standard, error of estimate = 0.10 

1990: alachlor = 0.00 + 0.64 ELISA. R2 =

0.76; standard, error of estimate = 0.08 

1991: alachlor = 0.00 + 0.84 ELISA. R2 =

0.86; standard, error of estimate = 0.09 

These equations were used to estimate concentrations 

of atrazine and alachlor in about 3,700 samples 

(including quality assurance samples) analyzed only 

by ELISA. The measured or estimated concentrations 

and precipitation amounts were then used to calculate 

weekly deposition of these two herbicides at each 

sampling site which were expressed as micrograms 

per square meter (ng/m2) per week.

RESULTS AND DISCUSSION

Ten of the 13 herbicides and metabolites 

analyzed in this study were detected in precipitation 

samples from the 81 sampling sites in the study area 

(table 1 ). Atrazine and alachlor were detected (at 

concentrations of 0. 1 to 0.2 ng/L or greater) by ELISA 

in 26.2 percent and 9.0 percent of the samples, 

respectively. These two herbicides were detected (at 

concentrations of 0.05 ng/L or greater) in 33.9 and 

21 .5 percent, respectively, of the samples selected for 

analysis by GC/MS. However, it should be stated that 

this is a biased result since the samples for GC/MS

analysis were previously screened by ELISA for the 

presence or absence of herbicides. Consequently these 

results are not typical of the frequency of occurrence 

of these herbicides in the 6,000 samples collected 

during the period of this study. Desethylatrazine was 

the third most frequently detected compound in the 

ELISA-screened samples (19.5 percent), followed by 

metolachlor (14.9 percent), and cyanazine 

(8.1 percent). Five other herbicides and metabolites 

(table 1) were detected in 3 percent or fewer of the 

samples. Only about 1 percent of the samples had 

herbicide concentrations exceeding 1 ng/L. Because 

of the proximity of some sampling sites to cropland, 

spray drift cannot be ruled out as a possible source for 

some of the high concentrations. Atrazine and alachlor 

were detected in concentrations near the analytical 

reporting limit (see table 1) in a small number of 

samples from sites located in the background area 

(figl.).

The spatial pattern of occurrence and temporal 

distribution of atrazine and alachlor in precipitation is 

similar to the pattern observed in streams throughout 

the Midwest (Goolsby and others, 1991; Thurman and 

others, 1992). The frequency at which these two herbicides were detected each week at sampling sites 

throughout the study area (fig. 1) began to increase in 

mid-April following application of herbicides to cropland, and peaked in mid-June or early July (fig. 2). 

During this period, atrazine was detected each week at 

60 to 75 percent of the sites sampled and alachlor was 

detected at 30 to 40 percent of the sites. During the 

remainder of the year, these herbicides generally were 

detected at less than about 20 percent of the sites sampled. Samples were not collected at all sites each week 

because of periods of no precipitation (fig. 2). Atrazine 

and alachlor were detected most frequently at sites in 

the Midwest, but at least one sample (confirmed by 

GC/MS) from all 23 States with sampling sites had 

detectable concentrations of one or both herbicides. 

These sites included sites in areas remote from cropland such as Maine and Isle Royale in northern Lake 

Superior. The processes by which these and other herbicides are transported into the atmosphere include 

volatilization (primarily from moist soils), attachment 

to soil and dust particles that are blown into the atmosphere, and spray drift during application. Of these 

processes, volatilization is believed to be predominant 

(Tierney, 1993). Atrazine is about 10 times less volatile than alachlor and has a longer soil half-life (about 

60 days) than alachlor (about 15 days). Consequently, 

it volatilizes into the atmosphere at a slower rate and is 

available for volatilization from the soil for a longer 

period of time than is alachlor. Because of these 

properties, atrazine is present in precipitation for a 

longer time than is alachlor (fig. 2).

414
Table 1. Statistical summary of herbicide concentrations measured in precipitation samples from 81 National Atmospheric 

Deposition Program/National Trends Network sites during March 1990 through September 1991

[Mg/L, micrograms per liter; <, less than; not detected: ametryn, prometryn. terhutryn]

Reporting 

Herbicide limit

ELISA 1 Analysis (all samples)

atrazine 0. 1

alachlor 0.2

Number 

of 

samples

4,762

4,764

Percent 

detections

26.2

9.0

Concentration, in Mg/L, for indicated percentiles

50

(median)

<0.10

<.2

75

0.10

<.2

90

0.26

<.2

95

0.47

.34

99

1.3

1.3

100

(maximum)

16

16

GC/MS2 Analysis (samples pre-scrcened by ELISA)

atrazine 0.05

aiachlor .05

dcsethylatrazine .05

metolachlor .05

cyanazine .05

deisopropylatrazine .05

simazine .05

metribuzin .05

prometon .05

propazine .05

1,848

1.849

1.849

1,849

1.849

1,849

1,849

1.849

1.849

1.849

33.9

21.5

19.5

14.9

8.1

3.0

1.6

.7

.5

.1

<.()5

<.05

<.05

<.05

<.05

<.05

<.05

<.()5

<.05

<.05

.09

<.05

<.05

<.()5

<.()5

<.05

<.05

<.05

<.05

<.()5

.25

.13

.10

.09

<.05

<.05

<.05

<.05

<.()5

<.05

.45

.29

.16

.18

.08

<.05

<.05

<.05

<.()5

<.()5

I.I

.97

.40

.68

.29

.18

.07

<.()5

<.()5

<.()5

10.9

3.2

.75

3.0

2.0

1.2

1.5

.18

.20

.12

'ELISA, enzyme-linked inuminosorbent assay 

2GC/MS. gas chromatography/mass spectrometry.

100

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40

20

1I|III!II1100

sites 

sampled

alachlor 

detections

80g

a

60

40

20

JFMAMJJASONDJFMAMJJASOND 

1990 1991

Figure 2. Frequency of atrazine and alachlor detections, by week, for 81 sites in study area (midwestern and northeastern 

United States). Includes detections by immunoassay and gas chromatography/mass spectrometry. Immunoassay reporting 

limits are 0.1 and 0.2 micrograms per liter respectively for atrazine and alachlor and 0.05 micrograms per liter for gas 

chromatography-mass spectrometry.

Atrazine and alachlor concentrations were 

highest and detections were most frequent at sites in 

the Midwest (fig. 3) where these herbicides are used 

most extensively. Concentrations as high as 3 u,g/L 

were detected in individual samples at some sites. In 

areas where the use of atrazine and alachlor are similar 

such as Iowa, Illinois, and Indiana, alachlor

concentrations generally were higher than atrazine 

concentrations (fig. 3), in part, because of the greater 

volatility of alachlor relative to that of atrazine. In 

remote areas such as Maine and parts of Michigan, 

herbicide concentrations were very low and detections 

were infrequent (fig. 3). The highest herbicide 

concentrations tended to occur during small

415
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PRECIPITATION, IN MILLIMETERS PER WEEK
precipitation events after herbicide application, 

because the first part of a precipitation event flushes 

most of the herbicides from the atmosphere (Capel, 

1991). Subsequent precipitation dilutes the 

concentrations from the earlier precipitation. 

Therefore, it is difficult to make meaningful 

comparisons of herbicide concentrations among sites 

or over time on the basis of individual weekly 

samples. A better comparison can be made with 

precipitation-weighted concentrations, which are 

calculated by summing the product of precipitation 

amounts and herbicide concentrations over a period 

of time, such as 3 months, and dividing by the total 

amount of precipitation. Precipitation-weighted 

concentrations of atrazine and alachlor calculated for 

a 13-week period from mid-April through mid-July 

1991 are shown in figure 4; the map was constructed 

using a geographic information system (CIS). 

Precipitation-weighted concentrations of 0.2 to 

0.4 jag/L are typical throughout the Midwest and 

weighted concentrations of 0.4 to 0.9 |ig/L were 

detected at sites in Iowa, Illinois, and Indiana. The 

spatial pattern of weighted concentrations reflects the 

intensity of use of these two herbicides. It also 

provides tentative evidence for long-range 

atmospheric transport of herbicides.

The estimated masses of atrazine and alachlor 

deposited in precipitation during January through 

September 1991 are shown in figure 5. The 

deposition rates range from more than 200 |ig/m2 for 

both herbicides at a few sites in the Midwest to less

^

than 10 ng/m in the Northeast. Most of the 

deposition occurs during April through July. The 

deposition patterns are somewhat similar for the two 

herbicides, except in the western part of the study 

area (Nebraska and Kansas) where alachlor 

deposition is considerably less than atrazine 

deposition, reflecting the different usage patterns of 

these herbicides. The total mass of herbicides 

deposited in precipitation over the entire study area 

during 1991 is estimated to be 140,000 kg for 

atrazine and 82,000 kg for alachlor. These amounts 

represent about 0.6 and 0.4 percent of the atrazine 

and alachlor applied to cropland in the study area. 

These percentages are smaller than volatilization 

losses reported in the literature (Taylor and Glotfelty, 

1988). There are many possible reasons for the 

relatively small measured amounts including dry 

deposition of herbicides between precipitation 

events, photochemical degradation of herbicides in

the atmosphere, especially atrazine (Gates and 

others, 1991; Pelizzetti and others, 1990), 

atmospheric transport of herbicides beyond the study 

area, and inadequate analytical detection limits, 

especially for the ELISA methods.

The amount of atrazine deposited annually in 

precipitation appears to be somewhat smaller than 

the amount transported to streams in surface runoff, 

whereas the annual deposition of alachlor appears to 

be somewhat larger than the amount transported in 

surface runoff. For example, in a recent study 

(Battaglin and others, 1993) estimated that 

321,000 kg of atrazine and 33,700 kg of alachlor 

were discharged from the Mississippi River basin to 

the Gulf of Mexico in streamflow during a 1-year 

period, while the amounts of these herbicides used in 

the basin were approximately equal. In comparison, 

it is estimated from this study that about 110,000 kg 

of atrazine and 68,000 kg of alachlor are deposited 

annually on the Mississippi River basin in 

precipitation. These quantities could be considerably 

larger if the amounts of dryfall were known. These 

results indicate that more alachlor is deposited 

through precipitation in the Mississippi River basin 

than is transported out of the basin in streamflow, 

whereas the opposite relation could be true for 

atrazine.

SUMMARY

Results from this study indicate that significant 

amounts of atrazine and alachlor are lost through 

volatilization and that these losses can be comparable 

to losses transported in surface runoff. Because of its 

comparatively higher volatility and shorter soil halflife, considerably more alachlor appears to volatilize 

than is transported by surface-runoff losses. 

Additional unmeasured amounts of herbicides were 

likely present in dry deposition. Further research is 

needed to determine the relative amounts of 

herbicides, insecticides, and other organic 

compounds, including photochemical degradation 

products, deposited in wet and dry deposition. 

Several other extensively used herbicides such as 

2,4-D, EPTC, and trifluralin are considerably more 

volatile than atrazine and alachlor and might be 

present in atmospheric deposition in substantial 

amounts.

417
EXPLANATION

PRECIPITATION-WEIGHTED 

CONCENTRATION, IN 

MICROGRAMS PER LITER

LESS THAN 0.05 

0.05 - 0.10 

0.11 - 0.20 

0.21 - 0.40 

MORE THAN 0.40

Background 

Area fStudy Area

i i

500 MILES

I I I

500 KILOMETERS

Figure 4. Precipitation-weighted concentrations of atrazine and alachlor throughout the midwestern and northeastern United 

States, April 16 through July 16, 1991.

418
T

EXPLANATION

DEPOSITION, IN 

MICROGRAMS PER SQUARE 

METER PER YEAR

LESS THAN 10

10 - 25

26 - 50

51 - 100

MORE THAN 100

Background 

Area Study Area

I I I

500 MILES

500 KILOMETERS

Figure 5. Estimated deposition of atrazine and alachlor in precipitation throughout the midwestern and northeastern United 

States, January through September 1991.

419
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Baker, D.B., and Richards, R.P., 1989, Herbicide

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cultivated farmlands of northwestern Ohio, in 

Pesticides in the Terrestrial and Aquatic Environments, 

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11-12, 1989, Blacksburg, Va., p. 103-120.

Battaglin. W.A., Goolsby, D.A., and Coupe, R.H., 1993, 

Occurrence and annual transport of agricultural 

chemicals in the Mississippi River, 1991-92, in 

Goolsby, D.A. and others, eds.. Selected papers on 

agricultural chemicals in water resources of the 

Midcontinental United States: U.S. Geological Survey 

Open-File Report 93-418, p. 26-38.

Buser, Hans-Rudolf, 1990, Atrazine and other s-triazine 

herbicides in lake and in rain in Switzerland: 

Environmental Science and Technology, v. 24, no. 7, 

p. 1049-1058.

Capel, P.D., 1991, Wet deposition of herbicides in

Minnesota, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxics Substances Hydrology 

Program Proceedings of technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 334-337.

Eisenreich, S.J., Looney, B.B. and Thornton, J.D., 1981, 

Airborne organic contaminants in the Great Lakes 

ecosystem: Environmental Science and Technology, 

v. 15, no. 1, p. 30-38.

Gates, P.M.. Foreman, W.T., DcGiacomo, W.G., and 

Anderson, L.G., 1991, Estimates of gas-phase 

photolysis for atrazinc, simazine, alachlor, and 

metolachlor from effective absorption cross-section 

measurements, in Mallard, G.E., and Aronson, D.A., 

eds., U.S. Geological Survey Toxics Substances 

Hydrology Program Proceedings of technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 219-222.

Gianissi, L.P., and Puffer, C.M., 1991, Herbicide use in the 

United States: Resources for the future, 127 p.

Glotfelty, D.E., Williams, G.H., Freeman, H.P., and Leech, 

M.M., 1990, Regional atmospheric transport and 

deposition of pesticides in Maryland, in Kurtz, D.L., 

ed., Long range transport of pesticides: Chelsea, 

Michigan, Lewis Publishers, p. 199-221.

Glotfelty, D.E., Seiber, J.N., and Lilijahl, L.A., 1987, 

Pesticides in fog: Nature, v. 325, p. 602-605.

Goolsby, D.A., Coupe, R.H., and Markovchick, D.J.,1991, 

Distribution of selected herbicides nitrate in the 

Mississippi River and its major tributaries. April 

through June 1991: U.S. Geological Survey WaterResources Investigations Report 91-4163, 35 p.

Hallberg, G.R., 1989, Pesticide pollution of groundwatcr in 

the humid United States: Agriculture, Ecosystems and 

Environment, v. 26, p. 299-367.

Holden, L.R., Graham, J.A., Whitmore, R.W., Alexander, 

W.J., Pratt, R.W., Liddle, S.K., and Piper, L.L., 1992, 

Results of the National Alachlor Well Water Survey: 

Environmental Science and Technology, v. 26, no.5. 

p. 935-943.

Kolpin, D.W., Burkart, M.R., and Thurman, E.M., 1994, 

Herbicides and nitrate in near-surface aquifers in the 

mid-continental United States, 1991: U.S. Geological 

Survey Water-Supply Paper 2413, 34 p.

National Atmospheric Deposition Program, 1986, Directory 

of precipitation monitoring sites: National Resource 

Ecology Laboratory, Colorado State University. Fort 

Collins, Colo., v. 2.

National Atmospheric Deposition Program, 1992, 

NADP/NTN annual data summary, precipitation 

chemistry in the United States, 1991: Natural 

Resource Ecology Laboratory, Colorado State 

University, Fort Collins, Colo.. 475 p.

Nations, B.K., and Hallberg, G.R., 1992, Pesticides in Iowa 

precipitation: Journal of Environmental Quality, v. 21, 

p. 486-92.

Pelizzetti, Ezio, Maurino, Valter, Mincro, Claudio, Carlin, 

Vilma, Pramauro, Edmondo, Zcrbinati. Orfco, and 

Tosato. M.L., 1990, Photocatalytic degradation of 

atrazine and other s-triazine herbicides: 

Environmental Science and Technology, v. 24, 

p. 1559-1565.

Pomes, M.L., Thurman, E.M.. and Goolsby, D.A.. 1991. 

Comparison of microtitre-platc, enzyme-linked 

immunosorbent assay (ELISA) and gas 

chromatography/mass spcctromctry (GC/MS) for 

analysis of herbicides in storm runoff samples, in 

Mallard, G.E.. and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program: Proceedings of technical meeting, Monterey, 

California, March 11-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 

91-4034, p. 572-575.

Richards, P.R., Kramer, J.W., Baker, D.B., and Kreigcr, 

K.A., 1987, Pesticides in rainwater in the northeastern 

United States: Nature, v. 327, p. 129-131.

Taylor, A.W. and Glotfelty, D.E., 1988, Evaporation from 

soils and crops, in Grover, R., ed.: Boca Raton, Fla., 

CRC Press, Environmental chemistry of herbicides, 

v. I, p. 89-129.

420
Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1992, A reconnaissance study of herbicides and 

their metabolites in surface water of the Midwestern 

United States using immunoassay and gas 

chromatography/mass spectrometry: Environmental 

Science and Technology, v. 26, no. 12, p. 2440-2447.

Thurman E.M., Meyer, M., Pomes, M., Perry, C.E., and 

Schwab, A.P., 1990, Enzyme-linked immunosorbcnt 

assay compared with GC/MS for the determination of 

herbicides in water: Analytical Chemistry, v. 62, 

p. 2043-2048.

Tierney, D.P., 1993, Biological assessment of atrazine and 

mctolachlor in rainfall: Ciba-Geigy Corporation, 

Greensboro, N.C., Technical Paper 1-1993, 16 p.

Trevisan, M., Montepiani, C, Ragozza, L.. Bartolctti, C., 

loannilli, E., and Del Re, A.A.M., 1993, Pesticides in 

rainfall and air in Italy: Environmental Pollution, 

v. 80, p. 31-39.

Williams, A.L., Sweet, C.W., and Peters, Cathy, 1992. 

Herbicide concentrations in air and rain: in 

Proceedings of the 85th Annual Meeting and 

Exhibition of the Air and Waste Management 

Association, 1992, p. 2-12.

Wu, T.L., 1981, Atrazine residues in estuarine water and the 

arcal deposition of atrazine into Rhode River, 

Maryland: Water, Air, and Soil Pollution, v. 15, 

p. 173-184.

421
422
Atrazine Transport and Degradation in a Pristine 

Watershed: The Fate of Atrazine Deposited by

Precipitation

ByAron E. Cromwell andE. Michael Thurman1

Abstract

Atrazine used in agricultural areas is being 

transported atmospherically and deposited by precipitation onto pristine watersheds. The fate of 

atrazine was studied at Isle Royale National Park, 

an island park located in Lake Superior on the 

United States-Canadian border. Samples of rainfall, soil water, surface water, and soils were analyzed by combining solid-phase extraction (SPE) 

with enzyme-linked immunosorbent assay 

(ELISA). This SPE-ELISA combination enabled 

the field analysis of water samples in which the 

concentrations of atrazine were as small as 5 ng/L 

(nanograms per liter). The SPE-ELISA results 

were confirmed using gas chromatography/mass 

spectrometry (GC/MS) with isotope dilution. 

Maximum atrazine concentrations in rainfall 

occurred in late spring, approaching the U.S. 

Environmental Protection Agency Maximum 

Contaminant Level (MCL) for drinking water of 

3.0 micrograms per liter. By mid-summer, rainfall concentrations of atrazine had decreased to 

less than 5 ng/L. Atrazine was found in small 

concentrations in water from all lakes that were 

sampled. Field data indicate that atrazine 

degrades rapidly in soil environments but more 

slowly in aquatic environments. This slow degradation rate in water has important implications for 

the quality of lakes in pristine areas receiving 

atrazine-contaminated rainfall because of the 

potential for accumulation of atrazine in the ecosystem.

U.S. Geological Survey, Lawrence, KS.

INTRODUCTION

Since World War II, American farmers have 

greatly improved crop yields by using increased 

amounts of chemically manufactured herbicides. The 

most frequently used herbicide in the Midwest is a 

triazine called atrazine (6-chloro-N-ethyI-N1 - (1-methylethyl) -l,3,5-triazine-2,4-diamine; 2-chloro-4-ethylamino-6-isoproplyamine-s-triazine) (Goolsby, 

Thurman, and Kolpin, 1990). Atrazine generally is 

applied as a preemergent herbicide to control broadleaf weeds in corn (Zea mays L.) and grain sorghum 

[Sorgum bicolor (L.) Moench].

Due to concern for water quality, the amount 

and fate of atrazine transported by agricultural runoff 

to waters of the Corn Belt have been studied to great 

lengths (Humenik and others, 1987; Thurman and others, 1992). Additionally, it has been discovered that 

significant amounts of atrazine can be transported aerially (Glotfelty and others, 1983; Richards and others, 

1987). This finding indicates that atrazine could be 

transported atmospherically from agricultural areas 

and then deposited by precipitation onto pristine areas.

The concentration of atrazine in precipitation 

can be determined accurately and inexpensively by 

using a method of environmental analysis called 

enzyme-linked immunosorbent assay (immunoassay, 

or ELISA) (Bushway and others, 1988; Goolsby 

Thurman, Clark, and Pomes, 1990; Thurman and 

others, 1990). When ELISA is coupled to a concentration process like solid-phase extraction (SPE), the 

detection level is lowered (D. S. Aga and E. M. 

Thurman, U.S. Geological Survey, written commun., 

1993).

This lower detection level facilitates the study 

of the fate of atrazine transported atmospherically into 

a pristine environment. Precipitation, lake-water, and 

soil-water samples were collected and analyzed in a 

pristine watershed in order to determine the occurrence

423
and fate of atrazine deposited by precipitation. The 

fate of atrazine is of great importance because it will 

affect aquatic plant and animal communities if concentrations accumulate to sufficiently high levels 

(deNoyelles and others, 1982; Jones and Winchell, 

1984).

DESCRIPTION OF STUDY AREA AND 

SAMPLE COLLECTION

The study area was Isle Royale National Park, 

Michigan. The 550-km2 (square kilometer) island is 

located in Lake Superior on the United StatesCanadian border. The park is protected from excessive human influence by its Congressional designation 

as a Wilderness Area and its United Nations designation as an International Biosphere Reserve. Isle Royale has never had any history of herbicide application 

and, being an island, is isolated from inflow from 

herbicide-contaminated areas. These facts suggest 

that all of the atrazine found in the hydrologic system 

of Isle Royale would be from precipitation. Also, Isle 

Royale is located approximately 1,000 km (kilometers) northeast of the Corn Belt. The northeast relation 

of the island to the Corn Belt places the island directly 

in the path of spring storms carrying the triazines.

A previous study by D. A. Goolsby (U. S. 

Geological Survey, written commun., 1993) identified 

atrazine in rain samples from Isle Royale. In order to 

confirm this finding and to determine the fate of atrazine in a pristine watershed, two data-collection sites 

were established at Isle Royale. The sites were located 

near two lakes on Isle Royale, Lake Richie and 

Wallace Lake (fig. 1). Rainfall, soil water, and surface 

water were collected from the two sites from mid-May 

through mid-July 1992.

Bulk precipitation was collected using low-density polypropylene funnels connected to 4-L (liter) 

baked amber-glass bottles. After each rain, bulk precipitation was transferred from the collectors to 

125-mL (milliliter) baked amber-glass bottles. The 

collectors then were washed in soap and deionized 

water and rinsed in deionized water before reuse. 

Three of these collectors were placed at each of the 

two field sites. Approximately 100 rainfall samples 

were collected during the 1992 season.

Soil water was collected in suction-cup lysimeters. The polyvinyl-chloride lysimeters had porous 

porcelain cups at their base. Twelve lysimeters were 

installed near Wallace Lake, and six lysimeters were 

located near Lake Richie. At both sites lysimeters

were paired, one 15-centimeter (cm) with one 30-cm 

depth lysimeter. After each rain, a hand pump was 

used to apply a vacuum to the lysimeters. The following day, the water in the lysimeters was pumped into 

125-mL baked amber-glass bottles. Precleaned 

Teflon 1 tubing was used for the transfer.

Surface water was collected and analyzed in 

order to determine the extent of herbicide contamination in the watershed. During the herbicide deposition 

season (from May to July) a total of 20 depth-integrated samples were collected from Wallace Lake and 

Lake Richie. In August of 1992, surface-water samples were collected from lakes across Isle Royale. 

This survey entailed the sampling of 12 major lakes on 

Isle Royale (fig. 1) including Wallace Lake and Lake 

Richie, which were also sampled earlier in the season. 

Samples were collected in 125-mL and 1,000-mL 

baked amber-glass bottles. Results from the August 

1992 lake sampling are located in Table 1.

METHODS

Reagents

Methanol and ethyl acetate were the pesticidegrade solvents used. The SPE cartridges that were 

used contained 360-mg (milligrams) of 40-|uim 

(micrometer) C ]8 -bonded silica. Standard stock solutions were prepared in methanol, and phenanthrenedlO was used as an internal standard for GC/MS 

analysis.

Solid-Phase Extraction

The SPE procedure was performed manually at 

Isle Royale but was automated when performed in the 

laboratory. The manual method used a vacuum manifold to pass liquids through the C| 8 cartridges. The 

automated method used a computerized robotic workstation to process the samples. In both methods, the 

cartridges were washed sequentially with 2 mL methanol, 6 mL ethyl acetate, 2 mL methanol, and 2 mL 

distilled water. One hundred (100) mL of sample were 

passed through the prewashed cartridge for the automated method, while 123 mL of sample were used in 

the manual procedure. The sample passed through the 

cartridge at a flow rate of 10 mL/min (milliliters per 

minute) or slower. The cartridge then was eluted with 

3 mL ethyl acetate followed by a transfer step to

424
89*i<r 89*OCT 88*50- 68-40- 68-30-

48*10'

LAKE SUPERIOR

1M£ «»CH/£ 

CH/PPBVAHAflBOfl 

NTERUEDIATE U&E 

 LAKE WHTTIESEY 

WOOOUXE 

SISKMflTLAKE

0 ISLEROYALE 

NATIONAL PARK

16 MILES 

_J 1 I

16KILOMRTERS

ISLEROYALE 

NATIONAL PARK

Index map

Figure 1. Location of study area.

425
Table 1 . Concentrations of atrazine on surface-water 

samples from Isle Royale National Park, Michigan, 

August 1992

Data collection sites (fig. 1)

Average 

concentration 

(nanograms

per liter)

Chippewa Harbor 6.6

Lake Whittlcsey 3.6

Wood Lake 10.1

Siskiwit Lake (9 sites) 10.6

Lake Desor (4 sites) 13.2

Intermediate lake (2 sites) 10.3

Lake Richie (4 sites) 6.6

Lake LeSage (2 sites) 9.0

Lake Livermore 5.2

Lake Chickenbone 6.5

Wallace Lake 6.9 

Lake Feldtmann surface (6 sites) 11.8 

Lake Feldtmann depth 1.0 feet (6 sites) 11.7

remove the ethyl acetate (top layer) from the residual 

water (bottom layer) in the eluate. The ethyl acetate 

then was evaporated to dryness using a hot-water bath 

and a stream of air when processed at Isle Royale or 

using a stream of nitrogen when processed in the laboratory. The dried sample was reconstituted with 1 mL 

of 20/80 (% v/v) methanol/water for analysis by 

ELISA (D. S. Aga and E. M Thurman, U.S. Geological 

Survey, written commun., 1993).

Gas Chromatography/Mass Spectrometry

GC/MS analysis of the eluates was performed 

on a mass selective detector (MSD). Operating conditions were identical to those described by Thurman 

and others (1990).

RESULTS AND DISCUSSION

Ensyme-Linked Immunosorbent Assay

The ELISA analysis was performed with assay 

kits using polyclonal antibodies coated on paramagnetic beads. The samples were analyzed according to 

the directions supplied with the kits. The optical densities were read using the photometric analyzer, and 

the analyte concentrations were calculated on the basis 

of a linear regression of linear/logit-transformed data. 

The calibration standards for atrazine were prepared at 

concentrations of 5.0, 1.0, 0.10, and 0 |ig/L (micrograms per liter) using 20/80 (% v/v) methanol/water 

solvent, the same solvent used to reconstitute the 

analyte.

Atrazine in Rainfall

Atrazine was detected at significant levels 

(above 50 ng/L) in the rainfall samples collected from 

Isle Royale between mid-May and late-June 1992. 

Atrazine concentrations in rainfall samples collected 

during 1992 are shown in figure 2. Also represented 

in figure 2 is the concentration of atrazine in the rainfall sample multiplied by the amount of rainfall associated with that rain sample (in millimeters). This 

calculated number is related to the mass of atrazine 

associated with a particular rain. The maximum atrazine concentration measured in rainfall occurred on 

June 3 at about 1.9 jag/L, and the maximum atrazine

426
b cc -J < 

cc 2 in o 

a. w

O cc

CC lit

g QH

18

1!

o w

o

1992

Figure 2. Concentration and mass of atrazine in rainfall samples from Isle Royale National Park, Michigan, as determined by 

ELISA and SPE-ELISA. Detection limit 0.005 (ig/L.

mass was measured on June 6. These values were 

much larger than anticipated, so GC/MS analysis was 

performed on these samples for confirmation. The GC/ 

MS results agreed with the ELISA results. Additionally, the GC/MS analysis enabled the testing for additional herbicides and metabolites. These results 

identified the presence of the herbicide cyanazine as 

well as two triazine metabolites, deethylatrazine and 

deisopropylatrazine. On June 3, concentrations of 

cyanaiine in rainfall approached the U. S. Environmental Protection Agency Health Advisory Level of 

1.0 [ig/L for drinking water (U.S. Environmental Protection Agency, 1992). Furthermore, on June 3 and 

June 6, GC/MS analysis revealed the presence of the 

triazine metabolites deethylatrazine and deisopropylatrazine at concentrations between 0.10 and 0.20 (ig/L. 

After June 6, the concentration of atrazine in the rainfall decreased, and by mid-July no appreciable mass of 

atrazine was present.

Atrazine Degradation and Transport in Soil The 

concentration of atrazine in soil water in one lysimeter 

near Wallace Lake is shown in figure 3. Because soil 

types and rates of transport can differ from one lysimeter to the next it was not appropriate to combine the 

data from different ly si meters. Samples from most 

lysimeters followed the general concentration trend of 

the lysimeter depicted in figure 3, which is similar to

the shape of the curve that depicts mass shown in figure 2 for rainfall samples. The small concentrations of 

atrazine in the soil water indicate that degradation of 

the herbicide takes place rapidly in the soil medium. 

Although concentrations of atrazine in soil water are 

considerably smaller than the concentrations in the 

rainfall, the degradation process is not complete.

£V 

LU

^ h- ''

o^-

:ENTRATION NANOGRAMi n o

0~

o

0 

, , ^ , ,

/ \

/ \

/ \

- \

j \ Detection Limit

y \

*^

1 1 1 ! 1 f

^ ^ \^ .n> n?> .^ ^ sq

1992

Figure 3. Concentration of atrazine in soil-water samples 

collected from a lysimeter near Wallace Lake, Isle Royale 

National Park, Michigan. Concentrations determined by 

SPE-ELISA. Detection limit 5 ng/L.

427
Atrazine Concentrations in Surface Water

Surface, inflow, and outflow samples collected 

from Lake Richie and Wallace Lake had concentrations that ranged from 5 to 15 ng/L. In August of 

1992, a survey of major lakes across Isle Royale 

(Fig. 1) revealed atrazine concentrations similar to 

those identified earlier in the season at the two field 

sites. Average concentrations ranged from 3.6 to 

13.2 ng/L (nanograms per liter). Concentrations of 

atrazine in surface water collected during the August 

1992 lake survey are listed in table 1.

CONCLUSIONS

Results from this study indicate that atrazine is 

capable of being transported atmospherically and 

deposited in significant amounts onto pristine areas. 

Once deposited on pristine areas, atrazine is detectable 

in surface water. The probable methods of transport 

include atrazine-laden precipitation falling directly 

onto lakes, surface runoff of atrazine-contaminated 

rainfall flowing into lakes, and possibly even groundwater inflow into lakes. Atrazine persisted in the lakes 

of Isle Royale at least several months after deposition 

by precipitation. Furthermore, the concentrations of 

atrazine in surface water did not decrease from the 

time of deposition (June) through August. This slow 

degradation of atrazine in surface water indicates that 

the mass of atrazine in the aquatic system may 

increase from year to year. Thus, ecosystems not 

affected by atrazine at this time should be evaluated to 

determine the risk to the systems if accumulation 

should occur. Finally, future studies may consider 

the inclusion of analysis for persistent metabolites of 

atrazine, such as deethylatrazine and deisopropylatrazine, in surface and soil water and to determine the 

amount and fate of these important compounds in the 

aquatic system. Also the discovery of cyanazine in the 

rainfall warrants an additional study of the amount of 

this toxic compound and it's metabolites being transported atmospherically as well as its fate in the pristine 

watershed.

REFERENCES CITED

Bushway, R. J., Perkins, Bill, Savage, S.A., and Lekouski, 

S. L., 1988, Determination of atrazine residues in 

water and soil by enzyme immunoassay: Bulletin of 

Environmental Contamination and Toxicology, v. 40, 

p. 647-654.

deNoyelles, Frank, Kettle, W.D., and Sinn, D.F., 1982, The 

responses of plankton communities in experimental 

ponds to atrazine, the most heavily used pesticide in 

the United States: Ecology, v. 63, p. 1285-1293.

Glotfelty, D.E., Taylor, A.W., Zollcr, W.H., 1983, Atmospheric dispersion of vapors Are molecular properties 

important?: Science, v. 219, p. 843-845.

Goolsby, D.A., Thurman, E.M., Clark, M.L., and Pomes, 

M. L., 1990, Immunoassay as a screening tool for triazine herbicides in streams, comparison of gas chromatographic-mass spectrometric methods, in 

Vanderlaan, M., Stanker, L.H., Watkins, B.E., Roberts, 

D. W., eds.; Immunoassays for trace chemical analysis: 

Washington, DC., ACS Symposium Series 451, p 86- 

99.

Goolsby, D.A., Thurman, E.M., and Kolpin, D.W., 1990, 

Geographic and temporal distribution of herbicides in 

surface waters of the upper midwestern United States, 

1989-90: U.S. Geological Survey Water-Resources 

Investigations Report 91 -4034, p. 183-188.

Humenik, F. J., Smolcn, M. D. and Dressing, S. A., 1987, 

Pollution from nonpoint sources: Environmental 

Science and Technology, v. 21, p. 737-742.

Jones, T W., and Winchell, L., 1984, Uptake and photosynthetic inhibition by atrazine and its degradation 

products on four species of submerged vascular plants: 

Journal of Environmental Quality, v. 13, p. 243-247.

Richards, R. P., Kramer, J.W., Baker, D. B., and Krieger, 

K.A., 1987, Pesticides in rainwater in the northeastern 

United States: Nature, v. 237, p. 129-131.

Thurman, E. M., Goolsby, D. A., Meyer, M. T, Mills, M.S., 

Pomes, M.L., and Kolpin, D. W., 1992, A reconnaissance study of herbicides and their metabolites in surface water of the midwestern United States using 

immunoassay and gas chromatography/mass spectrometry: Environmental Science and Technology, v. 26, 

p. 2440-2447.

Thurman, E. M., Meyer, M.T., Pomes, M.L., Perry, C.A., 

and Schwab, A. P., 1990, Enzyme-linked immunosorbent assay compared with gas chromatography/ 

mass spectrometry for the determination of triazine 

herbicides in water: Analytical Chemistry, v. 62, 

p. 2043-2048.

U.S. Environmental Protection Agency, 1992, Drinking 

water regulations and health advisories: Washington 

D.C., Office of Water, 13 p.

428
Annual Use and Transport of Agricultural Chemicals in

the Mississippi River, 1991-92

By William A. Battaglin 1 , Donald A. Goolsby 1 , and Richard H. Coupe2

Abstract

The presence of agricultural chemicals 

(herbicides, insecticides, and nutrients) in the 

Mississippi River and several tributaries follows 

an annual cycle. Herbicide concentrations are 

generally highest during periods of storm runoff 

following application in the spring and early summer. Nitrate concentrations are generally highest 

in the winter and spring and lower during the 

summer and fall.

Concentrations and mass transport of 

agricultural chemicals in rivers are generally 

related to the amounts of the chemicals used 

within the river drainage basins. A geographic 

information system (GIS) was used to analyze 

relations among county-level estimates of the 

mass of agricultural chemicals applied (kilograms 

of active ingredient applied per county) and the 

annual transport of dissolved chemicals in rivers. 

Relations were studied for the Mississippi River 

and several tributaries in the midwestern United 

States. County-level estimates of herbicide and 

nitrogen use were developed into a series of GIS 

data layers and used to estimate the mass of 

agricultural chemicals used annually within specific drainage basins. Analytical data from 

periodic water-quality sampling and daily streamflow data were used to estimate the mass of 

agricultural chemicals transported out of specific 

drainage basins by rivers. These data sets were 

used to develop statistical models for estimating 

annual transport of agricultural chemical as a

'U.S. Geological Survey, WRD, MS-406, Denver, Colo. 

80225

2 U.S. Geological Survey, Urbana, Illinois

percentage of estimated annual agricultural 

chemical use in the basins of the Mississippi 

River and several tributaries. Results indicated 

that, in 1991, estimated masses equivalent to 

about 15 percent of the commercial nitrogen 

fertilizer, 1.6 percent of the atrazine and cyanazine, 0.8 percent of metolachlor, and 0.2 percent 

of the alachlor applied in the drainage basins 

studied were transported out of the basins by 

rivers.

INTRODUCTION

The U.S. Geological Survey presently (1993) 

is conducting studies to determine the distribution, 

transport, and persistence of selected agricultural 

chemicals (herbicides, insecticides, and inorganic 

nutrients) in the Mississippi River and several tributaries. Previous studies indicated that transport of 

agricultural chemicals in runoff to streams is seasonal; 

the largest contributions occur during periods of late 

spring and early summer runoff (Baker and Richards, 

1990; Goolsby and others, 199la; Thurman and 

others, 1991; Goolsby and others, 1991b; Thurman 

and others, 1992; Goolsby and Battaglin, in press). 

These studies also indicate that agricultural chemicals 

are present in surface waters in concentrations that 

exceed Federal health-based levels for drinking water 

at certain times of the year (U.S. Environmental 

Protection Agency, 1992). However, these high 

concentrations generally do not persist throughout the 

year, and the health-based levels are based on average 

annual concentrations, not concentrations of shortduration (Goolsby and Battaglin, in press).

The Mississippi River Basin contains some 

of the most productive cropland in the world (Spalding 

and Exner, 1991). Agricultural chemicals are used 

extensively in this region to increase yields of agricul429
tural crops. In the Mississippi River Basin, an 

estimated 20,600 metric tons of atrazine, 18,500 

metric tons of alachlor, 15,800 metric tons of 

metolachlor, 8,170 metric tons of cyanazine, and 

6,263,000 metric tons of nitrogen fertilizer are used 

annually (Gianessi and Puffer, 1991; U.S. Environmental Protection Agency, 1990). Agricultural 

chemical use in the Mississippi River Basin accounts 

for more than one-half of the total use in the United 

States (Gianessi and Puffer, 1991; U.S. Environmental 

Protection Agency, 1990). Most of this agricultural 

chemical use is associated with the production of corn, 

soybeans, sorghum, and wheat.

Many agricultural chemicals are partially 

soluble in water and have the potential to leach to 

ground water or to run off to surface water. The 

presence of agricultural chemicals in surface waters of 

the midwestern United States is of concern because 

approximately 18 million people in the midwestern 

United States depend on the Ohio, Missouri, and 

Mississippi rivers and their tributaries for drinkingwater supply (Howard Perlman, U. S. Geological 

Survey, written commun., 1991). Conventional watertreatment practices do not remove most of these 

agricultural chemicals, and costly alternative treatment 

methods, such as carbon filtration or denitrification, 

might be required to treat waters that do not meet 

health-based limits (U.S. Environmental Protection 

Agency, 1989; Adams and others, 1990; Madison 

and Brunett, 1985). An improved understanding of 

the spatial and temporal distribution agricultural 

chemicals in midwestern rivers is needed so that 

water-managers and suppliers can address the problems resulting from agricultural chemical contamination of water supplies.

The purposes of this paper are to (1) describe 

the methods used to estimate agricultural-chemical 

transport in the Mississippi River and several of its 

tributaries, and agricultural chemical use within the 

drainage basins associated with specific sampling 

points; and (2) present the results of simple statistical 

models that were used to estimate the annual dissolved 

transport of agricultural chemicals as a percentage of 

the estimated annual use of those chemicals. The 

results presented in this paper are for a particular year; 

the effects of climatic conditions and other factors that 

might change from year to year have on results are not 

addressed in this paper.

Eight large basins with drainage areas of 29,400 

to 2,914,000 km (fig. 1) were studied. The Mississippi 

River was sampled at three locations, and the Missouri,

Ohio, Platte, Illinois, and White Rivers were each sampled at one location (fig. 1). All these river basins are 

predominantly within the corn and soybean producing 

region of the mid-continental United States.

METHODS OF INVESTIGATION

Streamflow from the eight basins was sampled 

periodically from April 1991 through March 1992. 

Except for the Ohio River, all sites were sampled 

every 3 to 4 days during May through mid-July, once 

per week during the remainder of the summer and fall, 

and once every 2 weeks during winter. The Ohio River 

was sampled once per week, except during winter 

when samples were collected every 2 weeks (Goolsby 

and others, 1991b). Samples were collected by depthwidth integrating methods (Edwards and Glysson, 

1988) and composited in stainless steel or glass 

containers. Samples for pesticide analysis were filtered 

through 0.7-|im pore-diameter glass-fiber filters into 

pre-cleaned glass bottles. Samples for nitrate analysis 

were filtered through 0.45-|im-membrane filters.

All samples from the eight basins were analyzed 

for herbicides by gas chromatography/mass spectrometry (GC/MS) following solid-phase extraction on 

Cj 8 cartridges (Sandstrom and others, 1991). Nitrite 

and nitrite plus nitrate concentrations for all sites were 

determined by automated colorimetric procedures 

(Fishman and Friedman, 1989). Nitrate concentration 

was calculated as the difference in concentration 

resulting from these two analytical determinations.

Daily dissolved transport of atrazine, alachlor, 

cyanazine, metolachlor, and nitrate as nitrogen was 

calculated for the eight basins as the product of 

measured or estimated daily concentrations and daily 

mean Streamflow. Concentrations were estimated by 

linear interpolation on days when no samples were 

collected. The estimated daily transport was accumulated for the period April 1, 1991 to March 31, 1992, 

to obtain estimates of annual transport.

County-level herbicide use estimates were 

constructed by Gianessi and Puffer (1991) by ~

(1) compiling statistics by State and by crop, on

percentage of acres treated with a given herbicide 

and average annual application rate of the 

herbicide from surveys sent to Extension Service 

weed scientists in 1989 and 1987;

(2) augmenting survey data with published surveys 

and reports from specific States (U.S. Department 

of Agriculture, 1989);

430
\

EXPLANATION \

A GAGING AND SAMPLING STATION 

 DRAINAGE BASIN BOUNDARY

0 100 200 300 MILES

I i l 1 '

0 100 200 300 KILOMETERS

Figure 1. Gaging and sampling station locations and associated drainage basins for the Mississippi River and several 

tributaries.

(3) establishing herbicide-use profiles, by State and 

by crop, containing the percentage of acres treated 

and average annual applications rates;

(4) multiplying county-level crop acreage estimates 

from the 1987 Census of Agriculture (Bureau of 

the Census, 1989) by percentage of acres treated 

and average annual application rates to get 

herbicide use estimates; and

(5) tabulating use of active ingredient of herbicides

used by crop and by county.

The total use of individual herbicides on all crops for 

each county was calculated prior to entering the herbicide-use estimates into the GIS.

County-level nitrogen fertilizer sales for the 

1991 fertilizer year (July 1, 1990 through June 30, 

1991), were estimated (Gerald Fletcher, West Virginia 

University, written commun., 1992; U.S. Environmental Protection Agency, 1990). The fertilizer-sales 

estimates are reported as tons of actual nutrient (inorganic nitrogen, phosphate, and potash). The fertilizer 

sales estimates do not account for the use of manure. 

County-level nitrogen-fertilizer sales estimates were 

constructed by ~

(1) compiling annual State fertilizer-sales data

reported as tonnages to the National Fertilizer and

431
Environmental Research Center of the Tennessee 

Valley Authority;

(2) calculating the ratio of expenditures on

commercial fertilizers by county to expenditures 

on commercial fertilizers by State from the 1987 

Census of Agriculture (Bureau of the Census, 

1989); and

(3) computing annual county-level fertilizer sales, in 

tons, by multiplying estimates of annual State 

sales by the ratio of county expenditures to State 

expenditures.

The county-level estimates of nitrogenfertilizer sales in tons of actual nutrient were entered 

directly into the CIS.

An area-weighted transfer algorithm was 

programed in the GIS and used to generate estimates 

of the masses of agricultural chemicals used within the 

drainage basins studied. The transfer algorithm 

accounts for cases where both the entire county is 

within a drainage basin, and where only a portion of a 

county is within a drainage basin. In the latter case, the 

value of the attribute being transferred is weighted by 

the ratio of the area of the county within the basin to 

the total county area. For example, if 25 percent of 

county X falls within basin Y, and atrazine use is 

estimated to have been 4,000 kg in county X, then 

1,000 kg of atrazine from county X is transferred to 

basin Y.

USE AND TRANSPORT OF 

AGRICULTURAL CHEMICALS

Graphs of concentration for atrazine, alachlor, 

cyanazine, metolachlor, and nitrate in water samples 

collected at the sampling site on the Mississippi River 

at Thebes, 111., (fig. 2) show the temporal patterns in 

agricultural-chemical concentrations that were 

observed at all eight sampling sites. The spring-flush 

phenomenon previously reported by Thurman and 

others (1991) is readily apparent. In the Mississippi 

River, concentrations change more slowly, peak 

concentrations are lower, and elevated concentrations 

are sustained for a longer period of time than they are 

in smaller streams. These differences can be attributed 

to the fact that these large rivers integrate the 

agricultural-chemical input from many smaller 

streams. The year-round occurrence of herbicides, 

such as atrazine, in the Mississippi River indicates 

that some agricultural chemicals are stored and subsequently released from surface-water and ground-water

reservoirs (Goolsby and Battaglin, in press; Squillace 

and Thurman, 1992)

The estimates of annual agricultural chemical 

transport (table 1) represent only transport of dissolved chemicals. Earlier research by Pereira and 

Rostad (1990) indicated that 99.5 percent of the 

atrazine, cyanazine and metolachlor was in the dissolved phase and less than 0.5 percent was in the 

suspended phase in the lower Mississippi River. 

Squillace and Thurman (1992) estimated that 1 

percent or less of the total atrazine was present in the 

suspended phase in samples from the Cedar River in 

Iowa. The estimates of annual transport of agricultural 

chemicals in the Mississippi River are similar to those 

reported by Goolsby and others (1991) but significantly larger than those reported by Pereira and others 

(1989).

Graphics showing the relative intensity of 

agricultural-chemical use were created to help 

visualize the spatial distribution of agricultural 

chemical use across the United States. Examples of the 

herbicide and nitrogen-fertilizer use data are shown in 

figure 3. For these examples, county-level estimates of 

atrazine use, in kilograms, or nitrogen-fertilizer use, in 

metric tons, have been divided by county land area. A 

statistical distribution was used to select six class 

intervals for the graphic displays of agricultural chemical use. The empty (white) class represents counties 

where data on agricultural-chemical use are missing or 

zero. The first shaded class (lightest grey) represents 

use data that are greater than zero and as large as the 

25th percentile of the data, the second grey-shaded 

class represents use data that are greater than the 25th 

percentile and as large as the median of the data, the 

third grey-shaded class represents use data that are 

greater than the median and as large as the 75th 

percentile of the data, the fourth grey-shaded class 

represents use data that are greater than the 75th 

percentile and as large as the 75th percentile plus two 

standard deviations of the data, and the fifth greyshaded class represents use data that are greater than 

the 75th percentile plus two standard deviations.

Linear regressions were used to investigate 

relations between estimates of agricultural chemical 

transport (table 1) and use (table 2), and to estimate 

transport of agricultural chemicals as a percentage of 

estimated annual agricultural-chemical use. For the 

Mississippi River and several tributaries, significant 

relations (p < 0.05) were determined when estimates 

of annual transport were regressed with estimates of

432
Atrazine 

Alachlor 

cyanazine 

Metolachlor

*3

r i \ r

A 

1991

o N F 

1992

500,000

400,000

O 

55 O

O O 300,000 

? uj

200,000

100,000

J J A s

1991

(c)

J F M 

1992

Figure 2. Temporal distribution of (a) herbicides: Atrazine, alachlor, cyanazine, and metolachlor concentrations (b) nitrogen 

concentration, and (c) discharge in the Mississippi River at Thebes, Illinois.

433
Table 1. Estimated transport of atrazine, alachlor, cyanazine, metolachlor, and nitrate in metric 

tons, by the Mississippi River and several tributaries, April 1991 through March 

1992

Sampling 

station name

Mississippi at Baton Rouge, La.

Mississippi at Thebes, 111.

Missouri River at Hermann, Mo.

Ohio River at Grand Chain, 111.

Mississippi River at Clinton, Iowa

Platte River at Louisville, Nebr.

Illinois River at Valley City, Dl.

White River at Hazelton, In.

Estimated transport

Atrazine Alachlor Cyanazine Metolachlor

321.

189.

68.4

57.0

15.1

9.28

35.8

5.68

33.7

42.9

7.87

4.97

5.64

2.35

8.79

1.04

127.

113.

31.3

13.4

8.94

5.91

19.8

1.71

123.

87.2

24.7

20.2

8.10

2.47

18.9

2.11

Nitrate

967,000

651,000

79,900

182,000

157,000

5,640

107,000

11,500

annual use for atrazine, alachlor, cyanazine, metolachlor, and nitrate as N (table 3).The regressions with 

alachlor and nitrate do not appear to fit well at the 

smaller values of transport and use (fig. 4). However, 

there are other sources of nitrogen besides chemical 

fertilizers, such as domestic and animal wastes, 

legumes, and natural sources that are not accounted for 

in the use term of the nitrate regression equation.

Values for the Pearsons product-moment correlation coefficient (Davis, 1986), multiple R-square, 

P(F) value, and slope of the regression line for the 

regressions equations are given in table 3. Both 

agricultural-chemical transport and use are highly 

correlated to basin area, which could inflate the goodness of fit statistics in table 3. Also, the Mississippi 

River at Baton Rouge site has a strong influence on all 

of the regressions because this site represents the 

largest drainage area (fig. 4).

Estimates of annual agricultural chemical 

transport as a percentage of estimated annual use for 

the eight drainage basins are shown in table 4. The 

results in tables 3 and 4 indicate that, on an annual 

basis, estimated masses equivalent to about 15 percent 

of the commercial fertilizer, 1.6 percent of the atrazine 

and cyanazine, 0.8 percent of metolachlor, and 0.2 of

the alachlor used in large drainage basins in the midwestern United States are transported out of the basins 

by rivers. However, table 4 indicates that there is 

significant basin-to-basin variability in the estimates 

of transport as a percentage of use. The maximum 

values of all herbicide transport as a percentage of use 

are for the Illinois River. The estimates of chemical 

transport as a percentage of estimated annual use for 

cyanazine are likely to be to smaller than those listed 

in table 4, since cyanazine use estimates were 

constructed using 1989 data, and the manufacturer of 

cyanazine reports a 24.3 percent increase in cyanazine 

sales between 1989 and 1991 (Patti Tillotson, DuPont 

Agricultural Products, written commun., 1993).

Previous investigations have demonstrated a 

relation between herbicide use and time-weighted 

mean herbicide concentrations (Baker and Richards, 

1990) and between herbicide use and transport (Hall 

and others, 1972; Squillace and Thurman, 1992). 

Wauchope (1978) summarized the results of several 

studies investigating the loss (transport from field to 

stream) of herbicides from agricultural areas and 

proposed numbers representing expected herbicide 

losses as a percentage of use that could be useful for 

water-quality planning for large regions. The results

434
ATRAZINE USE,

IN KILOGRAMS PER

SQUARE KILOMETER

MISSING OR 0 

0.01 - 0.41 

0.42 - 2.36 

2.37 - 9.17 

9.18 - 28.67 

MORE THAN 28.67

35°

NITROGEN FERTILIZER 

USE, IN METRIC TONS 

PER SQUARE KILOMETER

MISSING OR 0 

0.01 - 0.35 

0.36 - 1.16 

1.17 - 2.83 

2.84 - 7.09 

MORE THAN 7.09

BOO KILOMETERS

Figure 3. Estimated county-level: (a) atrazine use; (b) nitrogen fertilizer use, July 1,1990, to June 30,1991.

435
Table 2.--Estimated annual application of atrazine, alachlor, cyanazine, metolachlor, and nitrogen 

as commercial fertilizer in metric tons, in the Mississippi and several tributary basins

Sampling 

station name

Upstream drainage area Estimated application, in metric tons

in square kilometers Atrazine Alachlor Cyanazine Metolachlor Nitrogen

Mississippi at Baton Rouge, La. 2,914,000 20,600 18,500 8,170

Mississippi at Thebes, 111. 1,847,000 13,400 12,200 6,210

Missouri River at Hermann, Mo. 1,357,000 6,280 4,660 1,970

Ohio River at Grand Chain, 111. 526,000 5,060 4,900 1,440

Mississippi River at Clinton, Iowa 222,000 1,440 2,040 1,590

15,800 6,263,000

11,000 4,055,000

3,490 1,930,000

3,440 1,133,000

1,630 607,000

Platte River at Louisville, Nebr.

Illinois River at Valley City, Dl.

White River at Hazelton, In.

222,000

69,000

29,000

1,600

1,960

711

1,030

1,880

851

464

715

214

444

1,760

424

425,000

450,000

143,000

Table 3. Goodness of fit statistics for regression relations of agricultural chemical transport, 

April 1991 through March 1992, with annual use

Atrazine Alachlor Cyanazine Metolachlor Nitrate

Pearson product-moment

correlation coefficient

Multiple R-square

P(F)-value

Slope of regression line

0.994

0.989

0.0000

1.56

0.889

0.790

0.0032

0.22

0.983

0.967

0.0000

1.68

0.997

0.994

0.0000

0.79

0.968

0.936

0.0001

15.47

from this study can be compared with results from 

these other studies. Hall and others (1972) estimated 

atrazine transport in runoff water was 2 percent of the 

amount applied at the recommended rate. Squillace 

and Thurman (1992) estimated that 1.5 to 5 percent of 

the atrazine applied was transported from the Cedar 

River basin in 1984. Wauchope (1978) estimated that 

2 to 5 percent of herbicides applied as wettable

powders (includes atrazine and cyanazine) and 1 

percent of other herbicides (including alachlor), were 

lost in runoff. The results of the present study indicate 

slightly smaller loss of herbicides than estimated by 

Wauchope (1978) and Hall and others (1972). The 

difference could be the result of basin-to-basin 

variability, climatic conditions particular to the year of 

study, or changing agricultural management systems

436
400

O 300

I

g 200

100

(a)

Y = -12+ 0.0156'X

10.000 20.000 

ATRAZINE USE. IN METRIC TONS

25.000

50

40

Z 30 

oc

^ 20

£ 

DC O

(b)

Y = 1 + 0.0022*X

+

+

+

+

10.000 

ALACHLOR USE. IN METRIC TONS

20,000

140

W

§ 120

100

80

60

20

(c)

2,000 4,000 6.000 e.ooo

CYANAZINE USE. IN METRIC TONS

NOTE: The estimates of cyanazine use are 

likely to be larger than shown on this 

graph, since use estimates are constructed 

using 1989 data, and the manufacturer of 

cyanazine reports a 24.3 percent increase 

in cyanazine sales between 1989 and 

1991 (Patti Tillotson, DuPont Agricultural 

Products, written commun., 1993)

140

P 120

100

80

60

20

10.000

Y = -2 + 0.0070'X

10.000 

METOLACHLOR USE, IN METRIC TONS

20.000

10 20 30 40 50 60 

NITROGEN USE, IN HUNDRED THOUSAND METRIC TONS

Figure 4. Graphs showing agricultural chemical transport as a function of use in the Mississippi River and several tributaries 

for: (a) atrazine, (b) alachlor, (c) cyanazine, (d) metolachlor and (e) nitrate.

437
Table 4. Estimated transpon, April 1991 through March 1992, as a percentage of use for atrazine, 

alachlor, cyanazine, metolachlor, and nitrate, in the Mississippi River and several 

tributaries

Sampling

station same

Mississippi at Baton Rouge, La.

Mississippi at Thebes, 111.

Missouri River at Hermann, Mo.

Ohio River at Grand Chain, 111.

Estimated transpon as a

Atrazine

1.56

1.41

1.09

1.13

Mississippi River at Clinton, Iowa 1.05

Platte River at Louisville, Nebr.

Illinois River at Valley City, 111.

White River at Hazelton, In.

0.58

1.83

0.80

Alachlor

0.18

0.35

0.17

0.10

0.28

0.23

0.47

0.12

Cyanazine

1.55

1.82

1.59

0.93

0.56

1.27

2.77

0.80

percentage of use

Metolachlor

0.78

0.79

0.71

0.59

0.50

0.57

1.07

0.50

Nitrate

15.44

16.05

4.14

16.06

25.86

1.33

23.78

8.04

Value from regression of 

transport with use, all basins 1.56 0.22 1.68 0.79 15.47

(including changes in application method and rate); 

However, the difference is likely to be a function of 

the area studied and the design of the experiments. The 

experiments summarized by Wauchope (1978) were 

mostly field-scale studies with study plots less than 2 

ha, not basin-scale studies as used in the present study. 

In the study by Hall and others (1972), field runoff was 

maximized by experimental design and plot construction.

Estimates of nitrogen transport as a percentage 

of nitrogen use calculated for this study (table 4) are 

well within the range of values (1.5 to 64 percent) 

observed during 1979-90, but are somewhat lower 

than the reported long-term average (24.7 percent) in a 

small basin in Iowa (Lucey and Goolsby, 1993). In 

their review of nitrate contamination in the United 

States, Spalding and Exner (1991) indicated that 

nitrate in surface water is more of a problem in the 

Northeast, where runoff is high, and in the Midwest, 

where tile drainage contributes to surface-water flow, 

and less of a problem in the High Plains States. Results 

from this study also indicate that annual nitrogen 

transport, as a percentage of nitrogen use, is greater in

the Ohio and Illinois river basins than for the Platte 

and Missouri river basins (table 4).

Possible causes of variations in nitrate transport 

from east to west include differences in agricultural 

practices, population density, climatic conditions, and 

surficial materials. These possible causes need to be 

tested in future studies. The strength of the relations 

between agricultural chemical use and agricultural 

chemical transport (table 3) indicates that rates of 

chemical use could be as important in determining 

regional vulnerability to surface-water contamination 

as other hydrogeological factors, such as aquifer type, 

soil type, and agricultural management systems.

SUMMARY

Estimates of agricultural chemical use compiled 

using a GIS and estimates of agricultural chemical 

transport calculated from periodic water-quality sampling and daily mean stream flow data were compared 

using linear regression models for the Mississippi 

River and several of its tributaries in the midwestern 

United States. Results indicated that, during April

438
1991 thorough March 1992, estimated masses equivalent to about 15 percent of the commercial fertilizer, 

1.6 percent of the atrazine and cyanazine, 0.8 percent 

of metolachlor, and 0.2 of the alachlor used in the 

basins were transported in the dissolved phase out of 

the basins by rivers.

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chromatography/mass spectromctry with selected ion 

monitoring: U.S. Geological Survey Open-File Report 

91-519, 26 p.

Spalding, R.F.. and Exner, M.E., 1991, Nitrate Contamination in the Contiguous United States, in Bogardi, I. and 

Kuzelka. R.D., eds.. Nitrate Contamination, NATO 

ASI scries, Vol. G 30: Berlin, Springer-Vcrlag, 

p. 13-48.

Squillacc, Paul J., and Thurman, E.M., 1992, Herbicide 

transport in rivers: Importance of hydrology and 

geochemistry in nonpoint-source contamination: 

Environmental Science and Technology, v. 26, no. 3, 

p. 538-545.

Thurman, E.M., Meyer, M., Pomes, M., Perry, C.E.. and 

Schwab, A.P., 1990, Enzyme-linked immunosorbcnt 

assay compared with GC/MS for the determination of 

herbicides in water: Analytical Chemistry, v. 62, p. 

2043-2048.

439
Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1991, Herbicides in surface water of the midwestern United States The effects of the spring flush: 

Journal of Environmental Science and Technology, 

v. 25, no. 10, p. 1794-96.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., Mills, 

M.S., Pomes, M.L., and Kolpin, D.W., 1992, A 

reconnaissance study of herbicides and their 

metabolites in surface water of the midwestern 

United States using immunoassay and gas 

chromatography/mass spectrometry: Journal of Environmental Science and Technology, v. 26, no. 12, 

p. 2440-2447.

U.S. Department of Agriculture, 1989, Agricultural

Resources, Situation and Outlook Report: Washington, 

D.C. Economic Research Service, February.

U.S. Environmental Protection Agency, 1989, Drinking 

Water Health Advisory Pesticides Office of Drinking Water Health Advisories: Chelsea, Michigan, 

Lewis Publishers, Inc., 819 p.

U.S. Environmental Protection Agency, 1990, County-level 

fertilizer sales data: Office of Policy, Planning, and 

Evaluation (pm-221).

U.S. Environmental Protection Agency, 1992, Drinking 

Water Regulations and Health Advisories: Office of 

Water, 13 p.

Wauchope, R. D., 1978, The pesticide content of surface 

water draining from agricultural fields a review: 

Journal of Environmental Quality, 1978, v. 7, n. 4, 

p. 459-472.

440
Relation of Nitrate Concentrations in Surface Water to 

Land Use in the Upper-Midwestern United States,

1989-90

By David K. Mueller1 , Barbara C. Ruddy1 , and William A. Battaglin 1 

Abstract INTRODUCTION

As part of a study on contamination from 

agricultural chemicals, nitrate data were collected 

during several synoptic surveys at a large number 

of surface-water sites in 10 midwestern states during 1989-90. These data were analyzed using 

logistic regression to relate discrete categories of 

nitrate concentrations to land use in the drainage 

basins upstream from the sampling sites. The 

nitrate data were divided into three categories 

representing background concentrations, elevated 

concentrations, and concentrations that exceeded 

the U.S. Environmental Protection Agency maximum contaminant level for drinking water. Landuse data were derived from spatial-digital data 

available from several sources in national data 

bases. The explanatory variables selected for the 

best-fit model were percent! le of streamflow at the 

time of sampling, acreage of the basin in corn, 

acreage in soybeans, density of cattle, and population density. All these variables have qualitative 

relations to nitrate sources, mobilization, or transport. Classification of nitrate categories from this 

model was 80 percent accurate in comparison to 

observed categories. The accuracy of the model 

was better for classification into categories that 

represented lower concentrations; however, incorrect classifications were not biased either high 

or low. Results from this study indicate that landuse data can be useful in analyses of water-quality 

conditions in large regions and that logistic 

regression is a valuable technique for use in such 

analyses.

'U.S. Geological Survey, Lakewood, CO.

Large quantities of agricultural chemicals, 

including herbicides and nitrogen fertilizers are 

applied each year to farmland in the upper-midwestern 

United States. These applications, in conjunction with 

the moderately large solubility and mobility of some 

herbicides and nitrate, create the possibility for 

substantial contamination of surface and ground water. 

In 1989, the U.S. Geological Survey (USGS) began a 

regional study of herbicides and nitrate in surface 

waters of a 10-State area (fig. 1). Synoptic sampling 

surveys were conducted in the spring and summer of 

1989 and 1990 and in the fall of 1989. Analysis of 

regional synoptic data is an important aspect of several 

programs within the USGS, including the National 

Water Quality Assessment Program (NAWQA) and 

the Toxic Substances Hydrology Program. The results 

of herbicide analyses from these samples have been 

presented previously (Goolsby and others, 1991, 

Thurman and others, 1991).

This paper presents results of the nitrate-data 

analysis specifically, the concept that nitrate concentrations in the synoptic samples are related to land-use 

data derived from large-scale geographic data bases. 

Also, statistical models that can be used to estimate 

nitrate concentrations from land-use data are 

evaluated.

SOURCE OF DATA

Surface-water samples were collected at 

141 sites in Illinois, Indiana, Iowa, Kansas, Minnesota, 

Missouri, Ohio, Nebraska, South Dakota, and Wisconsin (fig. 1). Sampling sites were selected to be proportional to corn production in each State (D.A. 

Goolsby, U.S. Geological Survey, written commun., 

1989, 1990). Within States, site selection was based

441
100°

45 °-| South 

Dakota

EXPLANATION 

~""~ STATE LINE

Ohio

s/

" i *

/ 0 100 200 300 KILOMETERS

BASIN BOUNDARY

STREAMFLOW-GAGING 

STATION

Figure 1. Location of selected stream-sampling sites and upstream drainage basins 

in the upper midwestern United States.

on geographic distribution and drainage-basin area. 

Most sites were located at existing USGS streamfiowgaging stations.

Samples were collected three times in 1989: 

(1) spring (March-May), before the application of herbicides, but not necessarily prior to application of 

nitrogen fertilizers; (2) summer (May-July), during 

the first major runoff event after application of herbicides; and (3) fall (October-November), after the first 

killing frost. Samples were analyzed for selected herbicides and for dissolved nitrite plus nitrate (hereinafter referred to as nitrate). Field measurements of 

temperature, pH, and specific conductance were made 

and streamflow was either measured or derived from a 

rating curve.

Fifty sites were resampled during the spring and 

summer of 1990. Selection of these sites was based on 

ranking the samples collected during the summer of 

1989 according to the total herbicide concentration. 

The sampling sites were divided into three equal 

groups. Twenty-five sampling sites were randomly 

selected from the group that contained the highest concentrations; 13 sampling sites were randomly selected 

from the middle group; and 12 sampling sites were

randomly selected from the group that contained the 

lowest concentrations. Two additional sampling sites 

were selected in northeast Iowa from basins that were 

not sampled in the summer of 1989 because of 

drought.

Geographic data for the study area were 

obtained from several sources. Agricultural data, 

including land use, crop types, and livestock, were 

collected for the 1987 Census of Agriculture (U.S. 

Bureau of Census, 1989). Data on nitrogen fertilizer 

sales during 1989 and 1990 were obtained from the 

U.S. Environmental Protection Agency (USEPA) 

(1990). Although these data are for sales, they were 

considered to represent fertilizer application as well. 

Population data were collected for the 1990 Census of 

Population and Housing (U.S. Bureau of Census, 

1990). All data were retrieved from Geographic Information System (GIS) data bases, stored as 

1:2,000,000-scale digital maps of the coterminous 

United States. The agricultural and fertilizer data were 

stored by county. The population data were stored by 

census geographic units (block groups).

Digital data extracted from these sources were 

used to compute values for the drainage basin

442
upstream from each surface-water sampling site. 

A computerized GIS procedure was used to areally 

weight the extracted data and sum it by basin. When 

basins were nested (a large basin can contain several 

smaller basins), the data for the smaller basins were 

summed to derive the information for the large basin. 

The final step in development of the data was to 

convert the basin data to units that were independent 

of basin size. For the geographic data, this conversion 

was done by dividing the data values by basin area. 

Streamflow at the time of sampling was converted to a 

percentile value on the basis of the flow-duration relation for daily streamflow at the site. Percentile of flow 

has been shown to be preferable to actual streamflow 

in making comparisons among basins of different sizes 

and, consequently, different flow regimes (D.R. 

Helsel, U.S. Geological Survey, written commun., 

1993). Daily streamflow records were not available 

for five sampling sites; therefore, flow percentiles 

could not be computed and those basins were omitted 

from subsequent analyses. Also, one basin was

omitted because no samples were collected during the 

spring or summer sampling periods in either 1989 or 

1990. The resultant data set contained values for 135 

basins, having drainage areas that ranged from 89 to 

about 19,000 mi2 (square miles). The variables in this 

data set are listed in table 1.

STATISTICAL METHODS

The initial attempt to relate nitrate concentrations at each site to land use in the upstream drainage 

basin was made by using multiple linear regression 

(MLR). However, no satisfactory MLR models were 

identified. Logistic regression was selected as an 

alternative method.

Logistic regression commonly is used when the 

response variable is discrete or categorical, rather than 

continuous (Helsel and Hirsch, 1992, p. 393). The 

logistic regression model is similar to the MLR model 

in that a set of explanatory variable is used to estimate 

the value of a response variable. The response variable

Table 1. Variables (attributes) included in the final data set

[mg/L, milligrams per liter; acre/mi2, acres per square mile; number/mi2, number per square mile; ton N/mi2, tons nitrogen per square mile]

Variable (attribute) Units Minimum Median Maximum

Dissolved nitrite plus nitrate, as nitrogen

Percentile of flow

Total harvested cropland

Acreage in corn, including silage

Acreage in soybeans

Acreage in grain (wheat, oats, and rye)

Acreage in hay

Pastureland, including pastured woodland

Woodland, not pastured

Cattle

Hogs

Poultry (chickens and turkeys)

Fertilizer application

Population density

mg/L

none

acre/mi2

acre/mi2

acre/mi2

acre/mi2

acre/mi2

acre/mi2

acre/mi2

number/mi2

number/mi2

number/mi2

ton N/mi2

number/mi2

<0.1

.5

57.1

3.06

.51

1.69

4.25

6.04

1.09

5.09

.99

.05

1.38

1.58

2.5

67.5

345

149

126

15.9

25.4

49.4

11.3

55.3

98.3

32.3

14.3

32.8

19

99.5

480

267

239

192

158

397

75.5

176

448

6,280

27.2

1,960

443
in logistic regression is the log of the odds ratio, 

p/( 1 -p), where p is the probability of a data value being 

in one of the possible categories (Helsel and Hirsch, 

1992, p. 395-396). The logistic regression equation is

log

where, bo = the intercept, X= the vector of/: explanatory variables, and b = the vector of slope coefficients 

for each explanatory variable, so that bX= bjXj + 

b2X2 + bKXk .

The slope coefficients are fit to the categorical 

data by the method of maximum likelihood (Helsel 

and Hirsch, 1992, p. 397). This method optimizes the 

likelihood that the observed data will be estimated 

from a given set of slope coefficients.

To create a discrete variable, observed concentrations of nitrate (as nitrogen) in each sample were 

divided into three categories: less than 3 mg/L, 3 to 

10 mg/L, and greater than 10 mg/L. A total of 359 

samples from the 135 sites during four synoptic sampling periods (spring and summer, 1989 and 1990) 

were included in this data set. More than one-half of 

the observations (192) were in the first category, which 

was considered to include background concentrations 

of nitrate in the study area. Observations in the second 

category (138) were considered elevated but were less 

than the USEPA maximum contaminant level (MCL) 

of 10 mg/L for drinking water (U.S. Environmental 

Protection Agency, 1986). Observations in the third 

category (29) exceeded the MCL. The categorized 

nitrate data for samples collected during the spring and 

summer of 1989 are shown in figure 2.

Best-fit logistic regression models were selected 

to estimate nitrate concentrations from land-use data 

by using a stepwise procedure. Explanatory variables 

were added to the model in order of significance, if the 

significance level (p-value) of the slope coefficient 

was less than or equal to 0.15. As variables were 

added, previously entered variables could be removed 

if their p-values increased to greater than 0.20

To decrease the possibility of collinearity (cross 

correlation among the independent variables), the list 

of explanatory variables used in model fitting was 

restricted. Fertilizer application and total cropland 

were strongly correlated with corn and soybean acreage; therefore, fitting procedures were applied to two 

sets of explanatory variable: one set included fertilizer 

and total cropland data, and the other included corn 

and soybean acreage data. All the other explanatory

variables listed in table 1 were included in both sets. 

Slope coefficients were determined separately for 

spring and summer data. Separate coefficients also 

were determined for classification between the low 

(concentrations less than 3 mg/L) and medium (3 to 

10 mg/L) nitrate categories and between the medium 

and high (greater than 10 mg/L) categories. Overall, 

four sets of coefficients were determined using the 

same selected group of explanatory variables. A 

model consisted of the selected variables and the four 

sets of coefficients.

Models were compared on the basis of their 

capability to classify nitrate concentrations correctly. 

First, the model was used to estimate the probabilities 

of nitrate concentrations being in a particular category 

for each sample. Estimated probabilities were computed for the same data set used to fit the model. The 

classified category was selected to be the one with the 

maximum probability. The accuracy of the model was 

determined by comparing the classified category to the 

observed category.

RESULTS AND DISCUSSION

Two best-fit models were selected one from 

each set of explanatory variables. For the variable set 

that included fertilizer and total-cropland data but no 

data on individual crop types, the selected explanatory 

variables were percentile of flow, fertilizer application, 

and population density. The logistic regression fit to 

these explanatory variables is referred to as model 1 in 

this paper. Comparisons of nitrate categories estimated by model 1 to observed nitrate concentrations 

are listed in table 2. The overall accuracy of the model 

is determined by the total percentage of correct classifications. For model 1, 73 percent of the classifications were correct. However, only 34 percent of the 

observations greater than 10 mg/L were correctly classified. Also, model 1 might be biased because most of 

the incorrect classifications were low.

Model 2 was fitted by use of the set of explanatory variables that included crop types, but not fertilizer application or total cropland. The selected 

explanatory variables from this set were percentile of 

flow, acreage in corn, acreage in soybeans, cattle, and 

population density. Accuracy results for model 2 also 

are listed in table 2. The overall accuracy of correct 

classifications was 80 percent, which is a slight 

improvement over that of model 1. Most of this 

improvement was attributed to better accuracy in

444
"l,f #

;

Vm

^ik

»"-\ 

^'

\

1

/

}

1

1

/

^\

1

r

r

t&X'^

Spring 1989

EXPLANATION

CONCENTRATION, 

IN MILLIGRAMS 

PER LITER

-% /" 0 100 200 300 KILOMETERS 

'V

Summer 1989

445
Table 2. Observed nitrate categories and category classifications from the logistic regression models

[mg/L, milligrams per liter, - -, not applicable]

Category 

(mg/L)

Classified correct Classified low Classified high

Observed

Number Percent Number Percent Number Percent

Model 1 (percentile of flow, fertilizer application, and population density)

<3 192 164 85 -- -- 28 15 

3-10 138 87 63 42 30 97 

> 10 29 10 34 19 66 

Total 359 261 73 61 17 37 10

Model 2 (percentile of flow, acreage in corn, acreage in soybeans, cattle, and population density) 

<3 192 166 86 -- -- 26 14 

3 - 10 138 104 75 27 20 75 

>10 29 16 55 13 45 

Total 359 286 80 40 11 33 9

classifying the higher concentrations. Classifications 

from model 2 were correct for 75 percent of the 

observed nitrate concentrations in the medium (3 to 

10 mg/L) category and for 55 percent of the observed 

nitrate concentrations in the high (greater than 10 mg/L) 

category. This accuracy is a substantial improvement 

over model 1. Also, the incorrect classifications were 

about evenly distributed between high and low values, 

so model 2 does not seem to be biased.

The spatial distribution of the classification 

accuracy using model 2 for data from the spring and 

summer of 1989 is shown in figure 3. Incorrect classifications, both low and high, are distributed in a relatively uniform pattern throughout the study area. This 

result indicates that the model classifications are not 

geographically biased. Therefore, the accuracy of the 

model does not seem to depend on correlation of the 

explanatory variables with regional geography.

Generally, model 2, which is based on crop-type 

data, is superior to model 1, which is based on fertilizer data. Model 2 also is reasonably accurate in categorizing nitrate concentrations in the outflow from a 

variety of basins on a regional and seasonal scale. 

Model classifications do not seem to be biased in 

either magnitude or geographic distribution. These

results lead to two primary conclusions. First, the level 

of nitrate contamination in midwestern streams is most 

strongly related to streamflow and to several characteristics of the upstream basin, including the areal 

extent of corn and soybean production, the density of 

cattle, and the population density. This list seems logical because each variable has a qualitative relation to 

nitrate. Streamflow results from basin runoff, which 

provides the mechanism for mobilization and transport 

of nitrate. In the model, as the percentile of flow 

increases, the probability increases that nitrate concentration will be in a higher category. Corn and soybeans are the major crops in the region, and their 

extent in the basin logically should be related to the 

use of all fertilizers, including commercial fertilizers 

and manure. Cattle also could be related to fertilizer 

use because they are a primary source of manure. Perhaps if data were available on actual fertilizer application, rather than only on fertilizer sales, a better model 

based on fertilizer instead of crop-type and livestock 

data could be developed. Population density is related 

to outflow from sewage-treatment plants, which generally are point sources of nitrate or of ammonia (a nitrogen species that can readily oxidize to nitrate). Even

446
Spring 1989

EXPLANATION

ACCURACY OF 

ESTIMATED NITRATE 

CATEGORIES

V°0 100 200 MILES

* I r-H H

t" 0 100 200 300 KILOMETERS

'T

Summer 1989

aa ispring and summer of 1989.

atsus447
in rural areas, population density likely is related to 

nitrate contamination from septic system leachate.

A second conclusion is that an adequate model 

can be developed by relating synoptic water-quality 

data to regionally derived geographic data. All the 

explanatory variables in the model were retrieved 

from large GIS data bases that had, at best, a countyscale resolution. Yet these data were adequate to classify 80 percent of the observed nitrate concentrations 

correctly in outflow from a variety of basins.

The implication of these conclusions is that 

land-use data can be used to analyze water-quality 

conditions within large regions. Reasonable models 

can be developed for relating water quality to land use. 

These models can be adequate, even if the land-use 

data are extracted from large-scale data bases. Refinement of the models might be possible by collecting 

more precise land-use data in the basins upstream 

from sampling sites. More precise explanatory data 

could improve the accuracy of model classifications. 

The models could be used to identify other basins in 

the region where surface-water contamination might 

be a problem. Models from different regions could be 

compared to identify similarities and differences in the 

land-use factors that affect water quality in the 

regions. Logistic regression is a valuable technique 

for the development of these models.

REFERENCES

Goolsby, D.A., Thurman, E.M., and Kolpin, D.W., 1991, 

Geographic and temporal distribution of herbicides in 

surface waters of the upper midwestern United States. 

1989-90, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. Geological Survey Water-Resources Investigations Report 

91-4034, p. 183-188.

Helsel, D.R., and Hirsch, R.M., 1992, Statistical methods in 

water resources: New York, Elsevier, 522 p.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1991, Herbicides in surface water of the midwestern United States The effects of spring flush: 

Environmental Science and Technology, v. 25, no. 10, 

p. 1794-1796.

U.S. Environmental Protection Agency, 1986, Quality criteria for water 1986: EPA 440/5-86-001, Washington, 

D.C. Office of Water Regulation and Standards, 

unnumbered pages.

U.S. Environmental Protection Agency, 1990, County-level 

fertilizer sales data: PM-2221, Washington, D.C., 

Office of Policy, Planning, and Evaluation.

U.S. Department of Commerce, Bureau of the Census,

1989. Census of agriculture, 1987 Final county file: 

Washington, D.C., Data Users Services Division. 

U.S. Department of Commerce, Bureau of the Census,

1990. Census of population and housing, 1990: Washington, D.C., Data Users Services Division.

448
Reconnaissance Data for Selected Herbicides and Two

Atrazine Metabolites in Surface Water of the Midwestern

United States: Chemical Analysis by Immunoassay and

Gas Chromatography/Mass Spectrometry

By Elisabeth A. Scribner1 , E. Michael Thurman1 , and Donald A. Goolsby2

Abstract

Water-quality data were collected from 147 

rivers and streams during 1989-90 for an 

assessment of selected preemergent herbicides 

and two atrazine metabolites in 10 Midwestern 

States. All water samples were collected by 

depth-integrating techniques at three to five 

locations across each stream. Sites were sampled 

three times in 1989: before application of 

herbicides, during the first major runoff after 

application of herbicides, and in the fall during a 

low-flow period when most of the streamflow was 

derived from ground water. About 50 sites were 

selected by a stratified random procedure and 

resampled for both pre-and post-application 

herbicide concentrations in 1990 to verify the 

1989 analytical results. Laboratory analyses 

consisted of enzyme-linked immunosorbent assay 

(ELISA) with confirmation by gas 

chromatography/mass Spectrometry (GC/MS). 

The data have been useful in (1) studying 

herbicide transport, (2) comparing the spatial 

distribution of the post-application concentrations 

of 11 herbicides and 2 atrazine metabolites 

(deethylatrazine and deisopropylatrazine) in 

streams and rivers at a regional scale, (3) 

examining the annual persistence of herbicides 

and their metabolites in surface water, and (4) 

assessing whether the two atrazine metabolites 

can be used as indicators of surface- and groundwater interaction.

'U.S. Geological Survey, Lawrence, Kans. 

2U.S. Geological Survey, Denver, Colo.

INTRODUCTION

During 1987-89, about 136 million pounds per 

year of four major herbicides were applied in a 10- 

State region (table 1). The herbicides alachlor, 

atrazine, cyanazine, and metolachlor accounted for 

about 73 percent of the herbicides applied (Gianessi 

and Puffer, 1990). The intense use of these herbicides 

with their partial water solubility and mobility cause 

them to leach into ground water, to run off in surface 

water, as well as to be transported aerially and to occur 

in precipitation (Goolsby and others, 1990). Runoff 

from fields immediately after herbicide application 

substantially increases herbicide concentrations in 

streams and rivers as reflected by increased 

concentrations of these herbicides in the Mississippi 

River (Goolsby and others, 199la). Potentially serious 

problems concerning nonpoint-source contamination 

of surface- and ground-water supplies are associated 

with this agricultural chemical use. Drinking-water 

quality also is affected because conventional watertreatment practices do not remove these herbicides 

because they are soluble (Goolsby and others, 1990).

This paper presents the water-quality data 

collected during the reconnaissance study of surface 

water in the midwestern United States. This paper (1) 

describes the occurrence, distribution, and concentrations 

of nitrite plus nitrate and selected preemergent herbicides 

and two metabolites, (2) describes the geographic and 

seasonal distribution of nitrite plus nitrate and commonly 

used herbicides in streams of different size throughout 

the 10-State area, and (3) examines the usefulness of a 

low-cost immunoassay analysis for determining 

herbicide concentrations in a regional-scale 

reconnaissance study (Goolsby and others, 1990). The 

study area includes the following states: Illinois, Indiana, 

Iowa, Kansas, Minnesota, Missouri, Nebraska, Ohio, 

South Dakota, and Wisconsin (fig. 1).

449
Table 1 . Quantities of four major herbicides applied in 10 

agricultural Midwestern States, 1987-89

[From Gianessi and Puffer, 1990; values shown are in millions of pounds 

of active ingredient per year]

State Alachlor Atrazine

Illinois 8.0 8.5

Indiana 6.9 5.7

Iowa 6.4 5.6

Kansas 1.9 4.7

Minnesota 4.0 1 .5

Missouri 1.8 3.1

Nebraska 3.8 7. 1

Ohio 3.7 3.8

South 

Dakota 1 .9 0.5

Wisconsin 1 .3 2.7

Total 39.7 43.2

Total of four major herbicides: 

135.6

Cyanazine

3.1

1.7

3.7

0.2

2.8

1.2

1.8

1.6

0.3

1.7

18.1

Metolachlor

8.1

3.5

8.5

2.2

2.4

1.6

1.9

4.0

1.0

1.4

34.6

METHODS

Sampling Design

The 147 sampling sites were selected at U.S. 

Geological Survey (USGS) streamflow-gaging

stations by a stratified random-sampling procedure 

designed to ensure geographic distribution and 

regional-scale interpretation of the data. The 

number of sites per State was proportional to the 

amount of corn and soybean production in each 

state, and sites were chosen randomly by county. 

The drainage area of the hydrologic units in which 

streams were sampled ranged from 66 to more than *y

700,000 mi , with a median drainage area of 770

*j

mi . Most of the streams were sampled three times 

in 1989: before application of herbicides (March or 

April); after application and during the first major 

runoff (May or June); and in the fall during a lowflow period (October or November) when most of 

the streamflow was derived from ground water.

One-third of the streams sampled in 1989 

were resampled during 1990 before herbicide 

application (March or April) and after herbicide 

application (May or June). The streams sampled in 

1989 were ranked from largest to smallest total 

herbicide concentrations; 50 percent of the streams 

sampled in 1990 were randomly selected from the 

upper one-third, 25 percent from the middle onethird, and 25 percent from the lower one-third of the 

streams according to the ranked concentration.

400 ~w

100°

North 

Dakota

X88" 

) Minnesota /Tate^Superior

uWisconsirfS

EXPLANATION

H SAMPLED 1989 

H SAMPLED 1989-90

 BOUNDARY OF

HYDROLOGIC UNIT

 RECONNASSANCE 

SAMPLING SITE

"T "~~ -)

Base from U.S. Geological Survey digital survey. 12.000.000.1972 t 

Abers Equal-Area Conic Projection

Standard parralleb 39°3Cr and 43°301 . Central meridian 90*30

0 100 290 390 MILES 

0 100 200 300 KILOMETERS

Figure 1. Location of study area, hydrologic units, and sites sampled in 1989 and 1990, in the midwestern United States. 

(From Thurman and others, 1992, fig. 1A.)

450
Sample-Collection Techniques

All water samples were collected by depthintegrating techniques at three to five locations across 

each stream (Thurman and others, 1992). The 

herbicide samples were collected, composited in large 

glass containers, and filtered through 1-m pore 

diameter, glass-fiber filters into baked glass bottles for 

shipment to the laboratory.

Analytical Procedures

Measurements were made onsite for specific 

conductance and pH. Samples, preserved with 

mercuric chloride, were analyzed for nitrate plus 

nitrite at the National Water Quality Laboratory, 

Arvada, Colorado (Fishman and Friedman, 1989). 

Herbicides were analyzed by enzyme-linked 

immunosorbent assay (ELISA) at the USGS 

laboratory in Iowa City, Iowa (Goolsby and others, 

1991b). Most samples were also analyzed by gas 

chromatography/mass spectrometry (GC/MS) at the 

USGS laboratory in Lawrence, Kansas. Solvents used 

for analyzing samples included pesticide-grade 

methanol, ethyl acetate, and isooctane. Deionized 

water was charcoal filtered and glass distilled prior to 

use. Internal quantitative standard solutions were 

prepared in methanol and phenanthrene-djQ. Assay 

kits were used for the immunoassay analysis. The kits 

used polyclonal antibodies coated on the walls of 

polystyrene test tubes and an atrazine-enzyme 

conjugate prepared by covalently binding atrazine to 

horseradish peroxidase by a modified carbodiimide 

technique (Bushway and others, 1988). The 

immunoassay procedure is described by Goolsby and 

others (199 Ib).

HERBICIDE OCCURRENCE AND 

DISTRIBUTION

Analyses results indicate that large 

concentrations of herbicides were flushed from 

cropland and transported through the surface-water 

system in response to late spring and early summer 

rainfall. Of the 10 herbicides and metabolites 

analyzed, atrazine, alachlor, metolachlor, and 

cyanazine were present in the largest concentrations. 

During the 1989 post-application sampling period, 

atrazine was detected in water samples from 98 

percent of the stream sites. Alachlor, metolachlor, and

deethylatrazine, an atrazine metabolite, were detected 

in water from more than 80 percent of the stream sites, 

and cyanazine was detected in water samples from 

more than 60 percent of the stream sites. Herbicide 

concentrations during the post-application period 

generally were larger than those measured before 

application and in the fall during low streamflow. The 

median concentration of atrazine during the postapplication period was 3.8 (ig/L compared with about 

0.2 |j,g/L during the pre-application and fall low-flow 

periods. Similar seasonal distributions were observed 

for the other major herbicides (table 2).

During the late spring and summer, 

concentrations of one or more herbicides exceeded 

U.S. Environmental Protection Agency (USEPA) 

Proposed Maximum Contaminant Levels (PMCL) or 

lifetime Health Advisories Levels (HAL) for drinking 

water (U.S. Environmental Protection Agency, 1988; 

Federal Register, 1989, 1990) in more than one-half of 

the streams sampled. On the basis of results from 

storm-runoff samples collected in 1990, 

concentrations of some herbicides exceeded PMCLs 

during runoff for periods of several weeks to several 

months after application.

Analysis results indicate that atrazine was the 

most frequently detected and the most persistent of the 

major herbicides, followed by deethylatrazine and 

metolachlor. Alachlor and cyanazine were detected 

primarily during the late spring and summer. Atrazine 

was detected in three-fourths of the streams sampled 

during the fall, indicating that aquifers contributing 

base flow to the streams are contaminated by 

herbicides. Resampling of about 50 streams in 1990 

before and after herbicide application produced results 

similar to those obtained in 1989. These results 

indicate an annual cycle of herbicide application 

followed by a series of flushing events during which 

herbicides are transported into streams and ground 

water, and by periods of base flow during which 

persistent herbicides in ground water slowly reenter 

the streams.

These are significant findings because they 

indicate that some of the parent herbicides persist 

from year to year in soil and water. Degradation 

products, such as deethylatrazine, are both persistent 

and mobile, and the data suggest that the ratio of 

deethylatrazine to atrazine (called the DAR), which 

has been used as an indicator of nonpoint-source 

contamination of ground water, may be used also 

to trace ground-water discharge into rivers.

451
Table 2. Herbicide compounds analyzed and percentage of detections in pre-application, post-application, and 

fall low-flow samples during 1989-90

, micrograms per liter; N, number of samples analyzed; <, less than. 1 990 samples are from a stratified random subset of site samples in 1 989 and are 

biased toward a slightly higher detection rate; ametryn, prometryn, terbutryn were not detected]

Herbicide or metabolite

Alachlor

Atrazine

Cyanazine

Deethylatrazine

Deisopropylatrazine

Metolachlor

Metribuzin

Prometon

Propazine

Simazine

Reporting 

limit 

(HQ/L)

0.05

.05

.20

.05

.05

.05

.05

.05

.05

.05

Percentage of detections

1989 

pre-application 

(N=55)

18

91

5

54

9

34

2

0

0

7

1989 

post-application 

(N=132)

86

98

63

86

54

83

53

23

40

55

1989 

fall low flow 

(N=145)

12

76

0

47

0

44

0

6

<1

3

1990 

pre-application 

(N=52)

50

96

8

86

44

67

0

10

0

15

1990 

post-application 

(N=52)

92

1 00

81

98

90

96

58

17

65

67

Adams and Thurman (1991) found that DAR values 

for atrazine transport through the unsaturated zone 

were larger than 1.0, whereas DAR values for atrazine 

transported off the field by surface runoff were much 

less than 1.0. Thus, the magnitude of the DAR may be 

indicative of ground-water recharge by atrazine-laden 

water.

The ELISA data indicate that water samples 

from 55 percent of the streams in the hydrologic units 

studied exceeded the USEPA Maximum Contaminant 

Level (MCL) for atrazine in drinking water during the 

post-application sampling in 1989. It is important to 

note that ELISA cross reacts with ametryn, prometon, 

prometryn, propazine, simazine, and terbutryn, to a 

varying degree, but not with cyanazine and the 

degradation products of atrazine at the concentrations 

typically found in water samples (Thurman and others, 

1990). The GC/MS data indicate that cyanazine and 

deethylatrazine frequently were detected along with 

trace concentrations of propazine and simazine, but 

there were no detections of ametryn, prometon, 

prometryn or terbutryn. In fact, atrazine was detected 

in 98 percent of the post-application stream water 

samples. The reporting limit of the ELISA method 

was set at 0.20 ng/L. At this concentration, neither 

false-positive nor false-negative detections were 

found. The ELISA correlation with GC/MS was good,

 ^

with a correlation (r ) of 0.86 using a simple 

regression with a nonliner fit for 127 surface-water

samples. The difference between the measured and 

projected concentrations was less than 0.15 |ig/L for 

50 percent of the water samples and less than 0.8 (ig/L 

for 80 percent of the samples. Thus, ELISA is a viable 

method for triazine surveys of surface water, ground 

water, and precipitation in the Midwest.

SUMMARY

The results of this reconnaissance study indicate 

that runoff clearly has a regional effect on water 

quality after spring application of herbicides. This 

study also indicates that more needs to be learned 

about the duration of large herbicide concentrations in 

streams during periods of storm runoff and the 

possible implications of storing this water in reservoirs 

for long-term use. Furthermore, the large number of 

water samples in which a herbicide MCL was 

exceeded reinforces the need for environmental health 

studies to examine the combined and synergistic 

effects of the major herbicides and their degradation 

products. Finally, future studies should investigate the 

persistent degradation products of herbicides, such as 

deethylatrazine and deisopropylatrazine, and their role 

as hydrologic tracers in studies of the interaction of 

surface water and ground water.

452
REFERENCES CITED

Adams, C.D., and Thurman, E.M., 1991, Formation and 

transport of deethylatrazine in the soil and vadose 

zone: Journal of Environmental Quality, v. 20, no. 3, 

p. 540-547.

Bushway, R.J., Perkins, Bill, Savage, S.A., Lekousi, S.L., 

and Ferguson, B.S., 1988, Determination of atrazine 

residues in water and soil by enzyme immunoassay: 

Bulletin of Environnmental Contamination and 

Toxicology, v. 40, p. 647-654.

Federal Register, 1989, National Primary and Secondary 

Drinking Water Regulations: Washington, D.C., v. 54, 

no. 97, May 22, 1989, p. 22062-22160.

___1990, National Primary and Secondary Drinking 

Water Regulations: Washington, D.C., v. 55, no. 143, 

July 25, 1990, p. 30370-30447.

Fishman, M.J., and Friedman, L.C., 1989, Methods for the 

determination of inorganic substances in water and 

fluvial sediments: U.S. Geological Survey Techniques 

of Water-Resources Investigations, book 5, chap. Al, 

545 p.

Gianessi, L.P., and Puffer, C.M., 1990, Herbicides use in the 

United States: Washington, D.C., Resources for the 

Future, National Summary Report, revised April 1991, 

128 p.

Goolsby, D.A., Coupe, R.C., and Markovchick, D.J., 199la, 

Distribution of selected herbicides and nitrates in the 

Mississippi River and its major tributaries April-June 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4163, 35 p.

Goolsby, D. A., Thurman, E. M., Clark, M.L., and Pomes, 

M.L., 1991b, Immunoassay as a screening tool for 

atrazine herbicides in streams Comparison with gas 

chromatographic/mass spectrometric methods, in 

Vanderlaan, Martin, Stanker, L.H., Watkins, B.E., and 

Roberts, D.W., eds., Immunoassays for trace chemical 

analysis: American Chemical Society, ACS 

Symposium Series 451, chap. 8, p. 86-99. 11

Goolsby, D. A., Thurman, E. M., Kolpin, D. W., 1990, 

Geographic and temporal distribution of herbicides in 

surface waters of the upper midwestern United States, 

1989-90: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 86-99.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., Mills, M.S., 

Pomes, M.L., and Kolpin, D.W., 1992, A 

reconnaissance study of herbicides and their 

metabolites in surface water of the midwestern United 

States using immunoassay and gas 

chromatography/mass spectrometry: Environmental 

Science and Technology, v. 26, 2440-2447.

Thurman, E.M., Meyer, M.T., Pomes, M.L, Perry, C.A., and 

Schwab, A.P., 1990, Enzyme-linked immunosorbent 

assay compared with gas chromatography/mass 

spectrometry for the determination of triazine 

herbicides in water: Analytical Chemistry, v. 62, 

p. 2043-2048.

U.S. Environmental Protection Agency, 1988, Drinking 

water health advisory Pesticides: Chelsea, Mich., 

Lewis Publishers, 819 p.

453
454
Chemistry, Degradation, and Transport of Triazine 

Herbicide Metabolites in Surface Water

By E.M. Thurman1 , M.S. Mills 1 , ancfM.T. Meyer1

Abstract

Deethylatrazine and deisopropylatrazine are 

two major metabolites of the triazine herbicides 

that occur in surface water of the midwestern 

United States. They may originate from the 

decomposition of several original parent 

compounds. Atrazine, the major triazine herbicide 

used in the Midwest, degrades systematically to 

both deethylatrazine and deisopropylatrazine with 

a ratio of approximately 0.40 for runoff water from 

cornfields. Ground water usually contains only 

trace concentrations of deisopropylatrazine 

compared to deethylatrazine because of the more 

rapid decomposition of deisopropylatrazine in the 

unsaturated zone. Thus, the deisopropylatrazineto-deethylatrazine ratio (D2R) may be used to 

distinguish surface-water runoff from ground 

water that discharges to surface water. However, 

other parent herbicides may affect the D2R. For 

example, cyanazine may degrade to deisopropylatrazine by the loss of the cyano-isopropyl group. 

Because cyanazine is the fourth most frequently 

used herbicide in the Midwest, the D2R in surface 

water often is considerably greater than 0.40 (from 

0.60-0.80). Simazine, another parent triazine 

herbicide, also may degrade to deisopropylatrazine. Simazine is used for weed control along 

highways and in orchards where considerably 

larger concentrations are applied. The resulting 

runoff contains deisopropylatrazine as a major 

dealkylated degradation product. Propazine, a 

parent triazine herbicide that is used only occasionally on grain sorghum, degrades to yield the

'U.S. Geological Survey, Lawrence, Kaiis.

deethylatrazine metabolite. Finally, terbuthylazine, a triazine not marketed in the United States, 

but used extensively in Europe, also may degrade 

to deisopropylatrazine via the loss of a t-butyl 

group. Thus, these parent compounds should be 

considered when using metabolite ratios as indicators of flow path.

INTRODUCTION

Agricultural practices may cause widespread 

degradation of water quality in the midwestern United 

States (Humenik and others, 1987; Thurman and others 

1991; 1992). Approximately three-fourths of all preemergent herbicides used in the United States are 

applied to row crops in a 10-State area, called "the 

Corn Belt" (Gianessi and Puffer, 1986). Because herbicides are water soluble, there is the potential for leaching into surface and ground water (Hallberg, 1989; 

Thurman and others, 1991), as well as aerial transport 

and occurrence in precipitation (Glotfelty and others, 

1983; Richards and others, 1987). Monitoring studies 

in the Midwest have shown widespread detection of 

herbicides and their metabolites in ground water (Hallberg, 1989) and in surface water (Wauchope and others, 1978; Glotfelty and others, 1984; Leonard, 1988; 

Pereira and Rostad, 1990; Thurman and others, 1991; 

1992). Runoff from fields immediately after herbicide 

application substantially increases herbicide concentrations in streams that eventually are reflected in 

elevated concentrations in rivers, such as the Mississippi (Pereira and Rostad, 1990). Furthermore, conventional water-treatment practices do not remove these 

herbicides (Kearney and others, 1988; Miltner and 

others, 1989), which then can occur in drinking water.

In spite of these problems, the use of herbicides is 

required for successful farming, and elimination of 

herbicide use would create an economic hardship to 

farmers. Thus, it is important to understand the effects

455
EXPLANATION

U SAMPLED 1989 

H SAMPLED 1989-90

 BOUNDARY OF

HYDROLOGIC UNIT

RECONNASSANCE 

SAMPLING SITE

STORM-RUNOFF 

SAMPLING SITE

37°

Base from U.S. Geological Survey digital survey, 12.000,000,1972

Abers Equal-Area Conic projection

Standard parallels 39°3Cr and 43°301 . central meridian -90°30-

 , jr Illinois

o 100 200 300 MILES

0 100 200 300 KILOMETERS

Figure 1. Location of stream-sampling sites in the midwestern United States.

of herbicides on water quality. An approach to understanding the effects of herbicides on water quality is to 

identify molecules as tracers of surface- and groundwater flow. A possible set of tracers is the two major 

dealkylated metabolites of triazine herbicides, deethylatrazine and deisopropylatrazine. The concentrations 

of these two compounds may be compared to the 

parent herbicide or to each other as ratios. These ratios 

include the deethylatrazine-to-atrazine ratio (DAR) 

and the deisopropylatrazine-to-deethylatrazine ratio 

(D2R).

The objective of the research described herein was 

to examine herbicide metabolites as tracers of surfaceand ground-water flow. The specific objectives were to: 

(1) review previous work on the amounts of herbicides 

and dealkylated metabolites in the spring flush of 

streams on a regional scale and to examine the persistence of these metabolites in surface water, (2) illustrate 

how the parent herbicides may degrade to deethylatrazine and deisopropylatrazine, the major dealkylated 

metabolites of several important triazine herbicides, and 

(3) hypothesize how metabolites of atrazine may be used 

as indicators of surface- and ground-water interactions.

DESCRIPTION OF STUDY AREA AND 

SAMPLE COLLECTION

Water samples were collected at 149 sites in 122 

river basins in a 10-state area during three time periods 

(fig. 1). These states include: Illinois, Indiana, Iowa, 

Kansas, Minnesota, Missouri, Nebraska, Ohio, South 

Dakota, and Wisconsin. Water samples were collected 

during March and April (pre-planting), May and June 

(post-planting), and October and November (harvest) 

of 1989 at U.S. Geological Survey streamflow-gaging 

stations. One-third of the sites were resampled during 

1990 at pre-planting and post-planting times. A stratified random-sampling procedure was used to ensure 

geographic distribution. Drainage areas of the selected 

basins ranged from 260 km2 to more than 160,000 km2 , 

and, collectively, the basins drained more than 500,000 

km2 of the Midwest (fig. 1).

All water samples from the reconnaissance were 

collected by depth-integrating techniques from three 

to five locations across each stream. The samples were 

collected and composited in large glass containers and 

filtered through glass-fiber filters (1-micron pore

456
diameter) into baked glass bottles for shipment to the 

laboratory. The sampled drainage areas ranged from 

200 to 5,000 km2 . All samples were analyzed for 

specific conductance, pH, nitrite plus nitrate, and triazine herbicides by enzyme-linked immunosorbent 

assay. Twenty-five percent of the samples were 

selected for analysis by gas chromatography/ 

mass spectrometry (GC/MS).

EXPERIMENTAL METHODS 

Reagents and Supplies

Methanol, ethyl acetate, and iso-octane were 

pesticide-grade solvents. Stock solutions of triazine 

herbicides and their deionized water were charcoal filtered and glass distilled prior to use. The C\ 8 (carbon-18) 

cartridges for solid-phase extraction contained 360 mg 

of 40 (um C )8-bonded silica. Standard solutions were 

prepared in methanol, and d- 10 phenanthrene was used 

as an internal GC/MS quantification standard.

Solid-Phase Extraction

An automated workstation was used for solidphase extraction (SPE) of the analytes. C 18 cartridges 

were preconditioned sequentially with 2 mL each of 

methanol, ethyl acetate, and distilled water. Each 10-mL 

water sample was spiked with a 100 mL of a surrogate 

standard, terbuthylazine (2.4 ng/fiL), and pumped 

through a cartridge at a rate of 20 mL/min by robotic 

probe. Analytes were eluted with ethyl acetate and 

spiked automatically with d- 10 phenanthrene (0.2 ng/ 

fiL, 500 f^L). The ethyl-acetate layer was transferred 

automatically from a small co-eluted water layer to a 

centrifuge tube for nitrogen evaporation. The robotic 

probe was washed between samples by immersing in 

4 mL of ethyl acetate and bubbling air through the 

probe to ensure thorough removal of any herbicide or 

spike residues adhering to the outside of the probe. 

Each sample required approximately 13 minutes for 

SPE. Finally, the extract was evaporated to 50 fiL by 

a turbovap at 45 °C under a nitrogen stream.

Analysis by Gas Chromatography/Mass 

Spectrometry

Automated GC/MS analyses of the eluates were 

performed with a mass selective detector. Operating 

conditions were: ionization voltage, 70 electron volts; an 

ion-source temperature of 250 °C; electron multiplier, 

2,200 volts; direct capillary interface at 280 °C, tuned

daily with perflurotributylamine; and a 50-millisecond 

dwell period. Separation of the herbicides was carried 

out using a fused-silica 12 m x 0.2 mm capillary 

column of methyl silicone with a film thickness of 

0.33 um. Helium was used as the carrier gas at a flow 

rate of 1 mL/min and a head pressure of 35 kPa. The 

column temperature was held at 50 °C for 1 minute, 

then ramped at 6 °C per minute to 250 °C, where it was 

held for 10 minutes. The injector temperature was 280 °C. 

The filament and multiplier were not turned on until 5 

minutes into the analysis. Quantification of the base 

peak of each compound was based on the response of 

the 188 ion of the internal standard, d- ]0 phenanthrene. 

Confirmation of the molecular ion was based on the 

presence of two confirming ions with a retention time 

match of ±0.2 percent relative to d-[ 0 phenanthrene.

RESULTS AND DISCUSSION 

Spring-Flush Phenomenon

Atrazine was detected in 98 percent of the postplanting samples (table 1). The most striking feature of 

the reconnaissance data (table 1; fig. 2) was that large 

concentrations of herbicides were flushed from croplands and were transported through the surface-water 

system as pulses in response to late spring and early 

summer rainfall. The significance of the reconnaissance is that the pulse effect occurs at a regional scale 

throughout the midwestern United States and appears 

to be linked to regional agricultural land use. Herbicides are applied in the spring as a pre-emergent treatment. Fall application is minimal, and this fact is 

reflected by the small herbicide concentrations during 

the post-harvest sampling. Herbicide concentrations in 

only one basin exceeded U.S. Environmental Protection Agency Maximum Contaminant Levels (MCLs) 

for drinking water in the post-harvest samples. None of 

the pre-planting samples exceeded the MCLs.

Box plots of the GC/MS data from the three sampling periods are shown in figure 2. Median concentrations of the four major herbicides (alachlor, atrazine, 

cyanazine, and metolachlor) increased by an order of 

magnitude, or more, from the pre-planting to the postplanting sampling periods, and then decreased to near 

pre-planting concentrations at the harvest sampling 

(fig. 2A, 2B). For example, the median concentration 

of atrazine during the post-planting period was 3.8 ng/ 

L compared with 0.23 ng/L during the pre-planting and 

post-harvest periods. The total herbicide concentration, 

defined here as the sum of the concentrations of the 11 

herbicides and 2 metabolites determined by GC/MS,

457
Table 1. Summary of herbicide concentrations detected in pre-planting, 

post-planting, and harvest samples, 1989 

[ng/L, micrograms per liter. Concentrations were determined by gas 

chromatography/mass spectrometry with detection limits of 0.05 ng/L for all 

herbicides except cyanazine, which is 0.20

Number Concentrations

Herbicide

Alachlor

Atrazine

Deethyl-atrazine 

Deisopropyl-atrazine

Cyanazine 

Metolachlor

Metribuzin

Propazine 

Prometon

Simazine

Alachlor

Atrazine

Deethyl-atrazine 

Deisopropyl-atrazine

Cyanazine 

Metolachlor

Metribuzin

Propazine 

Prometon

Simazine

Alachlor

Atrazine

Deethyl-atrazine 

Deisopropyl-atrazine

Cyanazine 

Metolachlor

Metribuzin

Propazine

of Detections 

samples (percent)

55

55

55 

55

55 

55

55

55 

55

55

132

1 32

1 32 

1 32

1 32 

1 32

132

1 32

1 32

132

145

145

145 

145

145 

145

145

145

Pre-Planting

18

91

54 

9

5 

34

2

0 

0

7

Post-Planting

86

98

86

54

63 

83

53

40 

23

55

Harvest Samples

12

76

47 

0

0

44

0

<1

Median

(ng/L)

<0.05

0.23

0.05 

0.05

<0.20 

<0.05

<0.05

<0.05 

<0.05

<0.05

0.92

3.8

0.28 

0.09

0.97 

1.30

0.14

<0.05 

<0.05

0.07

0.05

0.23

0.05 

O.05

O.20 

O.05

O.05

0.05

Maximum

(ng/L)

0.44

1.70

0.39 

0.59

0.52 

0.53

0.16

O.05 

O.05

8.70

51

108

4.4 

3.2

61 

40

7.6

1.4 

0.93

7.0

0.5

3.1

0.56 

O.05

O.20 

3.4

0.05

0.08

was calculated for each sample. The results (fig. 2C) 

show that the median concentration of total herbicides 

in the 1989 post-planting samples was 8.7 ug/L, or 

more than 20 times larger than the pre-planting concentrations. Furthermore, resampling studies in one-third 

of the basins during 1990 verified these concentrations 

(fig. 2C).

The apparent order of decreasing stability of the 

herbicides and metabolites on the basis of results 

shown in figure 2 and listed in table 1 is atrazine, 

deethylatrazine, metolachlor, cyanazine, deisopropylatrazine, and alachlor. This stability relation may be 

interpreted from the number of detections of herbicides 

in both the pre-planting and post-harvest samples and 

in the range of the concentrations during those two 

sampling periods (figs. 2A and 2B). Decomposition 

rates in soil from the published literature support these

findings (Nash, 1988). For example, atrazine has a 

reported half-life of 140 days; metolachlor, 90 days; 

cyanazine, 85 days; and alachlor, 50 days. Half-lives 

were not known for deethylatrazine and deisopropylatrazine, but it is known that deethylatrazine will 

remain in the soil from 1 year to the next (Mills, 1991; 

Mills and Thurman, 1994).

The temporal trends in the number of detections of 

various herbicides are significant for several reasons. 

First, trends indicate that some of the compounds persist from year to year in soil and water. Second, degradation products, such as deethylatrazine, are both 

persistent and mobile in water. Third, the data suggest 

that the DAR, which has been proposed as an indicator 

of nonpoint-source contamination of ground water 

(Thurman and others, 1994), might also be used as a 

tracer of ground-water discharge into rivers.

458
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EXPLANATION

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YEAR

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and 1990 for pre-planting and post-planting periods.

459
Cl

Cl

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NH A.A"N 

C2H5

 » »

A.A

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Deethylatrazine 

CH3

CH

Explanation

C = Carbon 

H = Hydrogen 

N = Nitrogen 

Cl = Chlorine 

C2H5= Ethyl 

CH3 = Methyl

Figure 3. Degradation pathway for dealkylation reactions of major triazine herbicides to deethylatrazine and 

deisopropylatrazine.

Degradation of Triazine Herbicides

The degradation pathway for the dealkylation 

reactions of major triazine herbicides to deisopropylatrazine and to deethylatrazine is shown in figure 3. 

The triazine herbicides atrazine, cyanazine, simazine, 

and terbuthylazine may degrade to the same metabolite, deisopropylatrazine. By similar fashion, atrazine 

and propazine may degrade to deethylatrazine. Field 

studies (Mills, 1991; Mills and Thurman, 1994) show 

that atrazine degrades to both deethylatrazine and 

deisopropylatrazine at relatively constant rates, resulting in a D2R of approximately 0.40 (Thurman and

others, 1994). Mills and Thurman (1994) showed that 

there is preferential loss of deisopropylatrazine with 

depth because of the ease of removing an ethyl group 

over an isopropyl group. Thus, deethylatrazine will be 

transported downward in soil and eventually will be 

detected in ground water, whereas D2R is much smaller 

in ground water than in surface water. Thus, these two 

compounds may be used to follow rate of flow of herbicides through the unsaturated zone. Slow transport 

through the unsaturated zone would remove deisopropylatrazine but not deethylatrazine. Rapid transport 

(macropore flow) would result in the presence of both 

metabolites. Also, the D2R may be used to indicate

460
surface-water flow because of the relatively constant 

ratio of the metabolites in surface water. This result 

suggests that degradation of both metabolites is slow in 

surface water, a result born out in studies by Pereira and 

Rostad (1990) on the Mississippi River.

Adams and Thurman (1991) found that atrazine 

transport through the unsaturated zone produced a 

DAR greater than 1.0, whereas the DAR for atrazine 

transported from the field in surface runoff was much 

less than 1.0. Furthermore, deisopropylatrazine 

degraded rapidly in the unsaturated zone but was an 

important metabolite in surface runoff from fields 

(Mills, 1991; Mills and Thurman, 1994). Thus, the 

DAR may be indicative of ground-water recharge by 

atrazine-laden water. For example, the median DAR 

was 0.30 in the pre-planting samples, less than 0.10 in 

the post-planting samples, and 0.4 in the harvest samples (fig. 2B). It is hypothesized that the large decrease 

in the DAR from 0.3 to less than 0.10 indicates that 

runoff of recently applied herbicides contains a small 

ratio of degradation product, in spite of the increased 

detections and greater concentrations of deethylatrazine (fig. 2B). The large increase in the DAR from less 

than 0.10 to 0.40 in the post-planting to the harvest 

samples may reflect the discharge of alluvial ground 

water that provides stream base flow and the flux of 

herbicides into surface water.

Contamination of Alluvial Aquifers

A process for surface-water contamination of 

alluvial ground water is shown in figure 4. The preplanting samples had a large DAR (0.30), but they 

also contained a few detections of deisopropylatrazine 

(9 percent). These results suggest both a surface- and 

ground-water origin of herbicides during this sampling 

period as do the discharge records. The surface component likely is due to recent spring melting of snow 

and ice. The post-planting samples collected during a 

period of runoff had a small DAR (less than 0.10) and 

the greatest number of detections of deisopropylatrazine (54 percent), which indicates that surface runoff is 

the major contributor of herbicides at this time. Finally, 

the post-harvest sampling during a low streamflow 

period had the largest DAR (0.40) and no detections of 

deisopropylatrazine, which indicates that alluvial 

ground water was likely the major source of herbicide 

at that time (Squillace and Thurman, 1992). Measurements of streamflow (fig. 4) indicate that base-flow 

conditions predominated during the harvest sampling.

Pre-Planting

DAR = 0.30

Deisopropylatrazine = 9 percent detections

Land surface

GW

SW Input (Small)

GW Input (Large)

Post-Planting

DAR = less than 0.10

Deisopropylatrazine = 54 percent detections

Land surface Sfream SW Input (Large)

- __ T

Harvest

DAR = 0.40

Deisopropylatrazine = 0 percent detections 

Land surface . SW Input (Small)

___________T 

GW Input (Large)

NOT TO SCALE

EXPLANATION 

 1_ GROUND-WATER TABLE 

GW GROUND WATER 

SW SURFACE WATER 

 +- DIRECTION OF FLOW

Figure 4. Deethylatrazine-to-atrazine ratios (DAR) as a result 

of surface-water and ground-water interactions during three 

sampling periods.

Surface water contaminated during the spring 

runoff is hypothesized to contribute significantly to the 

contamination of alluvial ground water. Mechanisms 

include a rapid increase in stream stage that occurs in 

spring when rivers are bankfull. During this time, 

ground-water gradients may reverse and cause streamwater to flow into adjacent alluvial aquifers (fig. 4). 

Contaminated surface water also may enter alluvial 

aquifers by recharge from floodwaters and runoff from 

areas that discharge onto the alluvial plain. Later, this 

discharge becomes a recharge for the alluvial aquifer.

461
Furthermore, large alluvial plains typically are planted 

to row crops in the "Corn Belt," where herbicides are 

applied directly to the alluvial soils with the potential 

for leaching to the shallow ground water. Studies of 

metabolite degradation continue to improve the understanding of these various transport processes.

REFERENCES CITED

Adams, C.D., and Thurman, E.M., 1991, Formation and 

transport of deethylatrazine in the soil and vadose zone: 

Journal of Environmental Quality, v. 20, p. 540-547.

Gianessi, L.P., and Puffer, C.M., 1986, Use of selected 

pesticides for agricultural crop production in the United 

States, 1982-1985: Springfield, Va., National Technical 

Information Service, 490 p.

Glotfelty, D.E., Taylor, A.W., Isensee, A.R., Jersey, J., and 

Glen, S., 1984, Atrazine and simazine movement to the 

Wye River Estuary: Journal of Environmental Quality, 

v. 13, 115-121.

Glotfelty, D.E., Taylor, A.W., and Zoller, W.H., 1983, 

Atmospheric dispersion of vapors Are molecular 

properties important?: Science, v. 219, p. 843-845.

Hallberg, G.R., 1989, Pesticide pollution of ground water in 

the humid United States: Agriculture Ecosystems Environment, v. 26, p. 299-367.

Humenik, F.J., Smolen, M.D., and Dressing, S.A., 1987, 

Pollution from nonpoint sources: Environmental 

Science and Technology, v. 21, p. 737-742.

Kearney, PC., Muldoon, M.T., Somich, C.J., Ruth, J.M. and 

Voaden, D.J., 1988, Biodegradation of ozonated 

atrazine as a waste-water disposal system: Journal of 

Agriculture and Food Chemistry, v. 36, p. 1301-1306.

Leonard, R.A., 1988, Herbicides in surface water, in

Grover, R. ed., Environmental chemistry of herbicides: 

Boca Raton, Fla., CRC Press, chap. 3, p. 45-88.

Mills, M.S., 1991, Field dissipation of encapsulated 

herbicides Geochemistry and degradation: 

Lawrence, University of Kansas, Master's Thesis, 98 p.

Mills, M.S, and Thurman, E.M., 1994, Reduction of

nonpoint-source contamination of surface and ground 

water by starch-encapsulated atrazine: Environmental 

Science and Technology, v. 28, p. 73-79.

Miltner, R.J., Baker, D.B., Speth, T.F., and Grank, C.A., 

1989, Treatment of seasonal pesticides in surface 

waters; Journal of American Water Works Association, 

v. 81, p. 43-52.

Nash, R.B., 1988, Dissipation from soil, in Grover, R., ed., 

Environmental chemistry of herbicides; Boca Raton, 

Fla., CRC Press, chap. 5, p. 131-170.

Pereira, W.E., and Rostad, C.E., 1990, Occurrence, distributions, and transport of herbicides and their degradation 

products in the lower Mississippi River and its tributaries: Environmental Science and Technology, v. 24, 

p. 1400-1406.

Richards, R.P., Kramer, J.W., Baker, D.B., and Krieger, 

K.A., 1987, Pesticides in rainwater in the northeastern 

United States: Nature, v. 237, p. 129-131.

Squillace, P.J., and Thurman, E.M., 1992, Herbicide transport in rivers: Importance of hydrology and geochemistry in nonpoint-source contamination: Environmental 

Science and Technology, v. 26, p. 538-545.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1991, Herbicides in surface waters of the 

midwestern United States The effects of the spring 

flush: Environmental Science and Technology, v. 25, 

p. 1794-1796.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., Mills, M.S., 

Pomes, M.L., and Kolpin, D.W., 1992, A reconnaissance study of herbicides and their metabolites in 

surface water of the midwestern United States using 

immunoassay and gas chromatography/mass spcctrometry: Environmental Science and Technology, v. 26, 

p. 2440-2447.

Thurman, E.M., Meyer, M.T., Mills, M.S., Zimmerman, 

L.R., Perry, C.A., and Goolsby, D.A., 1994, Formation 

and transport of deethylatrazine and deisopropylatrazine in surface water: Environmental Science and 

Technology, v. 28, p. 2267-2277.

Wauchope, R.D., 1978, Pesticide content of surface-water 

draining from agricultural fields: Journal of Environmental Quality, v. 7, p. 459-472.

462
The Transport and Degradation of Cyanazine

Metabolites in Surface Water of the

Midwestern United States

By M.T. Meyer1 and E.M. Thurman 1

Abstract

Cyanazine was detected in 73 percent, 

cyanazine amide in 65 percent, deethylcyanazine 

in 36 percent, and deisopropylatrazine in 52 

percent of the samples collected during the first 

storm runoff after herbicide application in a reconnaissance study of stream water from 122 basins in 

a 10-State area in the Midwestern United States in 

1989. Furthermore, these compounds were 

detected in as many as 30 percent of the samples 

collected in the spring before planting and in the 

fall of the year during low streamflow. Analysis of 

samples collected from several stream sites during 

storm-runoff throughout the spring and early 

summer of 1990 indicate that cyanazine degrades 

to deisopropylatrazine and that the deisopropylatrazine-to-deethylatrazine ratio (D2R) is an indicator of the amount of cyanazine relative to atrazine 

that is used in the basin.

INTRODUCTION

Contamination of surface and ground water by 

chemicals applied for agricultural usage is a significant 

problem in the midcontinent region (Humenik and 

others, 1987). Pre-emergent herbicides used for weed 

control comprise a substantial amount of this agricultural usage. Studies have shown that in the Midwestern 

United States herbicide contamination of surface water 

is a periodic and persistent problem (Spalding and 

Snow, 1989; Stamer and others, 1990; Goolsby and 

others, 1991) and now there is concern that herbicide 

metabolites also may be a potential contamination

'U.S. Geological Survey, Lawrence, Kans.

problem in surface water. Most herbicide transport in 

surface water occurs in the spring and early summer 

(Thurman and others, 1991; Squillace and Thurman, 

1992). Also, it has been demonstrated that the effect of 

the spring flush on herbicide concentration in surface 

water is a regional phenomenon in the Midwestern 

United States (Goolsby and others, 1989; Thurman and 

others, 1991, 1992).

Recent studies show that herbicide metabolites can 

be useful in determining transport and contamination 

processes (Adams and Thurman, 1990; Thurman and 

others, 1991, 1992). Thurman and others (1991, 1992) 

proposed that the ratio of deethylatrazine, a metabolite 

of atrazine, to atrazine is an indicator of alluvial 

ground-water contributions to surface water and that 

the more rapidly degraded metabolite of atrazine, 

deisopropylatrazine, can be an indicator of surfacewater runoff. Thus, the presence of deisopropylatrazine 

in ground water, along with a low deethylatrazine-toatrazine ratio (DAR), may indicate the interaction of 

surface water with ground water.

The potential for using herbicide metabolites as 

indicators of transport processes is a relatively new 

concept that might be extended to the other major 

herbicides found in surface water, such as cyanazine, 

alachlor, and metolachlor. Cyanazine, one of the most 

widely used herbicides in the Midwest, readily metabolizes to form a suite of compounds that could be used 

as indicators. However, little is known about the occurrence of cyanazine metabolites in surface water. 

The objectives of this paper are to (1) describe the 

occurrence of cyanazine metabolites in surface water, 

(2) present evidence that cyanazine degrades to deisopropylatrazine, and (3) demonstrate that cyanazine and 

metabolite ratios are indicators of transport. To accomplish these objectives, this paper presents the results 

from a herbicide reconnaissance study of stream water 

in the Midwestern United States conducted in 1989 and 

from a storm-runoff sampling study of stream water 

conducted in 1990.

463
STUDY DESIGN AND METHODS OF 

ANALYSIS

The design of the herbicide-reconnaissance study 

of stream water in the Midwest has been described in 

Goolsby and others (1989) and Thurman and others 

(1992). Briefly, 149 sampling sites in 122 river basins 

throughout a 10-State area were selected using a stratified random-sampling procedure. Each site was sampled in the spring before the application of herbicides, 

after application of herbicides during the first storm 

runoff, and in the fall during a period of low flow.

The storm-runoff study, conducted in 1990, was 

designed to assess the effects of successive storms on 

the concentration of herbicides in stream water from 

spring through early to midsummer. For this study, 

automated water samplers were installed at nine stream 

sites in five States (Illinois, Iowa, Kansas, Nebraska, 

and Ohio). The auto-samplers were programmed to 

collect samples every 1 to 2 days during low flow and 

every 1 to 2 hours during periods of storm runoff. 

Approximately 20 percent of the samples collected 

from each site were analyzed for herbicides by gas 

chromatography/mass spectrometry (GC/MS).

Samples were analyzed for triazine herbicides 

using solid-phase extraction, GC/MS after the methods 

of Thurman and others (1990) and Meyer and others 

(1993). GC/MS analyses for the cyanazine metabolites, 

deethylcyanazine and cyanazine amide, were performed 

by GC/MS using a capillary GC column with a 5-percent phenyl phase. The quantitation limit for cyanazine, cyanazine amide, and deisopropylatrazine was 

0.05 u,g/L, and for deethylcyanazine amide, 0.2 |ig/L.

RESULTS AND DISCUSSION

Reconnaissance Study Variation by Sampling 

Round

Samples from the herbicide reconnaissance study 

in 1989 were analyzed to determine the occurrence of 

cyanazine metabolites in stream water in the Midwestern United States. In the first round, samples were 

collected from mid-March through mid-April, before 

herbicides had been applied. Cyanazine was detected 

in 14 percent, cyanazine amide in 20 percent, and 

deethylcyanazine in § percent of the samples (table 1). 

The data from this sample round suggest that cyanazine 

metabolites can be available for transport into surface 

water throughout the year.

Table 1. Summary of surface-water herbicide reconnaissance data for cyanazine 

and cyanazine metabolites, 1989 

[|.ig/L, micrograms per liter; <, less than]

Sample 

round

1

2

3

Quantitation limit (ng/L)

Percent detections

Number of analyses

Avrage (u.g/L)'

Standard deviation (u.g/L)

Range (ng/L)

Percent detections

Number of analyses

Avrage (ng/L)

Standard deviation (ng/L)

Range ng/L)

Percent detections

Number of analyses

Avrage (ng/L)

Standard deviation (ng/L)

Range (ng/L)

Cyanazine

0.05

17

57

.24

.14

0.10-0.60

73

142

7.1

11

0.05-61

29

142

.15

.10

0.07-0.45

Compound

Cyanazine 

amide

0.05

17

54

.13

.06

0.08-0.20

65

105

1.6

2.1

0.09-11.1

34

139

.11

.07

0.05-0.38

Deethylcyanazine

0.20

(j)

54

.30

<.2()

<.20

36

105

.72

.61

0.05-2.5

0

139

<20

<.20

<20

limit.

Average concentration based only on samples for which the compound exceeded the quantitation

464
The samples for round 2 were collected during the 

first large storm runoff occurring in each basin after the 

majority of herbicides had been applied to the fields. 

Round-2 samples were collected between mid-May 

and mid-August. In this sample round, cyanazine was 

detected in 73 percent, cyanazine amide in 65 percent, 

and deethylcyanazine in 36 percent of the samples 

(table 1). These data show that cyanazine metabolites 

are readily generated in the soil and transported into 

surface water shortly after herbicide application. The 

average concentration (calculated only for compounds 

that were detected at quantifiable concentrations) for 

the round-2 samples was 7.1 ug/L for cyanazine, 

1.6 ug/L for cyanazine amide, and 0.72 ug/L for 

deethylcyanazine (table 1). The U.S. Environmental 

Protection Agency (USEPA) Health Advisory Level 

(HAL) for cyanazine is 1.0 ug/L (U.S. Environmental 

Protection Agency, 1992).

Round-3 samples were collected from late September through late-November. The stream sites were sampled during low flow when most of the stream water 

could be attributed to ground-water discharge. In this 

round, cyanazine was detected in 29 percent, cyanazine 

amide in 34 percent, and deethylcyanazine in 0 percent 

of the samples. The results from this round are consistent with the round 1 data in that some cyanazine and 

cyanazine amide persisted and was still detectable in 

many streams. Because the samples from round 3 were 

collected during low flow, the occurrence of cyanazine 

and cyanazine amide most likely represents ground 

water contributions to stream water; whereas samples 

collected in round 1 were not necessarily collected 

during low-flow conditions. Therefore, the presence of 

cyanazine and cyanazine amide in round-1 samples 

could reflect either overland-flow runoff or ground 

water contributions to stream water. The average concentration of each compound for the first and third 

rounds was much smaller than it was for the round-2 

samples. For example, the average concentration for 

cyanazine was 0.24 and 0.15 ug/L; cyanazine amide, 

0.13 and 0.11 ug/L; and deethylcyanazine 0.2 less than 

0.20 ug/L for rounds 1 and 3, respectively (table 1).

Deisopropylatrazine: A Degradation Product 

of Cyanazine

Deethylatrazine and deisopropylatrazine are 

important, but not exclusive, degradation products of 

atrazine. For example, propazine and simazine degrade 

to deethylatrazine and deisopropylatrazine, respectively; however, about 15 to 20 times more atrazine

than simazine or propazine is applied in the United 

States (Gianessi and Puffer, 1991). Therefore, the 

presence of deisopropylatrazine and deethylatrazine in 

surface water and ground water is ascribed almost 

exclusively to the degradation of atrazine. However, 

there is evidence that cyanazine also degrades to deisopropylatrazine. Sirons and others (1973) noticed that, 

when cyanazine was applied to field plots to which 

atrazine previously had been applied, the concentration 

of deisopropylatrazine increased. Also, a study by 

Benyon and others (1972) indicates that cyanazine 

degrades to deisopropylatrazine. This is important 

because atrazine use is only three times that of cyanazine in the United States (Gianessi and Puffer, 1991). 

Therefore, in areas where cyanazine is used in approximately the same quantities as atrazine, a substantial 

amount of the deisopropylatrazine produced and transported into surface water might be from the degradation of cyanazine.

Field-dissipation studies of atrazine have shown 

that the deisopropylatrazine-to-deethylatrazine ratio 

(D2R) is about 0.3 to 0.5 (E. M. Thurman, U.S. Geological Survey, written commun., 1993). The D2R 

obtained from surface-water samples where atrazine is 

the primary triazine herbicide are in good agreement 

with the atrazine field-dissipation studies (that is, the 

D2R generally ranges from 0.3 to 0.5). However, in 

surface-water samples where the concentration of 

cyanazine is about the same concentration as that of 

atrazine, the D2R increases from 0.7 to larger than 1.0.

This point can be illustrated by examining data 

from water samples collected from a basin where much 

more atrazine was applied than cyanazine, and in 

another basin where approximately the same amounts 

of atrazine and cyanazine were applied. For example, 

figures 1A and IB shows that water samples collected 

from the Delaware River storm-runoff site had much 

larger concentrations of atrazine than cyanazine under 

low-flow and high-flow conditions. The average concentration for atrazine was 8.9 ug/L and for cyanazine 

was 0.29 ug/L. In contrast to the Delaware River, samples from the Iroquois River storm-runoff site (fig. 1D) 

had cyanazine concentrations that were mostly equal to 

or less than five times the concentration of atrazine. For 

example, the average concentration for atrazine was 

2.6 ug/L and for cyanazine was 1.7 ug/L for samples 

collected from the Iroquois River storm-runoff site. 

Numerous samples from both the Delaware and 

Iroquois River sampling sites exceeded the USEPA

465
o 10,000

So 8,000 

£ <*

3 £ 6,000

2 5 4,000

DC

0.01

Delaware River, Kansas Iroquois River, I

HAL = Cyanazine^

A Atrazine _ 

A Cyanazine -:

Ui 

CU

V)

9

bo

0

1990

u

o>

1o>

a.

a> 

C/3

t- 03

3 S

1990

are not regressed.
Maximum Contaminant Level (MCL) for atrazine of 

3.0 ng/L (U.S. Environmental Protection Agency, 

1992) except during low flow (figs. 1A and 1C). 

However, only one sample from the Delaware River 

exceeded the HAL of 1.0 ug/L for cyanazine, whereas 

the concentration of cyanazine exceeded the HAL in 

19 samples from the Iroquois River.

Figure ID also shows that the concentrations of 

atrazine and cyanazine in water from the Iroquois 

River storm-runoff site decreased in a fairly linear fashion from mid-June through August. Furthermore, the 

concentration of cyanazine appears to have decreased 

more rapidly than the concentration of atrazine. This 

is consistent with observations in the literature that 

cyanazine degrades more rapidly than atrazine. The 

systematic decrease in the concentration of herbicides 

indicates that 4 to 5 weeks after herbicide application 

the amount of herbicide available for transport from the 

field into surface water diminishes rapidly. Therefore, 

the production of metabolites by cyanazine relative to 

atrazine should decrease as a function of time after herbicide application. The evidence presented here indicates that one of the degradation products of cyanazine 

is deisopropylatrazine; field studies are underway to 

confirm this hypothesis.

Cyanazine Ratios as Transport Indicators

To look at the relation between the amount of 

cyanazine relative to atrazine and its effect on the 

production of deisopropylatrazine, it is instructive to 

examine the atrazine-to-cyanazine ratio (ACR) and the 

D2R. Figure 2A shows that the ACR in samples collected from the Delaware River was much larger than 

5. For these same samples, figure 2B shows that the 

D2R for most of the samples was less than 0.50 and, 

except for one sample, all had a D2R of less than 0.55. 

Furthermore, the average D2R for these samples was 

0.44. The variation of the D2R in the samples from the 

Delaware River are in good agreement with the results 

from field-dissipation studies of atrazine (E.M. Thurman, U.S. Geological Survey, oral commun., 1993). 

Therefore, the D2R indicates that atrazine is the primary source of the deethylatrazine and deisopropylatrazine that were transported into the Delaware River. 

This is further supported by the fact that the concentrations of simazine and propazine were much less than 

that of cyanazine for the samples collected from this 

storm-runoff sampling site.

In contrast, in the Iroquois River Basin contains 

approximately equal amounts of atrazine and cyanazine are applied (fig. 3A), and the D2R is elevated (fig. 

3B) from less than 0.5 to between 0.7 and 1.1 shortly 

after the application of herbicides. Furthermore, for the 

first 1.5 months after application when the majority of 

cyanazine, transport occurred (fig. 10), most ACR's 

were less than 2.0. The ACR shows a steady increase 

from mid-May through August (fig. 3A), indicating 

that cyanazine degrades more rapidly than atrazine in 

soil, as reflected in changing concentrations of atrazine 

and cyanazine in water samples from successive stormrunoff events (fig. ID). Thus, the decreasing availability of cyanazine for transport into surface water relative 

to atrazine was reflected by the increasing ACR in the 

surface-water samples with time.

Figure 3B shows that the D2R of most of the 

samples from the Iroquois River were larger than 0.50. 

During the first 1.5 months after herbicide application 

when the ACR of most of the samples was less than 

2.0, the D2R of most of the samples was larger than 

0.70. The high D2R in these samples indicates that 

cyanazine was responsible for at approximately onethird of the deisopropylatrazine transported into the 

Iroquois River during this time. From July through 

August (fig. 3B) there was a marked decrease in the 

D2R, which correlated with an increase in the ACR. 

These data suggest that cyanazine is a major source of 

deisopropylatrazine in the Iroquois River and that, as 

the concentration of cyanazine relative to atrazine 

decreases, the D2R decreases because atrazine then 

becomes the primary source of deisopropylatrazine.

The data from the Delaware and Iroquois Rivers 

shows that the D2R can be used as an indicator of transport of other herbicides, such as cyanazine or simazine 

and propazine. Furthermore, the D2R could be used as 

an indication that further analyses are needed, such as 

analyses for the polar metabolites of cyanazine.

CONCLUSIONS

Data from the herbicide reconnaissance study 

show that cyanazine amide and deethylcyanazine are 

important degradation products of cyanazine and are 

readily transported into surface water. Furthermore, 

cyanazine and cyanazine amide were detected in water 

samples from the pre-herbicide application and fall 

sampling periods of the reconnaissance study.

467
Delaware River, Kansas

w

2o

200

100

10

1.0

0.10

A -2

ACR = 5

1.5

1.0

0.50

0.0

B

= 0.50

ua

a> VI

3

1990

u0>

.0

E

a> C/3

Figure 2. Discrimination graphs of (A) atrazine-to-cyanazine ratios (ACR) and (B) deisopropylatrazineto-deethylatrazine ratios (D2R), in water samples from the Delaware River, April-September 1990. 

For (B), the D2R graph, black squares correspond to when the ACR was greater than 5; for the 

unfilled triangles, there was no corresponding ACR. Minor tick marks are in weeks.

This demonstrates that cyanazine and cyanazine amide 

can be available for transport into surface water long 

after herbicides have been applied. Data from the 

Delaware River in Kansas show that the D2R in surface 

water from basins where much more atrazine than 

cyanazine is applied is similar to the D2R from atrazine 

field-dissipation studies. However, for basins where

atrazine and cyanazine are applied in approximately 

equal amounts, such as the Iroquois River Basin in 

Illinois, the D2R is considerably larger. Data from the 

Iroquois River showed that, when ACR is less than 2, 

atrazine was responsible for at least one-third of the 

deisopropylatrazine transported into the river. The 

increased contribution of deisopropylatrazine by

468
Iroquois River, Illinois

1

fc

LU ^ 

Q CC

LU O

z: =

1.5

1.0

0.50

0. DC 

0 fe cc <

0. 

O </)

LU 

Q

0.0

B

2 D^R = 0.70

2/

D2R = 0.50

 c

& I

Ui V

JD

1990 a,

CO

Figure 3. Discrimination graphs of (A) atrazine-cyanazine ratios (ACR) and (B) deisopropylatrazineto-deethylatrazine ratios (D2R), in water samples from the Iroquois River, April-September 1990. For 

(B), the D2 R graph, the open, gray and black squares correspond to when the ACR was less than or 

equal to 2, greater than 2 and less than or equal to 5, and greater than 5, respectively. Minor tick marks 

are in weeks.

cyanazine was reflected by an increase in the D2R to 

values larger than 0.70. Furthermore, the samples from 

the Iroquois River show that, as the amount of cyanazine relative to atrazine decreases with time, atrazine 

becomes the primary source of deisopropylatrazine, 

which is reflected by a decrease in the D2R. The data

from these two storm-runoff sites indicate that cyanazine degrades to deisopropylatrazine. Finally, the D2R 

is a sensitive indicator of whether the degradation of 

herbicides other than atrazine contributes a significant 

amount of the deisopropylatrazine that is transported 

into surface water.

469
REFERENCES CITED

Adams, C.D., and Thurman, E.M., 1990, Formation and 

transport of deethylatrazine in the soil and vadose zone: 

Journal of Environmental Quality, v. 20, no. 3, p. 540- 

547.

Benyon, K.I., Stoydin, G., and Wright, A.N., 1972, A

comparison of the breakdown of the triazine herbicides 

cyanazine, atrazine and simazine in soils and maize: 

Pesticide Biochemistry and Physiology, v. 2, p. 153- 

161.

Gianessi, L.P., and Puffer, C.M., 1991, Herbicide use in the 

United States: Washington, D.C., Resources for the 

Future, National Summary Report, 128 p.

Goolsby, D.A., Thurman, E.M., and Koplin, D.W., 1991, 

Geographic and temporal distribution of herbicides in 

surface waters of the upper Midwestern United States, 

1989-1990: U.S.Geological Survey Water Resources 

Investigations Reports 91-4034, p. 31.

Goolsby, D.A., Thurman, E.M., Koplin, D.W., and Detroy, 

M.G., 1989, A reconnaissance for the triazine herbicides in surface waters of the upper Midwestern United 

States, in U.S. Geological Survey 2nd National Symposium on Water Quality Abstracts of the Technical 

409, p. 31.

Humenik, F.J., Smolen, M.D., and Dressing, S.A., 1987, 

Pollution from nonpoint sources: Environmental Science and Technology, v. 21, p. 737-742.

Meyer, M.T., Mills, M.S., and Thurman, E.M., 1993, Automated solid-phase extraction of herbicides from water 

for GC/MS analysis: Journal of Chromatography, 

v. 629, p. 55-59.

Sirons, G.J., Frank, R., and Sawyer, T., 1973, Residues of 

atrazine, cyanazine and their phytoxic metabolites in a

clay loam: Journal of Agricultural Food Chemistry, 

v. 21, no. 6, p. 1016-1020.

Spalding, R.F., and Snow, D.D., 1989, Stream levels of 

agrichemicals during a spring discharge event: Chemosphere, v. 19, no. 8-9, p. 1129-1140.

Squillace, P.J., and Thurman, E.M., 1992, Herbicide transport in rivers Importance of hydrology and geochemistry in nonpoint-source contamination: Environmental 

Science and Technology, v. 26, no. 3, p. 538-545.

Stamer, J.K., Pope, L.M., and Zelt, R.B., 1990, Occurrence 

and distribution of pesticides in the lower Kansas River 

Basin in Kansas and Nebraska: Proceedings of the 7th 

annual Water and the Future of Kansas Conference, 

p. 25-26.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Koplin, 

D.W., 1991, Herbicides in surface waters of the 

Midwestern United States-Trie effect of the spring 

flush: Environmental Science and Technology, v. 25, 

p. 1794-1796.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., Mills, M.S., 

Pomes, M.L., and Koplin, D.W., 1992, A reconnaissance study of herbicides and their metabolites in 

surface water of the Midwestern United States using 

immunoassay and gas chromatography/mass spectrometry: Environmental Science and Technology, v. 26, 

p. 2440-2447.

Thurman, E.M., Meyer, M.T., Pomes, Michael, Perry, C.A., 

and Schwab, A.P., 1990, Enzyme-linked immunosorbent assay compared with gas chromatography/ 

mass spectrometry for the determination of herbicides 

in water: Analytical Chemistry, v. 62, p. 2043-2048.

U.S. Environmental Protection Agency, 1992, Drinking 

water regulations and health advisories: Washington, 

D.C., Office of Water, 13 p.

470
Transport and Concentrations of Selected Herbicides 

and Nitrate in the Sangamon River, Illinois, 

April 1991 -March 1992

ByG.P. Johnson1 anofR.H. Coupe1

Abstract

The Sangamon River, located in eastcentral Illinois, drains some of the most intensively used agricultural land in the Midwest. 

Effects of agricultural practices in the drainage 

basin on stream-water quality were investigated 

by collecting surface-water samples from 

April 1991 through March 1992 at Monticello, 

Illinois. Samples were analyzed for concentrations of selected herbicides and nitrate. Total load 

was calculated and compared to estimated 

application totals of each compound in the drainage basin above Monticello.

During the study period, 1,607 pounds of 

atrazine were transported by the river; this mass 

represents 1.4 percent of the total annual application of 113,700 pounds. The river transported 

669 pounds of cyanazine~a mass equal to 1.6 percent of the total annual application of 41,900 

pounds. About 353 pounds of alachlor were transported by the river (0.31 percent of the 112,300 

pounds applied). About 12,100 tons of nitrogen 

fertilizer were applied, and 3,393 tons of nitrate as 

nitrogen were transported by the river (28 percent 

of the total amount applied). However, other natural and human-derived sources of nitrate could 

have contributed to this total load.

The timing of chemical application and the 

quantity of rainfall and runoff after herbicide 

application affected the concentrations. Herbicide concentrations were highest in early spring

U.S. Geological Survey, Urbana, Illinois

during storms immediately following application. 

Peak concentrations for atrazine, cyanazine, and 

alachlor occurred in May; these concentrations 

were 21, 16, and 8.8 micrograms per liter, respectively. Nitrate concentrations were at or above the 

U.S. Environmental Protection Agency's maximum contaminant level of 10 milligrams per liter 

from the beginning of the study until June 11, 

1991. Nitrate concentrations then remained less 

than the maximum contaminant level until 

November 1991.

INTRODUCTION

Stream-water-quality criteria for agricultural 

chemicals are expressed in terms of chemical concentrations. Concentrations of these chemicals in streams 

provide an indicator of whether the water quality is 

suitable for the designated uses of the stream. Agricultural-chemical concentrations, especially in small 

streams, are affected by the timing, intensity, and duration of rainfall, which can vary from year to year. 

Therefore, the effect of agricultural practices on 

stream-water quality cannot be assessed directly from 

concentration data. The effects of agricultural practices on stream-water quality can be better understood 

by measuring the transport of agricultural chemicals in 

streams. If the amounts of these chemicals applied 

within a drainage basin are known, annual losses of 

these chemicals from agricultural fields into a stream 

can be determined.

Agricultural chemicals applied to croplands can 

undergo physical, chemical, and biological transformations. Depending on the properties of a chemical 

and on conditions such as rainfall, wind, soil moisture, 

soil type, and temperature, a chemical can sorb to soil

471
particles, be transported through the soil column, 

undergo biological or chemical transformations, and 

(or) volatilize to the atmosphere (Biggar and Seiber, 

1987). Transport of these chemicals into rivers can be 

caused by storm runoff from cropland, tributary flow, 

ground-water discharge, and (or) atmospheric 

deposition.

Atrazine and alachlor are the most heavily used 

agricultural chemicals in Illinois and in the United 

States. Atrazine and cyanazine are preemergent and 

postemergent herbicides used on corn, and alachlor is 

a preemergent herbicide used on corn and soybeans 

(Pike and others, 1990). Application of these chemicals usually is accomplished by surface spraying in the 

spring prior to planting.

Nitrogen fertilizer is applied to croplands as 

either anhydrous ammonia, urea, ammonium nitrate, 

or as a nitrogen solution (University of Illinois Cooperative Extension Service, 1990). Most of the nitrogen 

is applied by injection of anhydrous ammonia into the 

soil in March and April before planting. Nitrogen also 

can be applied in late fall and early winter in any of the 

different forms.

The Sangamon River is located in east-central 

Illinois and has a drainage area of 550 square miles at 

Monticello, 111. (fig. 1). The land use of the basin

EXPLANATION

 BASIN BOUNDARY

10 IS MILES

0 5 10 15 KILOMETERS

Figure 1. Sangamon River Basin upstream from Monticello, 

Illinois.

above Monticello is predominantly row-crop agriculture mostly corn and soybeans. No major communities are upstream from Monticello. The Sangamon 

River Basin above Monticello lies in the Bloomington 

Ridged Plain, which is characterized by low, broad 

ridges separated by wide, gently undulating moraines 

and wooded areas in the river valleys (Leighton and 

others, 1948).

The purpose of this paper is to describe the total 

load and concentrations of selected herbicides and 

nitrate in the Sangamon River. To measure the transport of agricultural chemicals in the Sangamon River 

in east-central Illinois, surface-water samples were 

collected from the Sangamon River at Monticello 

from April 12, 1991, through March 25, 1992, and 

were analyzed for concentrations of dissolved atrazine, dissolved cyanazine, dissolved alachlor, and 

nitrate. Total annual loads were calculated for each 

compound and compared with application amounts. 

These annual loads were compared to loads calculated 

from an earlier study of the same basin. Water 

discharges were determined from records collected at 

a streamflow-gaging station at the sampling site.

METHODS

An automatic sampler was installed in the 

streamflow-gage house on the Sangamon River at 

Monticello to collect surface-water samples. The automatic sampler was programmed to collect water samples at a frequency associated with a given water-level 

(stage) in the stream or a rate-of-change in stage. At 

base flow, samples were collected once every 48 

hours. During high flow, samples were collected three 

times per day. Sample collection was increased up to 

four times daily with a rapid rise in stage.

Water samples were pumped through Teflon 

tubing into clean 350-mL glass bottles. Samples were 

picked up at least twice weekly and more often during 

high-flow periods. The bottles were capped with aluminum foil and transported to the U.S. Geological 

Survey (USGS) office in Urbana, 111., for processing. 

Water samples analyzed for herbicide concentrations 

were filtered through a glass-fiber filter and shipped on 

ice to the USGS office in Lawrence, Kan. The herbicide samples were analyzed using solid-phase extraction with Gas Chromatography/Mass Spectrometry 

(GC/MS) as described in Goolsby and others (1991). 

Water samples for nitrate were filtered through a 

0.45-micron pore-size filter, preserved with mercuric

472
chloride (HgCl2), and then shipped to the USGS 

National Water Quality Laboratory (NWQL) in 

Arvada, Colo., for analyses.

After October 1991, the automatic sampler was 

removed because of the potential of samples freezing 

and the samples were collected manually. Because 

herbicide detections are less frequent, and concentrations are much lower in the fall and winter when compared to spring and summer (Thurman and others, 

1991), sampling frequency was approximately twice 

monthly from October 1991 through March 1992.

Daily mass transport of the selected herbicides 

and nitrates were calculated from concentration and 

streamflow data. For each compound, a graph showing 

the instantaneous concentration as a function of time 

was plotted. Adjacent points on these graphs were 

connected by straight lines, and this trace was digitized. A linear-interpolation algorithm was then used 

to estimate concentrations at 4-hour intervals. Discharge data, which were available at 15-minute intervals, were averaged to estimate the mean discharge for 

the 4-hour period. This mean discharge was multiplied 

by the concentration and a unit conversion factor that 

estimated the mass transport rate for that 4-hour 

period. These rates were multiplied by the duration of 

the period to estimate the load (total mass transported),

1,800

and the loads were summed to estimate the total load 

for each month.

RESULTS

The monthly mass transports of atrazine and 

nitrate as nitrogen during the study period are shown 

in figure 2. The monthly mean streamflow for October 

1990 through March 1992 and the historical monthly 

mean streamflow also are shown. Large amounts of 

atrazine and nitrate were transported by the Sangamon 

River at Monticello at the beginning of the study. 

Transport rates during April 1991 were 166 pounds 

per month of atrazine and 547 tons per month of 

nitrate as nitrogen. In May 1991, transport rates 

increased and peaked at 1,211 pounds of atrazine and 

891 tons of nitrate per month. Transport rates for all 

compounds analyzed were largest in May 1991 and 

were smallest in September 1991 during an extended 

dry period. Nitrate transport rates and streamflow 

increased beginning in November 1991. Atrazine 

transport rates also increased as streamflow increased 

in November, but the increase was small compared to 

the transport rates for April, May, and June 1991.

Table 1 lists the masses of atrazine, cyanazine, 

alachlor, and nitrate as nitrogen transported by month. 

The largest monthly transport rates for atrazine, cyanaMonthly mean streamflow, in cubic feet per second 

Mean monthly streamflow, in cubic feet per second (1908-91) 

Monthly nitrate load as nitrogen, in tons 

B-U Monthly atrazino load, in pounds

200 -

Figure 2. Monthly mean streamflow and long-term mean monthly streamflow (1908-91) for 

October 1990 to March 1992 and monthly loads of atrazine in pounds and nitrate in tons, for 

April 1991 to March 1992 at Sangamon River at Monticello, Illinois.

473
Table 1. Mass transport of agricultural chemicals in the Sangamon River at Monticello, I 

April 1991-March 1992 and April-August 1990

[Q, water discharge; ft3/s, cubic feet per second; Ibs, pounds; NO2+NO3, nitrite+nitrate; -, not applicable]

April 1991-March 1992

Mean daily Q 

Month (ft-'/s)

April

May

June

July

August

September

October

November

December

January

February

March

TOTAL

710

940

242

33

10

6.0

40

250

545

240

360

370

Atrazine 

(Ihs)

166

1,211

148

4.4

.X4

.15

3.3

20

23

6.4

14

9.1

1,607

Cyana/.ine 

(Ibs)

84

558

25

35

.07

.04

.10

.31

.18

.43

.27

.07

669

Alachlur 

(Ibs)

34

288

18

.43

.08

.08

.80

2.1

3.6

.26

1.3

2.5

353

Nitrate 

(tons)

547

891

193

3.7

.22

.15

4.0

139

608

227

377

403

3,393

April-August 1990

Mean daily Q Atrazine NO2+NOj

(ft-Vs) (Ibs) (ions)

411 68 327

1,524 797 1,489

1,744 1,155 1,415

887 227 764

120 8.5 46

.

.

.

.

-

2,256 4,040

zine, and alachlor were in May 1991, representing 75, 

83 and 81 percent, respectively, of the total mass transport during the study period. The largest nitrate transport rate also was in May 1991, but the mass 

transported represented only 26 percent of the total 

mass of nitrate transported during the entire period. 

Mass transport of nitrite plus nitrate as nitrogen 

(NC>2 + NO3 ) and atrazine determined in a previous 

study at the same station (Coupe and Johnson, 1991) 

during the period April through August 1990 also is 

shown in table 1. The previous study was similar to 

this one, except that samples were analyzed for 

NO2 + NO3 instead of only nitrate and most of the 

atrazine concentrations were determined by an 

immunoassay method (Thurman and others, 1991). 

Only a few samples were analyzed by GC/MS to 

determine the concentration of atrazine.

Ambient stream-water-quality data show that 

the contribution of nitrite to the NO2 + NO3 concentration in the Sangamon River usually is very small 

less than 1 percent (Richards and others, 1992). Therefore, the mass of NC>2 + NO3 transported in 1990 can 

be compared with the nitrate transport computed for 

1991. The results of the immunoassay analysis provide 

a good indication of the atrazine concentration and 

were used to estimate the atrazine concentrations 

when values determined by GC/MS were not available. The mass transport for the 1990 study was 

estimated using the same method as the 1991 -92 study.

Most of the atrazine for the study period was 

transported between the beginning of April 1991 and 

the end of August 1991 (table 1). More than 95 percent

of the total atrazine load transported by the river at 

Monticello between April 1991 and March 1992 was 

transported during April through August 1991. Thus, 

the data presented in table 1 for the 1990 study probably represent an estimate of most of the total atrazine 

load for that year.

The total mass of NO2 + NO3 and atrazine 

transported during the 5-month study in 1990 

exceeded that for nitrate and atrazine transported during the entire 12-month 1991-92 study. This could be 

because of the high streamflows during the 1990 study. 

In the spring of 1990, the largest monthly mass transport for atrazine was in June; in the spring of 1991, the 

largest monthly mass transport of atrazine was in May. 

The monthly mean discharge for June 1990 was 

1,744 ft3 /s; for June 1991, it was 242 ft3 /s. Nitrate load 

at this site was dependent on streamflow. Regardless 

of season, an increase in streamflow was accompanied 

by an increase in nitrate mass transport. A plot of 

monthly mean discharge against monthly nitrate mass 

transport for data from all 12 months of the study plus 

data from the 5-month previous study is shown in 

figure 3.

The total mass transport of selected herbicides 

and nitrate in the Sangamon River at Monticello, as 

well as the total amount of these compounds applied in 

the drainage basin and the percentage transported in 

the Sangamon River, are shown in table 2. Nitrogenapplication rates were obtained from the U.S. Environmental Protection Agency (USEPA) (1990) and 

estimated herbicide application rates were obtained 

from Gianessi and Puffer (1991). The mass transport 

of nitrate is shown for reference as a percentage of the

474
1.600

y - -7.9277 + 0.89758x 

correlation coefficient - 0.976

400 600 800 1.000 1.200 1,400

DISCHARGE, IN CUBIC FEET PER SECOND

1.600 1.800

Figure 3. Monthly mean discharge and nitrate as nitrogen load, Sangamon 

River at Monticello, Illinois, April-August 1990.

Table 2. Mass transport and amounts of agricultural chemicals 

applied in the Sangamon River Basin above Monticello, Illinois, 

April 199-March 1992 and April-August 1990

[Ibs, pounds, , no data collected]

April 1991-March 1992 April-August 1990

Compound

Atrazine

Cyanazine

Alachlor

Nitrate

Annual

application

11 3.700 Ihs

4 1.900 Ibs

11 2.300 Ibs

12,100 tons

Total load

1.607 Ibs

669 Ibs

353 Ibs

3,393 tons

Percentage

or pesticides

applied

1.4

1.6

.31

28

Percentage

of pesticides

Total load applied

2.256 Ibs 2.0

-

4.040 tons 33

Source: Gianessi and Puffer. 1990

amount applied even though other sources of nitrate, 

both natural and human derived, can contribute to the 

mass transported.

Atrazine transported in the Sangamon River for 

the 1990 and the 1991-92 studies was 2.0 and 1.4 percent, respectively, of the amount applied. Wauchope 

(1978) indicates that losses up to 5 percent can be 

expected for herbicides, depending on the slope of 

the land and the timing of rainfall following their 

application. The Sangamon River is a secondary tributary to the Mississippi River. Goolsby and others 

(1991) calculated that 0.9 percent of the applied atrazine in the Mississippi River Basin flowed into the 

Gulf of Mexico. Although cyanazine was applied in 

smaller amounts than was atrazine or alachlor, more 

cyanazine was transported than the other two herbicides. The amount of alachlor transported was small 

compared with the amounts of atrazine and cyanazine. 

The maximum, minimum, and median concentrations, by month, for each compound are listed in 

table 3. Atrazine and nitrate were detected in every 

sample throughout the study period. Alachlor was 

detected in at least one sample each month throughout 

the study period except in January 1992. Cyanazine

475
Table 3. Summary of agricultural chemical concentrations by month, in the Sangamon 

River at Monticello, Illinois, April 1991-March 1992

[ug/L, micrograms per liter; mg/L, milligrams per liter; <, below detection limit; -, not applicable]

ATRAZINK

Month

April

M.iy

June

July

August

September

October

November

December

January

February

March

Number of

samples

11

KV

22

20

22

6

12

1

2

2

5

5

maximum

6.0

21

13

1.6

1.0

.27

I.I

.91

.39

.IX

.35

.26

(ug/L)

minimum

0.11

.12

.XI

.31

.12

.05

.08

-

.18

.13

.10

.10

median

2.1

7.5

1.3

.XI

.42

.14

.53

.29

.16

.21

.11

CYANAZINE

maximum

2.3

Hi

2.f)

.13

.12

.07

.10

<.05

<.()5

<.05

<.()5

<.()5

Oig/L)

minimum

<0.05

-.05

.07

<.()5

<.05

<.05

<.()5

-

<.05

<.()5

<.05

<.05

median

0.28

2.1

.If)

<.()5

<.05

.05

<.05

..

<.05

<.()5

<.()5

<.05

ALACHLOR

maximum

4.7

8.X

3.4

.37

.21

.79

.21

.14

.06

<.()5

.06

.14

(ug/L)

minimum

<0.05

 -.05

<.()5

<.()5

<.05

<.05

.05

-

<.05

<.05

<.()5

<.05

median

O.X4

.')')

.15

.OX

<.05

<.05

.09

-

<.()5

<.()5

<.()5

<05

NITRATE

maximum

12

n

12

I.X

.72

.41

2.9

8.1

14

12

14

14

(mg/L)

minimum

7.0

1?.

5.4

.07

.04

.21

.54

5.9

12

II

10

12

median

10

ft.O

9.6

1.3

.22

.39

1.7

6.2

14

II

14

13

was detected in at least one sample each month until 

November 1991, after which cyanazine was not 

detected.

The concentrations of atrazine and nitrate and 

the streamflow for April 12 through June 28, 1991, are 

shown in figure 4. Concentrations of atrazine reached 

a peak in mid-May. Atrazine concentrations exceeded 

the USEPA's maximum contaminant level of 3 ug/L 

for 13 days (May 15-27, 1991), and for shorter periods 

throughout April, May, and June 1991. Atrazine concentrations did not exceed 3 ug/L during the rest of the 

study.

Three distinct peaks in atrazine concentration 

occurred on May 9, 16, and 23, 1991 (fig. 4). Each of 

these peaks occurred concurrently with a peak in the 

discharge hydrograph. A discharge peak on April 19, 

1991, exceeded the discharges on May 9, 16, and 23, 

but the atrazine-concentration peak associated with 

this discharge peak was only 4 ug/L. Atrazine concentrations were low during this early spring runoff event 

because only about 9 percent of the corn planting was 

completed by April 21, 1991 (Illinois Agricultural 

Statistics Service, 1991). The percentage of the corn 

planting completed is a good indication of the percentage of atrazine applied, because atrazine is applied to 

about 76 percent of all corn acreage (Pike and others, 

1990), and it usually is applied about the same time as 

planting. These results indicate that the highest concentrations of agricultural chemicals can be expected 

during the first few runoff events following chemical 

application, although, during 1990, high flows 

continued to flush atrazine from the fields in June even 

after planting was complete. The highest daily mean 

concentration of atrazine observed during the study

period was 16.7 ug/L on May 16, 1991. The corresponding daily load for that day was 140 pounds, 

which is the largest calculated daily load.

Nitrate concentrations at the beginning of the 

study were at the USEPA maximum containment level 

of 10 mg/L and remained at or above this concentration, except for brief periods when dilution from runoff lowered the concentrations, until June 11, 1991; at 

this time the nitrate concentration decreased with a 

reduction in streamflow and remained below the MCL 

until December 1991, when the nitrate concentration 

again exceeded 10 mg/L until the end of the study.

SUMMARY

Surface-water samples from the Sangamon 

River at Monticello, 111., were collected from April 12, 

1991, through March 25, 1992, and were analyzed for 

concentrations of selected herbicides and nitrate. Total 

load of each compound was calculated using concentration and streamflow data; these calculated masses 

were compared with masses calculated in an earlier 

study. Transport rates were largest in May 1991 for all 

compounds analyzed. The total mass of each compound transported was compared to estimated application totals in the drainage basin above Monticello. 

During the study period, 1,607 pounds of atrazine 

were transported by the river; this mass represents 

1.4 percent of the total annual application of 113,700 

pounds. The river transported 669 pounds of cyanazine a mass equal to 1.6 percent of the total annual 

application of 41,900 pounds. About 353 pounds of 

alachlor were transported by the river (0.31 percent of 

the 112,300 pounds applied). About 12,100 tons of

476
o s 

uj O

11 (El

U. S. Environmental Protection Agency

Maximum contaminant level

lor drinking water

Z~OC <o

Q£ §1 <°l- 11 '

O ^ 6 -

25

Z 2 -

U. S. Environmental Protection Agency

Maximum contaminant level

lor drinking watar

June 30

Figure 4. Atrazine and nitrate concentrations and stream discharge, April 1991- 

June 1991, Sangamon River at Monticello, Illinois.

nitrogen fertilizer were applied, and 3,393 tons of 

nitrate as nitrogen were transported by the river 

(28 percent of the total amount applied). However, 

other natural and human-derived sources of nitrate 

could have contributed to this total load.

The timing of chemical application and the 

quantity of rainfall and runoff after herbicide application affected their concentrations. Herbicide concentrations peaked in early spring during storms after the 

herbicides were applied. Peak concentrations of atrazine, cyanazine, and alachlor occurred in May; these 

concentrations were 21, 16, and 8.8 ug/L, respectively. 

Nitrate concentrations were at or above the USEPA 

MCL of 10 mg/L from the beginning of the study 

period until June 11, 1991, when nitrate concentrations 

decreased to below the maximum containment level 

until November 1991 because of reduced streamflow.

Nitrate concentrations again exceeded 10 mg/L from 

December 1991 until the end of the study.

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water draining from agricultural fields a review: Journal of Environmental Quality, v. 7, p. 459-472.

478
Temporal Distributions of Herbicide and Nitrite-Nitrate 

Concentrations and Loads in the West Fork of the Big 

Blue River, Nebraska, 1990 and 1991-92

ByA.D. Druliner1 , D.A.Goolsby2, andM. Meyer3

Abstract

The West Fork of the Big Blue River in 

Nebraska drains a relatively flat basin of about 

1,200 square miles in which the dominant land 

use is rowcrop agriculture. From April through 

July 1990 and April 1991 through march 1992 

water samples were collected from the West Fork 

of the Big Blue River and analyzed principally for 

concentrations of selected herbicides, nitritenitrate, and suspended sediment. The purpose of 

the investigation was to determine the occurrence, 

temporal distribution, and persistence of selected 

herbicides and nitrite-nitrate in surface water 

within the West Fork of the Big Blue River Basin 

and to determine the effects of agricultural practices on the concentrations of these constituents.

Atrazine, alachlor, metolachlor, and cyanazine were the herbicides most commonly detected 

and present in the largest concentrations in the 

West Fork of the Big Blue River. Large concentrations of these herbicides were present in the 

river during the first major runoff events after 

spring chemical application in the fields . The 

magnitude of the concentrations diminished with 

successive runoff events. The maximum 

observed concentrations of the four herbicides 

were 116, 43, 26, and 8.6 micrograms per liter, 

respectively. Persistent concentrations of atrazine 

and some of its daughter products in base flow 

indicate that ground water is contributing atrazine

'U.S. Geological Survey, Lincoln, Nebraska. 

U.S. Geological Survey, Denver, Colo. 

3 U.S. Geological Survey, Lawrence, Kans.

to the river. On the basis of estimated herbicide 

application and estimated herbicide loads in the 

river in 1990 and 1991-92, about 0.4 to 1.3 percent of the total amount of herbicides applied are 

transported annually in the West Fork of the Big 

Blue River. The 3- to 4-fold increase in concentrations and loads of selected herbicides from the 

1990 to 1991-92 sampling periods probably 

reflect the timing of chemical application and 

periods of major precipitation.

Nitrite-nitrate concentrations in the West 

Fork of the Big Blue averaged about 2.2 milligrams per liter and unlike herbicides showed limited variation between periods of high and low 

flow. The timing of nitrite-nitrate loads in the 

river indicates that the principal source of nitrogen is probably surface runoff of nitrogen fertilizer. Ground water also may be contributing to 

low-flow concentrations of nitrite-nitrate 

observed in the river. The load of nitrite-nitrate in 

the river is estimated to be about 0.6 percent of 

the nitrogen applied to croplands in the basin 

during each of the two years.

INTRODUCTION

Nebraska is one of many Midwestern States that 

relies on large quantities of chemical fertilizers and 

herbicides to sustain crop production. This is especially true for rowcrop production such as corn, sorghum, and soy beans. In 1987, more than 28 million 

pounds of active herbicidal ingredients (Baker and 

others, 1990) and about 752,000 tons of nitrogen 

chemicals (Nebraska Department of Agriculture, 

1989) were applied to agricultural land in Nebraska.

479
To determine the potential effect of chemical fertilizers 

and herbicides on surface-water quality the U.S. 

Geological Survey, through its Toxic Substances 

Hydrology Program, began a series of regional studies 

in the Midwest in 1989. The first of the studies investigated 149 randomly selected streams throughout the 

Midwest that were sampled three times during 1989 

for the presence of agricultural chemicals (Goolsby 

and others, 1991). The study revealed large concentrations of four herbicides atrazine, alachlor, cyanazine, 

and metolachlor--in storm runoff at most sampling 

sites following the application of herbicides in the 

spring.

The West Fork of the Big Blue River, which is 

located in southeastern Nebraska, was among nine

sites that were selected for detailed study in 1990 and 

1991 (Goolsby and others, 1991). The principal 

objectives of detailed study were to determine the 

occurrence, temporal distributions, and persistence of 

selected herbicides, nitrite-nitrate, and other related 

constituents in the surface water and to determine the 

effects of agricultural practices on concentrations of 

these chemicals in the surface water. This paper summarizes some of those findings for the West Fork of 

the Big Blue River.

DESCRIPTION OF THE STUDY AREA

The basin drained by the West Fork of the Big 

Blue River is located in the southeastern part of

Hydrolooic Unit 10270203/ i

EXPLANATION

Water Quality Sampling Site

0 6 10 15 20 MILES

0 5 10 15 20 KILOMETERS

Figure 1. Location of the West Fork of the Big Blue River, Nebraska.

480
Nebraska (fig. 1), in a region of relatively flat uplands. 

The basin has an area of about 1,200 square miles and 

includes parts of eight counties. The average discharge for the 33 years (1958-present) of record from 

a gaging station just above the rivers' confluence with 

the Big Blue River is 185 ft3/s or 134,000 acre-ft/yr 

(Boohar and others, 1991). Typically the flow is less 

than 100 ft3 /s until late March or April, when periodic 

thunderstorms and surface runoff can increase the discharge to 2,000 to 3,000 ft3/s. The river usually 

returns to low-flow conditions in either August or 

September. Ground-water discharge is the principal 

source of water in the river during low-flow conditions, although irrigation-return flows also can 

contribute to the river during the mid- to late summer.

Thick, loamy soils that formed from loess 

deposits mantle most of the basin. The dominant land 

use in the basin is agriculture. About 75 to 80 percent 

of the basin is planted in corn, sorghum, and soy 

beans. Crop acreages and estimates of nitrogen and 

selected herbicides used in the basin during the 1990 

and 1991 growing seasons are shown in table 1. 

About 75 percent of the cropland in the basin is irrigated. Most of the irrigation water comes from the 

basin's nearly 5,000 registered irrigation wells. A 

limited amount of surface water from the West Fork of 

the Big Blue River is diverted for irrigation.

Table 1. Crop acreage and estimated agricultural-chemical 

use comparison for the West Fork of the Big Blue River 

Basin for the 1990 and 1991 growing season

Corn (acres)

Sorghum (acres)

Soy beans (acres)

Atrazine (pounds) 1

Alachlor (pounds) 1

Metolachlor (pounds) 1

Cyanazine (pounds) 1

Nitrogen (tons)2

1990

402,899

107,896

88,315

511,000

230,000

140,000

124,000

46,384

1991

432,191

96,037

85,234

528,000

243,000

143,000

144,000

47J65

1 Estimates of total herbicide use were made by combining estimates of herbicide use for each major crop type in 1987 for the eastern 

and southeastern crop-reporting areas of Nebraska by Baker and others 

(1990) with estimates of acres planted to each crop type in 1987 

(Nebraska Department of Agriculture, 1989). The herbicide-use figures 

were estimated by multiplying the rates as determined above by the 

number of acres planied in each of the major crops in the basin during 

1990 and 1991.

^ Estimates of nitrogen use were made by multiplying the county 

nitrogen sales records for 1990 and for 1991 for each county in the basin 

(Nebraska Agricultural Statistics Service, 1991, 1992) by the percentage 

of that county included in the basin, and then totaling these values.

The broad-leaf pre-emergent herbicides, such as 

atrazine, alachlor, metolachlor, and cyanazine, are

applied in corn fields at planting time by injection into 

the soil. Post-emergent herbicide applications frequently are made through center-pivot irrigation systems. Most of the nitrogen fertilizer is applied to fields 

cropped in corn as a preplant in March and April and 

is usually in the form of anhydrous ammonia that is 

injected into the soil. Urea-ammonium nitrate solutions also are used later in the spring and summer. 

Typically, rates of nitrogen fertilizer application on 

corn fields range from about 120 to 220 pounds of 

nitrogen per acre.

METHODS

Stream-water samples were collected by a combination of manual sampling and an automatic sampler 

from the West Fork of the Big Blue River from April 

through July 1990 and from March 1991 through April 

1992 at a gaging station just above the confluence with 

the Big Blue River. The automatic sampler pumped 

samples from the river about every 2 to 4 hours when 

the river stage was 4 or more feet above base level. 

Water samples also were collected with the automatic 

sampler several times per week during periods of base 

flow using the sampler's manual mode. Manual samples were collected after late summer 1991. Daily discharge and suspended-sediment data also were 

collected.

During the 1990 sampling, all water samples 

were analyzed for specific conductance, pH, and concentrations of nitrite plus nitrate (nitrite-nitrate), and 

triazine and acetanilide herbicides. The triazine and 

acetanilide herbicides were analyzed using an 

enzyme-linked immunosorbent assay procedure 

(ELIS A) (Pomes and others, 1991). Twenty percent of 

the herbicide samples also were analyzed by gas chromatography/mass spectrometry (GC/MS) using a 

method developed by Thurman and others (1990). 

The GC/MS analyses included atrazine, desethylatrazine, desisopropylatrazine, cyanazine, alachlor, metolachlor, metribuzin, simazine, propazine, prometon, 

prometryn, ametryn, and terbutryn. Atrazine concentrations were estimated for all samples from ELIS A 

data through linear regression with GC/MS atrazine 

concentrations. Water samples collected in 1991 and 

1992 were analyzed for the same constituents as in 

1990; however, all samples were analyzed for herbicides by GC/MS. Samples also were analyzed for 

concentrations of dissolved nitrite-nitrate, ammonia, 

and orthophosphate.

481
RESULTS AND DISCUSSION

All of the herbicides analyzed, with the exception of terbutryn and prometryn, were present in water 

samples collected during high-flow conditions in the 

late spring and early summer for both the 1990 and 

1991-92 sampling periods. The herbicides detected 

the most frequently and in largest concentrations were 

atrazine, alachlor, metolachlor, and cyanazine with 

peak concentrations of 116, 43, 26, and 8.6 (ig/L, 

respectively.

The changes in atrazine concentrations over 

time (fig 2) generally were similar to those observed 

for the other herbicides. The herbicide concentrations 

typically were near or below the applicable analytical 

detection limit prior to chemical application (usually 

mid-April to early May). Herbicide concentrations 

tended to peak almost simultaneously with peaks in 

stream discharge during the first major runoff event 

after chemical application. The height of the herbicide-concentration peaks diminished with each successive runoff event as the source of available 

herbicides decreased during the growing season. By 

late summer to early fall, concentrations of most of 

the herbicides decreased to below the detection limit. 

In the case of atrazine, small concentrations (0.1 to

0.5 u.g/L) persisted during base flow, suggesting that 

ground-water discharge to the stream is responsible 

for these small atrazine concentrations.

Fluctuations in nitrite-nitrate concentrations 

were minor (<0.1 to 3.7 mg/L, as nitrogen) in response 

to changes in river discharge compared to the large 

fluctuations in herbicide concentrations that were 

observed. Nitrite-nitrate concentrations increased during periods of increased stream discharge associated 

with spring and summer runoff events (fig. 3). Nitritenitrate dilution during periods of very large stream discharge is probably responsible for the difference in 

times between some of the nitrite-nitrate and discharge 

peaks. The change in nitrite-nitrate concentrations 

over time suggests that much of the large nitrogen 

concentrations observed in the spring and early summer result from the transport of nitrogen fertilizer to 

the river in surface runoff. The late-summer decline in 

nitrite-nitrate is probably a combination of biological 

consumption of nutrients in the stream and diminished 

runoff and the amount of transported fertilizer available for transport to the river. The winter increase in 

nitrite-nitrate concentrations in the river during the 

second year of data collection probably reflects a combination of the discharge of ground water containing

1.000 7,000

Apr May

Figure 2. Temporal distribution of atrazine concentration and stream flow in the West Fork of the Big Blue River near 

Dorchester, Nebraska, March 1991 - March 1992.

482
6,500

- 6,000

Figure 3. Temporal distribution of nitrite-nitrate concentration and stream flow in the West Fork of the Big Blue River near 

Dorchester, Nebraska, March 1991-March 1992.

low concentrations of nitrite-nitrate into the river and 

reduced biological consumption of nutrients.

The temporal distribution of dissolved orthophosphate concentrations in water samples collected 

during the 1991 -92 sampling period resembles a sine 

curve. Peak orthophosphateconcentrations of about 

0.4 mg/L (as phosphorous) occur in the late spring 

and early summer, followed by a decrease to about 

0.1 mg/L during the fall, and a late increase to over 

0.6 mg/L during the late winter. The peaks in orthophosphate concentration coincide with periods of 

phosphorous-fertilizer application in the basin. The 

large orthophosphate peak in the late winter is most 

like the result of surface application of the fertilizer in 

the fall and winter rather than its incorporation in the 

soil as is done in the spring application.

Herbicide concentrations and, most notably, 

atrazine concentrations in the river were associated 

closely with suspended-sediment concentrations. 

Correlation coefficients for the log of atrazine concentration and for total herbicide concentrations, which is

a sum of all herbicide concentrations for each sample, 

with a log of the suspended-sediment concentration 

were 0.675 and 0.731, respectively, for the period from 

March 1991 trough April 1992. These results, coupled 

with the relatively moderate solubility of these herbicides in water (Weed Science Society of America, 

1983), indicate that the herbicides are not strongly 

sorbed to the sediment and that they readily partition 

into the water during runoff events, thereby allowing 

both peaks in suspended sediment and dissolved herbicide concentrations to occur nearly simultaneously. 

Somewhat similar results were observed by Baker and 

Richards(1990) in Ohio rivers.

Comparison of the loads of atrazine and nitritenitrate during two 4-month periods (April through 

July) in 1990 and 1991 reveals a large increase in 

atrazine load (about 200 percent) in 1991 and a slight 

decrease in nitrite-nitrate load (about 9 percent) 

(table 2). During the same time period, the suspendedsediment load increased about 4 percent and the total 

stream discharge decreased about 10 percent.

483
Table 2. Comparison of selected agricultural chemical and 

sediment loads, stream discharge, and precipitation data for 

the 1990 and 1991 sampling periods

[ND, not determined]

Parameters

Atrazine (pounds 

Alachlor (pounds) 

Metolachlor (pounds) 

Cyanazine (pounds)

Nitrite-nitrate as nitrogen (tons)

Suspended sediment (tons)

Stream discharge (acre-feet)

Average basin precipitation 

(inches)

1990 

(April-July)

2,250' 

ND 

ND 

ND

298

232,000

106,320

15.52

1991 

(April-July)

6,9 10Z 

1,6702 

1.6602

3332

267

241,000

95,170

14.10

Herbicide concentrations estimated through regression relation between ELISA data and GC/MS data.

2Herbicide concentrations determined by GC/MS.

Much of the difference in loads between the two 

sampling years can be explained by the relative timing 

of chemical application and major runoff events in the 

basin rather than actual amounts of precipitation 

received. In 1990, 15.5 inches of precipitation fell, but 

only modest runoff events occurred until mid-July, 

most of which generated stream discharges of only a 

few hundred ft /s. During one runoff event, the discharge exceeded 1,000 ft /s but the large discharge 

lasted only 2 days. Most of the high flow (>1,000 ft3/s) 

that year occurred during the last 2 weeks of July 

when reduced amounts of herbicides were available 

for transport to the river. This resulted in only moderate concentrations and loads of atrazine and probably 

other herbicides being removed from the basin by 

streams in 1990. By contrast, all the runoff events in 

1991 that generated stream discharges of 1,000 ft3/s or 

more occurred in late May and early June 1 to 1 1/2 

months earlier than in 1990. Consequently, large 

amounts of herbicides were available for transport to 

the river. The slight increase in suspended sediment 

load in 1991 also mirrors runoff from the early precipitation events when little ground cover crop was available to reduce soil erosion and surface runoff.

A comparison of the estimated amounts of 

selected herbicides and nitrogen applied to crop land 

in the basin (table 1) with removal of the loads from 

the basin in the West Fork of the Big Blue River, 

shows that relatively small amounts of parent chemicals are being removed by surface runoff. Only about 

0.4 and 1.3 percent of the applied atrazine in the basin 

for 1990 and 1991, respectively, discharged to the 

river in runoff. About 0.6 percent of the estimated

nitrogen applied within the basin during both years 

was accounted for in the streamflow. The parent compounds of alachlor, metolachlor, and cyanazine measured in the water samples collected during 1991 

accounted for about 0.7, 1.1, and 0.2 percent, respectively, of total amounts of these chemicals applied in 

the basin. It should be noted, however, that significant 

amounts of herbicide metabolites, most of which were 

not analyzed for in this study, were most likely present 

in the stream (Baker and others, 1993), and that their 

consideration in the mass balance of herbicides would 

be expected to increase significantly the fraction of 

applied herbicides accounted for in the streamflow.

Analyses for desethylatrazine and desisopropylatrazine, two atrazine metabolites, were performed on 

water samples collected during the 1991 to 1992 sampling period. The concentrations of both metabolites 

increased during periods of increased discharge from 

mid-May through early June and tended to mirror atrazine concentrations observed during this time period, 

but at concentrations 3 to 50 times lower than those of 

atrazine. Desisopropylatrazine, which has a relatively 

short half-life, was detected primarily when peak atrazine concentrations were observed and was not 

detected after the end of September. Desethylatrazine, 

however, was detected throughout the sampling period 

and appeared to maintain a concentration of about 

0.10 fj,g/L during periods of base flow. The desethylatrazine: atrazine ratio (DAR) in streamflow maintained 

an almost uniform value of about 0.36 except during 

the period of increased discharge and increased atrazine concentrations during May through early June 

when the median DAR value was 0.10. The persistence of low concentrations of desethyl atrazine and 

large DAR's during periods of base flow suggest that 

ground water is contributing atrazine to the river in 

base flow. This conclusion is consistent with those of 

Thurman and others (1992) who found similar relations between low flow and desethylatrazine concentrations in other small midwestern river basins.

REFERENCES

Baker, M, Peterson, Nancy, an Shripat, K.T., 1990, Pesticide use on crops in Nebraska 1987: Institute of 

Agricultural and Natural Resources, University of 

Nebraska-Lincoln, Research Bulletin 311, 34 p.

Baker, D.B., and Richards, R.P., 1990, Transport of soluble 

pesticides through drainage networks in large agricultural river basins, in Kurtz, D.A., ed., Long range 

transport of pesticides, Lewis Publishers, Chelsea, 

Michigan, p. 240-241.

484
Apr May

6.500

- 6,000

Figure 4.

the West

Temporal distribution of desethylatrazine concentration, desethylatrazine to atrazine ratio (DAR) and stream flow in 

Fork of the Big Blue River near Dorchester, Nebraska, March 1991-March 1992.

Baker, D.B., Bushway, R.J., Adams, S.A., Macomber, C., 

1993, Immunoassay screens for alachlor in rural wells: 

false positives and an alachlor soil metabolite: Environmental Science and Technology, v. 27, no. 3, p. 

562-564.

Boohar, J.A., Hoy, C.G., and Steele, G.V., 1991, WaterResources Data Nebraska, Water Year 1990: U.S. 

Geological Survey Water-Data Report NE-90-1, 325 p.

Goolsby, D.A., Thurman, E.M., and Kolpin, D.W., 1991, 

Geographic and temporal distribution of herbicides in 

surface waters of the upper Midwestern United States, 

1989-90, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. Geological Survey Water-Resources Investigations Report 

91-4034, p. 183-188.

Nebraska Department of Agriculture, 1989, 1988 Nebraska 

agricultural statistics: Nebraska Agricultural Statistics 

Service, 126 p.

 - 1991, 1990-91 Nebraska agricultural statistics, 

Nebraska Agricultural Statistics Service, 148 p.

 1992, 1991-92 Nebraska agricultural statistics,

Nebraska Agricultural Statistics Service, 146 p.

Pomes, M.L., Thurman, E.M., Goolsby, D.A., 1991, Comparison of microtitre-plate, enzyme-linked immunosorbent assay (ELISA) and gas chromatography/mass 

spectrometry (GC/MS) for analysis of herbicides in 

storm runoff samples, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 572-575.

Thurman, E.M., Meyer, Michael, Pomes, Michael, Perry, 

C.A., and Schwab, A.P., 1990, Enzyme-linked immunosorbent assay compared with gas chromatography/ 

mass-spectrometry for the determination of triazine 

herbicides in water: Analytical Chemistry, v. 62, 

no. 18, p. 2043-2048.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., Mills, M.S., 

Pomes, M.L., and Kolpin, D.W., 1992, A reconnaissance study of herbicides and their metabolites in surface water of the Midwestern United States using 

immunoassay and gas chromatography/mass spectrometry: Environmental Science and Technology, v. 26, 

no. 12, p. 2440-2447.

Weed Science Society of America, 1983, Herbicide handbook (5th ed.): Champaign, Illinois, 515 p.

485
486
Persistence of Herbicides in Selected Reservoirs in the 

Midwestern United States: Some Preliminary Results

By Donald A. Goolsby 1 , William A. Battaglin 1 , James D. Fallen2 , Diana S. Aga2 , 

Dana W. Kolpin3 , and E. Michael Thurman2

Abstract

Preliminary results from a study of herbicides in 76 midwestern reservoirs show that some 

herbicides and metabolites of atrazine and 

alachlor are detected more frequently throughout 

the year in reservoirs than in streams. Except for a 

short period after application to cropland, herbicide concentrations also are generally higher in 

reservoirs than in streams. Herbicides or their 

metabolites were detected in 82 to 92 percent of 

the reservoirs sampled during four periods from 

late April through early November 1992. Atrazine 

was detected most frequently and in highest concentrations, followed by an alachlor metabolite 

(alachlor ethanesulfonic acid), and two atrazine 

metabolites (desethylatrazine and deisopropylatrazine). The longer persistence of some herbicides and metabolites in reservoirs than streams is 

attributed to longer half lives for these compounds in the water column than in the soil where 

concentrations of organic matter and microorganisms are much higher and contribute to 

rapid biodegradation of herbicides. A second 

contributing factor is long-term storage of water 

in reservoirs that originates as spring and summer 

storm runoff from cropland and which contains 

high concentrations of herbicides.

INTRODUCTION

Reservoirs are an important part of hydrologic 

systems in the Midwestern United States. According

U.S. Geological Survey, Lakewood, Colorado 

U.S. Geological Survey, Lawrence, Kansas 

3U.S. Geological Survey, Iowa City, Iowa

to data compiled by Ruddy and others (1990), about 

440 large reservoirs (normal storage capacity greater 

than 5,000 acre-feet) in 11 upper Midwestern States 

discharge streamflow to the Mississippi River by way 

of tributaries. The primary function of these reservoirs 

is to impound surface water for many uses, including 

flood control, hydropower, recreation, and aquatic life 

habitat. These large reservoirs and numerous smaller 

reservoirs also can serve as sources of drinking water 

for public supplies. In addition to storing surface 

water, reservoirs also can store undesirable substances 

such as sediment and toxic chemicals including pesticides. Most of the sediment entering reservoirs is permanently trapped and deposited on the bottom of the 

reservoir. However, chemicals such as soluble herbicides generally remain in the water column and are 

stored only temporarily until they are flushed from the 

reservoir or removed from solution by biotic and 

abiotic processes.

Storage of herbicides is a potential problem in 

reservoirs that receive drainage from agricultural areas 

in the upper Midwest. Recent studies by the U.S. Geological Survey (USGS) (Thurman and others, 1991, 

1992; Goolsby and others, 1991) have shown that 

most streams in the upper Midwest contain herbicides 

at some time during the year. Large quantities of herbicides are flushed from agricultural fields each spring 

and summer during rainfall following application of 

herbicides. Median concentrations of the herbicides 

atrazine, alachlor, cyanazine, and metolachlor in 

streams increased by at least an order of magnitude 

from March and April 1989 to May and June 1989. 

For example, the median concentrations of herbicides 

in Midwestern streams, in 1989, ranged from less than 

0.3 ng/L before planting to as much as 3 ^ig/L after 

planting, and the maximum concentrations in a few 

streams reached 100 jig/L (Thurman and others, 

1991). During late spring and early summer, concentrations of atrazine can exceed the U.S. Environmental

487
Protection Agency's maximum contaminant level 

(MCL) for drinking water of 3 jj.g/L for several weeks 

to several months in both small streams and large 

rivers, such as the Mississippi River.

Because reservoirs collect and store water, they 

can be affected by storm runoff that contains large 

concentrations of herbicides for a much longer period 

of time than the streams that supply the reservoirs 

(Stamer and Zelt, 1992). This can substantially affect 

the water quality of streams downstream from reservoirs. The length of time that reservoirs discharge 

water with elevated concentrations of herbicides 

depends on a number of factors including residence 

time of water in the reservoir, timing of inflow to the 

reservoir, land use and herbicide use in the contributing drainage area, and the timing and intensity of rainfall. Unregulated streams exhibit the flush effect 

(Thurman and others, 1991) which can produce high 

concentrations of herbicides for short periods of time. 

In contrast, peak concentrations of herbicides in 

streams regulated by reservoirs are much lower, but 

elevated concentrations (near or above MCLs) can 

persist for much longer periods of time. Stamer and 

Zelt (1992) have shown that atrazine concentrations in 

Perry Lake, Kans., remained near or above the MCL 

of 3 ng/L from March 1989 through October 1989 and 

above 1 ng/L through February 1990, whereas atrazine 

concentrations in the principal tributary to Perry Lake 

exceeded the MCL for only a few months in late 

spring. However, atrazine concentrations in some of 

the samples from tributaries exceeded 10 ng/L during 

this period and much of the water in Perry Lake was 

replaced with containing these atrazine concentrations. 

Because little additional inflow to Perry Lake occurred 

after early summer, this "herbicide rich" water was 

stored in the reservoir until the next spring, when the 

cycle was repeated.

The process of storage and attenuation of herbicides documented in Perry Lake likely occurs in most 

other Midwestern reservoirs to greater or lesser degree 

depending on physical and hydrologic characteristics 

of the reservoirs and land use in the reservoir drainage 

basin. Atrazine concentrations in mid-winter samples 

during 1990-92 from several large reservoirs in Illinois, Iowa, Kansas, and Missouri are listed in table 1. 

Water samples from many of these reservoirs had 

atrazine concentrations of at least 2 ng/L during this 

time, which likely reflected the storage of herbicideladen inflow originating from the "spring flush." In 

general, it appears that the larger the volume-todrainage area ratio of the reservoir, the greater the 

atrazine concentration.

The way in which unregulated streams in the 

Midwest respond to the seasonal application of herbicides has been documented (Thurman and others, 

1992; Goolsby and others, 1991). However, the temporal distribution of herbicides in streams regulated by 

reservoirs has not been examined. In addition, the process(es) by which physical, hydrologic, and land use 

characteristics of reservoirs and their contributing 

drainage areas interact to affect herbicide concentrations in the outflows from reservoirs has not been 

studied. To develop an understanding of these processes in reservoirs, a study was begun as part of the 

USGS's Toxic Substances Hydrology Program in 

April 1992. The study is still in progress at the present 

time (May 1993). The purpose of this paper is to 

describe the study plan and to summarize some of the 

preliminary study results.

STUDY AREA AND PLAN OF 

INVESTIGATION

The study area (fig. 1) was defined as all hydrologic units in parts of 11 states (Illinois, Indiana, Iowa, 

Kansas, Minnesota, Missouri, Nebraska, North 

Dakota, Ohio, South Dakota, and Wisconsin) that 

drain to the Ohio, Upper Mississippi, and Lower 

Missouri Rivers. This area comprises about 450,000 mi2 

and is virtually the same area covered by the 1989-90 

reconnaissance for herbicides in streams (Thurman 

and others, 1991; 1992).

The primary objectives of the study are to 

(1) determine the occurrence and temporal distribution 

of selected herbicides and herbicide metabolites in the 

outflow from selected reservoirs in the upper Midwest, 

and (2) determine if the persistence of large concentrations (greater than about 1 ng/L) of herbicides in reservoir outflow can be quantified on the basis of reservoir 

and drainage-basin characteristics, hydrology, land use, 

herbicide use, and climate. Some specific hypotheses to 

be tested are 

1. Herbicides will be detected in Midwestern 

reservoirs for a longer period of time than in 

unregulated streams, but peak concentrations will 

be lower in the reservoirs than in these streams.

2. The duration of herbicide concentrations in

reservoir outflow above a threshold value can be 

explained (statistical model) by reservoir and 

drainage-basin characteristics, land use, herbicide 

use, rainfall (intensity, timing and amount), and

488
Table 1. Atrazine concentrations in water samples from selected midwestern reservoirs during 

winter months, 1990-92

[Vol/DA, volume to drainage area ratio; acre-ft, acre-feet; GC, gas chromatography; ELISA, immunoassay; U.g/L, micrograms per liter, -, no data]

Atrazine Atrazine 

concentration concentration

Reservoir

Sample 

date

Vol/DA 

(acre-ft/acre)

byGC 

(ug/L)

by ELISA 

(ug/L)

Carlyle Lake outflow 1 - 3-92

Lake Decatur outflow 1- 8-92

Rend Lake Spillway 1 - 2-92

Lake ShelbyviUe outflow 1- 8-92

Lake Springfield at Sugar Creek 1 -30-92

Lake Springfield at 1-30-92 

Spaulding Dam

Iowa

Coralville Lake 2-21-92

Gory don Reservoir winter, 1992

Rathbun Reservoir -12-90

Rathbun Reservoir 2-20-92

Red Rock Reservoir 2-12-92

Saylorville Lake 2-12-92

Kansas

Perry Lake 2- 3-91

Missouri

0.14 

.05

.59 

.31 

.34 

.34

0.01

.58 

.58 

.01 

.02

.34

3.7

3.9

2.3

.2

.6

1.1

2.5

4.0

.2 

10

2.8 

.2 

.1

Long Branch Reservoir 

Smithville Reservoir

--12-90 

--12-90

.50 

1.06

2.0 

3.6

stability (half-lives) of individual herbicides. 

Consequently, the probability that a herbicide such 

as atrazine will persist in a reservoir all year above 

a specified concentration can be predicted.

The occurrence and concentrations of herbicides 

and herbicide metabolites in the anoxic 

hypolimnia of reservoirs during summer 

stratification differ from those in the aerobic 

epilimnia.

Reservoirs for study were selected from the reservoir data base compiled by Ruddy and others (1990). 

The principal criterion for selection of reservoirs was 

that data on reservoir volume and reservoir discharge 

must be obtainable so that the residence time of water 

in the reservoir and the timing of outflow can be determined. In addition, the reservoir outflow must be 

accessible for sampling. The reservoir data base was 

screened to determine which reservoirs met these

489
100

North Dakota

O I

45^

South Dakota

Nebraska

v

,,. \ Wisconsin Minnesota - x

r

40

K_*** \

r' m. . i Indiana ', Ohio

> Illinois A ^tt |

^F^ i ^

\

Kansas | 

EXPLANATION

 LOCATION OF RESERVOIR 

SELECTED FOR STUDY

300 MILES

» v

r\'-^'*>/

Missouri

STUDY AREA

/

i 

300 KILOMETERS

Figure 1 . Location of study area and selected reservoirs.

criteria. As a result of this screening, 74 of the 440 reservoirs in the reservoir data base were selected for 

sampling. Two additional reservoirs, Lakes Monona 

and Waubesa in Wisconsin, which form a chain of 

reservoirs receiving outflow from Lake Mendota, also

were selected. These three reservoirs will be treated as 

a single unit. Locations of these 76 reservoirs are 

shown in figure 1.

The outflow from each reservoir was sampled 

six times (approximately bimonthly) from April 1992

490
through March 1993, and a seventh time in midsummer 1993. Samples are collected near the centroid of flow or other outflow point by methods that 

provide a representative sample of dissolved herbicides and nutrients in the outflow from the reservoir. 

During August 1992, herbicide samples, dissolved 

oxygen profiles, and temperature profiles were collected near the deepest point in 19 selected reservoirs 

to examine the effect, if any, of chemical stratification 

on herbicide concentrations. Herbicide samples were 

collected near the surface and near the bottom of each 

reservoir.

All samples are analyzed for 11 herbicides 

(alachlor, atrazine, ametryn, cyanazine, metolachlor, 

metribuzin, propazine, prometon, prometryn, 

simazine, and terbutryn) and at least 5 herbicide 

metabolites (desethylatrazine, desisopropylatrazine, 

deethylcyanazine, cyanazine amide, and deethylcyanazine amide) by gas chromatography/mass spectrometry (GC/MS). A metabolite of alachlor, [(2,6- 

diethylphenyl)(methoxymethyl) amino-2-oxoethane 

sulfonic acid], (ESA) is analyzed by immunoassay 

following isolation on Cjg cartridges (Aga, 1994). 

Selected ESA samples are confirmed by high-performance liquid chromatography. Samples are also analyzed for nitrite, nitrate, ammonia, orthophosphate, 

and silica.

Ancillary data including land use, herbicide use, 

rainfall, and reservoir characteristics, are obtained 

from the following sources and stored in a geographic 

information system (GIS):

Data Type 

Land Use

Herbicide Use

Rainfall

Reservoir characteristics

Source

1987 Census of Agriculture 

data.

Gianessi and Puffer, 1990. 

National Weather Service.

U.S. Geological Survey and 

Corps of Engineers data 

bases.

PRELIMINARY RESULTS AND 

DISCUSSION

Analytical results from samples collected during 

the first four sampling periods, during April through 

November 1992 indicate that a number of herbicides 

and/or their metabolites, are present in many Midwestern reservoirs for long periods of time. The four 

sampling periods include pre-planting (late April-early

May), post-planting (late June-early July), late summer (late August-mid-September), and fall (mid 

October-early November). Herbicides were detected 

in 82 to 92 percent of the 76 reservoirs during all four 

sampling periods. Four compounds (atrazine, desethylatrazine, desisopropylatrazine, and metolachlor) were 

detected in more than half the reservoirs during the fall 

(October-November sampling; table 2), whereas only 

atrazine was detected in more than one half the 

streams sampled in the fall of 1989 (table 2). One of 

the most notable differences between the occurrence 

of herbicides in reservoirs and streams is the much 

higher frequency of detection of cyanazine and desisopropylatrazine in reservoirs. A possible explanation 

(hypothesis) for this observation is that these two compounds are much more stable in the water column of 

lake and streams than in soil, where organic matter and 

microorganisms promote rapid biodegradation. Consequently, late spring and summer runoff can flush large 

amounts of these two compounds into reservoirs, 

where they can persist in the water column for long 

periods of time. Neither cyanazine nor desisopropylatrazine was detected in streams during the fall (table 2) 

because these compounds are no longer present in significant amounts on the agricultural fields where they 

were applied. This hypothesis points to the need for 

data on the half-lives of herbicides and insecticides in 

water. Virtually all available data on the half lives of 

herbicides are for soils. Water-column-half lives are 

particularly important with regard to the persistence of 

herbicides in reservoirs, lakes, and estuarine systems.

The spatial distribution of the detections of 

herbicides and metabolites in the 76 reservoirs is 

shown in figures 2 and 3 for the four sampling periods. 

These figures also show which reservoirs contained 

herbicides in concentrations that exceeded MCL's and/ 

or health advisories (HA's) for drinking water. MCL's 

apply to average annual concentrations and are legally 

enforceable under the 1986 Safe Drinking Water Act, 

whereas HA's are not enforceable. Exceedence of 

MCL's or HA's is of concern because many Midwestern reservoirs are used for public water supply. 

Concentrations of one or more herbicides exceeded 

MCL's or HA's in 8 reservoirs during the first sampling period, in 16 reservoirs during the second sampling period (fig. 2), in 7 reservoirs in the third 

sampling period, and in 2 reservoirs during the fourth 

sampling period (fig. 3). More herbicides and metabolites were detected in reservoirs in areas where use of

491
Table 2. Herbicides analyzed and percent detections in Midwestern reservoirs during 1992, and in 

Midwestern streams during 1989

[|ig/L, micrograms per liter, ES A, ethanesulfonic acid metabolite of alachlor, -, no data <, less than; N, number of samples]

Percentage of detections greater than reporting limit:

76 Midwestern reservoirs in 1992 Midwestern streams in 1989

Reporting 

Herbicide limit 

(M/L)

alachlor

ametryn

atrazine

cyanazine

desethylatrazine

desisopropylatrazine

metolachlor

metribuzin

prometon

propazine

ESA

0.05

.05

.05

.05

! 0.2

.05

.05

.05

.05

.05

.05

.1

late Aprilmid-May

36

0

72

49

25

63

58

46

12

5

2

72

late Juneearly July

48

1

92

65

40

78

70

62

9

14

10

79

late August- late October- pre- post- fall 

early September early November application application low-flow 

(N=55) (N=132) (N=145)

26

3

86

56

33

74

63

52

5

15

5

77

16

1

80

46

26

70

62

51

0

14

1

64

18

0

91

-

5

54

9

34

2

0

0

~

86

0

98

-

63

86

54

83

53

23

40

~

12

0

76

-

0

47

0

44

0

6

<1

~

Reporting limit for Midwestern streams in 1989 was 0.2 }ig/L. Percent detections for both reporting limits are given for data 

from 76 Midwestern reservoirs.

herbicides is most intense-that is, the area from eastern Kansas and Nebraska to Ohio (figs. 2 and 3).

One of the most significant findings from this 

study to date is the abundance and persistence of 

herbicide metabolites in reservoirs. Data are presently 

available from this study on two metabolites of atrazine (desethylatrazine and desisopropylatrazine) and 

one metabolite of alachlor, ethanesulfonic acid. The 

occurrence of atrazine metabolites in streams and their 

use as indicators of surface-water/ground-water interaction has been reported previously by Thurman and 

others (1991, 1992). The presence of ESA in ground 

water was recently reported by Baker and others 

(1993) and Kolpin and others (1993). However, the 

present reservoir study is believed to be the first systematic effort to investigate ESA in surface water. The 

frequency of detection was greatest for atrazine, 

followed by three metabolites, ESA, desethylatrazine, 

and desisopropylatrazine in the 76 reservoirs during 

the four sampling periods (table 2). The overall 

median concentrations of these four compounds 

followed the same order. Cyanazine, metolachlor, and 

alachlor were fifth, sixth, and seventh, respectively 

with respect to frequency of detection and median 

concentration. Previous studies have shown that the 

herbicide, alachlor is not very persistent in streams 

(Thurman and others, 1991; 1992; Goolsby and others, 

1991) or in ground water (Kolpin and others, 1993). 

However, this does not appear to be the case for one of

492
Late April 

through 

mid May

O ; -L -,

0 r X °

so *-, ,

' }1> ^Q0o o M

\

. - - - -\ ®

Late June

L through

early July

300 MILES

_I

300 KILOMETERS

EXPLANATION

NUMBER OF HERBICIDES OR 

HERBICIDE METABOLITES DETECTED 

(14 DETECTIONS POSSIBLE)

o o 

O ! . 4

© 5-8

© MORE THAN 8

SOLID BLACK SYMBOL 

INDICATES CONCENTRATIONS OF 

ONE OR MORE HERBICIDES 

EXCEEDED HEALTH-BASED LIMIT

Figure 2. Generalized distribution of herbicide detections in Midwestern reservoirs and reservoirs in which concentrations of 

one or more herbicides exceeded a U.S. Environmenal Protection Agency maximum contaminant level or health-advisory 

level for drinking water during late April through mid May and late June through early July, 1992.

493
Late August

through

early September

% »: >,:y / ° «,,--

T

\ © © ^ © © ©

Late October 

>° o ^/- through

/ early November

y-'-ri © c$£ i

- © ;

©

O

"i

©^, o© ;

© ^' EXPLANATION

NUMBER OF HERBICIDES OR 

HERBICIDE METABOLITES DETECTED 

(14 DETECTIONS POSSIBLE)

300 MILES

_I i i i

300 KILOMETERS

O 1-4

© 5-8

© MORE THAN 8

SOLID BLACK SYMBOL 

INDICATES CONCENTRATIONS OF 

ONE OR MORE HERBICIDES 

EXCEEDED HEALTH-BASED LIMIT

Figure 3. Generalized distribution of herbicide detections in Midwestern reservoirs and reservoirs in which concentrations of 

one or more herbicides exceeded a U.S. Environmenal Protection Agency maximum contaminant level or health-advisory level 

for drinking water during late August through early September and late October through early November, 1992.

494
its metabolites, ESA, which apparently is both mobile 

and relatively persistent (stable) in surface water.

The temporal distribution of atrazine, alachlor, 

and three metabolites during the four reservoir 

sampling periods is shown in figure 4. Also shown for 

comparison purposes is the temporal distribution of 

these compounds (except ESA) during the 1989 reconnaissance of Midwestern streams (Thurman and others 

1991; 1992). These results indicate that concentrations of atrazine and its metabolites in streams shortly 

after herbicide application are higher than in reservoirs. However, at other times of the year, concentrations are somewhat higher in the reservoirs, 

particularly concentrations of the two metabolites of 

atrazine. Desisopropylatrazine was detected infrequently in streams prior to application and not at all in 

the fall of the year (fig. 4). In contrast, this metabolite 

of atrazine was detected in 58 to 70 percent of all samples collected during the four reservoir-sampling 

periods. As hypothesized previously, the reason for 

this large difference in frequency of detection is

probably the short half life for desisopropylatrazine in 

soil combined with its much longer half life in the 

water column, and long-term storage of this compound 

in the water mass within reservoirs.

With regard to alachlor there appears to be little 

difference between concentrations in streams and 

reservoirs except shortly after application when concentrations in streams are higher (fig. 4). Alachlor 

disappears rather quickly in streamflow and in reservoirs, consistent with data reported for Perry Lake in 

Kansas (Stamer and others, 1993). ESA appears to be 

a major soil metabolite of alachlor (Baker and others, 

1993); however, it is not known whether significant 

degradation of alachlor to ESA occurs in the water 

column of streams and reservoirs. The ESA concentrations in reservoirs were similar during all four 

sampling periods (fig. 4).

Data on herbicide and nutrient concentrations, 

reservoir inflow and outflow (residence time), rainfallpatterns, pesticide use and land use will be available 

upon the completion of this study.

495
100.

10.0

CA L°

&J 63J

rt 010

0*

(/}

^^ A fM

-

4

 1

_

J\

106M8/L

1

.

147 Midwestern Streams 76

^< U.U1 i 

O

Q 100.

Ctf

uHH

sz; iao

hH

£*

O

H L0

H

£

Cd 010

U

£

O

u

_

Psj o.oi 

t-K

O j 100'

CQ

^

b3S 10.0

0

9 L0

0

H 010

K

001

i

1989

atrazine

-

i

-

Midwestern Reservoirs - 1992

100.

10.0

1.0

010

nm

1 1

1 alachlor

-

i

-

i 

: j i , i 

- n -

i 1 1 I J i

147 Midwestern Streams 76 Midwestern Reservoirs - 1996

1989

1

desethyiatrazine

(atrazine metabolite)

i

»

i

i . i r

1 U

ilJU.

10.0

LO

010

nm

1 ~

ESA (ethanesulfonic acid metabolite of alachlor)

:

 * i .

 

no samples

 

_ i

Pre Post Fall April- June- Aug.- OcLPlant- Plant- Low- May July Sept Nov.

ing ing Flow

desisopropylatrazine

(atrazine metabolite)

i

J

-

t «

J

i i

__

SAMPLING PERIOD

EXPLANATION

 maximum concentration

I 95th percentile 

-L 75th percentile

 median

-j- 25th percentile

 5th percentile

Pro Port Fafl April- June- Aug.- Oct.- * reoortinz limit 

Plant- Plant- Low- May July Sept Nov. ' or^mmum

mg ing Flow concentration

SAMPLING PERIOD

Figure 4. Temporal distribution of atrazine, alachlor, and three metabolites in midwestern reservoirs during 1992 and in 

midwestern streams during 1989.

496
REFERENCES

Aga, D.S., Thurman, E.M., and Pomes, M.L., 1994, Determination of alachlor and its sulfonic acid metabolite in 

water by solid-phase extraction and enzyme-linked 

immunosorbent assay: Analytical Chemistry, v. 66, 

p. 1495-99.

Baker, D.B., Bushway, R.J., Adams, S.A., and Macomber, 

C., 1993, Immunoassay screens for alachlor in rural 

wells: false positives and an alachlor soil metabolite: 

Environmental Science and Technology, no. 3, 

p. 562-564.

Gianessi, L.P., and Puffer, C.M., 1990, Herbicide use in the 

United States: Washington, D.C., Resources for the 

Future, [Inc.], Quality of the Environment Division, 

127 p. [Revised 1991].

Goolsby, D.A., Coupe, R.H., and Markovchick, 1991, 

Distribution of selected herbicides and nitrate in the 

Mississippi River and its major tributaries, April 

through June 1991: U.S. Geological Survey WaterResources Investigations Report 91-4163, 35 p.

Kolpin, D.W., Burkart, M.R., and Thurman, E.M., 1994, 

Herbicides and nitrate in near-surface aquifers of the 

Mid-Continent United States: U.S. Geological Survey 

Water-Supply Paper 2413, 34 p..

Kolpin, D.W., Goolsby, D.A., Aga, D.S., and Iverson, J.L., 

and Thurman, E.M., 1993, Pesticides in near-surface 

aquifers: Results of the Midcontinental United States 

ground water reconnaissance, 1991-92, in Goolby, 

D.A., and others, eds., Selected papers on agricultural 

chemicals in water resources of the Midcontinental 

United States: U.S. Geological Survey Open-File 

Report 93-418, p. 64-74.

Ruddy, B.C., Hitt, K.J., and Kerie, J., 1990, Summary of 

selected characteristics of large reservoirs in the United 

States and Puerto Rico, 1988: U.S. Geological Survey 

Open-File Report 90-163, 295 p.

Stamer, J.K. and Zelt, R.B., 1992, Distribution of atrazine 

and similar nitrogen-containing herbicides in the lower 

Kansas River basin, U.S. Geological Survey Yearbook 

Fiscal Year 1991, U.S. Government Printing Office, 

Washington, D.C., p. 76-79.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1991, Herbicides in surface water of the midwestern United States: The effect of spring flush: 

Journal of Environmental Science and Technology, 

v. 25, no. 10, p. 1794-1796.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1992, A reconnaissance study of herbicides and 

their metabolites in surface water in the midwestern 

United States using immunoassay and gas chromatography/mass spectrometry: Environmental Science 

and Technology, v. 26, no. 12, p. 2440-2447.

497
498
Determining the Relative Age, Transport,

and Three-Dimensional Distribution 

of Atrazine in a Reservoir Using Immunoassay

By James D. Fallen1 andE.M. Thurman2

Abstract

The age, transport, and distribution of 

atrazine in a reservoir were determined by 

enzyme-linked immunosorbent assay. A pulse of 

stormwater runoff containing atrazine 

concentrations as much as nine times greater than 

the U.S. Environmental Protection Agency 

Maximum Contaminant Level (MCL) for 

drinking water (the MCL for atrazine is based on 

an annual average concentration of 3.0 

micrograms per liter) was monitored as it moved 

through Perry Lake, northeastern Kansas, during 

the 1992 growing season. The drainage basin of 

Perry Lake is the first Pesticide Management Area 

designated by the State of Kansas. The leading 

edge of the pulse marked the boundary and 

mixing zone between atrazine applied in previous 

years and freshly applied atrazine. 

Deethylatrazine-to-atrazine ratios (DAR) further 

defined the relative age of atrazine in the 

reservoir. Runoff entering the reservoir 

immediately after herbicide application was 

identified by its small DAR values (0.095 to 

0.134). Water with increasing DAR values (0.135 

to 0.254) entered the reservoir as the year 

progressed and gradually displaced water with 

smaller DAR values. Four hundred and twenty 

(420) samples from four detailed reservoir 

surveys (pre-application, post-application, 

summer, and autumn) were analyzed by

'U.S. Geological Survey, Mounds View, Minnesota 

2U.S. Geological Survey, Lawrence, Kansas

immunoassay to determine the distribution of 

herbicide concentrations in the reservoir. Also, 

weekly samples were collected from four fixed 

sites located upstream, within, and downstream 

from the reservoir. One hundred (100) of these 

samples were analyzed by gas 

chromatography/mass spectrometry to confirm 

immunoassay results and to determine 

deethylatrazine and deisopropylatrazine 

concentrations. A combination of immunoassay 

and DAR values could prove useful in developing 

reservoir-release strategies to mitigate atrazine 

concentrations in reservoirs and their outflows.

INTRODUCTION

Effects of herbicides on reservoir-water quality are 

not well known. Reservoirs are an important component 

of the supply and management of water resources in the 

midwestem United States. Their uses include drinkingand irrigation-water sources, flood control, recreation, 

and wildlife habitat. Reservoir-water quality is critical to 

many of these uses, especially drinking water. Many 

reservoirs used for drinking water drain basins whose 

primary land use is crop production. Herbicides 

generally are available for transport from agricultural 

areas into downstream reservoirs, and previous studies of 

stream water have shown that unregulated spring runoff 

can contain large concentrations of herbicides after their 

application. Furthermore, herbicide concentrations in 

streams have been shown to exceed the U. S. 

Environmental Protection Agency's MCL for drinking 

water for as long as 2 to 3 months during the cropgrowing season (Goolsby and others, 1991; Thurman and 

others, 1991). Because reservoirs are designed to retain 

water, whether for flood control or water supply, they also

499
may retain herbicides. In the Midwest, water from spring 

runoff often is held to attenuate spring flood peaks and to 

store water for use later in the year. This herbicide-laden 

water is released slowly, increasing herbicide 

concentrations downstream for a longer period of time 

than would normally be expected.

In northeastern Kansas, concern about 

concentrations of atrazine in water from the Delaware 

River, which supplies Perry Lake, prompted 

establishment of the first Pesticide Management Area in 

the State (Carole Jordan, Kansas State Board of 

Agriculture, written commun., March 30, 1992). Perry 

Lake is a Federal reservoir with a storage volume of 211 

x 106 m used for flood control, drinking water, 

recreation, and wildlife habitat. Conservative atrazineappHcation practices were exercised voluntarily in the 

basin with the goal of decreasing atrazine 

concentrations in stream water and in Perry Lake.

This paper briefly describes the transport and 

three-dimensional distribution of atrazine in Perry 

Lake at a more detailed spatial scale than previously 

attempted using enzyme-linked immunosorbent assay 

(ELISA). ELISA was used as an accurate, fast, and 

inexpensive analytical method for determining the 

concentrations of herbicides, such as atrazine. This 

method allows a much larger number of samples to be 

analyzed than would be economically feasible using 

conventional methods, such as gas 

chromatography/mass spectrometry (GC/MS). 

Furthermore, this paper describes a test of the 

hypothesis that, by using GC/MS to analyze a subset 

of the ELISA samples, the deethylatrazine-to-atrazine 

ratio, or DAR, (Thurman and others, 1991), could be 

used to improve definition of the age of atrazine in a 

reservoir.

SAMPLE COLLECTION

Four synoptic surveys were conducted on Perry 

Lake from March 30 to September 22, 1992. Water 

samples were collected to assess the three-dimensional 

distribution of atrazine and two metabolites, 

deethylatrazine and deisopropylatrazine. Synoptic 

surveys were conducted on the following dates: a preapplication survey, March 30 through April 10, 1992; 

a first-flush, post-application survey, June 16-25, 

1992; a summer survey, August 3 and 4, 1992, and a 

fall survey, September 16-22, 1992. The first two 

synoptic surveys were the most detailed, with 200 

water samples collected during each survey from

various depths at 100 sites throughout the reservoir. 

After these two surveys, it was apparent that atrazine 

concentrations varied more along the length of the 

reservoir than across the width. As a result, only about 

40 samples were collected for the third survey from 25 

sites along the north-south axis of the reservoir. No 

significant changes in atrazine concentration with 

depth were detected.

Sampling sites for the first two surveys were 

selected randomly from 500-m2 grid cells (fig. 1 A). 

When the water depth was less than 7.5 in, only one 

point in the water column was sampled. If the depth 

exceeded 7.5 m, two or more discrete samples were 

collected from different depths in the water column. 

Discrete analytical results from each depth within a 

cell were averaged together to produce figures 1 and 2. 

Water samples were collected with a Kemmerer 

sampler at selected intervals in the water column. 

Sample water was poured directly from the Kemmerer 

sampler into baked, 125-mL glass bottles and chilled.

ANALYTICAL METHODS

The immunoassay method used in this study 

was described previously (Thurman and others, 1990; 

Pomes and others, 1991) with minor differences in the 

arrangement of the samples and the rotation speed 

(200 revolutions per minute was used) during 

incubation. Triazine microtitre-plate kits used for the 

immunoassay contained polyclonal antibodies coated 

to the inside of the test-plate cell walls. Triazineherbicide molecules in the water sample competed 

with atrazine-enzyme conjugate molecules for binding 

sites on the antibodies. The enzyme conjugate 

consisted of atrazine molecules covalently bonded to 

horseradish peroxidase (Bushway and others, 1988). 

Theoretically an equal number of antibody sites was 

coated to each cell and an equal number of atrazine 

molecules in the enzyme conjugate was added to each 

cell, so the number of triazine molecules in the water 

sample was the determinate variable. If the number of 

triazine molecules in the water sample was small (that 

is, less than the number of enzyme-conjugate atrazine 

molecules), more enzyme-conjugate molecules would 

be able to bind to antibody receptor sites. If the 

number of triazine herbicide molecules in the water 

sample was large (that is, more than the number of 

enzyme-conjugate atrazine molecules), the molecules 

would preempt the enzyme-conjugate atrazine 

molecules from bonding to antibody-receptor sites.

500
EXPLANATION

MEAN TRIAZINE 

CONCENTRATION^1 

IN MICROGRAMS 

PER UTER

95°30'

 \ 

95°25'

COLUMN-NUMBERS 

14 8 12 / 16 20 24

Index map

Base map from U.S. Geological Survey

digital data. 1:100.000.1983 

Atxrs Equal-Area Conic projection /C4 A/O* o 

Standard parallels 33° and «° ^-=>^ *

A. Pre-application survey, March 30 through April 10,1992.

_____ l____________I

39°15'_ -

4.0 - 4.9 

Scale break 

5.0 - 9.9

95°30' 95°25'

COLUMN NUMBERS 

1 4 8 12/16 20 24 ''

14 8 12 16 20\ 24

39°05' - -

B. First-flush survey, June 16-25,1992.

_________I____________l

;..:..:.. i.i

; 

C. Summer s

g Figure 1. Mean concentrations of triazines from (A) pre-application, (B) first-flush, and (C) summer surveys at Perry Lake, north
All unfiltered samples were analyzed in the 

laboratory for triazine compounds by ELISA. 

Twenty-five percent of the samples were analyzed by 

GC/MS for confirmation of ELISA results and 

determination of atrazine metabolite concentrations. 

Water samples for confirmation of ELISA results and 

the determination of concentrations of deethylatrazine 

and deisopropylatrazine were analyzed by GC/MS by 

the method of Meyer and others (1993). Samples were 

filtered through a 0.7-|iim glass-fiber filter into baked, 

125-mL glass bottles. Analytes were extracted 

robotically from solution with solid-phase extraction 

cartridges. Eluates then were reduced to a volume of 

approximately 100 |iiL and transferred into 200-|iiL 

glass-lined polystyrene vials for GC/MS analysis .

RESULTS

Atrazine concentrations from the pre-application 

survey ranged from less than 1.0 |ng/L at the upstream end 

of Perry Lake to 3.0-5.0 |ng/L at the downstream end (fig. 

1 A). A gradual concentration change was identified 

throughout the length of the reservoir. In contrast, the 

first-flush survey showed the largest atrazine 

concentrations (5.0-25 |ng/L) in the upstream end of the 

reservoir (fig. IB). Atrazine concentrations in the 

remainder of the reservoir varied from 2.0-5.0 |ng/L. The 

summer survey identified smaller atrazine concentrations 

(1.0-3.0 |iig/L) entering the reservoir, and larger 

concentrations of atrazine (3.0-5.0 ^ig/L) in the 

downstream three-fourths of the reservoir (fig. 1C). No 

significant changes in atrazine concentration with depth 

were measured in any of the synoptic surveys. GC/MS 

analyses indicated that other triazines known to cross react 

with the immunoassay antibodies (Thurman and others, 

1990) were not present or were insignificant, confirming 

that all of the triazine concentrations determined by 

immunoassay were attributable to atrazine.

DAR values were used to improve definition of 

the age of the atrazine and to track the water's 

movement through the reservoir. Atrazine does not 

readily metabolize in surface water. A small DAR 

value (less than 0.100) indicated that atrazine was 

flushed from the field immediately after application 

because of the limited time it would have to metabolize 

into deethylatrazine. The longer the atrazine remained 

on the field before being flushed off, the larger the DAR 

values. Thunnan and others (1991) found that surfacewater samples from the midwestern United States had 

DAR values of 0.3 in preplant samples, <0.1 in 

postplant samples, and 0.4 in the harvest samples.

DAR values were found to be inversely 

proportional to atrazine concentrations. The 

respective locations of small and large DAR values 

reversed from the pre-application survey to the postapplication survey. During the pre-application survey, 

DAR values were largest (0.215-0.294) at the 

upstream end of the reservoir (fig. 2 A). This indicated 

that the most recent pulse of water to enter the 

reservoir was the previous year's post-harvest surfacewater runoff. The downstream end of the reservoir 

had the smallest DAR values (0.175-0.214), indicating 

this water had entered the reservoir during the 

previous growing season. During the post-application 

survey, the smallest DAR values (0.095-0.134) were 

found in the upstream half of the reservoir, and the 

largest DAR values (0.135-0.174) were found in the 

downstream half (fig. 2B). The smallest DAR values 

at the upstream end probably were from spring field 

runoff immediately after atrazine application which 

entered the upstream end of the reservoir as a pulse. In 

the summer survey, small DAR values were found in 

the downstream end of the reservoir and the outflow, 

whereas larger DAR values (0.135-0.174) were found 

entering the upper two-thirds of the reservoir (fig. 2C). 

No significant changes in DAR values with depth were 

observed in any of the synoptic surveys.

CONCLUSIONS

ELISA is an efficient and practical method for 

determining atrazine concentrations in reservoirs for 

large numbers of samples. Furthermore, by coupling 

ELISA with GC/MS on a small subset of samples, 

analytical results can be confirmed and relations 

among parent and metabolite compounds may be more 

readily explored. In this study, detailed spatial 

resolution of atrazine concentrations in reservoir water 

were determined by ELISA. Increases in atrazine 

concentration in the reservoir were detected following 

the first significant runoff after herbicide application. 

When data were combined with the subset of DAR 

values, analyses were proven to be effective in 

studying the transport of atrazine through Perry 

Lake. Also, an inverse relation was observed 

between the atrazine concentration in reservoir water 

and the DAR value. Similar relations may exist for 

other herbicides and their metabolites. These relations 

may be useful in assessing the water quality of 

reservoirs.

502
EXPLANATION sz/'pjVf^j:

DETHYLATRAZINE- , '.t~H±: .....>. 

TO-ATRAZINE RATIO :.|.;.J..i tl i i

0.095-0.134 44 ; H i i i;

0.255 - 0.294 tu : ;;; ? { : : ; ? "

Base map from U S. Geological Survey

digital data. 1:100.000,1983 

Abers Equal-Area Conic projection 

Standard parallels 33° and 45°.

A. Pre-application survey, March 30 through April 10, 1992.

39°15'_ -

39°05' ~

B. First-flush survey, June 16-25,1992.

95

CO

C. Summer surv

________I_

S Figure 2. Deethylatrazine-to-atrazine ratios from (A) pre-application, (B) first-flush, and (C) summer surveys at Perry Lake, nort
REFERENCES

Bushway, R.J., Perkins, L.B., Savage, S.A., Lekousi, S.L., 

and Ferguson, B.S., 1988, Determination of atrazine 

residues in water and soil by enzyme immunoassay: 

Bulletin of Environmental Contamination and 

Toxicology, v. 42, p. 899-904.

Goolsby, D.A., Thurman, E.M., and Kolpin, D.W., 1991, 

Geographic and temporal distribution of herbicides in 

surface waters of the upper midwestern United States, 

1989-90: U.S. Geological Survey Water-Resources 

Investigations Report 91 -4043, p. 183-188.

Meyer, M.T., Mills, M. S., and Thurman, E. M., 1993, 

Automated solid-phase extraction of herbicides from 

water for gas chromatographic-mass spectrometric 

analysis: Journal of Chromatography, v. 629, p. 55-59.

Pomes, M.L., Thurman, E.M., and Goolsby, D.A., 1991, 

Comparison of microtitre-plate, enzyme-linked 

immunosorbent assay (ELISA) and gas 

chromatography/mass spectrometry (GC/MS) for 

analysis of herbicides in storm runoff samples, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 572-574.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1991, Herbicides in surface waters of the 

midwestern United States~Thc effect of spring flush: 

Environmental Science and Technology, v. 25, no. 10, 

p. 1794-1796.

Thurman, E.M., Meyer, M., Pomes, M., Perry, C.A., and 

Schwab, A.P., 1990, Enzyme-linked immunosorbent 

assay compared with gas chromatography/mass 

spectrometry for the determination of triazine 

herbicides in water: Analytical Chemistry, v. 62, 

no. 18, p. 2043-2048.

504
Pesticides in Near-Surface Aquifers: Results of the

Midcontinental United States Ground-Water

Reconnaissance, 1991-92

By Dana W. Kolpin 1 , Donald A. Goolsby2 , Diana S. Aga3 , Jana L. Iverson2 ,

and E. Michael Thurman3

Abstract

Selected pesticides and metabolites were 

examined in near-surface aquifers in the corn- and 

soybean- producing region of the midcontinental 

United States to determine their hydrogeologic, 

spatial, and seasonal distribution. At least one 

herbicide or atrazine metabolite was detected in 

28.4 percent of the 303 wells sampled during the 

spring and late summer of 1991. During 1991, 

deethylatrazine was the most frequently detected 

compound followed by atrazine, deisopropylatrazine, and prometon. No herbicide concentration 

exceeded the U.S. Environmental Protection 

Agency's maximum contaminant levels for drinking water. During 1991, the highest frequency of 

herbicide detection was in the western part of the 

study region and lowest frequency of detection 

was in the eastern part of the study region. Unconsolidated aquifers were found to be more susceptible to herbicide contamination than bedrock 

aquifers on the basis of results of the 1991 study. 

During the summer of 1992, additional water 

samples were collected from 101 wells to examine the occurrence of agricultural chemicals not 

analyzed during the previous year. Water from 

62 percent of the resampled wells had detectable 

concentrations of either a pesticide or pesticide 

metabolite; 27 different compounds were 

detected. During the 1992 study, five of the

'U.S. Geological Survey, Iowa City, Iowa 

2U.S. Geological Survey, Denver, Colorado 

3 U.S. Geological Survey, Lawrence, Kansas

six most frequently detected compounds were 

herbicide metabolites.

INTRODUCTION

Almost 60 percent of the pesticides (Gianessi 

and Puffer, 1990) and fertilizers (U.S. Environmental 

Protection Agency, 1990) used in the United States are 

applied in the Midcontinent to enhance crop yields. 

Because of the potential for these contaminants to 

affect the quality of water resources, the Midcontinent 

has become a focus for research on agricultural-chemical contamination in all parts of the environment 

including streams, reservoirs, ground water, and rainfall. In 1991, the U.S. Geological Survey (USGS) 

through its Toxics Substances Hydrology Program, 

began a ground-water-reconnaissance of agriculturalchemical contamination in aquifers located within 

15 m of the land surface (near-surface) in the midcontinental United States. Near-surface aquifers represent 

hydrogeologic settings that have the greatest potential 

for herbicide contamination. The objectives of this 

reconnaissance are to (1) determine the hydrogeologic, 

spatial, and seasonal distribution of agricultural 

chemicals in near-surface aquifers of the midcontinent; (2) determine statistical relations between the 

occurrence of agricultural chemicals and selected 

human and natural factors; and (3) collect data on 

pesticides and their metabolites in ground water from 

geographic areas where few data previously were 

available.

Because of questions generated by results of the 

1991 sampling, additional data were collected during 

July and August 1992. The major objective of the 

additional sampling was to determine why reportable 

herbicide concentrations were not measured in more

505
than 70 percent of the near-surface aquifers sampled in 

1991. Possible explanations include the following: 

The wells sampled produce "old" water, which would 

not be expected to contain herbicides; herbicides are 

present in many of these aquifers but at concentrations 

below the reporting limits available during the 1991 

study; herbicides are present but are compounds that 

were not examined during 1991; and the herbicide 

metabolites might be more persistent and mobile than 

their parent compounds and are present in the aquifers 

sampled. To test these hypotheses, the number of compounds analyzed for during 1992 was increased and 

included additional herbicides and selected metabolites, insecticides and selected metabolites, detergents, 

volatile organic compounds, and surfactants. Only 

the results for pesticides and pesticide metabolites are 

discussed in this paper.

This paper was previously published (Kolpin 

and others, 1993b) and describes results of a regional 

reconnaissance of selected agricultural chemicals in 

near-surface aquifers that was conducted during 1991, 

including some initial results from the continued 

investigation of agricultural chemicals in near-surface 

aquifers conducted during 1992. A comprehensive 

discussion of the data and interpretive results derived 

from the 1991 reconnaissance study is presented in 

other USGS reports. ( Kolpin and others, 1993a; 

Kolpin and others, 1994).

STUDY DESIGN AND METHODS

Wells sampled during 1991 (fig. 1) were 

selected to ensure that sites were geographically and 

hydrogeologically representative of near-surface 

aquifers in the corn- and soybean-producing region of 

the Midcontinent. The wells were distributed geographically by State and hydrogeologically by aquifer 

class (unconsolidated as opposed to bedrock) and relative depth within the aquifer. A workplan outlining the 

design and plan for well selection, sample-collection 

protocol, and quality assurance was prepared to ensure 

the validity of collected data (Kolpin and Burkart, 

1991). Most of the wells selected were sampled twice 

during 1991. A total of 299 herbicide analyses were 

obtained during March or April (preplanting); and 

290 herbicide analyses were obtained during July or 

August (postplanting). Ancillary data on land use and 

other features near the wells were collected at the time 

of sampling.

In 1992, 101 wells were sampled during July or 

August after the application of pesticides. These wells 

were distributed geographically by State and hydrogeologically by aquifer class. This approach resulted 

in a widespread distribution of sampling sites across 

the Midcontinent (fig. 1).

All samples were collected by USGS personnel 

using equipment constructed of materials constructed 

of glass and stainless steel that would not leach or 

adsorb organic compounds. Decontamination procedures were implemented to prevent cross-contamination between wells and samples (Kolpin and Burkart, 

1991). Wells were purged before sampling until pH, 

water temperature, and specific conductance stabilized. A quality-assurance program using a scries of 

field blanks, field duplicates, and spikes verified effectiveness of the sampling protocol and the analytical 

procedures.

All samples collected during 1991 were 

analyzed for 11 herbicides and 2 atrazine metabolites 

by gas chromatography/mass spectrometry (GC/MS) 

(Thurman and others, 1990). This method had an 

analytical reporting limit of 0.05 ng/L (micrograms 

per liter). For the samples collected during 1992, 

several additional analytical methods also were used 

including a 1-liter solid-phase extraction GC/MS 

method (Steven D. Zaugg, U.S. Geological Survey, 

written commun., 1992). This method was used to 

analyze 47 pesticides with reporting limits as low as 

0.001 ng/L. A total of 45 samples collected during 

1992 were analyzed forchlorophenoxy acid herbicides 

(2,4-D for example) and DCPA. The selection of these 

45 samples was biased toward wells with herbicides 

detected during 1991. Sixty-six samples were 

analyzed for an ethanesulfonate alachlor metabolite 

(ESA) by the method of Aga and others (1996). The 

selection of these 66 samples also was biased toward 

wells with herbicides detected during 1991.

RESULTS AND DISCUSSION

1991 Sampling

Herbicides or atrazine metabolites were detected 

in 28.4 percent of the wells sampled during the two 

collection periods in 1991 (table 1), but no concentrations exceeded U.S. Environmental Protection Agency 

(USEPA) maximum contaminant levels (MCL) for 

drinking water. The relatively low frequency of

506
100°

North

 TV 94° 

^l..-

I I1VJIUI V

,' Dakola j Minnesota

 - - .. ^_ \ °

V«^\-.

South 

Dakota 3f- ...

/ (

/ Nebraska / 

EXPLANATION

WELLS SAMPLED DURING 

1991

WELLS SAMPLED DURING 

1991 AND 1992

Base from U.S. Geological Survey digital data. 12.000.000.1972

Abers Equal-Area Conic projection

Standard parallels 29°W and tyw. Central meridian -90°30'

200 MILES

100 200 KILOMETERS

Figure 1. Locations of wells sampled during 1991 and 1992.

Table 1. Frequency of herbicide and atrazinemetabolite detection in water from wells sampled 

during the spring and summer of 1991

[%, percent; N, number of wells sampled; 

, micrograms per liter]

Compound

Maximum

% detection1 concentration 

(N=303) (jig/L)

Any herbicide 

or metabolite

Deethylatrazine 

Atrazine

Deisopropylatrazine 

Prometon

Metolachlor

Alachlor

Simazine

Metribuzin

Cyanazine

28.4

21.4 

20.8

7.6 

6.3

4.0

2.0

1.6

1.3

1.3

2.32

2.32 

2.09

1.17 

0.86

1.46

1.05

0.27

0.57

0.68

Analytical reporting limit for all compounds

herbicide detection in near-surface aquifers strongly 

contrasts with conditions observed in streams for this 

same region; herbicides were detected in nearly all of 

the streams sampled, and some of the concentrations 

exceeded an MCL (Thurman and others, 1992). Herbicides were not distributed uniformly in near-surface 

aquifers across the study region (fig. 2). The highest 

frequency of detection was observed in the extreme 

western part of the region; the lowest frequency of 

detection was observed in the eastern third of the 

region.

Data on herbicide metabolites are important to 

understand fully the occurrence, distribution, and fate 

of herbicides in the environment. The most frequently 

detected herbicide compound in water from wells 

sampled during the 1991 study was deethylatrazine 

(table 1). The frequency of atrazine-residue detection 

(atrazine plus deethylatrazine plus deisopropylatrazine) was 25.7 percent, whereas the frequency of atrazine detection alone was 20.8 percent. It is probable 

that the frequency of detection for other herbicide residues also would be larger if analyses for the metabolites of those herbicides were available. Consequently, 

if herbicide metabolites are not quantified, the effects

507
Missou °

-f /'V

EXPLANATION

0 NO HERBICIDES OR ATRAZINE 

METABOLITES DETECTED

HERBICIDES OR ATRAZINE 

METABOLITES DETECTED

 .. ^^,- , « v^i"03' S'UrV°y d'9'1al dala< 1 2.WO.OOO. 1972 

StandardWtoh 3^^dSff. central meridian -90»30-

0 ^00 200 MILES 

0 100 200 KILOMETERS

Figure 2. Spatial distribution of herbicide or atrazine detections in wells sampled during 1991 (reporting limit of 0.05 micrograms 

per liter for all compounds).

of herbicide usage on the quality of water resources 

could be substantially underestimated.

The deethylatrazine-to-atrazine ratio (DAR) in 

water can provide information on the residence time of 

atrazine in the unsaturated zone (Thurman and others, 

1992). Atrazine can be degraded by soil microorganisms to deethylatrazine and other metabolites. Below 

the soil zone, the effects of decreased amounts of 

available oxygen, reduced microbial activity, and 

reduced amounts of organic carbon combine to greatly 

reduce the biodegradation rates for atrazine (Dhileepan and Schnoor, 1992). Thus, slow infiltration 

through the soil results in larger DAR values than does 

rapid infiltration. Thurman and others (1992) reported 

DAR values in midwestern streams during 1989-90 

generally were less than 0.1 during periods of spring 

runoff shortly after herbicide application but generally 

were 0.4 or larger in the fall during base-flow conditions. The median DAR value obtained from the 1991 

ground-water reconnaissance (Kolpin and others, 

1993a) was about 0.7 (fig. 3).

The frequency of herbicide detection in water 

from near-surface aquifers was significantly larger 

(p = 0.002; Mann-Whitney test) in unconsolidated 

aquifers (28 percent) than in bedrock aquifers (17 percent). Unconsolidated aquifers might be more susceptible to herbicide contamination than bedrock aquifers 

for several reasons. First, most of the unconsolidated 

aquifers sampled, being alluvial in origin, have a long 

and narrow planar geometry; accordingly, paths of 

ground-water flow from recharge areas are much 

shorter than ground-water flow paths in the more laterally extensive bedrock aquifers sampled for this study. 

Second, there is a general absence of low-permeability 

material overlying the unconsolidated aquifers. About 

78 percent of the unconsolidated aquifers sampled, but 

only 48 percent of the bedrock aquifers sampled, were 

under unconfined conditions. An aquifer that is confined in the vicinity of a properly constructed well is 

effectively isolated from the downward percolation of 

herbicides that might be applied near that well. Two 

lines of evidence support the conclusion that the

508
Ill 

z

cc

HI

z

HI 

HI 

Q 

LL

O

o

cc

50

10

1.0

0.1

0.01

GROUND-WATER 

RECONNAISSANCE

SURFACE-WATER 

RECONNAISSANCE

EXPLANATION

90TH PERCENTILE 

75TH PERCENTILE 

MEDIAN

25TH PERCENTILE

10TH PERCENTILE

Figure 3. Deethylatrazine-to-atrazine ratios in water from wells sampled during 1991 and for the surface water reconnaissance, 

Midcontinental United States during 1989-90. (Thurman and others, 1992.)

unconsolidated aquifers are unconfined more often 

than bedrock aquifers and contain mostly local paths 

of ground-water flow. First, a significant difference in 

the frequency of herbicide detection between preplanting (22.6 percent) and postplanting (34.1 percent) 

sampling periods was found only in water samples 

from unconsolidated aquifers. Differences in water 

quality during such a short time would be expected 

only in aquifers that are recharged rapidly or that are 

located a short distance from its recharge area. Second, 

the median DAR was 0.57 for water samples from 

unconsolidated aquifers and 1.25 for water samples 

from bedrock aquifers. Because the DAR can indicate 

residence time in the unsaturated zone, the comparatively small DAR for unconsolidated aquifers again 

indicate that the recharge rates were generally more 

rapid than for the bedrock aquifers sampled.

The frequency of herbicide detection differed 

significantly (p = 0.001; Kruskal-Wallis test) among 

intervals of depth to the top of the aquifer material 

(0 to 3 m, 28.4 percent; 3 to 6 m, 29.9 percent; 6 to

9 m, 23.9 percent; 9 to 12m, 16.9 percent; and greater 

than 12m, 11.1 percent). Generally, the greater the 

depth to the top of the aquifer material, the smaller the 

frequency of herbicide detection. This vertical distance from land surface to the top of the aquifer material gives a general indication of the proximity of an 

aquifer to recharge source. The greater this distance, 

the greater the opportunity for processes such as transport, transformation, or sorption of chemicals to take 

place.

The percentage of land in corn or soybean production within a 3.2-km radius surrounding the wells 

did not appear to affect the frequency of herbicide 

detection. Although this lack of an association was 

unexpected, it could have been affected by either the 

selection criteria for this study, which required at least 

25 percent of the land in a 3.2-km radius surrounding 

the wells be in corn or soybean production in the 1990 

growing season, or that many herbicides being used in 

the Midcontinent and their respective metabolites 

were not analyzed during the 1991 study.

509
The frequency of herbicide detection was 

related to crop irrigation. Water samples from wells 

located within 3.2 km of irrigated crops had almost 

double (35 percent) the frequency of herbicide detection than was observed in water samples from wells 

without nearby irrigation (19 percent). Almost 95 percent of the wells near irrigated areas were in unconsolidated aquifers, where some of the irrigation water 

typically recharges locally to aquifers. Under these 

conditions, irrigation could result in continuous leaching of chemicals to the aquifer throughout the irrigation season.

The frequency of herbicide detection was 

related to the proximity of a well to a stream. The frequency of herbicide detection in water samples from 

wells within 30 m of a stream was more than double 

(48 percent) that observed in water samples from wells 

with no stream nearby (22 percent). This difference 

can result from a hydrologic connection between aquifers and streams due to either natural or induced 

hydraulic gradients or a convergence of ground-water 

flow paths where ground water discharges to streams. 

Frequencies of detection and concentrations of herbicides are much larger in streams than in aquifers in the 

Midcontinent (Thurman and others, 1992). Thus, 

recharge to an aquifer by a stream could be a source of 

herbicide contamination to the aquifer (Squillace and 

others, 1993).

1992 Sampling

Herbicides, herbicide metabolites, or insecticides (table 2) were detected in water samples from 

62 percent of the wells resampled during 1992. Most 

of these detections were at concentrations of less than 

0.5 |ng/L; some concentrations were in the parts per 

trillion range. No concentration exceeded a USEPA 

MCL for drinking water. Although concentrations of 

some of these compounds can be very low, even trace 

amounts can be important in identifying flow paths 

and transformation patterns in the hydrologic system, 

because these compounds are not present in the environment under natural conditions. The frequency of 

herbicide detection was much larger in 1992 than in 

1991 because more herbicides and herbicide metabolites were analyzed (40 in 1992 as opposed to 13 in 

1991); 20 insecticides were included in the suite of 

compounds analyzed; analytical reporting limits for 

many compounds were about an order of magnitude 

lower for the 1992 study than for the 1991 study; and

the sample selection for some of the herbicides and 

herbicide metabolites analyzed for the 1992 study was 

biased toward wells where these compounds were 

detected during 1991.

A preliminary analysis of results indicates that 

the detections of pesticides in the 1992 study were 

much more evenly distributed across the Midcontinent 

(fig. 4) than determined from the 1991 sampling 

(fig. 2). In particular, spatial patterns of detections in 

the eastern part of the study region were similar to 

those in other parts of the region, in contrast to results 

from the 1991 study. The greater frequency of detection in the eastern part of the study region can be 

attributed to the lowered reporting limits for atrazine 

and simazine and to the analysis of alachlor metabolites. Spatial patterns of detections across the Midcontinent are apparent when specific compounds are 

examined. For example, more than 50 percent of the 

insecticide detections were in Nebraska and more than 

60 percent of the simazine detections were in Ohio.

Data obtained from the 1992 sampling reinforces the importance of metabolite information to 

understand fully the occurrence, distribution, and fate 

of pesticides in the environment. Of the six most frequently detected pesticide compounds during 1992, 

five were herbicide metabolites (table 2). In addition to 

alachlor, two alachlor metabolites also were detected: 

2,6-diethylanaline and ESA. ESA recently has been 

identified as a relatively persistent and mobile compound (Baker and others, 1993; Aga and others, 

1996). Analytical results on alachlor and the two 

aforementioned metabolites were available from 

66 samples. Of these samples, only about 8 percent 

contained detectable concentrations of alachlor; however, 21 percent contained detectable levels of 

2,6-diethylanaline and 47 percent contained ESA. The 

detection frequency of alachlor-residue (alachlor plus 

2,6-diethylanaline plus ESA) was 53 percent, or more 

than 5 times the detection frequency of the parent 

compound alone. Although no alachlor concentrations 

exceeded the USEPA MCL of 2.0 |ng/L, 4.5 percent of 

alachlor-residue concentrations exceeded 2.0 |ng/L, 

and 15 percent exceeded 1.0 |ng/L. These results show 

how the effects of pesticides on the quality of water 

resources can be substantially underestimated without 

data on metabolites.

510
Table 2. Frequency of pesticide and pesticide metabolite detections in water from wells sampled during 1992, 

Midcontinental United States

[jig/L, micrograins per liter; ESA, alachlor metabolite]

percent 

Compound detection

Any pesticide 

or metabolite

ESA

Atrazine

Deethylatrazine

Deisopropylatrazine

2,6-Diethylanaline

DCPA

Simazine

Metolachlor

Prometon

2,4-D

P,P' DDE

Alachlor

Dicamba

Picloram

Chlorpyrifos

Ethalfluralin

Cyanazine

2,4,5-T

EPTC

Trifluralin

Triallate

Benfluralin

Napropamide

Pcndimethalin

Propachlor

Tebuthiuron

Metribuzin

62.0

47.0

43.0

31.0

18.2

16.0

15.6

13.0

11.0

9.0

6.7

6.4

5.0

4.4

4.4

4.2

3.2

3.0

2.2

2.1

2.1

2.1

1.1

1.1

1.1

1.1

1.1

1.0

Number 

of 

samples1

100

66

100

100

66

94

45

100

100

100

45

94

100

45

45

94

94

100

45

94

94

94

94

94

94

94

94

100

Maximum Reporting 

concentration limit 

(^ig/L) (^ig/L)2 Action/use3

4.95

1.03

1.79

.28

.022

2.22

.077

.76

1.35

.89

.03

.99

.01

.03

.024

.014

.02

.02

.003

.016

.007

.018

.008

.01

.002

.05

.05

0.100

.005

.015

.050

.003

.010

.005

.002

.010

.010

.006

.002

.010

.010

.004

.004

.008

.010

.002

.003

.001

.004

.003

.008

.007

.010

.005

Herbicide metabolite (alachlor)

Selective herbicide: corn, sorghum

Herbicide metabolite (atrazine)

Herbicide metabolite (atrazine)

Herbicide metabolite (alachlor)

Herbicide metabolite (dacthal)

Selective herb:icide corn,

Selective herbicide: corn, sorghum, soybeans

Nonselective herbicide

Selective herbicide:c orn, sorghum, pasture

Insecticide metabolite (DDT)

Preemergent herbicide: corn, soybeans

Herbicide

Systemic herbicide: deep-rooted plants

Insecticide: corn, soybeans

Selective preemergence herbicide: soybeans

Selective herbicide: corn

Discontinued herbicide

Selective herbicide: corn

Selective preemergence herbicide: soybeans

Selective preemergence herbicide: wheat

Herbicide: alfalfa, clover

Selective herbicide: orchards, tobacco

Selective herbicide: corn, soybeans

Herbicide

Herbicide: noncropland areas, rangeland

Herbicide: soybeans

1 When sample size is less than 94, the frequency of detection for that compound may be affected by the biased sample 

selection toward wells that had a herbicide detection during 1991

2Reporting limits of compounds not detected: alpha BHC (0.002), ametryn (0.05), bentazon (0.10), butylate 

(0.002),carbaryl (0.01), carbofuran (0.01), 2.4-DP (0.01), dacthal (0.002), diazinon (0.002), dieldrin (0.002), 

dimethoate (0.02), disulfoton (0.017), ethoprop (0.003), fonofos (0.003), linuron (0.01), malathion (0.005), methyl 

parathion (0.008), methyl azinphos (0.005), molinate (.004), parathion (0.005), pebulate (0.005), permethrin (0.01), 

phorate (0.003), prometryn (0.05), pronamide (0.004), propargite (0.02), propanil (0.005), propazine (0.05), silvex 

(0.01), terbacil (0.01), terbufos (0.013), terbutryn (0.05), thiobencarb (0.002)

3modified from Meister (1992)

511
EXPLANATION

° NO PESTICIDES OR PESTICIDE 

METABOLITES DETECTED

 PESTICIDES OR PESTICIDE 

METABOLITES DETECTED

Base from U.S. Geological Survey digital data, 12.000.000.1972

Afoers Equal-Area Conic projection

Standard parallels 39°30* and 43°301 . Central meridian -90°3Cr

200 MILES

100 200 KILOMETERS

Figure 4. Spatial distribution of pesticide or metabolite detections in water from wells sampled during 1992 (reporting limits of 

0.001 to 0.1 micrograms per liter for compounds analyzed).

SUMMARY

Herbicides or herbicide metabolites were 

detected in water samples from 28.4 percent of wells 

sampled in near-surface aquifers of the Midcontinent 

during 1991. During 1991, deethylatrazine was the 

most frequently detected compound followed by atrazine, deisopropylatrazine, and prometon. The highest 

frequency of detection was observed in the extreme 

western part of the region; the lowest frequency of 

detection was observed in the eastern third of the 

region. Unconsolidated aquifers were more susceptible to herbicide contamination than the bedrock aquifers sampled because of their general long and narrow 

planar geometry and the general absence of overlying 

confining units. The frequency of herbicide detection 

was inversely related to the depth to the top of the 

aquifer material. The greater this depth, the greater the 

opportunity for processes such as transport, transformation, and sorption of chemicals to occur. The 

frequency of herbicide detection almost doubled when

crops were irrigated within 3.2 km of the well. The 

frequency of herbicide detection more than doubled 

when a well was within 30 m of a stream, possibly 

because of either natural or induced hydraulic gradients or from a convergence of ground-water flow paths 

near a stream.

During 1992, a pesticide or pesticide metabolite 

was detected in 62 percent of the 101 wells resampled. 

The greater frequency of detection during 1992 was 

caused by an increased number of herbicides and herbicide metabolites analyzed, inclusion of insecticides 

in the suite of compounds analyzed, analytical reporting limits for many compounds that were lower (about 

an order of magnitude) in 1992 than in 1991, and biasing the sample selection for some of the herbicides 

analyzed toward wells in which herbicides were 

detected during 1991. Herbicide detections in the 1992 

study were much more evenly distributed across the 

region than in the 1991 study, probably because of the 

low reporting limits of atrazine and simazine and the 

analysis of alachlor metabolites.

512
Data on herbicide metabolites are important for 

a full understanding of the occurrence, distribution, 

and fate of herbicides in the environment. During 

1991, the most frequently detected compound was 

deethylatrazine, and the frequency of atrazine-residue 

detection was greater than the frequency of atrazine 

detection alone. During the 1992 resampling, five of 

the six most frequently detected compounds were herbicide metabolites. The frequency of alachlor-residue 

detection was more than five times the frequency of 

alachlor detection alone. These results show how the 

affects of pesticides on the quality of water resources 

can be substantially underestimated without data on 

metabolites.

REFERENCES

Aga, D.S., Thurman, E.M., and Pomes, M.L., 1996, Determination of alachlor and its sulfonic acid metabolite 

in water by solid-phase extraction and ELISA, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology Program Proceedings of the Technical Meeting, 

Colorado Springs, Colorado, September 20-24, 1993, 

Water-Resources Investigations Report 94-4015.

Baker, D.B., Bushway, R.J., Adams, S.A., and Macomber, 

Carol, 1993, Immunoassay screens for alachlor in rural 

wells: False positives and an alachlor soil metabolite: 

Environmental Science and Technology, v. 27, no. 3, 

p 562-564.

Dhileepan, N.R., and Schnoor, J.L., 1992, Effect of two 

electron acceptors on atrazine mineralization rates in 

soil: Environmental Science and Technology, v. 26, 

no. 11, p. 2298-2300.

Gianessi, L.P., and Puffer, C.M., 1990, Herbicide use in the 

United States: Washington, D.C., Resources for the 

Future, [Inc.], Quality of the Environment Division, 

127 p. [revised 1991]

Kolpin, D.W., and Burkart, M.R., 1991, Work plan for 

regional reconnaissance for selected herbicides and 

nitrate in ground water of the mid-continental United 

States, 1991: U.S. Geological Survey Open-File 

Report 91-59, 18 p.

Kolpin, D.W., Burkart, M.R., and Thurman, E.M., 1993a, 

Hydrogeologic, water-quality, and land-use data for the 

reconnaissance for herbicides and nitrate in nearsurface aquifers of the midcontinental United States, 

1991: U.S. Geological Survey Open-File Report 

93-114.

Kolpin, D.W., Burkart, M.R., and Thurman, E.M., 1994, 

Herbicides and nitrate in near-surface aquifers in the 

midcontinental United States, 1991: U.S. Geological 

Survey Water-Supply Paper 2413.

Kolpin, D.W., Goolsby, D.A., Aga, D.S., Iverson, J.L., and 

Thurman, E.M., 1993b, Selected papers on agricultural 

chemicals in water resources of the midcontinental 

United States, in Goolsby, D.A., Boyer, L.L., and 

Mallard, G.E., compilers: U.S. Geological Survey 

Open-File Report 93-418, p. 64-74.

Meister, R. T, editor, 1992, Farm Chemicals Handbook 

1992: Willoughby, Ohio, Meister Publishing Company, 

794 p.

Squillace, P.J., Thurman, E.M., and Furlong, E.T., 1993, 

Ground water as a nonpoint source of atrazine and 

deethylatrazine in a river during base-flow conditions: 

Water Resources Research, vol 29, no. 6, p. 1719- 

1729.

Thurman, E.M., Meyer, Michael, Pomes, Michael, Perry, 

C.E.,a nd Schwap, A.P., 1990, enzyme-linked immunosorbent assay compared with gas chromatography/ 

mass spectrometry for the determination of triazine 

herbicides in water: Analytical Chemistry, v. 62, 

no. 18, p. 2043-2048.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., Mills, M.S., 

Pomes, M.L., and Kolpin, D.W., 1992, A reconnaissance study of herbicides and their metabolites in 

surface water of the midwestern United States using 

immunoassay and gas chromatography/mass spectrometry: Environmental Science and Technology, v. 26, 

no. 12 p. 2440-2447.

U.S. Environmental Protection Agency, 1990, County-level 

fertilizer sales data: Washington, D.C., Office of Policy 

Planning and Evaluation, September 1990, computer 

disk.

513
514
Predicting Nitrate-Nitrogen and Atrazine Contamination 

in the High Plains Aquifer in Nebraska

ByA.D. Druliner1 ancfT.S. McGrath1

Abstract

Three statistical models, developed through 

the Toxic-Waste Ground-Water Contamination 

Program of the U.S. Geological Survey, were used 

to predict the concentrations of nitrate-nitrogen 

and atrazine and the probability of atrazine 

detections in ground water in Buffalo and Hall 

Counties, south-central Nebraska. The models use 

a combination of hydrochemical, hydrologic, 

soils, and land-use explanatory variables, and 

ARC/INFO techniques to generate maps showing 

predicted concentrations of the selected contaminants. Confirmational testing of the models and 

comparison of maps of predicted concentrations to 

maps of observed concentrations showed that the 

statistical models were reasonable predictors of 

nitrate-nitrogen and atrazine concentrations in 

ground water of the High Plains aquifer in southcentral Nebraska.

INTRODUCTION

Today, farmers in Nebraska, like farmers throughout most agricultural States across the Nation, rely on 

fertilizers and pesticides to maximize crop yields and 

to sustain productivity. Mineral fertilizers, such as 

nitrogen, and broadleaf herbicides, such as atrazine, 

have been applied annually over large areas of 

Nebraska for more than 30 years. In 1989, Nebraska 

farmers applied an estimated 662,000 tons of nitrogen 

(as N) fertilizers an amount slightly less than amounts 

applied earlier in the decade (Nebraska Department of 

Agriculture, 1980; 1990). In 1987, an estimated 28.6 

million pounds of herbicide was applied in Nebraska 

(Baker, and others, 1990)~an 18 percent increase over

'U.S. Geological Survey, Lincoln, Nebr.

1982 estimates even though planted acreages actually 

declined by about 13 percent during the 6-year period. 

As a result of the intensive use of agricultural chemicals (Gormly and Spalding, 1979), ground water in 

many areas of the State is contaminated with nitratenitrogen concentrations in excess of the U.S. Environmental Protection Agency's (1991) maximum contamination level of 10 mg/L for drinking water and 

contains trace concentrations of atrazine (Exner and 

Spalding, 1990). However, not all ground water in 

areas of Nebraska where these chemicals are regularly 

used is contaminated to the same extent (Chen and 

Druliner, 1987), suggesting that one or more combinations of physical factors are affecting the transport of 

these contaminants to the ground water.

The U.S. Geological Survey conducted a nonpointsource ground-water contamination study in Nebraska 

to investigate the relations among concentrations of 

nitrate-nitrogen and atrazine in ground water and 

quantifiable physical factors. During the first phase of 

the study, Chen and Druliner (1987) investigated the 

presence of selected agricultural chemicals in ground 

water in parts of the High Plains aquifer and performed 

a preliminary investigation of the effects of selected 

physical factors on ground-water contamination by 

nitrate-nitrogen and atrazine. One of the specific objectives of the second phase of the study was to develop 

techniques that could be used to recognize or predict 

areas of potential ground-water contamination by 

selected agricultural chemicals.

This paper describes three statistical models developed as part of the second phase of the study that relate 

nitrate-nitrogen and atrazine concentrations in ground 

water to hydrologic, land-use, and soils variables in 

Nebraska. The models are used to predict nitrate-nitrogen and atrazine concentrations and atrazine detections 

in ground water in Buffalo and Hall Counties, southcentral Nebraska (area 1, fig. 1), and the predictions are 

compared to observed concentrations of nitrate-nitrogen and atrazine in ground water in the same area.

515
SOUTH DAKOTA

**Mm&MvM^^^w$'F%

US Geoloflicil -Survey intJO.tXKJ 

StKe of Nefamke, 1984 

Projection Albeie Equel ATM Conic projection 

Stendard pereUe 28«30' and 4B«30', eentrd meridian 98

AREA UNDERLAIN BY HIGH 

PLAINS AQUIFER

EXPLANATION

BOUNDARY OF STUDY AREA

MODEL CONFIRMATION 

STUDY AREA

. WELL SAMPLED FOR WATER QUALITY

IRRIGATION WELLS USED FOR 

CONFIRMATION OF PREDICTIVE MODELS

Figure 1. Extent of the High Plains aquifer in Nebraska, the location of ground-water sampling sites in the six study areas, and 

the location of Merrick County.

APPROACH

Six study areas encompassing 12 counties of 

Nebraska were selected in an effort to investigate a 

variety of hydrologic, land-use, and soils factors that 

could affect ground-water contamination by agricultural chemicals (fig. 1). More than 70 explanatory variables were tentatively identified as possibly affecting 

ground-water contamination. Some examples of these 

variables are depth to water, average hydraulic conductivity, fertilizer application amounts, tillage type, soil 

slope, and soil permeability. Water-quality samples 

were collected from 268 irrigation, domestic, public, 

and stock wells in the six study areas during a 4-year 

period (1984-87). Data quantifying the explanatory

variables were compiled for each of the 268 well sites 

and compared to nitrate-nitrogen concentrations in the 

ground water at each site. Similarly, data describing 

explanatory variables for 210 of the well sites were 

compared with atrazine concentrations in the ground 

water. Data for all six areas, as well as for single areas, 

were used to construct a variety of statistical models 

that could be used to predict areas of present or potential ground-water contamination.

DESCRIPTION OF STUDY AREAS

The six study areas contain from one to three 

counties each, and all are underlain by the High Plains 

aquifer (fig. 1). A variety of locations was selected to

516
provide a wide range of hydrochemical, hydrologic, 

soils, and land-use conditions in addition to targeting 

some areas of existing and potential ground-water contamination by agricultural chemicals. A brief description of the hydrology, land-use, and soils variables 

follows.

The High Plains aquifer is generally unconfined 

and consists of silt, sand, and sandstone of Tertiary Age 

with some calcareous cement and local zones of coarse 

sand and gravel and (or) overlying clay, silt, and sand 

of Quaternary age. In the six study areas, the High 

Plains aquifer is from 100 to 400 ft thick and is underlain by relatively impermeable clay of Cretaceous age. 

Recharge, chiefly from precipitation, occurs throughout most of the High Plains aquifer in the study areas. 

Additional recharge is provided by seasonal irrigation 

return flows, losing reaches of streams and rivers, and 

leakage from canals. Discharge from the aquifer in all 

areas occurs primarily through irrigation-well pumpage; secondary losses occur from evapotranspiration 

and ground-water discharge to streams, lakes, and 

canals.

The soils overlying the High Plains aquifer vary 

with the local topography and underlying geology. 

Soils in the study areas range from loamy, sandy soils 

in the Platte River Valley (areas 1 and 2) and adjacent 

to the Sand Hills (area 6) to clayey and silty soils in 

area 4 (Kuzila and others, 1990).

Irrigated corn is the dominant crop in five of the six 

areas and typically receives larger application amounts 

of nitrogen fertilizer and pesticides than other crops 

grown in Nebraska. Nitrogen most commonly is applied 

as anhydrous ammonia, which is usually injected into 

the ground prior to planting time. During the 1980's, 

nitrogen sales in the six study areas fluctuated from 

100,000 to 150,000 tons/yr (Nebraska Department of 

Agriculture, 1980; 1990). The most commonly used 

pesticide on row crops in Nebraska is atrazine, a preemergent herbicide used to control broadleaf weeds in 

corn and sorghum. An estimated 13.4 million pounds of 

atrazine was applied in Nebraska in 1988 (Baker and 

others, 1990).

METHODS

Multiple linear-regression techniques as described 

by Minitab (1989) and SAS (1990) were used to generate statistical models with nitrate-nitrogen and atrazine 

concentrations as the dependent variables. Stepwise, 

stepwise forward, stepwise backwards, best-regression, maximum R, R-squared, and logistic regression 

methods were used. In the Minitab stepwise-regression 

efforts, a minimum F-statistic of 1.8 was used to determine which explanatory variables would remain in the 

models even though all models containing explanatory 

variables with F-statistics less than 4 (and with T-ratio 

less than 2) were later rejected. The maximum R and 

R-squared techniques in SAS were designed to maximize 

the R-square values (percentage of explained variation) 

for each model by stepwise selection of explanatory 

variables, or by considering all combinations of 

explanatory variables, respectively. The Lillifores test 

(Iman and Conover, 1983) was used to ensure that the 

residuals for each of the representative models were 

normally distributed at the 95-percent confidence level. 

Plots of the residuals as a function of the predicted values were examined for each model to reveal trends that 

might indicate the presence of untested variables.

Logistic regression methods were used to develop 

models predicting the probability of atrazine detections 

in ground water with atrazine data that were converted 

to detect or nondetect values. Through a maximumlikelihood methodology, this technique produced 

multiple-regression models that predicted the probability of the presence or absence of the dependent variable 

at the specified threshold value. The goodness-of-fit for 

the logistic models was determined by comparing the 

percentages of correct and incorrect predictions rather 

than through coefficients of determination. Models 

were considered acceptable if they represented logically plausible relations with the explanatory variables 

and if the probability of exceeding the Chi-square statistic was less than 0.05 for each explanatory variable 

included in the model.

RESULTS

A number of multiple linear-regression models 

were produced with nitrate-nitrogen and atrazine concentrations as the dependent variables. Models were 

generated using data for all of the sampled areas, and 

other models used subsets of the data based on individual study areas. The best of these models explained 

from 50 to 68 percent of the variation (coefficient of 

determination) observed in the dependent variables. 

Logistic models using just atrazine concentrations as 

the dependent variables yielded predicted accuracies 

up to 81 percent when compared to the input data. 

Examples of these models are presented in table 1.

517
Table 1. Regression models used to predict concentrations of nitrate-nitrogen and 

atrazine and the probability of atrazine detections in ground water

Explanatory variables Regression 

coefficients T-ratio P-value

Model A: Nitrate-nitrogen concentration (multiple-linear regression) 

Intercept 21.890 2.00 0.047 

Specific conductance .014902 6.01 0

Average hydraulic conductivity of the .080430 4.72 0 

unsaturated zone

Median well-completion date in 1-mile -.43650 1.96 .050 

radius

Coefficient of determination = 0.68

Model B: Atrazine concentration (multiple-linear regression)

Intercept

Nitrate-nitrogen concentration

Depth to water

Number irrigated acres in 1.7-mile radius

Coefficient of determination = 0.61

-1.46360

.0035944

-.006926

.00007743

-6.40

6.05

2.52

2.02

0

0

.015

.048

Model C: Probability of atrazine detection (logistic regression)

Explanatory variables Regression _,. _, .

* . Chi-square P-value coefficients ^

Intercept

Nitrate-nitrogen concentration

Water-table gradient

Percentage of clay in soil

Logarithm of well depth

Percentage total correct identifications: 81

-5.6503

.1822

-959.1

-.0559

-2.0797

6.6052

14.3439

11.0128

4.8374

4.8282

0.010

0

.001

.028

.028

Model A was formulated using only data from 

area 1. The model predicted nitrate-nitrogen concentrations using specific conductance, the average 

hydraulic conductivity of the unsaturated zone, and the 

median completion date of irrigation wells within a 1- 

mi radius of the sampled well. Specific conductance is 

believed to be an indirect measure of nitrate-nitrogen 

concentration, as nitrate-nitrogen was one of the principal anions in many of the collected water samples. 

The average vertical hydraulic conductivity of the 

unsaturated zone demonstrated that increased permeability of the sediments enhanced the likelihood of 

ground-water contamination by nitrate-nitrogen. The 

median completion date of irrigation wells within a 1- 

mi radius showed that sites that had been under irrigation the longest had an increased probability of groundwater contamination by nitrate-nitrogen.

Model B was formulated using only data from 

area 1. The model predicted atrazine concentrations 

using nitrate-nitrogen concentration, depth to water, 

and the number of irrigated acres in a 1.7-mi radius 

of the sampled well and explained about 61 percent

518
Table 2. Results of comparisons of observed nitrate-nitrogen and atrazine concentrations 

from sites in Merrick County with predicted concentrations and probabilities of detection from 

selected models 

[mg/L, milligrams per liter; ^g/L, micrograms per liter]

Model A: Model B: Model C:

1 Observed Predicted 

f: fiitraie-nifeO§sft nitrate-nitrogen 

= eoncemiariun concentration

i fe8g£Lji (mg/L)

1 28 4.1

I 21 4.4

I J& 18

g o *} ic 

g --p,P: ID

1 1$ 15

1 1& 6.8

1 21 12

| 4,1 7.2

P-valiKS from 

1 Wiicoxonrank 0.16 !

P ^W51 ., .. . 

W Observed :; ... ' ".^ '"''

-60

30

:i Not analyzed

B ^o

,60

.60

1.4

.30

Sr

,:,,,,;,,::, ;

Predicted 

atrazine 

n concentration 

(ug/L)

.26

m 

m .16

* Not determined

.06

S - 13

'SB* 11 .12

°S:^ .lO ?1

 || .04

::;ss 0.02

:*:?;5:;

:m

Probability of 

atrazine 

detection greater 

than 0.05 ug/L

1.0

.95

Not determined

.84

.97

.90

.97

.71

1 P-values greater than 0.05 are not significantly different at the 95-percent confidence level.

of the variation observed in atrazine concentrations. 

The presence of nitrate-nitrogen as an explanatory 

variable was not surprising given the overall correlation coefficient of 70 percent between nitrate-nitrogen and atrazine concentrations analyzed for this 

study. This supports the assumption that many of the 

factors affecting the transport of nitrate-nitrogen into 

the ground water also possibly affect the transport of 

atrazine. The depth-to-water variable demonstrates 

the greater likelihood of contamination with atrazine 

in areas with shallow water tables. The number of 

irrigated acres in a 1.7 mi radius of the sampled well 

is a measure of the intensity of irrigated agriculture 

and represents the effects of adjacent land use on 

water quality and the significance of lateral groundwater flow paths.

Model C was a logistic model using data from all 

study areas. The model predicted the probability of 

atrazine detections at or larger than the 0.01 fig/L 

detection limit using nitrate-nitrogen concentration, 

gradient of the potentiometric surface, percentage of 

clay in the soil, and the logarithm of well depth as

explanatory variables. The presence of the potentiometric gradient in the model logically maintains that 

areas with flat gradients and relatively slow rates of 

ground-water flow may tend to accumulate atrazine 

at the water table faster than areas with relatively 

steep gradients. The percentage of clay represents the 

importance of soil permeability, and the log of well 

depth again demonstrates that deeper water supplies 

are not as easily contaminated by agricultural chemicals as shallow water supplies.

The overall effectiveness of the models was 

tested in two ways. First, water from eight irrigation 

wells in Merrick County (fig. 1), which is adjacent to 

area 1, was collected and analyzed for nitrate-nitrogen and atrazine concentrations independently of the 

data used to produce the models. Data quantifying 

the explanatory variables for each of the models were 

collected for each ground-water sampling site and 

predicted concentrations, or probabilities of detection, were compared to observed concentrations 

(table 2). The Wilcoxon rank sum test (Minitab, 

1989), which compares population medians, showed

519
99 OO

9830

41 OO

4045 

41 OO

4045

99 OO

EXPLANATION

OBSERVED NITRATE-NITROGEN CONCENTRATIONS, IN MILLIGRAMS PER LITER 

20 30 40 50 60 70

Figure 2. Distribution of observed nitrate-nitrogen concentrations in Buffalo and Hall Counties, south-central Nebraska.

that the predicted nitrate-nitrogen concentrations 

were statistically similar to the observed concentrations at the 95-percent confidence level and that the 

atrazine concentrations predicted by model B were 

statistically smaller than observed atrazine concentrations. Model C predicted probabilities of atrazine 

detection of 0.71 to 1.0 for the seven irrigation wells 

for which atrazine data are available.

The second means of testing the models was 

through comparison of maps of observed and 

predicted ground-water concentrations of nitratenitrogen and atrazine (figs. 2-6). The predictive maps 

of nitrate-nitrogen and atrazine concentrations and 

probability of atrazine detection (figs 4-6) were

selected to represent each of the explanatory variables in the models. The surfaces for each model 

were overlaid, and the predictive equations were 

solved using the explanatory variable values for each 

cell. The maps predicting nitrate-nitrogen and atrazine concentrations and probabilities of atrazine 

detection were compared to maps showing observed 

nitrate-nitrogen and atrazine concentrations generated from data sources other than those used to 

produce these models (Exner and Spalding, 1990).

In comparing the maps of observed and predicted 

concentrations, note that both the predicted nitratenitrogen and atrazine concentrations and the probability of atrazine detections look quite similar to the maps

520
ta. horn US a-**" Sun«y DlgW D-* 1:100,000. ISM

Projection Alb«» _*» p." 

Stated P»«IWi 29*30 «! 46 . -ntr.) m-WUn -« 6 10 IE KILOMETERS

EXPLANATION

 - LINE OF EQUAL OBSERVED ATRAZ.NE CONCENTRATION !N GROUND WATER

-lntor«i In Mlerogr-n. P» life, b v-UW 

» LOCATION OF WELL FROM WHICH WATER WAS COLLECTED FOR ATRAZINE

Figure 3. Distribution of observed atrazine concentrations through 1989 in Buffalo and Hall Counties, south-central 

Nebraska.

of observed concentrations. Both maps of observed and 

predicted contamination show a generalized belt of 

ground-water contamination in the lowlands adjacent 

to the Platte River with some areas of larger concentrations superimposed on it. The upland areas north of the 

Platte River, where the depth to water is greater and 

where row-crop agriculture is less intense, show much 

smaller observed and predicted concentrations of 

nitrate-nitrogen and atrazine. Even in the upland areas, 

subtle effects, such as small stream drainages, are 

revealed by the predictive maps as being slightly more 

susceptible to ground-water contamination by agricultural chemicals than other upland areas. The fine detail 

shown on the predictive maps is the result of much

larger data sets describing the explanatory variables 

(usually an order of magnitude larger) than the 

observed nitrate-nitrogen and atrazine data sets.

CONCLUSIONS

Multiple-regression and logistic regression techniques appear to be well suited for the generation of 

predictive ground-water-quality models. Although the 

models explained only 50 to 68 percent of the variation 

in nitrate-nitrogen and atrazine concentrations in 

ground water, the predicted concentrations and probabilities of detection generally agreed with observed 

contaminant concentrations from the eight wells in

521
99OO

98 3O

41°00

404

4O45

EXPLANATION

NITRATE OONCENTTRATION, IN MILLIGRAMS PER LITER 

10 15 20 25 30

Figure 4. Distribution of predicted nitrate-nitrogen concentrations in ground-water in Buffalo and Hall Counties, south-central 

Nebraska.

Merrick County, and maps of predicted concentrations 

and probabilities of detection closely agreed with maps 

of observed concentrations of nitrate-nitrogen and 

atrazine in ground water in area 1.

The coupling of regression and logistic models that 

predict ground-water quality with a geographic information system, such as ARC/INFO, can be a powerful 

tool for evaluating models as they are generated and for 

utilizing the models after they have been completed. 

The modeling and graphic representation permits the

user to define basic relations between selected groundwater contaminants and physical factors that may be 

affecting those concentrations and to use the relations 

to better define areas of potential ground-water contamination beyond the immediate areas of observed 

water-quality data sites. In fact, the explanatory variable data sets are commonly orders of magnitude larger 

than the dependent variables and can yield much larger 

coverages with finer detail than is possible with most 

contaminant data bases.

522
99 00

98 3O

41*00

40 45-

41*00

4046

I I

0.00

EXPLANATION

ATRAZINE CONCENTRATION, IN MICROGRAMS PER LITER 

0.14 0.28 0.42 0.55

Figure 5. Distribution of predicted atrazine concentrations in ground-water in Buffalo and Hall Counties, south-central 

Nebraska.

REFERENCES

ARC/INFO, 1991, 6.0 Grid command references, 

ARC/INFO Command references: Environmental 

Systems Research Institute, Inc.

Baker, Maurice, Peterson, Nancy, and Kamble, 

Shripat, 1990, Pesticide use in Nebraska--1987: 

Agricultural Research Division, Institute of 

Agriculture and Natural Resources, University 

of Nebraska, Research Bulletin 311, 34 p.

Chen, H.H., and Druliner, A.D., 1987, Nonpointsource agricultural chemicals in ground water in 

Nebraska Preliminary results for six areas of the

High Plains aquifer: U.S. Geological Survey 

Water-Resources Investigations Report 86-4338, 

68 p.

Exner, M.E., and Spalding, R.F., 1990, Occurrence 

of pesticides and nitrate in Nebraska's ground 

water: Water Division, Institute of Agriculture and 

Natural Resources, University of Nebraska, WC1, 

34 p.

Gormly, J.R., and Spalding, R.F., 1979, Sources and 

concentrations of nitrate-nitrogen in groundwater 

of the Central Platte Region, Nebraska: Ground 

Water, v. 17, no. 3, p. 291-300.

523
88 OO

41OO

4O 46

8Be30

41 00

0.00

88 OO

EXPLANATION

PROBABILITY OF ATRAZINE DETECTION 

0.25 0.50 0.75

16 KILOMETERS

0.99

Figure 6. Distribution of probability of atrazine detection greater than or equal to 0.01 micrograms per liter in 

ground water in Buffalo and Hall Counties, south-central Nebraska.

Iman, R.L., and Conover, W.J., 1983, A modern 

approach to statistics: New York, John Wiley and 

Sons, 497 p.

Kuzila, M.S., Mack, A.M., Culver, J.K., and Schaefer, 

S.J., 1990, General soils map of Nebraska: Conservation and Survey Division, Institute of Agriculture and Natural Resources, University of 

Nebraska, scale 1:1,000,000.

Minitab, 1989, Minitab reference manual, Release 7. 

Nebraska Department of Agriculture, 1980, Nebraska

agricultural statistics: Annual Report 1978-1979,

172 p.

 1990, 1989/90 Nebraska agricultural statistics, 

Nebraska Agricultural Statistics Service, 148 p.

SAS, 1990, SAS/STAT user's guide-Volume 2, GLMVARCOMP: Gary, N.C., SAS Institute Inc., 

p. 892-1,686.

U.S. Environmental Protection Agency, 1991, 

National primary drinking-water regulations 

final rule: U.S. Federal Register, v. 56, no. 20, 

88 p.

524
Comparison of Simulated and Observed Movement of 

Bank-Storage Water Adjacent to the Cedar River, Iowa

By Paul J. Squillace1 and David W. Pollock2

Abstract

A ground-water flow model was constructed to describe quantitatively the movement 

of bank-storage water in an alluvial aquifer adjacent to the Cedar River, Iowa, during March 7- 

April 17, 1990, at the Palisades study area. The 

results from this two-dimensional model were 

compared to hydrologic and water-chemistry 

data. Hydrologic data consisted of 745 daily 

ground-water-level measurements made in 

27 wells during March 8-April 5, 1990. These 

745 measurements were used for model calibration and for comparing measured and simulated 

hydraulic heads and gradients. Water-chemistry 

data indicate that bank-storage water had a lower 

specific conductance and higher concentration of 

atrazine compared to the ambient ground water. 

Analysis of water chemistry and model results 

indicate that bank-storage water moved about 

30 meters into the aquifer at a depth of 6 meters 

below land surface. The model also showed that 

70 percent of the total bank-storage water moved 

through the river bottom while the remaining 

30 percent moved laterally through the riverbank. 

The flux of bank storage-water both into and out 

of the river bottom is largest near the river's edge 

and decreases substantially with distance from the 

river's edge to the center of the river. The model 

showed that it would take about 5 weeks for bankstorage water to discharge from the alluvial aquifer after the peak river stage.

'U.S. Geological Survey, Rapid City, South Dakota 

2 U.S. Geological Survey, Reston, Virginia

INTRODUCTION

During periods of snowmelt or rainfall, direct 

runoff to a river can cause an increase in river stage. 

Because of the increased river stage, river water can 

move through the river bottom and riverbank into the 

adjacent aquifer. As the stage declines, this same water 

is discharged back into the river. The volume of water 

so stored and released is referred to as bank storage.

Bank storage attenuates the flood wave in a river 

with permeable bank materials (Finder and Sauer, 

1971) and can contribute substantial discharge to 

rivers during base flow (Kondolf, and others, 1987; 

Cooper and Rorabaugh, 1963; Squillace and others, 

1993). Squillace and others (1993) have shown that 

atrazine is present in bank-storage water and that the 

discharge of this atrazine is probably the principal 

source of atrazine in the river during early base flow 

conditions. The movement of bank-storage water 

between surface water and ground water can be 

described analytically and by numerical flow models.

Analytical solutions that describe and quantify 

the movement of bank-storage water between surface 

water and ground water have been presented by many 

researchers (Morel-Seytoux, 1975; Dever and Cleary, 

1979; Hunt, 1990; Gill, 1985; Cooper and Rorabaugh, 

1963; Rorabaugh, 1963; Moench and others, 1974; 

Newsom and Wilson, 1988); however, assumptions 

made for some of these calculations could be questionable and could yield erroneous values (Sharp, 1977). 

These assumptions include Dupuit-Forcheimer conditions, homogeneity, full penetration of streams, and 

systems that are solely confined or unconfined.

Numerical flow models can simulate the complex movement of bank-storage water in the alluvial 

aquifer; however, accurate simulation of this movement requires accurate and ample field data. The purpose of this paper is to (1) describe quantitatively the 

movement of bank-storage water at the Palisades

525
study area adjacent to the Cedar River, Iowa, during 

March 8-April 17, 1990, by means of a ground-waterflow model, and (2) compare the model-simulation 

results with pertinent observed hydrologic and waterchemistry data. These water-chemistry data include 

specific conductance and the concentration of atrazine 

in ground water.

Acknowledgments

We would like to thank James Snyder for generously providing access to his land for the installation 

and monitoring of the observation wells.

DESCRIPTION OF STUDY AREA

The Palisades study area is located in an 

unfarmed flood plain of the Cedar River in east-central 

Iowa (fig. 1). The riverbank is approximately 4 m high 

at the water's edge during low-flow conditions and 

slopes steeply down to the river (fig. 2). Nested monitoring wells were installed at selected distances from 

the river's edge at the time of installation. As stage 

changes, the distance from the river's edge to the wells 

also changes because the bank is not vertical. Therefore, the distance of the monitoring wells from the 

river edge depends on river stage and can be considered accurate to within several meters over the range 

of river stages simulated. The screened interval of 

each well is 75 cm and is positioned 3 to 9 m below 

land surface.

The Cedar River alluvium consists of a typical 

fluvial sequence of coarse-grained sand and gravel at 

the base that grades upward to fine-grained sand, silt, 

and clay at the surface. Test drilling at the study site 

showed that the alluvium is underlain by 15 m of till. 

The alluvium is 15 m thick but decreases in thickness 

with distance from the river. These unconsolidated 

sediments were subdivided into seven units on the 

basis of grain size and clay content. Slug tests performed at 24 observation wells showed a range of 

horizontal hydraulic conductivity from a high of 71 

m/d in the coarse-grained alluvial sands to a low of 

0.3 m/d in the clayey-silty units. Estimates of the 

hydraulic conductivity of the coarse-grained sand 

ranged from 25 to 80 m/d on the basis of aquifer 

diffusivity, which was calculated by comparing the 

response of the aquifer to fluctuations in the river stage 

by the method of Finder and others (1969).

METHODS

The protocol for sampling the surface water and 

ground water is described by Squillace and others 

(1992). Atrazine was extracted from the water samples 

using solid-phase cartridges and then analyzed by gas 

chromatography/mass spectrometry according to the 

method of Thurman and others (1990).

RESULTS

Field Evidence for Bank Storage

In March 1990, direct runoff caused the river 

stage to increase almost 2 m and bank storage to occur 

because the normal ground-water hydraulic gradient, 

which was toward the river, was reversed. Hydraulic 

gradients were determined from 745 daily water-level 

measurements made during March 8- 

April 5, 1990, from both lines of wells along traces 

A-A' and B-B' shown in figure 1. Equipotential lines 

prepared from the water levels in the wells were parallel to the riverbank; therefore, the ground-water-flow 

paths during this time period were parallel to the two 

lines of wells shown in figure 1.

Bank storage of atrazine is demonstrated by 

water samples that were collected from wells and from 

beneath the river by means of a minipiezometer 

before, during, and after the March 1990, direct runoff 

event (fig. 2). Before direct runoff occurred, groundwater samples were collected during February 20-22, 

1990. Analysis of these samples showed that background concentrations of atrazine in the aquifer ranged 

from 0.1 to 0.4 [ig/L with one exception: a sample collected at the river's edge at a depth of 1 m below land 

surface had an atrazine concentration of 0.66 [ig/L. 

During direct runoff on March 20-22, 1990, the movement of bank-storage water into the aquifer was apparent because the concentrations of atrazine in bankstorage water were larger than ambient or background 

concentrations in the aquifer; the concentrations of 

atrazine in the river increased to almost 1.0 (ig/L when 

the river stage peaked on March 18, 1990, and these 

larger atrazine concentrations were carried into the 

aquifer with bank-storage water. Therefore, atrazine 

concentrations showed that bank-storage water moved 

30 m into the aquifer at a depth of 6 m below land surface. After the direct runoff event, water samples were 

collected for analysis of atrazine concentration after 

the river returned to base-flow conditions during

526
41°55'30"

41°55'20> 

Base from U.S. Geological Survey. 

Bertram. 1:24.000, 1982

200 METERS

EXPLANATION

 279 TOPOGRAPHIC CONTOUR-- 

Shows altitude of lane surface. 

Contour interval 3 r.ears. Datum 

is sea level

A A 4 TRACE OF SECTION-Shown below 

EDGE OF SURFACE WATER 

INTERMITTENT STREAM

 "' " " ' " MINIPIEZOMETER LINE 

......... GRAVEL ROAD

A. SURFACE-WATER SAMPLING SITE

30» OBSERVATION-WELL NEST-Number 

is distance from edge of river, in meters

CD

DEPTH BLOWLAND SURFACE. IN METERS D i/ioin 

4 A- B B'

I

:1 

, I I i j A«-v.um /

Glacial till

i i i ; ~ ~~, i i ged^ock

:! H ! ' ! :

Alluvium

Glacial tin

Bedrock

25 50 75 100 125 150 175 200 225 250 275 300 325 0 25 50 75 100 125 150

DISTANCE FROM EDGE OF RIVER. IN METERS

EXPLANATION

J WELL NEST-Each rectangle indicates screened 

t depth of individual observation wells in nest

Figure 1. (A) Location of observation-well nests at Palisades site, and (B) cross-sections showing depths of 

screened intervals.

April 3-5, 1990 (fig. 2). These results of these analysis 

show that large concentrations of atrazine were still 

present in the aquifer and that bank-storage water was 

still discharging to the Cedar River at that time. 

Squillace and others (1993) have shown that atrazine 

in this alluvial aquifer is in the dissolved phase and is 

not sorbed to the aquifer material in significant quantities.

Measurements of specific conductance shown in 

figure 3 confirm the movement of bank-storage water 

into the alluvial aquifer. The specific conductance in 

the river decreased from 840 u.S/cm during February 

20-22, 1990, to 580 ^S/cm during March 20-22, 1990. 

Bank storage of river water caused the specific conductance in the alluvial aquifer to decrease in an area 

within 30 m of the river's edge during March 20-22, 

1990.

527
Sampling date 8, February 20-22,1990

A Sampling dale 9, March 20-22, 1990

A Sampling dale 10. April 3-5, 1990

VERTICAL EXAGGERATION x 12 6 

DATUM IS SEA LEVEL

EXPLANATION

120 240 360 FEET

iQjl SANDY CLAYEY SILT

|^l SILTY CLAYEY FINE SAND

[^Jj MEDIUM SAND

Rj)1 COARSE SAND

Rs)] SILTY CLAYEY' MEDIUM SAND

Re)] SILTY CLAYEY COARSE SAND

RT)! GLACIAL TILL

 o i LINE OF EQUAL ATRAZINE CONCENTRATION-- 

Contour interval 0.1 micrograms per liter

 *- DIRECTION OF GROUND-WATER MOVEMENT 

__t_ WATER TABLE

^__£' TRACE OF SECTION SHOWN IN FIGURE 1 

 WELL SCREEN LOCATION

Figure 2. Hydrogeologic sections and distribution of atrazine in water in the alluvial aquifer at the Palisades 

site February through April 1990.

Simulation of Bank Storage

A two-dimensional ground-water-flow model 

was constructed using the U.S. Geological Survey's 

modular finite-difference ground-water flow model 

(McDonald and Harbaugh, 1988), with an option that 

allows model cells that have gone dry to become resaturated (McDonald and others, 1991). The model simulates ground-water flow in a two dimensional cross 

section of the Palisades study area along the line of 

wells labeled A-A' (fig.l and 4), which is perpendicular to the Cedar River and is oriented parallel to the 

direction of ground-water flow. The cross-sectional 

area simulated in the model is 440 m wide and 15 m 

deep and is represented by a uniform grid with 15

528
A Sampling date 8, February 20-22. 1990

A Sampling dale 9, March 20-22.1990

VERTICAL EXAGGERATION x 12 6 

DATUM IS SEA LEVEL

120 240 360 FEET

EXPLANATION

 500 LINE OF EQUAL OF EQUAL SPECIFIC CONDUCTANCE-Coutour interval 100 microsiemens 

per centimeter at 25 degrees Celsius

l(p| SANDY CLAYEY SILT

[(2)1 SILTY CLAYEY FINE SAND

RTF! MEDIUM SAND

|^1 COARSE SAND ' *" DIRECTION OF GROUND-WATER MOVEMENT

[If)] SILTY CLAYEY MEDIUM SAMD _i_ WATER TABLE

[(6)1 SILTY CLAYEY COARSE SAND A A'

[(Z)l GLACIAL TILL ' '

WELL SCREEN LOCATION

TRACE OF SECTION SHOWN IN FIGURE 1

Figure 3. Hydrogeologic sections and distribution of specific conductance in water in the alluvial 

aquifer at the Palisades site February through March 1990.

layers, each with 44 cells (fig. 4). The depth of each 

cell in the plane of the cross-section is 1 m.

The river stage was simulated using a general 

head-dependent boundary (fig. 4), and the left lateral 

boundary beneath the center of the river was simulated 

as a no-flow boundary. The mean daily river stages at 

the Palisades site during March 7-April 12, 1990, was 

used for heads in the general head-dependent boundary. The river stage from April 13-17, 1990, increased 

slightly because of direct runoff. However, the river 

stage in the model was estimated by extrapolating the 

recession curve for another 5 days, so that the movement of bank storage out of the aquifer could continue 

for a longer time period. The left-lateral model boundary beneath the river (fig. 4) is assumed to be a 

ground-water divide.

The cells in contact with the river were also 

used to simulate the 0.5 m-thick streambed sediment 

along the riverbank and bottom (fig. 4). The presence 

of a thin layer of fine-grained streambed sediment 

along the riverbank has been documented in many 

rivers (Schumm, 1960) and may affect ground-water 

flow at the Palisades site.

The right-lateral model boundary farthest from 

the river (fig. 4) is assumed to be a no-flow boundary 

because of the presence of thick deposits (about 15m) 

of glacial till in this area. The basal boundary is 

assumed to be a no-flow boundary because of the thick 

deposits (about 15 m) of glacial till that underlie the 

study area.

Recharge to the water table was treated differently in the steady state and transient simulations.

529
NOT TO SCALE

10 15

CELLS 

20 25 30 35 40 44

1

2 

3

4 

5

6

7 

8

9

10

11 

12

13 

14 

15

i i

- ©

-

Nfc - g<

II ^

-

| 

(*)

i i i

 

 . .

 

i i i i i i i

©

©

 .

©

1 . 1 1 1 1 1 1 1 1 1 1

tf

 

. 1 1 .

©

i i . i i i i i i i i i i i i

©

© '

© 

.

©

-

GC

n

z

o m

2

0

NO-FLOW BOUNDARY 

EXPLANATION

© 1 SANDY CLAYEY SILT

^1 SILTY CLAYEY FINE SAND

©1 MEDIUM SAND

Ql COARSE SAND

©1 SILTY CLAYEY MEDIUM SAND

©1 SILTY CLAYEY COARSE SAND

®| GLACIAL TILL

©| RIVER HEAD DEPENDENT BOUNDARY

AND BED SEDIMENT (WHEN SIMULATED)

Z)1 INACTIVE

'

TRACE OF SECTION SHOWN IN FIGURE 1

 OBSERVATION WELL USED IN MODEL 

CALIBRATION

Figure 4. Model grid, location of hydrogeologic subunits, and observations wells on hydrogeologic 

sections on line A-A' (fig. 1) used in simulation of bank storage.

However, the steady-state model was used to generate 

the starting heads for the transient simulation. In the 

steady-state simulation, a constant uniform recharge 

was applied in columns 5-44. In the transient simulation, this same recharge was used in columns 5-43; 

however, in column 44, the recharge was increased to 

simulate an ephemeral stream. About 19 cm of precipitation fell during March 1990 and supplied direct runoff to this ephemeral stream. Measured vertical 

hydraulic gradients in the streambed indicate that this 

stream recharges the alluvial aquifer.

Comparison of Model Simulation With Field 

Data

The model was compared to field data as 

follows:

(1) The largest horizontal hydraulic conductivities determined by the slug-test analysis and the floodwave analysis were similar to those used in the model.

(2) Daily ground-water-level measurements (a 

total of 745) made in 27 wells during March 8-April 5, 

1990 were compared in three ways to calibrate the 

model. First, the mean difference between the measured and calculated heads for the wells located in 

columns 4-27 is 3.2 cm. Additional water levels measured in the wells shown in column 37 (fig. 4) were 

used to calibrate the model but were not compared to 

the model-generated heads. These measured water 

levels were excluded because they do not lie within 

the principal water-bearing unit of the aquifer and the 

water levels in the deepest well represented in column 

37 were not at equilibrium because of sampling that 

occurred on March 20-22. Secondly, the vertical head 

differences from well nests at the Palisades site were 

compared to those in the model. The heads measured 

in the wells at 3 and 9 m depth correlate to heads in 

layers 4 and 10 of the model and are shown in figure 5. 

Thirdly, the horizontal ground-water hydraulic gradient between the wells 5 and 30 m from the riverbank 

were compared to the model-simulated hydraulic gradients in equivalent nodes.

(3) Field data show that the area within 30 m of 

the river's edge contained bank storage water and was 

characterized by low specific conductance and a high 

concentration of atrazine when compared that in the 

background (or ambient) ground water (fig. 2 and 3).

530
16 22 »

207.5

12 17

CO

DC 

UJ

UJ

2

UJ

o

z

UJ

o

UJ 

QC 

UJ

o

Q

UJ

O

_l

15 22 29

207.5 207.5

- 207.0

1990 EXPLANATION

 River stage

* Measured vertical hydraulic-head difference

D Modeled vertical hydraulic-head difference

Figure 5. Vertical hydraulic-head difference between layers 4 and 10 in model with those measured in wells screened 3 and 9 

meters below land surface at selected distances from Cedar River's edge, March-April 1990.

This area also was identified as containing bank-storage water from preliminary analysis using a using a 

particle tracking post-processing package (D.W. 

Pollock, U. S. Geological Survey, written commun., 

1993).

(4) Atrazine concentrations in samples collected 

during April 3-5, 1990 were used to estimate the areal 

extent of bank-storage water at that time. This area 

was also identified as containing bank-storage water

based on a preliminary analysis using a particle tracking post-processing package (D.W. Pollock, U. S. 

Geological Survey, written commun., 1993).

(5) The current calibrated model does not 

explain the comparatively large concentrations of atrazine at 50 and 80 m from the river's edge at a depth of 

3 m below the land surface. Additional model runs are 

needed to determine whether the volume of bank storage could have expanded up to 80 m from the river's 

edge.

531
(6) During base-flow conditions, the amount of 

ground water that directly discharges to the Cedar 

River is comparable to the discharge calculated by the 

model. Along a 48-km reach of the Cedar River, the 

average ground-water discharge from the alluvial 

aquifer to the Cedar River was indirectly measured to 

be 2.1 and 1.8 m /d/m length of river, on two different 

occasions. The model calculated ground-water discharge to be 1.9 m3 /d/m length of river.

Sensitivity Analysis

Selected input parameters were varied from the 

calibrated values to determine the effects of variations 

in these parameters on model results. Table 1 shows 

the values that were changed in the sensitivity analysis. The range of values does not represent the absolute minimum and maximum values but are believed 

to include fairly extreme values.

The results of the sensitivity analysis show that 

the accumulation and release of bank- storage water is 

most sensitive to aquifer hydraulic conductivity, aquifer storage, and recharge. However, changes made in 

these parameters caused the heads to be significantly 

different from those calculated in the calibrated model 

and from field measurements. Changes in the other 

parameters did not cause the accumulation and release 

of bank-storage water calculated by the model to differ 

significantly from what was calculated by the calibrated model, even though the heads were somewhat 

different from those measured in the field. Increasing 

and decreasing the aquifer vertical anisotropy yielded 

calculated heads that were substantially different from 

those measured in the field. The hydraulic characteristics of the confining units do not substantially affect 

bank storage or the heads within the alluvial aquifer.

The vertical hydraulic conductivity of the 0.5- 

m-thick river bottom can be 50 percent less than the 

underlying aquifer and still not appreciably affect the 

movement of bank-storage water. Increasing the riverbottom hydraulic conductivity so that movement of 

river water into the aquifer is restricted only by the 

hydraulic conductivity of the aquifer resulted in little 

change in the accumulation and release of bank-storage water.

Movement of Bank Storage Water in the 

Model

The model showed that 70 percent of the total 

bank-storage water moved through the river bottom, 

whereas the remaining 30 percent moved laterally 

through the riverbank. More bank storage water 

moved through the river bottom because the river 

bottom has a surface area that is about 10 times larger 

than the surface area of the riverbank. However, per 

unit area, the flux of bank storage water through the 

riverbank was about 4 times as large as the flux 

through the river bottom because the horizontal 

hydraulic conductivities through the riverbank are 

larger than the vertical hydraulic conductivities 

through the river bottom.

The model showed that the flux of bank-storage 

water both into and out of the river bottom is largest 

near the river's edge and decreases substantially with 

distance from the river's edge to the center of the river. 

The flux through the river bottom reached a maximum 

on March 11, 1990. On this particular day, 48 percent 

of the flux through the river bottom occurred within 

10 m of the river's edge and 71 percent of the flux 

moved through the river bottom within 20 m of the 

river's edge. The fluxes continued to decrease toward 

the center of the river.

Field measurements of the vertical and horizontal gradients of the alluvial aquifer under transient 

conditions were useful for calibrating the model. 

Model simulated hydraulic heads can closely match 

measured hydraulic heads, and yet still simulate the 

vertical movement of ground water erroneously. As an 

example, when the ground-water flux through the river 

bank is reduced to simulate streambed sediment the 

simulated hydraulic heads were similar to the measured heads. However, the vertical gradients in column 

5 were reversed from those measured in the field at an 

equivalent location.

Bank storage is very sensitive to the recharge 

component; therefore, conceptually, the area of an 

alluvial aquifer that receives the smallest amount of 

recharge from precipitation and is located farthest 

from the effects of recharge in the upland areas has the 

greatest potential for bank storage and will retain the 

bank-storage water for the longest time, all other conditions being equal. Recharge from precipitation falling directly on the flood plain only supplies a small 

percentage of the ground water that discharges to the 

river.

532
Table 1. Calibrated values and range of values used in sensitivity analyses of the ground-water flow model

Model input variable Calibrated values Range of values used

in sensitivity analysis

Confining unit horizontal hydraulic conductivity 

(meters per day)

glacial till 1.0 x 10'3 1.0 x W4 1.0 x 10'2 

clay and silt .01 .001 .1 

silty clayey sand 1.0 .1 10.0

Aquifer horizontal hydraulic conductivity 

(meters per day)

silty sand 50.0 25.0 100.0 

fine sand 1.0 0.5 2.0 

medium sand 50.0 25.0 100.0 

coarse sand 100.0 50.0 200.0

Confining unit vertical anisotropy

horizontal hydraulic conductivity/vertical hydraulic conductivity (meters per day) 

glacial till 20.0 4.0 100.0 

clay and silt 20.0 4.0 100.0 

silty clayey sand 20.0 4.0 100.0

Aquifer vertical anisotropy

horizontal hydraulic conductivity/vertical hydraulic conductivity (meters per day) 

fine sand ' 10.0 2.0 50.0 

medium sand 10.0 2.0 50.0 

coarse sand 10.0 2.0 50.0 

silty sand 10.0 2.0 50.0

Recharge (meters per day) columns 1-43 1.04 x 10'3 .52 x 10'3 2.08 x 10"3 

stream-column 44 .03 .01 .09

Aquifer storage (percent by volume)

Specific yield-clean sand units .2 .1 .3 

Specific yield-silty sand units .05 .05 .15 

Specific storage 1.0 xlO"4 1.0 xlO"5 1.0 xlO"3

Confining unit storage (percent by volume)

Specific yield .01 .01 .15 

Specific storage 1.0 x 10'3 1.0 x 10"4 1.0 x 10'2

River-bottom sediment

Horizontal and vertical hydraulic

conductivity (0.5-meter thick zone)(meters per day)5.0 .5 200.0

River-bottom and bank sediment

Changed the horizontal and vertical

hydraulic conductivity (0.5-meter thick zone)

in the river bottom and riverbank at

the same time.

River bottom /riverbank (meters per day) 5.0/200.0 200.0/0.5 200.0/5.0

533
The discharge of bank-storage water from the 

alluvial aquifer can continue for a long period of time. 

By extrapolation, the model showed that it would take 

about 5 weeks for bank-storage water to discharge 

from the alluvial aquifer at the Palisades study area 

after the river stage peaked on March 17, 1990.

REFERENCES

Cooper, H.H., and Rorabaugh, M.I., 1963, Ground-water 

movements and bank storage due to flood stages in surface streams: U.S. Geological Survey Water-Supply 

Paper 1536-J, p. 343-366.

Dever, R.J.Jr., and Cleary, R.W., 1979, Unsteady-state, twodimensional response of leaky aquifers to stream stage 

fluctuations: Advances in Water Resources, v. 2, no. 1, 

p. 13-18.

Gill, M.A., 1985, Bank storage characteristics of a finite 

aquifer due to sudden rise and fall of a river level: 

Journal of Hydrology, v. 76, p. 133-142.

Hunt, Bruce, 1990, An approximation for the bank storage 

effect: Water Resources Research, v. 26 no. 11, p. 

2769-2775.

Kondolf, G.M., Maloney, L.M. and Williams, J.G., 1987, 

Effects of bank storage and well pumping on base flow, 

Carmel River, Monterey County, California: Journal of 

Hydrology, v. 91, p. 351-369.

McDonald, M.G. and Harbaugh, A.W., 1988, A modular 

three-dimensional finite-difference ground-water flow 

model: U.S. Geological Survey Techniques of WaterResources Investigations, book 6, chap. Al, 586 p.

McDonald, M.G., Harbaugh, A.W., Orr, B.R., and Acherman, D.J., 1991, A method of converting no-flow cells 

to variable-head cells for the U.S. Geological Survey 

modular finite-difference ground-water flow model: 

U.S. Geological Survey Open-File Report 91-536, 

99 p.

Moench, A.F., Sauer,V.B., and Jennings, M.E., 1974, Modification of routed streamflow by channel loss and base 

flow: Water Resources Research, v. 10, no. 5, p. 963- 

968.

Morel-Seytoux, H.J.,1975, A combined model of water 

table and river stage evolution: Water Resources 

Research, v. 11, no. 6, p. 968-972.

Newsom, J.M. and Wilson, J.L., 1988, Flow of ground

water to a well near a stream-effect of ambient groundwater flow direction: Ground Water, v. 26, no. 6, 

p. 703-711.

Pinder, G.F., Bredehoeffi, J.K., and Cooper, H.H., Jr., 1969, 

Determination of aquifer diffusivity from aquifer 

response to fluctuations in river stage: Water Resources 

Research, v. 5, no. 4, p. 850-855.

Pinder, G.E., and Sauer, S.P, 1971, Numerical simulation of 

flood wave modification due to bank storage effects: 

Water Resources Research, v. 7, no. 1, p. 63-70.

Rorabaugh, M.I., 1963, Estimating changes in bank storage 

and ground-water contribution to streamflow: International Association of Scientific Hydrology, no. 63, 

p. 432-441.

Schumm, S.A., 1960, The shape of alluvial channels in relation to sediment type: U.S. Geological Survey Professional Paper 352-B, 13 p.

Sharp, J.M.,1977, Limitations of bank-storage model 

assumptions: Journal of Hydrology, v. 35, p. 31-47.

Squillace, P.J., Liszewski, M.J., and Thurman, E. M., 1992, 

Agricultural chemical interchange between ground 

water and surface water, Cedar River basin, Iowa and 

Minnesota A study description: U.S., Geological 

Survey Open-File Report 92-85, 26 p.

Squillace, P.J., Thurman, E.M., and Furlong, E.T., 1993, 

Ground water as a nonpoint source of atrazine and 

deethylatrazine in a river during base-flow conditions: 

Water Resources Research, v. 29, no. 6, p. 1719-1729.

Thurman, E.M., Meyer, Michael., Pomes, Michael, Perry, 

C.A., and Schwab, A.P., 1990, Enzyme-linked immunosorbent assay compared with gas-chromatography/ 

mass spectrometry for the determination of triazine 

herbicides in water: Analytical Chem., v. 62, p. 2043- 

2048.

534
Fate of Alachlor, Atrazine, and Bromide 

in a Corn Plot near Topeka, Kansas

By DA Eckhardt1 , R.J. Wagenet2, E.M. Thurman3 , and PL. Barnes4

Abstract

In April 1991, a bromide tracer and the 

herbicides atrazine and alachlor were applied to a 

0.75-hectare (ha) field plot of Eudora silt loam in 

the Kansas River Valley near Topeka, Kansas. 

The chemicals were immediately disked into the 

top 7.5 cm of soil, and the plot was planted with 

corn. Composite soil cores were collected from 12 

contiguous cells in the plot, each 15 meters (m) by 

15 m, at 15-centimeter (cm) depth intervals to 

90 cm. Cores were collected immediately after 

application, then after 30,60, and 120 days. Mean 

atrazine losses from the soil profile, relative to the 

immediate postapplication core results, were 75 

percent at day 30, 92 percent at day 60, and 96 

percent at day 120. Mean alachlor losses were 93 

percent at day 30, 99 percent at day 60, and nearly 

100 percent by day 120. The ratio of desethylatrazine metabolite to atrazine concentrations steadily 

increased in the top 15-cm soil layer during the 

120-day period. An analysis of the soil-water 

budget and bromide distribution in the soil profile 

indicates that downward movement of the chemicals was restricted by a lack of significant downward soil-water flux during the growing season, 

when evapotranspiration was greatest. Loss of the 

herbicides from the soil profile is attributed 

mainly to degradation and transformation, and 

loss of bromide is attributed mainly to uptake by 

the crop. Low but persistent herbicide concentrations 120 days after application indicate that

1 U.S. Geological Survey, Ithaca, NY 

2Cornell University, Ithaca, NY 

3 U.S. Geological Survey, Lawrence, KS 

4Kansas State University, Manhattan, KS

herbicide half lives are not constant and that 

herbicide residues could be tightly bound within 

solid-phase organic carbon and unavailable for 

biotic degradation.

INTRODUCTION

Agricultural herbicides are applied to soils 

throughout the world in increasing quantities for the 

control of weeds and grasses that would otherwise 

compete with crops. About 29 million kilograms of 

atrazine and about 25 million kilograms of alachlor are 

applied annually to soils in the United States, mainly 

in corn and soybean growing regions (Gianessi and 

Puffer, 1991). The ecological fate of these chemicals 

in soil is of interest to hydrologists, agriculturalists, 

and the public.

Agricultural herbicides are usually sprayed in 

liquid form just before spring planting, when soils are 

relatively moist and evapotranspiration rates are near 

minimum. Their fate is largely controlled by hydraulic 

properties of the soil and by weather conditions 

directly after application. When soil erosion is minimized and the herbicides are incorporated into shallow 

soil layers, the bulk of the chemicals will partition to 

organic material in the soil and will decay within the 

root zone as the crop matures. In some soils, however, 

especially under very wet spring conditions directly 

after application, soil water can flow relatively quickly 

in macropores through shallow soil zones and transport herbicide residues to deeper zones, thus minimizing residence time in zones of most active decay and 

increasing the potential for leaching into the ground 

water. Under dryer conditions, when soil-moisture 

content is less than field capacity and soil-water flow 

is commonly negligible, active herbicides are 

adsorbed and decay in shallow soil zones, where the 

organic-carbon content is generally the highest.

535
In 1991, the U.S. Geological Survey conducted 

a field study of alachlor and atrazine in unsaturated 

soils to document the spatial differences in chemical 

processes and transport rates that affect agricultural 

herbicides in soil. The study was intended to provide 

data to characterize the spatial and temporal variations 

in processes and rates in a numerical transport model 

of herbicide fate and movement within a 0.75-ha corn 

plot.

This paper describes the fate of atrazine and 

alachlor herbicides and a bromide tracer applied to the 

0.75-ha silt-loam corn plot in Kansas during the 1991 

growing season. It presents (1) results of analyses of 

soil cores taken throughout the plot at specified time 

and depth intervals after application and (2) describes 

results of preliminary model simulations of advectivedispersive solute movement that reflect the sorption 

and decay of herbicides in the unsaturated silt-loam.

METHODS

The field site is in the Kansas River Valley near 

Topeka, Kansas (fig. 1). The soil is a mesic profile 

with a 30-cm silt-loam A horizon (1 percent organic 

carbon (OC) over a silt-loam B horizon (0.4 percent 

OC). Depth to water table is about 6 m. Corn and soybean crops are rotated annually at the site. The 0.75 ha 

site consists of two adjacent rows of six 15-m by 15-m 

plots, each bounded by neutron-logging tubes for 

monitoring soil-water content.

In April 1991, alachlor (1.5 kg/ha active 

ingredient) and atrazine (1.5 kg/ha) herbicides and 

bromide tracer (70 kg/ha) were applied, and corn was 

planted. Soil cores were collected from six locations 

per plot cell and composited as one sample for each 

cell. Core samples were collected just before chemical 

application, immediately after application, and 30, 60, 

and 120 days thereafter in 15-cm lengths at depth 

intervals of 0 to 15 cm, 15 to 30 cm, 45 to 60 cm, and 

75 to 90 cm. The soil samples representing the four 

depths in each of the 12 cells were split three ways, 

and each subsample was analyzed for gravimetric 

soil-moisture content, bromide content, and herbicide 

content. Bromide was extracted through a washing 

procedure with 0.005M CaSC>4 solution and analyzed 

by ion-chromatography. Herbicide residues in soil 

samples were analyzed by C-18 solid-phase extraction 

and gas chromatography with mass spectrometry, as 

described by Mills and Thurman (1992).

Herbicide and tracer movement and fate were 

simulated by a one-dimensional numerical model 

(Wagenet and Hutson, 1992) that solves the Richards 

transient-state unsaturated flow equation and the 

advective-dispersive solute-transport equation with 

sorption, transformation, and root uptake by the growing crop. Soil-water movement was simulated from 

the hydraulic characteristics of soil as estimated 

through 20 onsite unit-gradient drainage experiments 

(Eckhardt and Barnes, 1991) and moisture-retention 

analysis of 160 soil cores collected from pits at the 

field site. The soil-water flow simulation was calibrated by adjusting evapotranspiration rates over 

time to reflect bromide distribution in soil cores and 

the changes in soil moisture content observed at 20 

neutron-logging tubes that border each of the 12 coresampling cells (fig. 1). Then, herbicide-transformation 

rates and sorption coefficients were adjusted to match 

observed herbicide mass balances and distributions in 

the soil profile over the 120-day growth period. Water 

and chemical movement and mass balances in an 

average soil profile 1 m thick were simulated. Growth 

of corn roots and the transpiration loss of water from 

the soil profile during the growing season were simulated with a root-growth model (Davidson and others, 

1964) that represents water and chemical extraction 

from the soil profile as a function of root depth and 

density and time.

RESULTS AND DISCUSSION

The soil profile received 713 mm of water 

during 28 rainstorms and 7 spray irrigations during the 

modeled period of March through September 1991, 

which was slightly dryer than normal. Total evaporation from the profile (estimated to be 326 mm) 

occurred mainly in June, before the crop cover became 

significant. As the corn grew and roots penetrated 

deeper into the profile, transpiration (414 mm) became 

the main cause of water loss. Total estimated evapotranspiration (740 mm) during the 120-day growing 

season more than offset the rain and irrigation 

(713 mm) and slightly depleted available soil water in 

the profile. The net downward movement of water and 

chemicals was thereby minimized, and drainage loss 

by deep percolation was virtually zero.

536
Figure 1. Location of measurement sites at the field plot.

Bromide

Bromide, which had been applied to the plot 

just before planting, was used as a tracer to indicate 

average downward soil-water velocity during the 

growing season and the degree of hydrodynamic 

dispersion of solutes in the soil profile. Average 

downward soil-water velocity in the soil was less than 

5 mm/d during the 120-day growing season; dispersivity was estimated to be 100 mm. During the 120

40°-'

02C 95 C

LOCATION OF 

KANSAS RIVER VALLEY 

EXPERIMENT FIELD 

AT TOPEKA

KANSAS

LOCATION MAP

EXPLANATION

| NEUTRON-ACCESS TUBE

GROUND-WATER MONITORING 

WELL

P EXCAVATION PIT 

H| SOIL-CHEMICAL CORE SITE

0 15 METERS

50 FEET

days after application, bromide was detected in all 

cores above the 30-cm depth, in 6 cores from the 

60-cm depth, and in only 1 core from the 90-cm depth 

(1.4 mg/kg). The bromide distribution (fig. 2A) supports water-budget estimates that indicate that most of 

the water entering the soil as rain or irrigation was 

removed by evapotranspiration during the growing 

season and that drainage from the profile was minimal.

The mass of the applied bromide tracer was 

not conserved in the soil profile because dissolved

537
A. BROMIDE

30 DAYS AFTER APPLICATION 60 DAYS AFTER APPLICATION 120 DAYS AFTER APPLICATION

0 

20

40

60

80 

100

 £*» 

t/X^

 OBSERVED 

__ SIMULATH)

D 5 10 15 20 25 3

0 

20

40

60

80

100

0 (

;i( - >/ ^

 OBSERVED 

__ SIMULATED

) 5 10 15 20 25 3

46°f/

80

100 

0

1* * OBSERVED

1 __ SIMULATH)

0 5 10 15 20 25 3C

CONCENTRATION IN SOIL, IN MILLIGRAMS PER KILOGRAM

I

c* 

a

a s 

£= °

ra 20

z

3 40

u

U. 60

BELOW LAND SL 8 g

I. ALACHLOR

0 DAYS AFTER APPLICATION

 2J*. f

 OBSERVED 

__ SIMULATED

} 20 40 6

CONCENTRAT1

C

0

20

40 

60

80

100 

0 (

ON IN

<Q DAYS AFTER APPLICATION

}-1. ' '

i

 OBSERVED

__ SIMULATH)

f f

) 20 40 6

SOIL, IN MICROGRAMS P

1;

0 

20

40 

60

80

100 

0 G

ERKIL

JO DAYS AFTER APPLICATION

i i

t

 OBSERVH)

__ SIMULATED

20 40 60

OGRAM

| 

1 C. ATRAZINE

3 

0

20

40

60

80 

ton

0 DAYS AFTER APPLICATION

 ^mi£^f**» <* f~*^

 OBSERVED 

__ SIMULATED

6

0

20

40

60

80 

100

0 DAYS AFTER APPLICATION

~1. .

r

-

 OBSERVED 

__ SIMULATED

12

20

40

60

80 

100

 0 DAYS AFTER APPLICATION

 OBSERVED 

__ SIMULATED

 \ 1 1 1 " " 

0 20 40 60 80 100 120 140 0 20 40 60 80 100 120 140 0 20 40 60 80 100 120 140

CONCENTRATION IN SOIL, IN MICROGRAMS PER KILOGRAM

Figure 2. The depth distribution of observed and simulated chemical concentrations in soil cores at three 

times after application in 1991: A. Bromide tracer, B. Alachlor, and C. Atrazine.

538
bromide was removed through uptake by the corn. At 

harvest in August 1991, less than 5 percent of the 

applied bromide remained in the soil. Analyses of corn 

grain and plant tissue indicate that more than half the 

applied bromide was taken up by the corn (table 1). 

Most of the bromide was concentrated in the stalks 

and leaves. The rest was probably in plant roots, which 

were not analyzed.

Bromide tracer in the simulated soil profile was 

lost mainly by crop uptake, which corresponded to 

root-growth and transpiration rates. Bromide mass in 

the simulated profile was relatively constant after 

application until about late June (60 days), when root 

development began to extract increased amounts of 

water and bromide from the rooting zone (fig 3). The 

discrepancy between low observed concentrations in 

soil cores 60 days after application and the higher simulated concentrations (fig. 2A) indicate that the rootgrowth model underpredicted the penetration and

Table 1. Bromide uptake by the corn crop after 120-day 

growing season, 1991

BROMIDE MASS IN PLANTS AT HARVEST:

(in milligrams per kilogram)

PLOT CORN GRAIN STALK AND LEAVES

A

B

C

D

MEAN

160

160

140

150

153

2,180

2,620

2,010

2,130

2,235

BROMIDE UPTAKE:

(in kilograms per hectare)

PLOT CORN GRAIN STALK AND LEAVES

A 

B 

C

D

MEAN

1.6

1.7 

1.7 

1.6 

1.65

38

41

34

29

36

density of roots in the soil profile between days 30 and 

60. An overprediction both in the soil-water content 

and bromide concentrations in the 15- to 30-cm layer 

on day 60 is further evidence that the roots at the field 

plot grew faster than the growth model predicted. 

Observed and simulated water contents at lower 

depths were in close agreement at all times.

Herbicides

Mean alachlor losses, relative to the immediate 

postapplication core results, were 93 percent at day 30, 

99 percent at day 60, and nearly 100 percent by day 

120. The partition coefficient (Koc) of alachlor was 

estimated to be 340 L/kg; therefore, its vertical movement in the soil profile was mainly within the top 15 

cm (fig. 2B). Alachlor degraded relatively quickly 

within the first 60 days in the shallow soil profile (0 to 

15 cm), where biological-decay activity is greatest and 

where alachlor readily partitions to the solid-phase 

organic carbon. During the first 60 days, the estimated 

half life of alachlor in the upper 15-cm soil layer was 

11 days.

Mean atrazine losses from the soil profile, 

relative to the immediate postapplication core results, 

were 75 percent at day 30,92 percent at day 60, and 96 

percent at day 120. Atrazine moved slightly deeper in 

the soil profile (0 to 30 cm) than alachlor, because it is 

adsorbed less readily than alachlor (fig. 2C). The Koc 

of atrazine was estimated to be 170 L/kg, one-half that 

for alachlor. Atrazine degraded more slowly than 

alachlor and remained in the top 30 cm of soil at concentrations nearly twice those of alachlor. During the 

first 60 days, the estimated half life of atrazine in the 

upper 30-cm soil layer was 23 days.

The atrazine metabolite desethylatrazine (DEA) 

was formed in the soil through loss of the ethyl functional group by the bacterial action. DEA was not 

applied to the plot, nor was it detected in soil cores 

immediately after application. It reached its maximum 

concentration of about 3 (ig/kg 30 to 60 days after 

application and was detected in nearly all samples 

from shallow soil layers (0-30 cm), where the atrazine 

was concentrated. The ratio of DEA to atrazine 

increased steadily after application for 60 days as 

DEA formed and atrazine was lost through hydrolysis 

and biodegradation (Adams and Thurman, 1991). A 

second atrazine metabolite, desisopropylatrazine, was 

detected infrequently at concentrations near the detection limit of 0.05 (J.g/kg.

539
80

0)

 oc

D

< o

S CO 

iM OC

9 S 

o

LU

a.

60

40

20

Application 30-days 60-days

I I

120-days

20

0) 

15 lo:

C5 "J

C/f LU 10 cn DC

LU

DC 

CD

APR MAY JUNE JULY

1991

AUG SEPT

ATRAZINE

EXPLANATION 

-- ALACHLOR BROMIDE

Figure 3. Simulated mass of bromide tracer, alachlor, and atrazine in the soil profile during the 1991 growing season.

Alachlor and atrazine in the simulated soil 

profile were lost mainly through degradation in 

shallow soil layers during the first 60 days (fig. 3), 

when biological activity was greatest and herbicide 

residue was sufficient for bacterial activity. After 60 

days, however, herbicide half lives apparently 

increased as herbicide concentrations decreased, and 

herbicide residues in soil cores remained, although at 

relatively low concentrations, 120 days after application (fig. 2). Residues of alachlor were detected 

120 days after application in 75 percent of the cores 

from above 30-cm depths (maximum 10 |j,g/kg); 

residues of atrazine were detected in all samples from 

above 30-cm depths (maximum 16 |J.g/kg). These 

low but persistent concentrations indicate that the 

herbicide residues could be tightly bound within solidphase organic carbon and unavailable for bacterial 

degradation. The uptake of herbicides by the growing

540
crop is estimated to be negligible because herbicides 

were largely degraded or partitioned to the solid phase 

by the time of significant root development.

SUMMARY AND CONCLUSIONS

In April 1991, bromide tracer and the herbicides 

atrazine and alachlor were applied to a 0.75-ha field 

plot of Eudora silt loam in the Kansas River Valley 

near Topeka, Kansas. The chemicals were immediately disked into the top 7.5 cm of soil, and the plot 

was planted with corn. Composite soil cores were collected from 12 contiguous cells in the plot, each 15 m 

by 15 m, at 15-cm intervals to the 90-cm depth.

Movement and fate of the herbicide and tracer 

were simulated with a one-dimensional numerical 

model that solves the Richards equation of unsaturated 

transient-state flow and the advective-dispersive 

solute-transport equation that describes sorption, 

transformation, and root uptake by the growing crop. 

An analysis of the soil-water budget and bromide 

distribution in the soil profile indicates that downward 

movement of the chemicals was restricted by a lack of 

significant downward soil-water flux during the growing season, when evapotranspiration was greatest. 

Most of the water that entered the profile as rain or 

irrigation was lost through evapotranspiration during 

the 120-day growing season; drainage loss by deep 

percolation was close to zero.

Bromide tracer mass was not conserved in the 

soil profile because dissolved bromide was removed 

from the soil through uptake by the corn. At harvest 

in August 1991, less than 5 percent of the applied 

bromide remained in the soil. Results of bromide 

analysis of corn grain and plant tissue show that more 

than half of the applied bromide was taken up by the 

corn crop, and the balance of 45 percent of the bromide is assumed to have been in the roots, which were 

not analyzed.

Alachlor degraded relatively quickly within the 

first 60 days after application and remained mostly in 

the shallow soil profile (0 to 15 cm), where bacterial 

activity is greatest and where alachlor readily partitions to the solid-phase organic carbon. The estimated 

alachlor half life in the upper 15 cm through the first 

60 days after application was 11 days. Atrazine moved 

slightly deeper (0 to 30 cm) than alachlor because it 

partitions less readily than alachlor, and it degraded 

less quickly; its estimated half life in the upper 30 cm 

through the first 60 days after application was 23 days. 

After 60 days, herbicide half lives apparently 

increased as herbicide concentrations decreased, and 

herbicide residues remained at relatively low concentrations 120 days after application. The low but persistent residue concentrations indicate that the herbicides 

could be tightly bound within solid-phase organic carbon and unavailable for bacterial degradation.

REFERENCES

Adams, C.D., and Thurman, E.M., 1991, Formation and 

transport of desethylatrazine in the soil and vadosc 

zone: Journal of Environmental Quality, v. 20, no. 3, 

p. 540-547.

Davidson, J.M., Graetz, P.S., Rao, P.S.C., and Selim, H.M., 

1964, Simulations of nitrogen movement, transformation, and uptake in plant root zones: U.S. Environmental Protection Agency, Report 600/3-78-029.

Eckhardt, D.A., and Barnes, P.L., 1991, Spatial variability 

in hydraulic conductivity of a silt-loam soil near 

Topeka, Kansas, in Mallard, G.E., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, California March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 210-213.

Gianessi, L.P., and Puffer, Cynthia, 1991, Herbicide use in 

the United States: Washington, D.C., Resources for the 

Future, Quality of the Environment Division, 128 p.

Mills, M.S., and Thurman, E.M., 1992, Mixed-mode

isolation of triazine metabolites from soil and aquifer 

sediments using automated solid-phase extraction: 

Analytical Chemistry, v. 64, p. 1985-1990.

Wagenet, R.J., and Hutson, J.L., 1992, LEACHM ~ Leaching estimation and chemistry model: Ithaca, NY, 

Cornell University, Soil, Crop, and Atmospheric 

Sciences Research Series 92-3, 125 p.

541
542
Effects of Farming Systems on Ground-Water Quality

at the Princeton, Minnesota, Management

Systems Evaluation Area, 1991

By Matthew K. Landon 1 , Geoffrey N. Delin 1 , John /\, Lamb2 , 

Robert H. Dowdy3 , and James L. Anderson2

Abstract

The 65-hectare Management Systems Evaluation Area (MSEA) near Princeton, Minnesota is 

one of five primary MSEA's in the Midwest Cornbelt. The Princeton MSEA program is a cooperative research effort among the U.S. Department of 

Agriculture-Agricultural Research Service, the 

University of Minnesota Soil Science Department, 

and the U.S. Geological Survey. The primary 

program objective is to evaluate the effects of 

ridge-tillage practices in corn and soybean farming systems on ground-water quality in a sandplain setting. A sweet-corn and potato farming 

system is also being evaluated. Analyses of 

samples collected at the beginning of the study 

during April 1991 indicated that some contamination of ground water with nitrate nitrogen (nitrateIs!), atrazine, and atrazine metabolizes had already 

occurred before implementation of the MSEA 

farming systems.

Chloride in potash fertilizers applied to the 

MSEA cropped areas in April 1991 was used as a 

tracer for ground water affected by the MSEA 

farming systems. Chloride reached ground water 

beneath all of the cropped areas. Analyses of 

ground-water samples collected during June, 

August, and December 1991 and during April 

1992 indicated that concentrations of chloride in 

the upper meter of the saturated zone beneath and 

downgradient of the cropped areas were 20 to

'U.S. Geological Survey, Mounds View, Minn.

2University of Minnesota, Department of Soil Science, 

St. Paul, Minn.

3U.S. Department of Agriculture, Agricultural Research 

Service, St. Paul, Minn.

50 mg/L (milligrams per liter) compared with 

background concentrations of 2 to 19 mg/L.

Because the nitrogen and potash fertilizer 

application rates on the potato-cropped area were 

1.4 and 4.5 times greater, respectively, than on the 

other cropped areas, nitrate-N and chloride 

concentrations were greatest beneath the potatocropped area. Nitrate-N concentrations in the 

upper meter of the saturated zone beneath the 

potato-cropped area (22-47 mg/L) were greater 

than concentrations measured upgradient (6-18 

mg/L) from the cropped area. Nitrate-N concentrations were similar upgradient, beneath, and downgradient of the other cropped areas. Thus, 

additions of nitrate-N as a result of MSEA activity 

were difficult to distinguish from background 

concentrations. Concentrations of nitrate-N in the 

saturated zone across the entire MSEA typically 

decreased from 15 to 25 mg/L near the water table 

to 5 to 10 mg/L two meters below the water table.

Atrazine and atrazine metabolites de-ethylatrazine (DEA) and de-isopropylatrazine (DIA) 

were detected in 75, 97, and 76 percent, respectively, of the 124 wells sampled at the Princeton 

MSEA during the four sampling periods from June 

1991 through April 1992. The concentrations of 

these compounds in the saturated zone varied 

temporally. Atrazine was detected in 51 percent of 

the 315 ground-water samples collected in 

concentrations ranging from less than 0.01 jag/L 

(micrograms per liter) to 0.54 (ag/L. The median 

concentration of atrazine was the detection limit of 

0.01 |ng/L. DEA was detected in 74 percent of the 

samples at concentrations ranging from less than 

0.03 ng/L to 1.00 |ng/L. The median concentration 

of DEA was 0.08 ng/L. DIA was detected in 10 

percent of the samples in concentrations ranging543
from less than 0.06 |ug/L to 0.66 |ug/L. The median 

concentration of DIA was less than the detection 

limit of 0.06 |ug/L. Alachlor was detected in 2 

percent of the samples. Concentrations of DEA in 

April 1992 increased with depth from less than 

0.06 |ng/L near the water table to 0.08 to 0.30 |ug/ 

L between 1 and 2 m below the water table. This 

trend with depth was not noticeable prior to April 

1992 and could indicate that water recharging the 

aquifer beneath the cropped areas during 1991 had 

lower DEA concentrations than older ground 

water.

Because some of the applied fertilizers and 

herbicides and herbicide metabolites existed in 

shallow ground water at the onset of this study, it 

is difficult to assess the affects of MSEA farming 

systems on water quality without additional data 

collected over several years. None of the detections of herbicides or herbicide metabolites can be 

conclusively linked to the MSEA farming systems 

at this time. Because some atrazine and atrazine 

metabolites were detected in ground water that 

was recharged through the MSEA cropped areas, 

however, the possibility that some of these chemicals leached from the MSEA farming systems to 

ground water cannot be discounted. Use of conservative tracers like chloride and bromide to delineate annual recharge fronts beneath the MSEA 

cropped areas should make it easier to identify the 

affects of MSEA farming systems on groundwater quality as older ground water affected by 

previous farming practices moves horizontally and 

vertically away from the cropped areas.

INTRODUCTION

The northern cornbelt sand-plains Management 

Systems Evaluation Area (MSEA) program is a multiagency, multistate initiative to evaluate the effects of 

farming systems on ground-water quality in a sandplain area in Minnesota and at satellite areas in North 

and South Dakota, and Wisconsin (Delin and others, 

1992a). The primary objective of this program is to 

evaluate the effects of ridge-tillage practices in a corn 

and soybean farming system on ground-water quality. 

The Minnesota MSEA is located in the Anoka Sand 

Plain near the town of Princeton (fig. 1). The Princeton 

MSEA program is a cooperative study primarily

between the U.S. Department of Agriculture-Agricultural Research Service, the University of Minnesota 

Soil Science Department, and the U.S. Geological Survey. The Minnesota Pollution Control Agency and the 

Department of Geology at the University of Minnesota 

are also cooperating in the evaluation of ground-water 

quality at the MSEA. This paper describes the effects 

of farming systems on ground-water quality at the Princeton MSEA during the first year of implementation 

(1991).

DESCRIPTION OF THE PRINCETON, 

MINNESOTA MSEA

Topographic relief at the 65-ha area is less than 

3 m. Mostly, the unsaturated zone consists of fine-tomedium-grained sand and the saturated zone consists 

of medium-to-coarse grained sand. The unsaturated 

and saturated zones contain discontinuous layers of silt 

and very-fine-grained sand up to 20 cm thick. A clayey 

till underlies the surficial aquifer (the till is less permeable than the aquifer). During 1991, the average depth 

to the water table was about 3.6 m below land surface, 

and the saturated thickness ranged from 4 to 16m 

across the study area. The average horizontal saturated 

hydraulic conductivity is about 0.04 cm/s. Ground 

water generally moves from west to east (fig. 1) at an 

average rate of about 8 cm/d. The method of hydrograph analysis of Rasmussen and Andreason (1959) 

was used to calculate ground-water-recharge rates. 

Recharge generally ranged from 10 to 20 cm in the 

study area during 1991.

In 1991, five 1.8- to 2.6-ha cropped areas were 

aligned parallel to the predominant direction of 

ground-water flow at the study site (fig. 1). These areas 

were used to evaluate three cropping systems (Anderson and others, 1991): (1) a corn-soybean annual rotation under ridge (conservation) tillage, split nitrogen 

(N) application, N credit for legumes, and banding of 

herbicides (application only over row and amounting to 

one-third of the broadcast application amount) (fig. 1, 

field corn and soybeans); (2) a sweet corn-potato 

annual rotation with conventional full-width (disk or 

chisel) tillage, banding of herbicides for sweet corn, 

and broadcast application of herbicides for potatoes 

(fig. 1, sweet corn and potatoes); and (3) field corn in 

consecutive years (continuous corn) under conventional full-width tillage and broadcast application of 

herbicides (fig. 1, continuous corn). Application rates 

of agricultural chemicals in 1991 are shown in table 1.

544
Linear-move 

Irrigation system

\.

750 1000 FEET

A

T»

2 METERS

4,

/

0 $/l \ \ ~~"~4n 42'3 

"0*12.2 X *X "-- Jiaa

D 13.4 "X °'~ 37/t

D 15.7 Potatoes ~~~ 14IF

YWater Table \

D _/' i 

J __ -~"3~8~.3 0 / D

0 352 D/ if

0 23.5 D 30 ,0 

IT ... 32.3'nX-^D

A

24.7

24.7

174 

11.7

300 METERS

EXPLANATION

Cropped area and 1991 crop type 

Research area, including cropped areas

Direction of ground-water 

flow, January, 1991

A A Line of section

.. 20- Approximate line of equal chloride

concentration, interval in

10 milligrams per liter

D 13.4

The research area is 

located In the northeast 

quarter of section 18, 

township T35N, range R26W.

Data point, number indicates diloride 

concentration, in milligrams per liter

Observation well, number indicates 

number of wells at site

Multiport well 

Building

Figure 1. Layout of the Princeton, Minnesota,Management Systems Evaluation Area 

and example illustrating contouring of chloride concentrations along a cross 

section beneath the potato-cropped area, April 1992.

545
Table 1. Application rates of agricultural chemicals on crops at the Princeton, Minnesota, Management Systems 

Evaluation area, 1991

[All application rates are broadcast rates; band, application of herbicide only over row, amount is one-third of the total amount per hectare 

for broadcast application, kg/ha, kilograms per hectare; --, not applied]

Potatoes Soybeans Sweet corn Field corn Continuous corn

Total Total Total Total Total

Agricultural rate Month(s) rate Month(s) rate Month(s) rate Month(s) rate Month(s)

chemical kg/ha applied kg/ha applied kg/ha applied kg/ha applied kg/ha applied

Fertilizer

Potash 504 April 112 April 112 April 112 April 112 April

(potassium

chloride)

Nitrogen 224 April -- -- 157 April 157 April 157 April

and June and June and June and June

Herbicide

Atrazine

Alachlor

Metribuzin 0.56 May

Metolachlor 1.12 May

--

2.25 May

band

0.56 May

band

--

1.70

band

2.25

band

--

May

May

--

1.70 May 1.70

band

2.25 May 2.25

band

__

--

May

May

--

A buffer area around and between the cropped areas 

(fig. 1) was planted with a mixture of timothy and 

smooth brome grass. Agricultural chemicals were not 

applied in this buffer area. The entire 65-ha field was 

planted in alfalfa during 1981-89 and in corn during 

1990, prior to the implementation of the MSEA fanning systems in spring 1991. Detailed records of fanning 

practices and chemical applications during this period 

were not available.

The ground-water-quality sampling network at the 

Princeton MSEA consists of 29 observation wells and 

22 multiport wells (fig. 1). In addition, 14 observation 

wells are located off the 65-ha field (not shown in 

fig. 1). Observation wells were used to measure water 

levels monthly and to determine background concentrations of agricultural chemicals. These wells are constructed of 5.1-cm inside-diameter (i.d.) galvanizedsteel or polyvinyl chloride (PVC) casing with 0.6-mlong screens located at the water table or 0.15-m-long 

screens installed deeper in the aquifer. The multiport 

wells are located 21 m upgradient, in the middle, at the 

downgradient edge, and 25 m downgradient (slightly

546

less than the distance ground water travels in 1 year) of 

each cropped area (fig. 1). Each multiport well consists 

of six, 0.6-cm i.d. stainless-steel tubes housed in a 

5.1-cm i.d. PVC casing; each tube has a 3-cm-long 

screened interval (port) which is external to the PVC 

casing. The sampling ports were installed at 0.5-m 

intervals with the uppermost port 0.5 m above the 

water table to allow sample collection if the water table 

rose.

Water samples were collected during 1991 at four 

different times from all multiport wells, selected onsite 

observation wells, and Battle Brook (fig. 1). These 

samples were collected before the application of agricultural chemicals (April), twice during the growing 

season (June and August), and in the late fall after 

crops were harvested (December). Sample-collection 

and laboratory-analysis quality-assurance/quality-control protocols were followed (MSEA Steering Committee, written commun., 1991). Specific conductance, 

pH, temperature, dissolved oxygen concentration, and 

oxidation-reduction potential of ground water were 

measured during pumping at each sampling site.
Water samples were collected once these properties 

stabilized. Alkalinity titrations were performed the 

same day in the field. Water samples were collected and 

analyzed for dissolved major cations and anions, nutrients, and selected herbicides and herbicide metabolites 

(atrazine, de-ethylatazine (DEA), de-isopropylatrazine 

(DIA), alachlor, chloroalachlor, 2,6-diethylanaline, 

metolachlor, and metribuzin). Gas chromatography/ 

mass spectrometry was used to determine herbicide 

concentrations (P.O. Capel, U.S. Geological Survey, 

written commun., 1991).

GROUND-WATER QUALITY, 1991

Concentrations of selected water-quality constituents in the upper 2 m of the saturated zone were contoured along a cross section extending beneath each 

MSEA cropped area (figs. 1-4). The contouring is 

somewhat speculative because of the wide horizontal 

spacing of the data points. Contour intervals were 

selected that best emphasized contrasts in constituent 

concentrations along each cross section.

Because potash (potassium chloride) fertilizer was 

applied only to the cropped areas (fig. 1 and table 1) 

and because chloride moves conservatively with 

ground water, chloride was used as a tracer of water 

recharged through the cropped areas. Where concentrations of chloride and nitrate nitrogen (nitrate-N) were 

clearly greater beneath a cropped area than upgradient, 

contours were drawn to reflect the interpretation that 

the source of the higher concentrations was recharge of 

chemicals from the overlying cropped area. For example, beneath the potato cropped area during April 1992 

(fig. 1), the 20-mg/L chloride contour was extended to 

the upgradient (west) end of this area because the 

source of the higher chloride concentrations was considered to be downward movement of chloride to 

ground water from the cropped area.

Most of the closed contours in figures 1-4 are artifacts of the contour intervals selected. The closed contour around the 32.3-mg/L chloride concentration in 

figure 1, for example, likely represents a volume of 

water that was recharged through the overlying potatocropped area. This volume of water could have become 

isolated within water that had lower chloride concentrations because of temporal variations in fluxes of 

water and chemicals to the saturated zone. Other closed 

contours in figures 1-4, particularly for the herbicides, 

likely represent mixing of waters in the aquifer and are 

not related to the MSEA farming systems. For atrazine

and DEA, the contour intervals chosen were the reporting and detection limits for each compound. Because 

there is uncertainty associated with the reproducibility 

of these herbicide analyses near their detection limits, 

there is also uncertainty associated with the locations 

of the herbicide-concentration contour lines (figs. 2, 3, 

and 4).

Chloride. Chloride concentrations (23-26 mg/L) 

greater than background (2-19 mg/L) were first 

detected in the upper 1 m of the saturated zone beneath 

the field corn-cropped area (fig. 2) during the June 

1991 sampling period, about 1.5 months after application. Chloride concentrations (33-37 mg/L) greater 

than background were first detected in the upper 1 m of 

the saturated zone beneath the potato- (fig. 3) and soybean-cropped areas during the August 1991 sampling 

period, about 4 months after application. Chloride concentrations (21-31 mg/L) greater than background 

were first detected in the upper 1 m of the saturated 

zone beneath the sweet corn and continuous corn during the December 1991 sampling period, about 7.5 

months after application. On the basis of these times of 

first detection of elevated chloride concentrations, the 

estimated time of travel of water and chloride moving 

through the unsaturated zone was 1 to 7.5 months. 

These traveltimes are maximum values because the 

sampling periods were not frequent enough to identify 

first arrival of the chloride tracer. In addition, relatively 

high background concentrations (2-19 mg/ L) of chloride partially conceal increases in chloride concentrations that could have resulted from recharge through 

the MSEA cropped areas. These traveltimes are comparable with a traveltime of 75 days determined in a 

dye-tracing study done during 1991 (Delin and others, 

1992b).

The April 1992 data reflect the affects of agricultural practices in 1991 and earlier on ground-water 

quality because agricultural chemicals had not yet been 

applied for the 1992 growing season. Additional inputs 

of chloride and continued spread of chloride through 

the aquifer occurred during spring 1992 (figs. 2-4) 

when approximately 6.4 cm of recharge occurred. 

Chloride concentrations were greatest beneath the 

potato-cropped area (fig. 4), reflecting that the chloride 

(potash fertilizer) application on the potato-cropped 

area was 4.5 times greater than that used with any other 

crop (table 1). Chloride concentrations in April 1992 in 

the upper 1 m of the saturated zone beneath the cropped 

areas ranged from 20 to 50 mg/L compared with concentrations of 2 to 19 mg/L away from the cropped

547
Chloride

2 METERS

t

l^ T r\ \

August 1991

D ' 

0

a 

a

1 D

\ D

kDecember 1991

April 1992

300 METERS

<10 10-19.9 20-29.9 30-39.9 40-50

Chloride concentration, in milligrams per liter 

Chloride applied during April 1991

Nitrate 

nitrogen

T

2 METERS

Jc.

n___ .____.-P. °

'Tl 4rv"iri5Q3'"'~--~~--------1-- -"" ---" O ~-Jl

300 METERS

<10 10-19.9 20-29.9 30-39.9 40-50

Nitrate nitrogen concentration, in milligrams per liter 

Nitrogen applied during April and June 1991

 Field cornAtrazine ^T

June 1991

August 1991____-^^v-^"-'~ ^'

hDecember 1991

T

2 METERS

JL

0

.°,

-S-Apr

% v

1 1992

D / 

0

0 '

D

V

D'5.

a

300 METERS

Quantitive reporting limit 0.04 ug/L 

Qualitative detection limit 0.01 ug/L

nd 0.01 - 0.039 >0.04

Quantitive reporting limit 0.06 ug/L 

Qualitative detection limit 0.03 ug/L

nd 0.03-0.059 >0.06

Atrazine concentration, in ug/L {micrograms per liter) 

Atrazine applied during May 1991

Lines of section shown in figure 1.

De-ethylatrazine concentration, in ug/L (micrograms per liter) 

Atrazine applied during May 1991

EXPLANATION

Water-table 

Approximate line of

equal chemical

concentration

N- -H Cropped area

August 1991 Sampling period date

nd Below the detection limit

D Data point

Figure 2. Chloride, nitrate nitrogen, atrazine, and de-ethylatrazine concentrations along a 

cross section beneath the field-corn cropped area during the June 1991 through April 1992 

sampling periods at the Princeton, Minnesota,Management Systems Evaluation Area.

548
 Potatoes -

Chloride /"*

9 <B

June 1991 %

o

I

~-^3S8!W?$$ffi!i

a

IVugust 1991 o

D

\ 

 '\\

-^r * \

-~- " a '

a

D

D

' 1 V ' \

O

December 1991 D

Q

 »»,,» >*"*'"~"*"tl

D

0

<10 10-19.9 20-29.9 30-39.9 40-50

Chloride concentration, in milligrams per liter 

Chloride applied during April 1991

-PotatoesNitrate 

nitrogen June 1991

2 METERS I n

1 n

D

August 1991

T "^ I

0 

o

o ~~~-_

0 ~""~--_...

<10 10-19.9 20-29.9 30-39.9 40-50

Nitrate nitrogen concentration, in milligrams per liter 

Nitrogen applied during April and June 1991

Atrazine

2 METERS

Quantitive reporting limit 0.04 ug/L 

Qualitative detection limit 0.01 ug/L

nd 0.01-0.039 >0.04

Quantitive reporting limit 0.06 ug/L 

Qualitative detection limit 0.03 ug/L

nd 0.03-0.059 >0.06

Atrazine concentration, in ug/L (micrograms per liter) 

Atrazine applied during May 1991

De-etnylatrazine concentration, in ug/L (micrograms per liter) 

Atrazine applied during May 1991

Lines of section shown in figure 1.

EXPLANATION

Water-table 

Approximate line of

equal chemical

concentration

Cropped area

August 1991 Sampling period date

nd Below the detection limit

D Data point

Figure 3. Chloride, nitrate nitrogen, atrazine, and de-ethylatrazine concentrations along a 

cross section beneath the potato-cropped area during the June 1991 through April 1992 

sampling periods at the Princeton, Minnesota,Management Systems Evaluation Area.

549
Chloride

2 METERS

10-19.9 20-29.9 30-39.9 40-50

Nitrate 

nitrogen

0 \ 

D \

o C&otinuous

O

o

a 

a

 - 0 a 

D B

' D «

 fa F»W--Coro_

o Sweet Corn

/ ^ 

 3B&'

;&,

O

o Soybean*

Q ' ' ---- -

0

. .: 0

5 a

..- B

a

 H

Q

a

a

a 

a

B

-

on°l\

al\ 

a I \

2 METERS

JL Potatoes

a a

0 ?-

0 /a

300 METERS

<10 10-19.9 20-29.9 30-40

Chloride concentration, in milligrams per liter 

Chloride applied during April 1991

Atrazine

H - w"*1

D

a

a

o

a

a

Continuous ,5^

Corn / ° \

a o

D '£

a o 

o .'6

Field Com

a Sweet Corn

CO.*

2 METERS

Nitrate nitrogen concentration, In milligrams per liter 

Nitrogen applied during April and June 1991

E NDe-ethylatrazine

H-

.S'"

a

D

~n- ' - ' '

&

a

i&nsr~ "" **

O

1D

a

0

o

\

a 

a

g

Potatoes «£;

300 METERS

Quantitive reporting limit 0.04 ug/L 

Qualitative detection limit 0.01 ug/L

nd 0.01-0.039 >0.04

Quantitive reporting limit 0.06 ug/L 

Qualitative detection limit 0 03 ug/L

nd 0.03-0.059 >0.06

Atrazine concentration, in ug/L (micrograms per liter) 

Atrazine applied during May 1991 

Lines of section shown in figure 1

EXPLANATION

De-ethylatrazine concentration, in ug/L (micrograms per liter) 

Atrazine applied during May 1991

-H Cropped area 

Water-table nd

Approximate line of equal chemical concentration a Data point 

Below the detection limit sweet com 1991 crop

550

Figure 4. Chloride, nitrate nitrogen, atrazine, and de-ethylatrazine concentrations along a cross 

section beneath cropped areas at the Princeton, Minnesota, Management Systems Evaluation 

Area, during the April 1992 sampling period.
areas and deeper than 1 m in the saturated zone. These 

results indicate that shallow ground-water quality 

reflects recent (May 1991 to April 1992) ground-water 

recharge. This conclusion is supported by chloride concentrations near the water table beneath the buffer areas 

(less than 10 mg/L) that were less than concentrations 

deeper in the saturated zone (10-20 mg/L) or beneath 

the MSEA cropped areas (fig. 2, 3, and 4). Because 

chloride was not applied in these buffer areas during 

1991, ground water near the water table beneath these 

areas likely was recharged during 1991. Ground water 

with chloride concentrations of 10 to 20 mg/L 1 to 2 m 

below the water table beneath the entire MSEA probably represent ground water unaffected by farming systems implemented for the MSEA study.

Nitrate-nitrogen. Ground-water samples were 

collected from 14 observation wells within the 65-ha 

MSEA during April 1991 before the implementation of 

the MSEA farming systems. Seven of the 14 wells 

were located within the area that was cropped before 

1991. Concentrations of nitrate-N in these 7 wells 

ranged from 2 to 23 mg/L; the median concentration 

was 10 mg/L.

Nitrate-N concentrations (22-47 mg/L) greater 

than those measured upgradient and 1 to 2 m below the 

water table (6-18 mg/L) were first detected at the water 

table beneath the potato-cropped area (fig. 3) during the 

December 1991 sampling period, about 7.5 months 

after application. Nitrate-N concentrations increased 

beneath the potato-cropped area by April 1992 compared to earlier sampling periods (fig. 3). Nitrate-N 

concentrations were greatest beneath the potatocropped area (fig. 4), reflecting an application rate on 

the potato-cropped area that was 1.4 times greater than 

on any other cropped area (table 1). Because nitrate-N 

concentrations were similar upgradient, beneath, and 

downgradient of the other cropped areas, additions of 

nitrate to ground water as a result of the MSEA farming 

systems research could not be clearly distinguished 

from background concentrations (figs. 2 and 4).

Elevated concentrations of nitrate-N (greater than 

20 mg/L) were measured near the water table beneath 

the field corn-cropped area in June 1991 and the sweet 

corn-cropped area in April 1992 (figs. 2 and 4). These 

nitrate-N concentrations were greater than typical 

background concentrations (0.1-18 mg/L) and were 

within the part of the saturated zone affected by 

recharge through the MSEA cropped areas, as delineated by the chloride data (figs. 2 and 4). However, 

these nitrate-N concentrations did not exceed the

highest background nitrate concentrations of 25 mg/L. 

Therefore, it cannot be concluded at this time whether 

the elevated nitrate-N concentrations beneath the sweet 

corn and field corn were due to inputs from the MSEA 

farming systems or were background concentrations.

Nitrate-N concentrations in the saturated zone 

across the entire MSEA typically decreased from 15 to 

25 mg/L near the water table to 5 to 10 mg/L 2 m below 

the water table. Similarly, dissolved oxygen concentrations typically decreased from near saturation with the 

atmosphere (from greater than 8 parts per million 

(ppm)) at the water table to less than 3 ppm 2 m below 

the water table. This decrease in dissolved-oxygen concentrations with depth implies that oxygen was being 

consumed by biochemical oxidation-reduction reactions faster than it could be replenished by advection of 

oxygen-rich water from the water table (Ronen and 

others, 1987; Champ and others, 1979). Concentrations 

of nitrate N were less than 0.2 mg/L 7 m below the 

water table, where the water was generally anoxic, 

indicating that the decrease in nitrate-N concentrations 

continued between 2 and 7 m below the water table. 

This decline in nitrate-N concentrations could indicate 

that nitrate-N was denitrified at depth within the 

aquifer (Postma and others, 1991; Trudell and others, 

1986).

Because nitrogen fertilizer was not applied to the 

buffer areas, the nitrate-N (15 to 25 mg/L at the water 

table) beneath the buffer areas during 1991 likely 

resulted from the residual effects of previous land use. 

Possible sources of nitrate-N reaching the ground water 

include decomposition of alfalfa roots and residues in 

the soil (from cropping with alfalfa during 1981 -1989) 

or application of nitrogen fertilizer to the corn grown 

during 1990.

Herbicides. During April 1991, before implementation of the MSEA farming systems, atrazine, 

DEA and DIA, were detected in 2, 9, and 3, respectively, of the 14 observation weds sampled within the 

65-ha MSEA. The median concentrations for atrazine, 

DEA, and DIA were less than 0.01, 0.10, and less than 

0.01 fj,g/L, respectively, and maximum concentrations 

were 0.17,2.30, and 0.98 fj,g/L respectively. No samples had concentrations of atrazine that exceeded the 

U.S. Environmental Protection Agency (1986) advisory level of 3 n,g/L.

Atrazine, DEA, and DIA were detected in 75, 97, 

and 76 percent, respectively, of the 124 wells sampled 

at the Princeton MSEA during the four sampling 

periods from June 1991 through April 1992. The

551
Table 2. Concentrations of herbicides and herbicide metabolites in 315 ground-water samples collected during four sampling 

periods from June 1991 through April 1992 at the Princeton, Minnesota, Management Systems Evaluation area, 1991

[DEA, de-ethylatrazine; DIA, de-isopropylatrazine; ng/L, micrograms per liter; <, less than; >, greater than. Number (#) of detections greater 

than the reporting limit or detection limit]

Atrazine 

Mg/L

DIA Alachlor 

Mg/L

2,6-

Diethylanaline

Mg/L

Chloralachlor 

Mg/L

Metolachlor

Metribuzin

Maximum 

Mnimum 

Median

0.54

.01

1.00

<.03

.08

0.66 

<.06 

<.06

1.08 0.02 0.31 

<.06 

<.06

0.04 0.45 

<.03 

<.03

>Detection limit

# of detections 161 233 32

Percentage of 51 74 10 

samples 

collected

>Reporting; limit1

# of detections 63 195 25 

Percentage of 20 62 8

samples

collected

10

3

1 The quantitative reporting limit for atrazine, DEA, DIA, alachlor, 2,6-diethylanaline, chloralachlor, metolachlor, 

and metribuzin analyses were 0.04,0.06,0.08,0.04,0.04,0.08,0.04, and 0.06 u,g/L, respectively. The qualitative 

detection limit for atrazine, DEA, DIA, alachlor, 2,6-diethylanaline, chloralachlor, metolachlor, and metribuzin 

analyses were 0.01,0.03,0.06,0.01,0.01,0.06,0.01, and 0.03 u,g/L, respectively.

concentrations of these compounds in the saturated 

zone varied both spatially and temporally. Summary 

statistics for herbicide and herbicide metabolize concentrations in 315 samples collected during these sampling periods are shown in table 2. Detections of 

atrazine (from less than 0.01 to 0.54 |ig/L, median 0.01 

fig/I,), DEA (from less than 0.03 to 1.00 u>g/L, median 

0.08 u,g/L), and DIA (from less than 0.06 to 0.66 ng/L, 

median less than 0.06 u,g/L) (table 2) beneath the buffer 

areas and beneath the cropped areas where atrazine was 

not applied (the potatoes, figs. 3 and 4, and soybeans, 

fig. 4) indicate that background concentrations of these 

compounds were present that were unrelated to the 

MSEA farming systems. DEA was the most frequently 

detected herbicide or herbicide metabolite and had the 

greatest median concentration. Concentrations of DEA

in April 1992 increased with depth from less than 0.06 

|ig/L near the water table to 0.08 to 0.30 u,g/L between 

1 and 2 m below the water table. This trend with depth 

was not noticeable prior to April 1992 and could 

indicate that water recharging the aquifer beneath the 

cropped areas during 1991 had lower DEA concentrations than older ground water. There were few clear 

spatial or temporal trends in the distribution of other 

herbicides or metabolites in the ground water from 

June 1991 through April 1992.

Atrazine in low concentrations (0.01 -0.47 u,g/L) 

was detected within the region of the saturated zone 

affected by recharge through the sweet corn-, field 

corn-, and continuous corn-cropped areas as delineated 

using the chloride data (figs. 2 and 4). Because similar 

atrazine concentrations were detected in ground water

552
recharged beneath areas where atrazine was not 

applied, these detections cannot be currently linked to 

the MSEA farming systems. However, the possibility 

still exists that some of the atrazine used in the MSEA 

farming systems may reach ground water. Additional 

data over a multi-year period will be needed to evaluate 

whether atrazine applied as part of the MSEA farming 

systems reaches ground water. Possible sources for the 

widespread background concentrations of atrazine and 

atrazine metabolites at the MSEA are application of 

atrazine to the field corn grown on the 65-ha MSEA 

during 1990, one year prior to implementation of the 

MSEA, or application before 1980 (prior to cropping 

with alfalfa).

Alachlor and its metabolites chloroalachlor and 

2,6-diethylanaline were infrequently detected in low 

concentrations at the MSEA (table 2). These compounds were detected both beneath cropped areas 

where alachlor was applied and in areas where it was 

not applied, indicating that there were background levels of these compounds. There were no detections of 

metolachlor or metribuzin beneath cropped areas 

where these chemicals were applied.

DISCUSSION AND SUMMARY

The total mass of chloride and nitrogen applied to 

the MSEA cropped areas that reached the saturated 

zone was estimated. Concentrations of chloride and 

nitrate-N in the upper 1 m of the saturated zone beneath 

the cropped areas that exceeded background concentrations were assumed to have resulted from leaching of 

chemicals applied to the cropped areas. The concentrations in excess of background concentrations were 

multiplied by the estimated volume of aquifer affected 

to estimate the mass of chemicals in the saturated zone 

derived from the MSEA farming systems. The total 

mass of MSEA-derived chloride in the upper 2 m of 

the saturated zone beneath the potato-cropped area 

(120 kg/ha) exceeded the mass beneath each of the 

other cropped areas (60 kg/ha) because 4.5 times more 

chloride was applied on the potato-cropped area than 

on the other cropped areas. About 32 percent of the 

chloride applied to the potato-cropped area and 71 percent of the chloride applied to the other cropped areas 

during April 1991 likely reached the saturated zone by 

April 1992. The percentage of applied chloride that 

reached the saturated zone beneath the potato-cropped 

area likely was less than for the other cropped areas 

because the flux of water that moved through the unsaturated zone was insufficient to transport more of the 

applied chloride to the water table. The residual chloride could indicate that carryover of chemical applications occurred from one growing season to the next. 

Residual chloride is likely to continue to leach from the 

unsaturated zone during subsequent recharge events. 

About 21 percent of the nitrogen applied to the potatocropped area during April and June 1991 reached the 

saturated zone by April 1992. Because the inputs of 

nitrate-N from the other MSEA cropped areas could 

not be distinguished from background concentrations, 

the percentage of applied nitrogen that reached the saturated zone could not be determined beneath these 

other cropped areas.

Use of chloride in potash fertilizers applied to the 

cropped areas as a tracer for water recharged through 

these areas allowed delineation of a region of the saturated zone affected by the MSEA farming systems. 

With the exception of comparatively high nitrate-N 

concentrations beneath the potato-cropped area concentrations of nitrate-N, herbicides, and herbicide 

metabolites within this region of the saturated zone 

were generally similar to background concentrations. 

This indicates that the effects of the MSEA farming 

systems on ground-water quality were no greater than 

the effects from previous cropping practices. The background concentrations of chloride, nitrate-N, herbicides, and herbicide metabolites, however, make it 

difficult to evaluate fully the effects of MSEA farming 

systems on ground-water quality without additional 

data over several years. Use of conservative tracers, 

such as chloride and bromide, to delineate annual 

recharge fronts beneath the cropped areas should make 

it easier to describe the effects of MSEA farming 

systems on ground-water quality as older ground water, 

affected by previous farming practices, moves horizontally and vertically away from the cropped areas.

REFERENCES

Anderson, Jl., Dowdy, R.H., and Delin, G.N., 1991, 

Ground water impacts from irrigated ridge-tillage, in 

Ritter, W.F., ed.. Proceedings of the American Society 

of Civil Engineers, 1991 National Conference on Irrigation and Drainage Engineering, Honolulu, Hawaii, 

July 23-26: p. 604-610.

Champ, D.R., Gulens, J., and Jackson, RE., 1979, Oxidation-reduction sequences in ground water flow systems: 

Canadian Journal of Earth Sciences, v. 16, no. 1, 

p. 12-23.

553
Delin, G.N., Landon, M.K., Anderson, J.L., and Dowdy, 

R.H., 1992a, Hydrologic research at the Princeton, 

Minnesota Management Systems Evaluation Area: 

U.S. Geological Survey Open-File Report 92-107, 2 p.

Delin, G.N., Landon, M.K., Healy, R.W., and Olsen, H.W., 

1992b, Preferential flow through the unsaturated zone 

beneath a corn field near Princeton, Minnesota, in EOS, 

Transactions, Proceedings of the American Geophysical Union spring meeting, Montreal, Canada, May 12- 

16,1992, v. 73, no. 14, p. 132.

Postma, D., Boesen, C., Kristiansen, H., and Larsen, F., 

1991, Nitrate reduction in an unconfined sandy aquiferwater chemistry, reduction processes, and geochemical 

modeling: Water Resources Research, v. 27, no. 8, 

p. 2027-2045.

Rasmussen, W.C., and Andreasen, G.G., 1959, Hydrologic 

budget of the Beaver Dam Creek Basin Maryland: U.S. 

Geological Survey Water Supply Paper 1472, 106 p.

Ronen, D., Margaritz, M., Almon, E., and Amiel, A.J., 

1987, Anthropogenic anoxification ("eutrophication") 

of the water table region of a deep phreatic aquifer, 

Water Resources Research, v. 23, no. 8, p. 1554-1560.

Trudell, MR., Gillham, R.W., and Cherry, J.A., 1986, An 

in-situ study of the occurrence and rate of denitrification in a shallow unconfined sand aquifer Journal of 

Hydrology, no. 83, p. 251-268.

U.S. Environmental Protection Agency, 1986, Quality 

criteria for water 1986: EPA-440/5-86-001, 100 p.

554
Effects of Topography on the Transport of Agricultural 

Chemicals near Princeton, Minnesota, 1992

By Geoffrey N. Delin 1 and Matthew K. Landon 1

Abstract

In 1991, the U.S. Geological Survey (USGS), 

with funding from the USGS Toxic Substances 

Hydrology Program, began studying the movement of water and agricultural chemicals at the 

Management Systems EvaluationArea (MSEA) 

near Princeton, Minnesota. The research is being 

conducted in a topographically low (lowland) site 

and a topographically high (upland) site within the 

northernmost cropped area at the 65-hectare Princeton MSEA. The sites are about 78 m (meters) 

apart and with a difference in land-surface elevation of 1.4 meters. Soils in the upper meter at both 

the upland and lowland sites are similar and are 

composed of about 95 percent sand and 5 percent 

silt and clay. The grain size is coarser between the 

1.0- and 2.0-m depths at the lowland site than at 

the upland site, however. Total organic carbon in 

the upper 20 cm of topsoil is about 1.0 percent at 

the lowland site and about 0.6 percent at the 

upland site. Water sampling and monitoring equipment were installed at the upland and lowland sites 

following a dye-tracing and trenching study that 

was conducted to identify zones of preferential 

and retarded water movement and to collect soil 

samples from the unsaturated-zone.

Infiltration tests were conducted during 1992 

at the upland and lowland sites following selected 

precipitation events or by the application of 2.5 cm 

of water from a linear-move sprinkler irrigation 

system. The movement of wetting fronts through 

the unsaturated zone to the water table were 

evaluated using time-domain reflectometry. 

Water samples were collected from suction lysim-

'U.S. Geological Survey, Mounds View, Minn.

eters as the wetting front passed a given plysimeter 

elevation, and from water-table weds after the wetting front reached the saturated zone.

Wetting fronts typically penetrated deeper into 

the unsaturated zone at the lowland site than at the 

upland site during 1992. In addition, the total flux 

of water into the soil at the lowland site was 

greater than at the upland site. Following application of about 8 cm of water beginning June 29, 

1992, for example, recharge through the unsaturated zone was about 1.0 cm at the lowland site 

compared to about 0.5 cm at the upland site, based 

on hydrograph analysis. For this recharge event, 

the estimated fluxes of nitrate nitrogen and atrazine through the unsaturated zone to the water 

table at the lowland site were 3.5 and 5 times 

greater, respectively, than the fluxes at the upland 

site. Differences in the movement of wetting 

fronts, agricultural chemicals, and tracers at the 

upland and lowland sites likely resulted from a 

combination of factors including differences in 

soil organic matter, grain size, porosity, sedimentary heterogeneities, antecedent moisture conditions, and topographic relief. Possible mechanisms 

that could cause focused recharge of water and 

agricultural chemicals at the lowland site include 

surface runoff and iron-rich layers that form a barrier to vertical water movement through the unsaturated zone at the upland site.

INTRODUCTION

Previous studies have indicated that greater 

ground-water recharge occurs in topographically low 

(lowland) areas compared to topographically high 

(upland) areas (Meyboom, 1966; Lissey, 1968). This 

increased flux of ground-water recharge in lowland 

areas has been termed focused recharge. The effects of

555
agricultural activities on ground-water quality have 

been investigated in numerous field studies at various 

scales. These range from regional studies covering 

thousands of square kilometers (Helgesen and others, 

1991; Burkart and others, 1991) to plot-scale studies of 

the effects of macropores, tillage, and crop rotation 

(Kanwar, 1991; Cameron and others, 1979). Little is 

known, however, about the effects of focused recharge 

on the concentrations and fluxes of agricultural chemicals in lowland areas. An improved understanding is 

needed of how topography and focused recharge affect 

the movement of agricultural chemicals. In 1991, the 

U.S. Geological Survey (USGS), with funding from 

the USGS Toxic Substances Hydrology Program, 

1 began a study of the movement of water and agricultural chemicals to the water table at the Management 

Systems Evaluation Area (MSEA) near Princeton, 

Minnesota (fig. 1).

The primary objective of this 4-year research 

study is to evaluate the effects of transient recharge, 

topography, and subsurface heterogeneities on the 

flux of water and agricultural chemicals to the water 

table. The approach used was to (1) install instruments to sample and measure the movement of water 

through the unsaturated zone beneath upland and 

lowland sites of a corn field at the Princeton, Minnesota MSEA; (2) conduct recharge and tracer tests in 

the field to evaluate the movement of water and agricultural chemicals at both topographic settings; and 

(3) simulate the field recharge and tracer tests in the 

laboratory (Delin and Landon, 1993). The purpose of 

this paper is to describe preliminary results of ongoing research on the effects of topography on water 

and chemical movement at the Princeton MSEA. 

Data from 1992 are presented, that describe differences in the movement of water and selected agricultural chemicals as related to topography.

DESCRIPTION OF RESEARCH SITES

The research is being conducted at an upland site 

and a lowland site within the northernmost area 

cropped in corn at the 65-ha Princeton MSEA (fig. 1). 

The upland and lowland sites are about 78 m apart and 

differ in land-surface elevation by 1.4 m (slope of about 

0.02). The unsaturated zone mostly is a fine- to 

medium-grained sand and the saturated zone mostly is 

a medium- to coarse-grained sand (Delin and others, 

1992a). The soils in the upper meter at both the upland 

and lowland sites are similar and are composed of 

about 95 percent sand and 5 percent silt and clay. The

grain size is coarser, however, between the 1.0- and 

2.0-m depths at the lowland site than at the upland site. 

Total organic carbon in the upper 20 cm of topsoil is 

about 1.0 percent at the lowland site and about 0.6 

percent at the upland site (J.A. Lamb, University of 

Minnesota Soil Science Department, written commune 

1993).

During October 1991, the water table was about 

4 m below land surface at the upland site and 2.8 m 

below land surface at the lowland site. The mean saturated horizontal hydraulic conductivity in the upper 

meter of the unsaturated zone at both sites was similar 

(about 3.6xlO'3 cm/s, (H.W. Olsen, U.S. Geological 

Survey, written commun., 1993)). Below the 1-m depth 

at the lowland site, where the grain size is coarser, the 

mean saturated horizontal hydraulic conductivity 

(11.9x 10"3 cm/s) is greater than that for the same depth 

interval at the upland site (4.8xlO~3 cm/s). On the basis 

of water-level hydrograph analyses, ground-water 

recharge rates were 10 to 20 cm per year at the Princeton MSEA and ground-water-flow gradients were 

0.0009 to 0.002 during 1991. Ground-water flow is 

from southwest to northeast beneath the upland and 

lowland sites at a velocity that varies seasonally from 

6 to 15 cm per day. Ground-water-flow directions, 

however, shift to a more northerly direction as a result 

of focused recharge in lowland areas (fig. 1).

The herbicides atrazine and alachlor were applied 

to the research sites during May 1991 and May 1992 at 

broadcast rates of 1.7 and 2.25 kg/ha active ingredients, 

respectively. Nitrogen fertilizer was applied during 

April to June 1991 and during June 1992 at rates of 

157 kg/ha and 123 kg/ha, respectively. Potash fertilizer 

was applied during April 1991 and during May 1992 at 

rates of 112 kg/ha and 45 kg/ha, respectively. Chloride 

in the potash fertilizer has been used as a tracer of water 

recharging through the MSEA cropped areas (Landon 

and others, 1993). The entire 65-ha field was planted in 

alfalfa from 1981 to 1989, prior to the implementation 

of the MSEA farming systems in spring 1991. Detailed 

records of farming practices and chemical applications 

during this period were not available.

METHOD OF STUDY

Before the unsaturated-zone instruments were 

installed, a dye-tracing and trenching study was done 

at the upland and lowland sites to identify zones of 

preferential and retarded water movement and to 

characterize the unsaturated-zone properties (Delin 

and others, 1992b). A 3-percent solution of rhodamine

556
Planimetric View

Vertical profile of n 

domain-reflectometry 

nine thermocouples, 

tensiometer;

^r%

Water-table 

well

of Instrument Locations at the

ne time- Vertical profile of 

r probes, tnree suction lysimeters 

and five > ^^**

5 \ / ^-^^

\..........^

 ::::::::: ' ' : ; :::x:x::::::::::;:: _

\jy w

__ Trench 

 - Dye-applicatio

Upland and

Multiport well

n area 0 

0 1

Lowland Sites

ion gage

5 10 FEET 

2 3 METERS

^f^||||^(iiibniii8teji^

^^^Mf^^^f 2

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250 500 750 1000 FEET

100 200 300 METERS

EXPLANATION

Boundary of cropped area 

Lowland areas

-298 Topographic contour. Shows altitude 

of land surface. Interval 0.25 

meters. Datum is sea level.

*4 Observation well, number indicates 

number of wells at site.

® Multiport well

The research area is located 

in the northeast quarter of 

section 18, township T35N, 

range R26W.

Figure 1. Instrument locations at the upland and lowland sites, land-surface topography, and 

lowland areas at the Princeton, Minnesota, Management Systems Evaluation area (MSEA).

557
WT dye was applied uniformly as a tracer to a 3.5-m by 

6-m area at each site (fig. 1). About 12 L of dye was 

applied at 10-day intervals from July 5 through September 13, 1991 to ensure complete penetration of the 

unsaturated zone. Total rainfall from July 5 through 

October 22, 1991 was about 30 cm. Beginning October 

22, 1991, a 3-m by 2-m trench was dug in the middle of 

each dye-application area. The depth of each trench 

was increased by intervals of 0.5-m to a total depth of 

2.0 m to locate the dye and to collect soil samples from 

the unsaturated zone. Del in and others (this volume) 

provide a detailed description of the dye-tracing and 

trenching study.

The following instruments were installed in two 

vertical profiles at the upland and lowland sites (fig. 1) 

on the basis of the presence or absence of dye: 

(1) thermocouples and time-domain reflectometry 

(TDR) probes, at depths of 0.2, 0.4, 0.6, 0.8, 1.0, 1.5, 

2.0, 2.5, and 3.0 m; (2) suction lysimeters, at depths of 

0.8, 1.8, and 2.8 m; and (3) tensiometers, at depths of 

0.5, 1.0, 1.5, 2.0, and 2.5 m. TDR is a technique in 

which electronic pulses are transmitted through a coaxial cable that is attached to a fixed-length probe embedded in the soil (Topp and others, 1980). The resultant 

waveform is affected by the moisture content of the soil 

surrounding the probe. Because the dielectric constant 

of water is much higher than that of most other materials, a signal in moist soil will propagate more slowly 

than in the same soil when dry. Thus, the moisture 

content can be calculated by measuring the waveform 

propagation time over the fixed-length probe. TDRmeasured moisture contents at the lowland site correlated well (R2=0.92) with neutron-probe measurements (M.D. Tomer, University of Minnesota, Soil 

Science Department, written commun., 1992).

A water-table well was installed upgradient of each 

trench so that concentrations of agricultural chemicals 

reaching the water table could be determined (fig. 1). 

Soil temperature, soil moisture, soil-water tension, 

ground-water levels, and precipitation were continuously recorded at both sites with data loggers.

Infiltration tests were conducted during 1992 at the 

upland and lowland sites following selected precipitation events or by the application of 2.5 cm of water 

from a linear-move sprinkler irrigation system. The 

movement of wetting fronts through the unsaturated 

zone to the water table were evaluated using TDR. 

Water samples were collected from the suction 

lysimeters and the water-table wells before each infiltration test to determine pre-test concentrations of the

agricultural chemicals. Samples were also collected of 

the irrigation water or precipitation. Following the 

application of irrigation water, water samples were collected from the suction lysimeters as the wetting front 

passed a given lysimeter elevation, and from the watertable well after the wetting front reached the water 

table.

The water samples were analyzed for dissolved 

major anions (including nitrate-nitrogen (nitrate-N), 

chloride, sulfate, and bromide) using ion chromatography and analyzed for selected herbicides and herbicide 

metabolites (atrazine, de-ethylatrazine (DBA), de-isopropylatrazine (DIA), alachlor, chloroalachlor, 2,6- 

diethylanaline, metolachlor, and metribuzin) using gas 

chromatography/mass spectroscopy (P.D. Capel, U.S. 

Geological Survey, written commun., 1991). In addition to the water sampling, soil samples were collected 

by the University of Minnesota Soil Science Department from the upland and lowland sites six times per 

year. The soil samples were analyzed for concentrations of nitrate-N and the herbicides atrazine, alachlor, 

and metolachlor (MSEA Steering Committee, written 

commun., 1991).

WATER INFILTRATION AT THE UPLAND 

AND LOWLAND SITES

The movement of wetting fronts was different at 

the upland and lowland sites during 1992. Wetting 

fronts typically penetrated deeper into the unsaturated 

zone at the lowland site than at the upland site. About 

24 hours after 2.5 cm of irrigation water was applied on 

June 9, for example, the wetting front at the upland site 

had moved to a depth of only 0.8 m below land surface 

(fig. 2). By comparison, the wetting front at the lowland site reached the water table (about 2.8 m below 

land surface) after the same 24-hour period resulting in 

ground-water recharge and a water-level rise of about 

0.3 cm. The increase in moisture content from the 

2.0-m depth to the 2.5-m depth at the lowland site 

(fig. 2) is because the TDR probe at the 2.5-m depth 

was located within the capillary fringe. The decrease in 

moisture content at the lowland site from June 10 to 

June 11 from the 0.8-m depth to the 2.0-m depth at the 

lowland site (fig. 2) was due to drainage of water from 

the unsaturated zone as a result of recharge to the saturated zone. The increase in the amount of water stored 

in the soil from June 9 to June 10 at the lowland site 

(3.7 cm) was greater than the amount stored in the soil 

at the upland site (2.3 cm). Focused recharge occurred

558
at the lowland site because the amount of water stored 

in the unsaturated zone (3.7 cm) was greater than the 

amount of water applied (2.5 cm). Additional evidence 

of focused recharge is that the water table rose at the 

lowland site in response to the irrigation but did not rise 

at the upland site.

During a recharge event beginning June 29, 1992 

about 8 cm of water (irrigation and rainfall) was 

applied to the upland and lowland sites over a 4-day 

period. Measurements of daily precipitation and irrigation, profiles of volumetric moisture content, and 

water-level fluctuations for the lowland site from June 

29 through July 8,1992 are shown in figure 3. Continuous measurements of volumetric moisture content were 

not measured at the upland site. The water applied during the first 3 days of the recharge event brought the 

soil moisture to field capacity. Field capacity is defined 

as the amount of water remaining in a soil 2 or 3 days 

after having been wetted and after free drainage is negligible (Cassel and Nielsen, 1986). The field capacity is 

different for each soil depth and at the 1.0-m depth, for 

example, corresponds to a volumetric moisture content 

of about 0.10 (fig. 3). The 4-cm of rainfall that occurred 

on July 2 caused the soil moisture to exceed the field 

capacity, resulting in the wetting front moving through 

the 2.8-m-thick unsaturated zone beneath the lowland 

site in about 4 days (93 hours). In response to this 

recharge, the water-table rose about 2 cm over a 3-day 

period at the lowland site compared to only about a 

0.6-cm rise at the upland site. On the basis of this 

water-table response and a specific yield of 0.3, 

recharge through the unsaturated zone at the lowland 

site for this event was about 1.0 cm, compared to about 

0.5 cm at the upland site. These results indicate that 

even when recharge occurred in both topographic settings, greater recharge occurred at the lowland site.

The apparent pattern of soil-water movement as 

indicated by rhodamine-WT distributions in the unsaturated zone did not reflect measured soil-water distributions during 1992. Soil-water movement in what was 

identified as a retarded flow zone during the dye-tracing study actually moved faster, not slower, than the 

soil-water movement identified as a preferential flow 

zone. This difference indicates that the zones of preferential flow observed on the trench walls did not extend 

greater than about 10 cm beyond the face of the trench 

or failed to encompass the entire length of the installed 

TDK probes. In addition, it is likely that the probes 

installed in what was considered a preferential flow 

zone might actually have penetrated an adjacent zone 

of retarded flow.

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VOLUMETRIC MOISTURE CONTENT

EXPLANATION

 June 9 

10

-- June 11

Figure 2. Soil moisture at the upland and lowland sites following application of 2.5 centimeters of water on June 9,1992, at 

the Princeton, Minnesota, Management Systems Evaluation 

area.

Differences in the movement of wetting fronts at 

the upland and lowland sites could have resulted from 

a combination of factors that include differences in 

organic content, grain size, porosity, sedimentary heterogeneities, antecedent moisture conditions, evapotranspiration, and topographic relief. Delin and others 

(this volume) provide a description of the differences in 

unsaturated-zone properties at the upland and lowland 

sites.

One mechanism that could cause focused recharge 

to occur at the lowland site is surface runoff. Runoff 

was observed at the Princeton MSEA during periods of 

intense or prolonged rainfall such as the one that began 

on June 29, 1992, when about 8 cm of irrigation water 

and rainfall were applied to the upland and lowland 

sites (fig. 3). Runoff was never observed, however, during recharge events such as the one that began June 9, 

1992, during which only about 2.5 cm of irrigation 

water was applied to each site during a 1-hour period 

(fig. 2). Thus, mechanisms other than runoff must be 

contributing to focused recharge at the research site 

during periods of limited water application.

559
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WATER 

TABLE

Figure 3. Daily precipitation and irrigation, profiles of volumetric moisture content, and watertable fluctuations for the lowland site at the Princeton, Minnesota, Management Systems Evaluation area, June 29-July 8, 1992.

560
Table 1: Concentrations of agricultural chemicals at the upland and lowland sites at the

Princeton, Minnesota Management Systems Evaluation Area, 1992

[ATR, atrazine; DEA, de-ethylatrazine; DIA, de-isopropylatrazine; mg/L, milligrams per liter; 

u,g/L, micrograms per liter1 ; #, number; <, less than; --, no samples analyzed]

Upland site Lowland site

Chloride, ATR, DEA, DIA, Nitrate- ATR_ DEA DIA

mg/L

Unsaturated zone

# of analyses

Maximum

Minimum

Median

Water table

# of analyses

Maximum

Minimum

Median

8

28.7

1.9

21.1

3

15.9

12.5

13.5

8

28.6

1.4

15.1

3

25.3

15.9

19.8

4

0.20

.04

0.08

3

0.03

<.01

.02

4

0.25

.08

0.21

3

0.14

<.03

.04

4

0.23

<.06

0.15

3

0.10

<.06

<.06

10

60.3

29.9

37.1

3

20.6

13.0

16.6

10

43.8

2.5

11.8

3

27.5

16.2

18.3

5

0.42

.07

0.20

5

0.08

.01

.02

5

1.43

<.03

0.06

5

0.06

<.03

<.03

5

0.49

<.06

0.09

5

<0.06

<.06

<.06

1. The qualitative detection limit for atrazine, de-ethylatrazine, and de-isopropylatrazine analyses were 0.01, 

0.03, and 0.06 u,g/L, respectively. The quantitative reporting limit for atrazine, de-ethylatrazine, de-isopropylatrazine analyses were 0.04,0.06, and 0.08 u,g/L, respectively.

The presence of iron-rich layers in the unsaturated 

zone at the upland site could also be a factor contributing to focused recharge at the lowland site. The 2 to 20 

cm thick layers are cemented with iron oxides, and 

contain an average of about 5 to 10 percent silt and 

clay, compared to less than about 5 percent in adjacent 

horizons. The layers were detected in soil cores and 

observed in the trench walls at the upland site, primarily between the 1- and 2-m depth intervals, and extend 

to below side-slope areas (M.D. Tomer, University of 

Minnesota, Soil Science Department, oral commun., 

1993). The layers were not observed in the trench walls 

at the lowland site. Because of the iron-oxide cement 

and the higher percentage of silt- and clay-sized 

particles compared to the surrounding soil, the layers 

form barriers that could retard the vertical flow of water 

through the unsaturated zone. Thus, the iron-rich layers 

likely retain water, and agricultural chemicals, that 

would otherwise penetrate deeper than this depth into 

the unsaturated zone at the upland site. The net effect 

likely is a reduction in the flux of water and agricultural 

chemicals through the unsaturated zone beneath the 

upland site. Some of the water and agricultural chemicals retained in the iron-rich layers could be used by

plants, further reducing the fluxes to the water table. 

Another possible effect of the iron-rich layers terminating beneath the side-slope areas is that water could 

move horizontally along these layers toward the lowland site over short distances of perhaps 1 to 2 m. The 

net effect of this horizontal movement along numerous 

iron-rich layers could be increased soil-water movement to the lowland site.

WATER QUALITY AT THE UPLAND AND 

LOWLAND SITES

Concentrations of nitrate-N in water samples in the 

unsaturated zone at the lowland site were greater than 

those in the unsaturated zone at the upland site during 

1992 (table 1, above). Nitrate-N concentrations in the 

unsaturated zone were generally greatest during April 

of 1992 at both sites, immediately following the application of agricultural chemicals, but changed little 

throughout the summer and fall. Concentrations of 

nitrate-N from the water table at the lowland site were 

slightly greater than concentrations from the upland 

site (table 1). By comparison, median concentrations of 

nitrate-N in soil samples collected during 1991 from

561
the 0-to 0.8-m depth interval at both the upland and 

lowland sites were 3.2 mg/kg J.A. Lamb, University of 

Minnesota, Soil Science Department, written commun., 1993).

Concentrations of chloride in water samples in the 

unsaturated zone at the lowland site were less than 

those in the unsaturated zone at the upland site during 

1992 (table 1). As with nitrate-N, chloride concentrations in the unsaturated zone were greatest during April 

of 1992 at both sites, immediately following the application of agricultural chemicals, but changed little 

throughout the summer and fall. Concentrations of 

chloride were similar in water samples from the water 

table at the upland and lowland sites during 1992 

(table 1).

Concentrations of atrazine in water samples collected from the unsaturated zone at the lowland site 

were slightly greater than concentrations at the upland 

site during 1992 (table 1). Concentrations of DEA and 

DIA from the unsaturated zone at the upland site were 

slightly greater than at the lowland site. The concentrations of atrazine, DEA, and DIA in samples from the 

water table were below or near the qualitative detection 

limits at both sites (table 1). As with data from the 

water analyses, the median concentration of atrazine in 

soil cores collected during 1991 from the 0- to 0.5-m 

depth interval at the lowland site (19.5 ng/g) was 

slightly greater than the median at the upland site 

(13.0 ng/g) J.A. Lamb, University of Minnesota, Soil 

Science Department, written commun., 1993). 

Atrazine was not detected above the detection limit of 

5 ng/g in soil samples below the 0.6-m depth from 

either site during 1991.

The concentrations of nitrate-N, chloride, atrazine, 

DEA, and DIA (table 1) were within the range of background concentrations at the MSEA (Landon and others, this volume). Because increased recharge at the 

lowland site should have caused the concentrations of 

agricultural chemicals to be diluted, however, it was 

expected that the concentrations would be lower at the 

lowland site than at the upland site. This hypothesis is 

supported by data from the unsaturated zone for chloride, DEA, and DIA, as well as by data from the water 

table for chloride and DEA (table 1). The concentrations of nitrate-N and atrazine, however, were greater 

at the lowland site than at the upland site in both the 

unsaturated zone and at the water table. The greater 

concentrations of nitrate-N and atrazine at the lowland 

site could have resulted from a concentrating of the 

chemicals due to runoff and along flow paths through

the unsaturated zone. The greater atrazine concentrations at the lowland site could have also resulted from 

greater biodegradation at the upland site because of a 

longer residence time in the unsaturated zone compared to the lowland site.

The fluxes of nitrate-N, chloride, atrazine, DEA, 

and DIA through the unsaturated zone to the water 

table were estimated for the recharge event beginning 

June 29,1993 (table 2). The calculated water fluxes for 

this recharge event of 0.5 and 1.0 cm/d were used for 

the upland and lowland sites, respectively. Median concentrations of the agricultural chemicals for unsaturated-zone samples collected during 1992 were used 

(table 1) in this analysis because water samples were 

not available for this recharge event. The median concentrations shown in table 1 are representative of the 

climate and farming practices during the growing season. The estimated fluxes for all of the agricultural 

chemicals, except DEA, were greater at the lowland 

site than at the upland site (table 2). The differences 

were greatest for nitrate-N and atrazine where the estimated fluxes at the lowland site were 3.5 and 5 times 

greater, respectively, than corresponding fluxes at the 

upland site. The greater fluxes at the lowland site are 

mostly the result of increased recharge rather than 

higher chemical concentrations.

The fact that atrazine and its metabolites were 

detected in water samples collected from most depths 

in the unsaturated zone and from the water table at the 

upland and lowland sites indicates that some of the 

applied herbicide moved through the unsaturated zone 

to the water table as a result of farming practices at the 

MSEA. Atrazine was not detected, however, in soil 

samples below the 0.6-m depth, primarily because the 

detection limit for atrazine in soil samples of 5 ng/g is 

too large to detect the lower concentrations of any herbicides present. The 5-ng/g detection limit for soil analyses corresponds to aqueous concentrations of 5 fiy/L. 

These results indicate that full evaluation of the movement of atrazine to the water table requires analyses of 

water samples from the unsaturated zone in addition to 

analyses of soil samples.

The limited number of chemical analyses from the 

upland and lowland sites during 1992 (table 1) limits 

the conclusions that can be made from these preliminary results. Additional analyses are needed to make 

definitive statements about differences in the concentrations and fluxes of agricultural chemicals at the 

upland and lowland sites.

562
Table 2. Estimated fluxes of water and agricultural chemicals to the 

water table at the upland and lowland sites for the June 29 through 

July 9,1992 recharge event at the Princeton, Minnesota Management 

Systems Evaluation area

[cm/d, centimeters per day; (kg/d)/ha, kilograms per day per hectare]

Lowland flux

Upland site Lowland site divided by

upland flux

Water flux, cm/d

Nitrate-nitrogen

flux, (kg/d)/ha

Chloride flux,

(kg/d)/ha

Atrazine flux,

(kg/d)/ha

De-ethylatrazine

flux, (kg/d)/ha

De-isopropylatrazine

flux, (kg/d)/ha

0.5

3.52

2.52

1.33X10'5

3.5xl(T5

2.5xlO'5

1.0

12.4

3.93

6.67xlO'5

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3.5

1.6

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1.2

CONCLUSIONS

Results of this study indicate that even small landsurface differences of only 1.4 m over a 78-m horizontal distance (slope of about 0.02) can result in focused 

recharge in lowland areas. Possible mechanisms that 

could cause focused recharge of water containing agricultural chemicals at the lowland site include surface 

runoff and iron-rich layers that form a partial barrier to 

vertical water movement through the unsaturated zone 

at the upland site. Based on analyses of few samples 

collected during 1992, the fluxes of most agricultural 

chemicals to the water table were greater at the lowland 

site than at the upland site, primarily because of 

increased recharge at the lowland site compared to the 

upland site. Additional analyses are needed to make 

definitive statements about the differences in the concentrations and fluxes of agricultural chemicals at the 

upland and lowland sites.

REFERENCES

Burkart, M R. Onstad, C.A., and Bubenzer, G.D, 1990, 

Research on agrichemicals in water resources, in EOS, 

Transactions, American Geophysical Union, v. 71, 

no. 29, p. 980-988.

Cameron, D.R., Kowalenko, C.G., and Campbell, C.A., 

1979, Factors affecting nitrate nitrogen and chloride 

leaching variability in a field plot: Soil Science Society 

of America Journal, v. 43, p.455-460.

Cassel, O.K., and Nielsen, D.R., 1986, Field capacity and 

available water capacity, in A. Klute and others, eds.,

Methods of soil analysis, physical and mineralogical 

methods, (2d ed.): Madison, Wise., American Society 

of Agronomy, no. 9, part 1, p. 901-926.

Delin, G.N., Landon, M.K., Anderson, J.L., and Dowdy, 

R.H., 1992a, Hydrologic research at the Princeton, 

Minnesota management systems evaluation area U.S. 

Geological Survey Open-File Report 92-107, 2 p.

Delin, G.N., Landon, M.K., Healy, R.W, and Olsen, H.W, 

1992b, Preferential flow through the unsaturated zone 

beneath a corn field near Princeton, Minnesota, in EOS, 

Transactions, American Geophysical Union, v. 73, 

no. 14, p. 132.

Delin, G.N., and Landon, M.K, 1993, Effects of focused 

recharge on the transport of agricultural chemicals at 

the Princeton, Minnesota Management Systems Evaluation Area, 1991 92: U.S. Geological Survey OpenFile Report 93-42, 8 p.

Helgesen, J.O., Stulken, L.E., andRutledge, A.T., 1991, 

Overview of nonpoint-source contamination of the high 

plains aquifer in south-central Kansas, in Mallard, GE., 

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water Resources Investigations Report 91-4034, p. 301-306.

Kanwar, R.S .,1991, Preferential movement of nitrate and 

herbicides to shallow ground water as affected by tillage and crop rotation, in Gish and Shirmohammadi, 

eds., Preferential Flow Proceedings of the National 

Symposium, December 16-17, 1991, Chicago, Illinois, 

American Society of Agricultural Engineering, 

p. 328-337.

563
Landon, M.K., Delin, G.N., Lamb, J.A., and Guo, L., 1993, Meyboom, P., 1966, Unsteady ground water flow near a

Ground-water quality at the Management Systems willow ring in hummocky moraine: Journal of HydrolEvaluation Area (MSEA) near Princeton, Minnesota, ogy, v. 4, p. 38-62.

1991: U.S. Geological Survey Open-File Report 93-43, Topp, G.C., Davis, J.L., and Annan, AS., 1980, Electromag8 P- nctic determination of soil water content, MeasurcLissey, A., 1968, Surficial mapping of ground-water flow ments in coaxial transmission lines: Water Resources

systems with application to the Oak River Basin, Research, v. 16, no. 3, p. 574-582.

Manitoba: Published PhD Thesis, University of

Saskatchewan, 141 p.

564
Spatial Variability of Unsaturated-Zone Properties in 

Relation to Topography in a Sand-Plain Setting near

Princeton, Minnesota

EyG.N. Delin 1 , M.K. Landon 1 , R.W. Healy2, and H.W. Olsen3

Abstract

The spatial distribution of preferential flow 

paths and unsaturated-zone properties in two topographic settings were determined from a dyetracing and trenching study done at the Management Systems Evaluation Area (MSEA) near 

Princeton, Minnesota. The topographic settings 

are upland and lowland sites about 78 m (meters) 

apart that differ in elevation by 1.4 m. A 3 percent 

solution of rhodamine-WT dye was applied uniformly as a tracer to a 3.5- by 6-m area at both sites 

at 10-day intervals from July 5 through September 

13,1991. After application of the dye, a 3- by 2-m 

trench was dug to a depth of 2 m in the middle of 

each dye-application area to locate the dye and to 

collect soil samples.

Water samples were collected periodically 

from a multiport well, located 2.5 m horizontally 

downgradient of each dye-application area, to estimate the time-of-travel of recharge water through 

the unsaturated zone. The dye was first detected in 

ground water about 100 days after application. On 

the basis of average ground-water velocity at the 

Princeton MSEA of 10 cm/d (centimeters per 

day), the transport velocity of dye through the 

unsaturated zone was calculated to be 3.7 cm/d at 

the lowland site and 5.3 cm/d at the upland site. 

The dye moved 2 m vertically through the saturated zone over a horizontal distance of about 9 m, 

whereas a steady-state ground-water-flow model 

predicted less than 0.2 m of vertical movement.

A total of about 450 soil samples were 

collected from the sides and bottom of the trenches

'U.S. Geological Survey, Mounds View, Minn. 

2U.S. Geological Survey, Lakewood, Colo. 

3U.S. Geological Survey, Golden, Colo.

for analyses of bulk density, dye fluorescence, 

grain-size distribution, hydraulic conductivity, 

moisture-retention characteristics, organic-carbon 

content, and volumetric moisture content. The distribution of dye through the unsaturated zone was 

highly variable at the upland and lowland sites. 

Dye movement was greatest beneath the furrows 

and least beneath the corn rows. Preliminary 

results indicate the dye moved preferentially in 

response to tillage patterns (corn rows and furrows), microtopography, presence of plant roots, 

differences in total organic carbon, and coarsergrained heterogeneities in the unsaturated zone. 

Visible dye distribution did not correlate strongly 

with bulk density, saturated hydraulic conductivity, moisture-retention characteristics, and 

volumetric moisture content.

INTRODUCTION

Two factors that affect water movement through 

the unsaturated zone are spatial variability of soil properties (Delhomme, 1979; Sharma and others, 1980; and 

Russo and Bresler, 1981) and topography (Meyboom, 

1966). Spatial variability of soil properties affects the 

pathways and rates of water movement. Among the 

important soil properties that can affect water movement are bulk density, grain-size distribution, hydraulic 

conductivity, moisture-retention characteristics, 

organic-carbon content, and volumetric moisture content. This study was designed to evaluate the effects of 

spatial variability of these soil properties on the distribution of preferential flow paths in upland and lowland 

areas in sandy soils.

This paper describes preliminary results of ongoing 

studies by the U.S. Geological Survey of unsaturatedzone properties at the Management Systems Evaluation Area (MSEA) near Princeton, Minn. (fig. 1).

565
Planimetric View Dye Application 

Area

Three-dimensional View

Corn Plant

Application 

Area

Corn Row

6 Meters 

 ¥ f V ¥ ¥ *"¥ *

Multiport 

- Well

Furrow

Trench 

NOT TO SCALE''

Trench

~7x-' \ Sample 

X' Location

%*f E

f;^:Utfj«iicf:;si!#

////M.

Linear-mow:.. 

Imgaijpn system:

250 500 750 1000 FEET

100 200 300 METERS

EXPLANATION

Cropped area

Direction of ground-water flow, 

January 1991

Observation well, number indicates 

number of wells at site

Multiport well 

Building

The research area Is located In 

the northeast quarter of section 18, 

township T35N, range R26W.

Figure 1. Layout of the Princeton, Minnesota, Management Systems Evaluation Area.

566
Preliminary results of a dye-tracing and trenching 

study are presented. The study is being done at a topographically high (upland) site and a topographically 

low (lowland) site within the northernmost cropped 

area at the 65 ha Princeton MSEA (fig. 1). The upland 

and lowland sites are about 78 m apart and differ in elevation by 1.4 m (slope of about 0.02). During October 

1991, the water table was about 4 m below land surface 

at the upland site and 2.8 m below land surface at the 

lowland site. On the basis of well hydrograph analyses, 

ground-water discharge rates range from 10 to 20 cm/y 

at the Princeton MSEA (Delin and others, 1992a). 

Ground-water-flow gradients ranged from 0.0009 to 

0.002 during 1991. Ground-water flow is from southwest to northeast beneath the upland and lowland sites 

at a velocity that varies seasonally between 6 and 

15cm/d.

METHODOLOGY

Instruments used to make measurements in the 

unsaturated-zone can be installed either vertically in a 

borehole or horizontally in a trench wall. Horizontal 

installation in a trench wall is preferred to minimize disturbance of the unsarurated zone above the instruments, 

and to facilitate identification of lateral heterogeneities 

in soil properties. Heterogeneities in soil properties can 

produce preferential flow paths along which water and 

solutes move more rapidly than through the soil matrix.

Most gross heterogeneities can be identified visually 

within a trench, such as a gopher burrow filled with 

coarse sand in silty soil. Other heterogeneities, such as 

subtle differences in grain size, are commonly difficult to 

visually identify. Instruments installed in a soil heterogeneity can yield data that are not representative of typical 

unsaturated flow for the site.

A dye-tracing and trenching study was done at the 

upland and lowland sites to (1) identify zones of preferential and retarded water movement, (2) facilitate 

collection of soil samples to characterize unsaturatedzone properties, (3) select installation sites for unsaturated-zone instruments, and (4) estimate the transport 

velocity of water and solutes moving through the 

unsaturated and saturated zones (Delin and others, 

1992b). Rhodamine WT dye was used in this study 

because it has been used successfully as a tracer of 

water and agricultural chemicals moving through 

sandy soils similar to the Princeton MSEA (Omoti and 

Wild, 1979; Trudgill, 1987; and Kung, 1990). Results 

of the Kung (1990) study indicated that the dye adsorbs 

to the soil matrix, allowing for tracing of dominant 

paths of water movement in the unsaturated zone. The 

dye has also been used as a tracer of ground water in the 

saturated zone of alluvial sand aquifers (Sabatini and 

Austin, 1991).

A 3 percent solution by volume of the dye was 

applied uniformly to a 3.5 by 6 m area at each site 

(fig. 1). About 12 L of dye was applied at 10-day

TRENCHING 

BEGINS

DYE APPLIED

CU MULATI VE 

PRECIPITATION

DATA 

POINT

DYE CONCENTRATION

8/1 9/1 10/1 11/1

1991

Figure 2. Cumulative precipitation and concentration of dye in ground water 1.0 meter below the water table and 

2.5 meters downgradient of the dye-application area at the lowland site, Princeton, Minn., 1991.

567
intervals from July 5 through September 13, 1991 

(fig. 2) to ensure dye visibility and complete penetration through the unsaturated zone. Dye was also 

applied near several aluminum neutron access tubes 

installed at each site. Total rainfall from July 5 through 

October 22, 1991, was about 30 cm (fig. 2).

Beginning on October 22,1991, a 3- by 2-m trench 

was dug in the middle of each dye-application area. 

The depth of each trench was increased by intervals of 

0.5 m to a total depth of 2.0 m. The long walls of each 

trench coincided with a corn row. Photographs of the 

sides and bottom of each trench were taken to document the visible distribution of dye in the unsaturated 

zone. Following dye application, water samples were 

collected periodically from a multiport well located 

2.5 m horizontally downgradient of each dye-application area (fig. 1). Each multiport well consisted of six 

0.6-cm diameter stainless-steel tubes housed in a 

5.1-cm inside diameter (i.d.) polyvinyl chloride (PVC) 

casing; each stainless-steel tube had a 3-cm-long 

screened interval (port) which was external to the PVC 

casing. Five sampling ports were installed below the 

water table at 0.5 m intervals. Water samples were collected periodically from each port and analyzed for dye 

concentration to estimate the transport time of water 

and solutes moving through the unsaturated zone.

Soil samples, 20 cm3 in volume, were collected 

from the sides and bottom of each trench for measurements of grain-size distribution, bulk density, total 

organic-carbon content, volumetric moisture content, 

and dye fluorescence. About half of the samples were 

located beneath the corn rows and half beneath the 

furrows. About 25 samples were collected at regularly 

spaced intervals of about 1.0 m horizontally and 0.5 m 

vertically from the northern and eastern walls of each 

trench. About 25 and 60 samples were collected at 

irregular spacing to facilitate geostatistical analysis of 

the data from the southern and western walls of each 

trench, respectively. About 20 samples were collected 

at regularly spaced intervals from the bottom of each 

0.5-m depth interval in the trench to determine the 

three-dimensional distribution of dye. Fourteen soil 

samples were collected adjacent to and 0.05 m away 

from a tube and analyzed for dye fluorescence to evaluate the effects of the neutron access tubes on preferential water movement.

Thirty-two undisturbed soil cores were collected 

from the west walls of the trenches and at several 

depths beneath the corn rows and furrows for measurements of saturated hydraulic conductivity. Most of 

these samples were obtained by pushing a 5-cm i.d. by 

568

8-cm-long thin-wall brass tube horizontally into the 

wall of the trench. A vertically oriented core was also 

obtained adjacent to three of the horizontally oriented 

cores to estimate the vertical hydraulic conductivity. 

Saturated hydraulic conductivities (horizontal and 

vertical) were measured in a triaxial system with the 

constant-flow method (Olsen and others, 1988; 1991). 

Effective stresses were varied between 0.35 and 2.81 

kg/cm2 Each sample also was analyzed for grain-size 

distribution and bulk density.

About 80 undisturbed cores were collected from 

the two trenches for measurements of moisture-retention characteristics. These samples were obtained by 

pushing a 5-cm i.d. by 8-cm-long thin-wall steel tube 

into the wall of the trench. The cores were placed in 

Tempe cells and moisture retention was determined at 

tensions of 25, 50, 75, 100,125, 150, 175, and 400 cm 

of water, according to the method of Richards (1965).

RESULTS AND DISCUSSION

The unsaturated zone at both the upland and lowland sites mostly consists of well-sorted fine- to 

medium-grained sand. Although the upper 1 m of soil 

at both sites is similar and composed of about 95 percent sand and 5 percent silt and clay, the unsaturated 

zone at the upland site is generally better sorted than at 

the lowland site (fig. 3; table 1). The grain size is 

coarser between the 1.0- and 2.0-m depths at the lowland site than at the upland site (fig. 3; table 1). Gopher 

burrows in the upper 0.5 m at the upland site are filled 

with sand grains that have diameters about 0.03 mm 

larger than grains in the surrounding matrix. Discontinuous iron-rich layers are present in the unsaturated 

zone between the 1 - and 2-m depths at the upland site. 

The layers are 2- to 20-cm thick, are cemented with 

iron oxides, and contain an average of about 5 to 10 

percent silt and clay compared to less than about 5 percent in adjacent horizons (M.D. Tomer, University of 

Minnesota, Soil Science Department, oral commun., 

1993).

The total organic carbon (TOC) content at every 

stratigraphic horizon in the upper 0.6 m at the lowland 

site was typically about twice as large as that at the 

upland site (table 1). The TOC ranged from 0.38 to 1.10 

percent at the lowland site and from 0.20 to 0.62 percent 

at the upland site (J.A. Lamb, University of Minnesota, 

Soil Science Department, written commun., 1993).

Most bulk densities ranged from 1.54 to 1.60 g/cm3 

throughout the 2-m profile and were generally similar 

at each site (table 1). No significant difference in bulk 

density was evident between areas of coarse-grained
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Table 1. Summary of unsaturated-zone properties for the upland and lowland sites at the

Princeton, Minn., Management Systems Evaluation Area

[cm/d, centimeters per day; number of analyses in parentheses, where appropriate]

Unsaturated zone property

Thickness, meters

Iron-rich, cemented layers present

Grain size in upper meter (median), millimeters

Grain size between 1 and 2 meters (median), millimeters

Total organic carbon content in upper 0.6 meters (median), 

percent

Bulk density in upper meter (median), October 22, 1991, 

grams per cubic centimeter

Volumetric moisture content in upper meter (median), 

October 22, 1991, dimensionless

Volumetric moisture content between 1 and 2 meters 

(median), October 22, 1991, dimensionless

Saturated hydraulic conductivity in upper meter (median), 

centimeters per second

Saturated hydraulic conductivity between 1 and 2 meters 

(median), centimeters per second

Approximate dye transport time through the unsaturated 

zone, July 5 through October 23, 1993, cm/d

Average depth of visible dye beneath corn rows, meters

Average depth of visible dye beneath furrows, meters

Upland Site

4.0

yes

0.19(27)

.20 (13)

.35(8)

1.56 (89)

.09 (89)

.08 (33)

4.3xlO-3 (5)

5.7xlO-3 (5)

5.3

.1

.4

Lowland Site

2.8

no

0.23 (57)

.40 (41)

.74(8)

1.58(57)

.12(57)

.06(41)

3.5xlO-3 (7)

7.2xlO'3 (5)

3.7

.04

.2

sand filling gopher burrows and the surrounding finegrained matrix.

Volumetric moisture contents in the upper 1 m, and 

particularly in the upper 0.6 m, at the lowland site were 

slightly greater than at the upland site (table 1). Most 

moisture contents ranged from 0.07 to 0.11 at the 

upland site and from 0.05 to 0.11 at the lowland site. 

Below a depth of 1 m, moisture contents at the lowland 

site were slightly less than those at the upland site 

(table 1). The smaller moisture contents below a 1-m 

depth at the lowland site are likely due to the coarser 

grain size compared to that at the upland site.

Saturated hydraulic conductivities (K) at the 

upland and lowland sites generally reflect the measured 

differences in grain-size distribution and TOC content. 

The median horizontal K in the upper 1 m of the unsaturated zone is slightly greater at the upland site than at 

the lowland site (fig. 3; table 1), probably because the 

TOC content is greater at the lowland site than at the

upland site. Because the grain size is coarser below 1 m 

at the lowland site, the median horizontal K is about 

twice as large as the horizontal K for the same depth 

interval at the upland site. The median vertical K of 

5.4xlO'3 cm/s, based on the samples collected vertically, is similar to the median horizontal K (table 1).

Preliminary results of moisture-retention studies 

are available only for the upland site. The average 

moisture-retention curve plotted from data for 31 samples is shown in figure 4 and is typical of a mediumgrained sand. The standard deviation decreases with 

decreasing tension and increasing moisture content 

between soil tensions of 0 and 25 cm. The standard 

deviation is greatest between soil tensions of 25 and 

75 cm and is relatively constant between soil tensions 

of 175 and 400 cm.

The distribution of dye in the unsaturated zone was 

highly variable at both sites. The average depth of 

visible dye beneath the corn rows and furrows at the

570
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LU

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en

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LU

5 0.3 -

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o

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Average moisture retention

100 200 300 

SOIL TENSION, IN CENTIMETERS

Figure 4. Average moisture-retention characteristics of 31 unsaturated-zone soil samples from the upland site, 

Princeton, Minn.

400

upland site was about twice as large as the lowland site 

(table 1). The dye was likely adsorbed to the organic 

matter, the content of which was greater at the lowland 

site than at the upland site. The maximum depth of visible dye in the soil matrix at both sites was about 1.0 m. 

Dye movement was greatest beneath the furrows and 

least beneath the corn rows. The greater dye movement 

beneath the furrows was largely due to runoff of the 

applied dye solution from the side-slope areas of adjacent corn rows. Because the dye exhibited hydrophobicity and did not easily infiltrate the soil surface, 

runoff from the corn rows was accentuated. Dye moved 

preferentially in the vicinity of corn plant roots. These 

areas of visible dye movement were narrower (3-10 cm 

in diameter) than the areas of dye movement beneath 

the furrows (20-40 cm in diameter).

The dye generally moved below land surface in 

vertically elongated pulses rather than as a uniform 

front to a single depth. Most of these pulses occurred 

below areas of topsoil that were about 1 to 3 cm lower 

than the surrounding topsoil. Because of runoff, dye 

collected in these microtopographic depressions, 

resulting in preferential movement of dye below land 

surface.

Dye visibly moved around heterogeneities in the 

unsaturated zone, such as the upland site gopher burrows that were filled with sand that was coarser than the 

surrounding soil matrix. The dye moved around gopher 

burrows that were less than about 8 cm in diameter. 

This phenomenon of unsaturated movement of water 

around coarser sand was first described by Haines 

(1930) and has been termed funnel flow (Kung, 1990). 

Water moves into the coarser sand when the suction of 

water within pores in the overlying finer sand is greater 

than it is in the coarse sand. This movement of water 

into the coarser sand is termed a Haines jump (Haines, 

1930). Dye movement indicated that a Haines jump 

occurred through the middle of gopher burrows that 

were greater than about 10 cm in diameter.

The median dye concentration adjacent to neutron 

access tubes was about 2 times greater than the median 

concentration at a radial distance of 0.05 m. The greater 

dye concentrations adjacent to the neutron access tubes 

probably resulted from development of plant roots 

around the tubes along which dye moved preferentially. Warmer air temperatures within each tube, compared to the ambient soil temperature, likely caused the 

roots to grow around the tubes. These results could 

indicate that soil-moisture contents determined by use

571
of the neutron access tubes are greater than the average 

volumetric moisture content in the surrounding soil 

matrix. Additional research is needed to evaluate these 

apparent differences in moisture content.

About 100 days after application, the dye was 

detected at all five sampling ports in the multiport well 

between 0.5 and 2.0 m below the water table at the 

upland and lowland sites. Figure 2 illustrates the concentration of dye in ground water about 1.0 m below 

the water table at the lowland site from July 5 through 

November 5,1991. Concentrations of dye ranged from 

0.01 to 0.8 ug/L from late October through December

1991. and the dye was not detected after January 9,

1992. The greatest concentration of dye was detected 

ori October 23, 1991, at the lowland site (0.5 ug/L) and 

on January 9, 1992, at the upland site (0.8 ug/L).

Assuming that the average ground-water velocity 

was about 10 cm/d and that some of the dye moved 

conservatively with the ground water, the travel time 

for the dye from the point where the dye reached the 

water table to the multiport well sampling ports (2.5 m) 

was about 25 days. Subtracting this 25-day travel time 

through the saturated zone, from the total travel time of 

about 100 days, yields an estimated 75-day conservative travel time for dye through the unsaturated zone. 

Therefore, the velocity of dye through the unsaturated 

zone was about 3.7 cm/d at the lowland site and 

5.3 cm/d at the upland site on the basis of unsaturatedzone thicknesses of 2.8 and 4.0 m, respectively (table 1).

The dye moved 2 m vertically through the saturated zone over a horizontal distance of about 9 m, 

whereas less than 0.2 m of vertical movement was predicted with a steady-state ground-water-flow model of 

the saturated zone. Detection of the dye in all sampling 

ports at both sites indicates a much steeper vertical 

component of ground-water flow than that predicted by 

the conceptual ground-water-flow model used for the 

MSEA study.

SUMMARY AND CONCLUSIONS

The distribution of dye in the unsaturated zone was 

highly variable at both sites. The average depth of 

visible dye beneath the com rows and furrows at the 

upland site was about twice as large as at the lowland 

site. The dye was likely adsorbed to the organic matter, 

which was greater at the lowland site than at the upland 

site. Dye movement was greatest beneath the furrows 

and least beneath the corn rows. The greater dye movement beneath the furrows was largely due to runoff of 

the dye solution from the side-slope areas of adjacent

corn rows. The dye generally moved below land surface in vertically elongated pulses rather than as a uniform front at one depth. Most of these pulses occurred 

below areas of topsoil that were about 1 to 3 cm lower 

than the surrounding topsoil. Dye moved preferentially 

in the vicinity of corn plant roots; these areas, however, 

were narrower than the areas of dye movement beneath 

the furrows. Dye moved around heterogeneities in the 

unsaturated zone at the upland site, such as gopher burrows that were filled with sand that was coarser than the 

surrounding soil matrix. Dye moved around gopher 

burrows that were less than about 8 cm in diameter. 

Dye movement indicated that a Haines jump occurred 

through the middle of gopher burrows that were greater 

than about 10 cm in diameter. The median dye concentration adjacent to neutron access tubes was about 2 

times greater than the median concentration at a radial 

distance of 0.05 m. These results may indicate that soilmoisture data collected from neutron access tubes are 

greater than the average volumetric moisture content in 

the surrounding soil matrix. Visible dye distribution 

was not strongly correlated with soil bulk density, 

saturated hydraulic conductivity, moisture-retention 

characteristics, or volumetric moisture content.

About 100 days after application, the dye was 

detected at all five sampling ports between 0.5 and 2.0 m 

below the water table at the upland and lowland sites. 

The transport time for the dye from the point where the 

dye reached the water table to the multiport well sampling ports (2.5 m) was about 25 days. The transport 

velocity of dye through the unsaturated zone was about 

3.7 cm/d at the lowland site and 5.3 cm/d at the upland 

site on the basis of unsaturated-zone thicknesses of 2.8 

and 4.0 m, respectively. The dye moved 2 m vertically 

through the saturated zone over a horizontal distance of 

about 9 m, whereas less than 0.2 m of vertical movement was predicted with a steady-state ground-waterflow model of the saturated zone.

REFERENCES

Delhomme, J.P., 1979, Spatial variability and uncertainty in 

groundwater flow parameters, a geostatistical 

approach: Water Resources Research, v. 15, p. 269-280.

Delin, G.N., Landon, M.K., Anderson, J.L., and Dowdy, 

R.H., 1992a, Hydrologic research at the Princeton, 

Minnesota, Management Systems Evaluation Area: 

U.S. Geological Survey Open-File Report 92-107, 2 p.

Delin, G.N., Landon, M.K., Healy, R.W., and Olsen, H.W., 

1992b, Preferential flow through the saturated zone 

beneath a corn field near Princeton, Minnesota: in EOS, 

Transactions, Proceedings of American Geophysical 

Union spring meeting, Montreal, Canada, May 12-16, 

1992, v. 73, no. 14, p. 132.

572
Haines, W.B., 1930, Studies in the physical properties of 

soils, V The hysteresis flow in capillary properties 

and the modes of moisture distribution associated 

therewith: Journal Agricultural Science, v. 20, p. 97- 

116.

Kung, K.-J.S., 1990, Preferential flow in a sandy vadose 

zone, field observation: Geoderma, v. 46, p. 51-58.

Meyboom, P., 1966, Unsteady groundwater flow near a 

willow ring in hummocky moraine: Journal of Hydrology, v. 4, p. 38-62.

Olsen, H.W., Gill, J.D., Willden, A.T., and Nelson, K.R., 

1991, Innovations in hydraulic-conductivity measurements: Geotechnical Engineering: National Research 

Council, Transportation Research Record 1309, 

p. 9-17.

Olsen, H.W., Morin, R.H., and Nichols, R.W., 1988, Flow 

pump application in triaxial testing: in Donaghe, R.T., 

Chancy, R.C., and Silver, M.L., eds., Advanced triaxial 

testing of soil and rock: Philadelphia, American Society 

for Testing and Materials, ASTM STP 977, p. 68-81.

Omoti, U., and Wild, A., 1979, Use of fluorescent dyes to 

mark the pathways of solute movement through soils 

under leaching conditions, field experiments: Soil 

Science, v. 128, no. 2, p. 98-104.

Richards, L.A., 1965, Physical condition of water in soil, in 

Black, C.A., and others, eds., Methods of soil analysis: 

Madison, Wise., American Society of Agronomy, 

p. 128-152.

Russo, D., and Bresler, E., 1981, Soil hydraulic properties 

as stochastic processes, I An analysis of field spatial 

variability: Soil Science of America Journal, v. 45, 

p. 682-687.

Sabatini, D.A., and Austin, T.A., 1991, Characteristics of 

rhodamine WT and fluorescein as adsorbing groundwater tracers: Ground Water, v. 29, no. 3, p. 341-349.

Sharma, M.L., Grander, G.A., and Hunt, C.G., 1980, Spatial 

variability of infiltration in a watershed: Journal of 

Hydrology, v. 45, p. 101-122.

Trudgill, S.T., 1987, Soil water dye tracing, with special 

reference to the use of rhodamine WT, lissamine FF, 

and amino G acid: Hydrological Processes, v. 1, 

p. 149-170.

573
574
Nitrogen Uptake and Soil Water Variability Across a

Sand Plain Landscape

Abstract

By J.A. Lamb1 , M.D. Tomer2, J.L Anderson3 , and R.H. Dowdy4

INTRODUCTION

Research at the Northern Cornbelt Sand Plains 

Management Systems Evaluation Area at Princeton, Minnesota, was conducted to characterize the 

variability of whole-plant nitrogen uptake and 

soil-water content under a continuous corncropping system. The soil was a Zimmerman fine 

sand (mixed, frigid, Alfic Udipsamment). Nitrogen fertilizer was applied uniformly in 1991 and 

1992. In 1991 and 1992, corn whole-plant samples 

were collected in 15-meter intervals along a 244- 

meter-long transect to determine nitrogen uptake. 

In 1992, plastic neutron-moisture-meter access 

tubes were placed to a depth of 2 meters every 20 

meters along the same transect. Nitrogen uptake 

ranged from 115 to 157 kilograms nitrogen per 

hectare in 1991 and 92 to 132 kilograms nitrogen 

per hectare in 1992. Mean uptake of nitrogen was 

24 kilograms per hectare greater in 1991 than it 

was in 1992. Soil-water content ranged from 10 to 

15 centimeters for the upper 1.7-meter depth interval on July 23, 1992. The area where nitrogen 

uptake in 1992 was least is a sideslope area where 

the surface soil was driest. Leaching may have 

occurred during recharge events during the 

growing season at this and positions in the terrain.

1 Associate Professor, Soil Science Department, University of 

Minnesota, St. Paul, MN.

2 Research Fellow, Soil Science Department, University of 

Minnesota, St. Paul, MN.

3 Professor and Director of Center for Agricultural Impacts 

on Ground Water Quality, Soil Science Department, University of 

Minnesota, St. Paul, MN.

4 Research Leader and Soil Scientist with U.S. Department of 

Agriculture-Agricultural Research Service, St. Paul, MN.

The Northern Cornbelt Sand Plains Management 

Systems Evaluation Area (MSEA) program is a multiagency, multistate initiative to evaluate the effects of 

present and modifications to present agricultural management systems on water quality in a sand plain areas 

in Minnesota, North Dakota, South Dakota, and 

Wisconsin. The Northern Cornbelt Sand Plains MSEA 

program is a cooperative study primarily between the 

U.S. Department of Agriculture-Agricultural Research 

Service, the University of Minnesota Soil Science 

Department, and the U.S. Geological Survey.

Ground water contamination from agricultural 

management systems is caused by numerous complex 

processes. An objective of the MSEA project is to evaluate agricultural systems with the goal of minimizing 

adverse effects of agricultural management practices 

on ground water. Nitrogen (N) fertilizer is necessary 

for profitable corn production; however, N fertilizers 

combined with N mineralized from soil organic matter 

can be major sources of ground-water contamination. 

N fertilizer traditionally has been uniformly applied on 

production fields. The Northern Cornbelt Sand Plains 

MSEA has been investigating the effect of landscape 

variability on N uptake by crops and on soil-water 

content. N in soils and fertilizer can be removed by processes such as plant uptake and removal, leaching of 

nitrate-N through the soil, erosion of soil, and denitrification by bacteria. Because of typically high infiltration rates for the sandy soils at the Northern Cornbelt 

Sand Plains MSEA, offsite losses of N resulting from 

erosion are negligible. The purpose of this paper is to 

examine the relation of N uptake by corn to soil-water 

content across a typical sand-plain terrain.

MATERIALS AND METHODS

The Northern Cornbelt Sand Plains MSEA, 

Princeton, Minn., site (45°31 1 34"N. and 93°371 8"W.) 

was established in 1991 on the Anoka Sand Plain

575
Princetonv Minh

«P-v .>i4 ;: >*l^sss^ 5 Kilometer^

«*liK-' **- .-*J4" t .

Unjeaf-rnove 

Irrigation system

1

t

250 500 750 1000 FEET

100 200 300 METERS

EXPLANATION

Cropped area 

Research area

A' Transect 

Building

The research area is located 

in the northeast quarter of 

section 18, township T35N, 

range R26W.

Figure 1. Princeton, Minnesota, Management Systems Evaluation Area and 

transect used for nitrogen-uptake and soil-water-content measurements.

located 90 kilometer (km) northwest of Minneapolis, 

Minn. The predominant soil is a Zimmerman fine sand 

(mixed, frigid, Alfic, Udipsamment). Topographic 

relief at the 64 hectare (ha) Princeton MSEA is less 

than 3 meters (m). Generally, the unsaturated zone 

consists of fine- to medium-grained sand, and the saturated zone consists of medium- to coarse-grained sand. 

Discontinuous layers of silt and very fine-grained sand 

as thick as 20 centimeters (cm) are present in the unsaturated and saturated zones. A predominantly clayey till 

underlies the surficial aquifer. During 1991, the average 

depth to the water table was about 3.7 m below land

surface, and the saturated thickness ranged from 4 to 

16 m. The direction of ground water flow generally 

is from west to east at about 8 centimeters per day 

(cm d" 1 ). Ground-water recharge rates generally range 

from 10 to 20 centimeters per year (cm yr~').

Five 1.8- to 2.7-ha cropped areas are aligned with 

the predominant direction of ground-water flow at the 

Princeton MSEA (fig. 1) where three agricultural 

management systems are being evaluated: (1) Cornsoybean crops are rotated under ridge (conservation) 

tillage, split N application, N application adjusted for N 

mineralized from previous legume crop (soybean), and

576
banding of herbicides (application of herbicide directly 

over the plant row. Application area is a third of area 

covered by a broadcast application, the area where herbicide is applied receives it at the same rate) in cropped 

areas A and C. (2) Sweet corn and potato crops are 

rotated with conventional full-width (disk or chisel) 

tillage and a similar "banded" application of herbicides 

for sweet corn and conventional full-width tillage and 

broadcast application of herbicides for potatoes in 

cropped areas B and D. (3) Corn is grown in consecutive years (continuous corn) under conventional fullwidth tillage and broadcast application of herbicides in 

cropped area E. The buffer around the cropped areas 

(fig. 1) was planted with a mixture of timothy and smooth 

brome grass to which agricultural chemicals were not 

applied. Before the implementation of the MSEA agricultural management systems in spring 1991, the entire 

67 ha field was planted in alfalfa during 1981 -89 and in 

corn during 1990. Additional details are in Landon and 

others (1993) and Anderson and others (1991).

The continuous corn system was used for this 

study. N fertilizer, irrigation water, and pesticides were 

uniformly applied in the 2.7-ha crop area. The corn was 

fertilized with a total of 156 kilograms of nitrogen per 

hectare (kg N ha' 1 ) in 1991 and 168 kg N ha-' in 1992. 

An adapted corn hybrid was planted during the first 

week of May in 1991 and 1992. A transect was established 9 m (10 crop rows) north of the southern boundary of the continuous corn-crop area (fig. 1). The 244- 

m-long transect was chosen to represent a variety of 

terrains. Plastic access tubes were installed to a depth 

of 2 m along the transect with 20-m spacing in 1992.

W 160

£ EH 150

o o

53 K 140

O «

9" 130 i

E2 120

§ 110 

O

H 100 +

H

>J53 90

O H

These access tubes were used to measure soil-water 

concentration on a volumetric basis with a neutron 

moisture meter in multiple-day sequences that began 

after previous significant rainfall or irrigation. The soilwater content for a specific soil layer was determined 

by multiplying the soil water concentration by the 

thickness of the soil layer from which the data were 

derived.

Samples of above-ground corn-plant biomass 

(grain plus stover) were collected at physiological 

maturity from 2-m by 1 -m areas located every 15 m 

along the transect. N concentration was measured by 

means of Kjeldahl digestion techniques. Whole-plant 

N uptake was then calculated by multiplying the 

biomass times N concentration.

RESULTS AND DISCUSSION

Whole-plant N uptake along the transect at physiological maturity for 1991 and 1992 is illustrated in 

Jigure 2. N uptake ranged from 115 to 157 kg N ha' in 

1991 and 92 to 132 kg N ha' 1 in 1992. Mean N uptake 

was 24 kg N ha' 1 greater in 1991 than it was in 1992 

(table I). This difference in N uptake was caused by the 

reduced N concentration in 1992 as a result of cooler 

growing conditions compared to those in 1991. Wholeplant biomass for both years was similar (table 2). The 

coefficient of variation for N uptake in 1992 (9.7 percent) 

was slightly greater than that in 1991 (9.1 percent). N 

uptake was not the same along the transect in each year 

although there were some similarities. Greater N 

uptake occurred in areas 35 to 40 m, 90 to 100 m, and

Table 1 . Descriptive statistics for corn whole-plant 

nitrogen uptake in 1991 and 1992 at the Northern 

Cornbelt Sand Plains Management Systems Evaluation 

Area near Princeton, Minnesota

Year

I99I 

I9Q2

Nitrogen uptake 

(kilograms per hectare)

Mean

1 36 

II2

Range

II5-I57 

92- I32

Coefficient 

of variation 

(percent)

Q.I 

Q.7

0 50 100 150 200 250 ?

DISTANCE FROM WESTERN END OF FIELD, IN METERS

Figure 2. Corn whole-plant nitrogen uptake in kilograms per hectare 

for continuous corn in 1991 and 1992 at the Princeton, Minnesota 

Management Systems Evaluation Area along transect A-A' (transect 

shown in fig. 1).

Table 2. Descriptive statistics for corn whole-plant 

biomass in 1991 and 1992 at the Northern Cornbelt Sand 

Plains Management Systems Evaluation Area near 

Princeton, Minnesota

Year

Plant biomass 

(megagrams per hectare)

Mean Range

Coefficient

of variation

(percent)

IQ9I 

IQQ2

14.3

14.5

II.6-I8.I 

10.7- 18.2

I2.I 

13.6

577
COf±

LU 

LU

LU 

LU

LJJ 

CO 

LJJ

o

CO

<

LJJ 

Q

0 25 50 75 100 125 150 175 200 225 250 

DISTANCE, IN METERS FROM WESTERN END OF FIELD

EXPLANATION

VOLUMETRIC WATER CONTENT, IN PERCENT

greater 

0-5 5-10 10-15 15-20 20-25 than 25

Figure 3. Cross section of the (a) wettest and (b) driest soil-water conditions measured 

during the 1992 growing season along transect A-A' (transect shown in fig. 1).

578
160 to 165 m from the western end of the field in both 

years, whereas less than average N uptakes occurred in 

the area 110 to 150 m along the transect (fig. 2). The N 

uptakes are local maximums in 1991 and local minimums in 1992 at three locations along the transect (23, 

69, and 177 m). The N uptakes are local maximums in 

1992, local minimums in 1991 at two locations along 

the transect (54 and 192 m). These results suggest that 

the large differences in N uptake occur in similar places 

each year. There is not a strong correlation between 

years and location for N uptake (r = 0.24).

Soil in the vadose zone is not homogeneous. Thin 

illuvial horizons (lamellae) are present in the 1- to 

1.7-m-depth interval throughout the Princeton, Minn.. 

MSEA, usually in topographically high terrain. These 

layers may retard or divert soil-water movement, 

thereby affecting potential loss of N by leaching and 

plant N uptake. The spatial variability of soil-water 

content illustrated in figure 3 is caused by the perching 

of soil water above the lamellae. The greatest perching 

of soil water occurred at the hilltop position in the area 

125 and 175 m along the transect (fig. 3). Rates of 

water loss caused by crop use or drainage after rainfall 

or irrigation were slower where the degree of perching 

was greatest. During dry periods in the growing season, 

soil-water content in the upper 1-m depth interval was 

least along sideslope positions, which coincide with 

areas of least N uptake, particularly the west-facing 

slope in the area 110 to 140 m along the transect. N 

uptake by plants may have been limited by poor 

surface- and subsurface-water retention relative to that 

at other locations, resulting in reduced crop growth and 

an increased potential for leaching of N.

After irrigation on July 23, 1992, soil-water content in the 1.7-m depth interval ranged from 10 to 

15 cm. Soil-water content was least at the western end 

of the transect, where subsoils are coarse and lamellae 

are absent, and greatest in upland positions where soilwater perching occurs. Soil-water content also was 

small on sideslope positions where the surface soil 

horizon is thinner than it is along the rest of the 

transect. Water loss from crop use and drainage during 

July 23 -August 3, 1992, ranged from 3.5 to 5.5 cm. If 

the differences in soil-water content and soil water 

losses from crop use and drainage were considered, 

there is a potential for variations in the amount of N 

that could be transported vertically to ground water,

even over a relatively short horizontal distance of the 

transect (244 m).

Considering soil-water content and N uptake, the 

area where plant uptake of N was least was also the 

area where the potential for leaching of N was greatest 

(sideslopes). The variability in plant N uptake and soilwater content indicate that the processes involved in 

the removal of N from the soil and the transport of N to 

the ground water are complex. Accordingly, monitoring the effects of an agricultural-management practice 

on ground-water quality and in identifying the predominant process affecting the movement of N to the 

ground water will be difficult. These results strongly 

suggest that knowledge of the location of a particular 

site with a managed area is important in interpreting 

effects of agricultural-management systems.

SUMMARY

Variability in corn N uptake and soil-water content 

was very complex along a 244-m-long transect at the 

Northern Cornbelt Sand Plains MSEA near Princeton, 

Minn. N uptake by corn had ranged from 115 to 157 kg 

ha' 1 in 1991, and 92 to 132 kg ha" 1 in 1992. The mean N 

uptake was 24 kg ha" 1 greater in 1991 than it was in 

1992, indicating that annual fluctuations in N uptake 

occur and were similar in 1991 and 1992 as measured by 

range in uptake and coefficient of variation. The correlation of N uptake and location along the transect during 1991 and 1992 was not good (r = 0.24). Soil-water 

content in the upper 1.7 m depth interval along the 

transect varied from 10 to 15 cm on July 23, 1992. The 

least N uptake occurred in 1991 and 1992 in a sideslope 

area where surface soils were dry because of a lack of 

subsurface lamella to slow vertical water movement. 

Therefore, the loss of N by leaching may be increased in 

sideslope areas.

REFERENCES

Anderson, J.L., Dowdy, R.H., and Delin, G.N., 1991, 

Ground water impacts from irrigated ridge-tillage, in 

Ritter, W.F., ed., Proceedings of the American Society 

of Civil Engineers: 1991 National Conference on 

Irrigation and Drainage Engineering, Honolulu, 

Hawaii, July 23-26, p. 604-610.

Landon, M.K., Delin, G.N., Lamb, J.A., and Guo, L., 1993. 

Ground-water quality at the management systems 

evaluation area near Princeton, Minnesota, 1991-1992: 

U.S. Geological Survey Open-File Report 93-80, 2 p.

579
580
Geophysical Investigations of Heterogeneity and Scale

at the Princeton, Minnesota, Management

Systems Evaluation Area

Abstract

By Jeffrey E. Lucius 1 and Gary R. Olhoeft 1

INTRODUCTION

The U.S. Geological Survey collected more 

than 1 gigabyte of ground penetrating radar (GPR) 

data in -1991 at the Princeton, Minnesota, Management Systems Evaluation Area (MSEA) to generate detailed images of the subsurface and to 

determine the three-dimensional spatial variability 

of hydrogeologic properties. The Princeton MSEA 

is located on the Anoka glacial outwash sand plain 

northwest of Minneapolis-St. Paul. The GPR 

system transmitted electromagnetic pulses that 

propagated through the ground. The pulses were 

partially reflected back to the GPR antennas when 

they encountered changes in electrical properties 

(which are controlled by water content, bulk 

density, lithology, and porosity). The GPR data 

were computer processed to produce geometrically correct images of the subsurface. Reflector 

continuity, amplitude, configuration, and spatial 

frequency were analyzed in the GPR images to 

determine lithologic structure, depositional 

processes, moisture content, stratification patterns, 

and grain-size distributions. Changes in watertable capillary fringe thickness, areas of possible 

focused recharge, areas consisting of eolian or 

other fine-grained deposition, and undulations of 

the till surface were identified. The GPR images 

were also correlated with lithology logs from 

sampling wells. For agricultural sites with very 

little or no clay near the surface, ground penetrating radar offers a fast, cost-effective, high-resolution method for extending information acquired at

1 U.S. Geological Survey; Denver, Colo.

Agricultural pesticides, herbicides, and fertilizers 

are intended for the plant and root zone of the shallow 

subsurface. Sometimes, however, these chemicals 

migrate downward into the water-saturated zone and 

affect ground-water quality. The MSEA program is 

part of a multiscale, interagency project to evaluate the 

effects of agricultural practices on ground-water quality. An objective of the MSEA program is to improve 

our understanding of the processes and factors affecting the transport and fate of agricultural chemicals in a 

particular hydrologic system. The U.S. Geological 

Survey (USGS) has been conducting investigations to 

determine the three-dimensional spatial variability of 

hydrogeologic properties by use of GPR. The overall 

objective was to determine the usefulness of geophysical methods to hydrologic investigations in an 

agricultural setting. The specific objectives were to 

(1) generate high-resolution images of the subsurface 

geology and (2) characterize the hydrogeologic heterogeneity in terms of length-scales and statistical distributions important factors in understanding threedimensional patterns of ground-water flow.

The investigations were conducted at the MSEA 

(fig. 1) located on the Anoka Sand Plain near Princeton, 

Minnesota (approximately 100 kilometers (km) northwest of Minneapolis and St. Paul) during January, April, 

and October of 1991. The near-surface material at the 

site is glacial outwash that consists of mixtures of fineto-medium and medium-to-coarse sand and gravel with 

discontinuous, thin silty layers (Delin and others, 1992). 

A loamy sand soil is present at the surface. The outwash is underlain by glacial till below the water table. 

Topography at the site is relatively flat; relief is about 

three meters (fig. 1). Average annual precipitation is 

about 76 centimeters (cm).

The GPR system used in this study was a Subsurface 

Interface Radar (SIR ) System-7 manufactured by 

Geophysical Survey Systems, Inc. The system

581
93°3?'30"W 93°37'l0.3"w

8

(O

45°31 / 49.5 //N

Explanation

r Well MC-11^ EDM station

? Trench Rl

- - Limits of north 

crop area

- - Limits of neutron 

access tube area

? Land surface

^\ topographic contour

45d 3l'23.4"N

Wooded 

Are

0 100 200 300 METERS

Princeton, Minn. 5 kilometers

Row 16

Row 8 Row 5B Row 1

lat. and long, from U.S. Geological 

Survey Princeton, Minnesota 

1:24,000 quadrangle map, 1982

0 250 500 750 1000 FEET

National Geodetic Vertical Datum of 1929 

Contour Interval 0.25 m

Figure 1. Land-surface topography at the Princeton, Minnesota, Management Systems Evaluation Area.
periodically transmits radio-frequency electromagnetic 

pulses into the ground. The pulses propagate through 

the earth and are partially reflected back to the GPR 

antennas when they encounter changes in relative 

dielectric permittivity and conductivity. These changes 

are dominantly controlled by water content, bulk density, lithology, and porosity which are in turn affected 

by sedimentological processes and hydrological conditions. Radar antennas having center-frequencies of 80, 

300, and 500 megahertz (MHz) were used. Depth of 

penetration of the electromagnetic pulses into the earth 

is approximately inversely proportional to antenna 

frequency. However, resolution of subsurface features 

approximately proportional to antenna frequency. Surveys were performed by slowly pulling the antennas in 

a pickup truck across the ground surface. The reflected 

signals were recorded on digital magnetic tape. A 

continuous recording acquired on a straight heading is 

called a traverse.

In January 1991, GPR data were collected on 40 

west-to-east traverses spaced 20 m apart across the 

truck-accessible parts of the MSEA using 300-MHz 

antennas to obtain a comprehensive view of the site 

subsurface. In April 1991, traverses using the 80-MHz 

antennas were conducted on selected areas of the site 

and on adjacent roads to partly determine the variability of the underlying till. A detailed survey of the northernmost crop area (fig. 1) was also performed using the 

300-MHz antennas. In October 1991, data from 300- 

MHz and 500-Mhz antennas were collected over one of 

the areas to be trenched (trench Rl, fig. 1). Data also 

were collected near the neutron-probe access tubes in 

the spatial variability plot (fig. 1) to compare GPR data 

to the moisture variations measured by neutron probes. 

In all, more than a gigabyte of GPR data was collected. 

More than 3,500 stations were topographically surveyed to record the locations of the traverses.

The purpose of this paper is to (1) present a sample 

of the subsurface images generated from GPR data that 

show local variations in hydrogeology, (2) compare 

GPR data to lithology logs of wells at the site, and (3) 

discuss the utility of GPR for agricultural-site characterization. Chi-square tests were performed on GPR 

reflector-length histograms of adjacent 20-m-long sections along selected traverses to quantify the heterogeneity of the subsurface and to identify sections of a 

traverse that are statistically similar. Those results were 

presented in Lucius and Olhoeft (1993) and are not discussed in detail here.

GROUND PENETRATING RADARGENERATED CROSS-SECTIONS

The GPR horizontal-sampling interval along the 

traverses was about 2 to 3 cm; this interval varied 

slightly according to antenna towing speed. The vertical sampling interval depends on how the GPR system 

was adjusted; and ranged from about 2 to 4 cm. Vertical 

resolution is a function of antenna frequency, signal 

wavelength in the ground, data-collection rate, and 

speed of the radar-wave propagating in the ground. 

Although the GPR system can detect thin layers, those 

layers must be separated by about one wavelength or 

constructive and destructive interference of closelyspaced reflected waveforms will complicate interpretation. The estimated radar wavelengths in the ground 

(using a relative dielectric permittivity of 4 measured 

in dry sand at the site) were 1.8 meter (m) for the 80- 

MHz antennas, 0.5 m for the 300-MHz antennas, and 

0.3 m for the 500-MHz antennas.

The GPR data were computer processed to correct 

for geometric distortions inherent in the data-collection 

process and to produce a more tractable data set for 

characterization of the subsurface or estimation of the 

hydraulic properties. A background trace (the average 

of all traces in a traverse) was subtracted from each raw 

trace to minimize system noise in the data. A runningaverage filter, approximately 0.5 in in length, was 

applied to the raw data to improve the signal-to-noise 

ratio. For display, data placement was horizontally corrected for variations in towing speed. Then a time-todepth conversion was determined (based on bulk 

relative dielectric permittivity of 4 above the water table 

and 6.5 below the water table and above the till) and 

used to convert the recorded time sections into approximate depth sections; static corrections were applied to 

adjust for elevation changes. The result was (approximately) geometrically correct images of the subsurface.

Gray-scale images of processed 300-MHz data 

along two west-to-east traverses are shown in figure 2 

(locations of the two traverses are shown in fig. 1). Elevations are approximate because the time-to-depth 

conversion (using a relative delectric permittivity of 4 

for the entire section) is approximate. The vertical 

exaggeration of 15 causes the dips of many features 

and the topography to appear much steeper than they 

really are. Reflections that are either very light or very 

dark in the GPR images indicate subhorizontal features. The banding in the top 1 in of each image is 

caused by near-field interference of the 300-MHz radar 

wave and is not due to subsurface reflections.

583
WEST

300

WT

285

100 200 300 400 500 

HORIZONTAL DISTANCE, IN METERS

DATUM IS SEA LEVEL, VERTICAL EXAGGERATION x!5

600

EAST

700 780

Figure 2. Gray-scale images of computer-processed 300-MHz ground penetrating radar data along west-to-east traverses, rows 1 and 

8, at the Princeton, Minnesota, Management Systems Evaluation Area. (Locations are shown in fig. 1.)
The strongest and most continuous reflector in these 

two images is from the water table, located at about 

294.5 m above sea level. The water-table reflector is 

easy to identify in the image of row 1 between 70 to 400 

m horizontal distance and in the image of row 8 between 

50 and 470 m horizontal distance. This suggests the capillary fringe is thin and that sediments near the water 

table are coarse-grained. East of these areas in both 

images, the water-table reflector becomes harder to identify because the relatively thick capillary fringe in finegrained sediment causes a gradual transition in dielectric 

properties and reduces the clarity of the reflection.

The remaining reflections in these two images are 

generated by small changes in water content above the 

water table and by changes in lithology and bulk density below the water table. Analysis of reflection facies 

(see, for example, Sangree and Widmier, 1979; Beres 

and Haeni, 1991) can be applied to these images. The 

area west of 400 m horizontal distance in row I is dominated by simple, layered reflections of moderate 

amplitude and low spatial frequency. This indicates 

thin to moderately thick, horizontally layered, coarse 

sand and gravel with only slight contrasts in electrical 

properties. East of 400 m, the reflections are more 

chaotic, having higher amplitude and spatial frequency, 

suggesting the presence of mixed layers of sand, silt, 

and gravel of variable thickness and extent.

For row 8, the water-table reflector between 160 

and 320 m horizontal distance, at an elevation of about 

294 m, appears to bulge upward near the low area at 

240 m horizontal distance, indicating this low area 

might be a source of enhanced or focused recharge. 

Delin and Landon (1993) have discovered this to be 

true for one lowland area on the site and indicate that 

focused recharge could be occurring at other lowland 

areas. These areas might contribute to focused transport of agricultural chemicals to the ground water. At 

600 m horizontal distance, there appears to be a 

"depression" in the layering 2 m below the surface. The 

solid gray area above the reflector indicates homogeneous deposition, perhaps the result of eolian deposition of fine-grained material in low-lying areas. Eolian 

silt and sand are found at the surface over several areas 

on the site.

Gray-scale images of processed 80-MHz data along 

a west-to-east traverse are shown in figure 3. The location of the traverse is shown in figure 1. The vertical 

exaggeration is 10 in the upper image. The lower image 

shows part of the traverse between 140 and 280 m horizontal distance without vertical exaggeration. The banding caused by near-field interference of the 80-MHz 

radar wave covers about 3 m at the top of each radar 

image, effectively concealing the water-table reflector at 

294.5-m elevation. The lowest strong reflector in each 

image is from the glacial till underlying the outwash 

sands and gravels. Undulations of the till surface are 

apparent. Comparing the upper and lower images also 

shows how vertical exaggeration distorts the GPR 

reflectors. The reflectors in the lower image indicate the 

structural complexity in the depositional sequence on 

this part of the Anoka glacial outwash plain.

CORRELATION OF REFLECTORS AND 

LITHOLOGY LOGS

Several of the GPR traverses were near USGS 

sampling wells. The driller's lithology logs were examined to determine how well they correlated with the 

GPR images. An example for well MC-11 near the start 

of GPR traverse row 16 is shown in figure 4 (see fig. 1 

for locations). A possible correlation of the log to the 

GPR reflectors is indicated by the lines connecting the 

log to the GPR image. "WT" is the water-table reflector. Ground surface at MC-11 is slightly higher than it 

is at the western end of Row 16. In general, the logs do 

not correlate well with the GPR image, in part, because 

the sediments are mixed by the power auger used to 

drill the wells and can only be approximately located. 

Beres and Haeni (1991) note that correlation commonly improves when the core is a continuous sample. 

There can be several other causes for the poor correlation. Variations in porosity, which change the water 

content and, therefore, the electrical properties, may be 

not be uniquely related to variations in lithology. In 

addition, constructive and destructive interference of 

the reflected GPR waves may hide closely spaced layers, create "false" layers, and/or combine several layers 

to appear as only one.

DISCUSSION

There are several scales of investigation in the 

MSEA program: core sample, small, field, and regional 

scales. The resolution and sampling of the GPR system 

are probably too coarse to be useful at the scale of a 

core sample (millimeters to several centimeters), and 

the size of the field site is too small for regional-scale 

investigations (several kilometers). However, GPR 

data can be useful at small-scale (less than I m) and 

field-scale (tens to hundreds of meters) investigations,

585
WEST ROW 5B EAST

300

275

0 200 300 400 500

HORIZONTAL DISTANCE, IN METERS

DATUM IS SEA LEVEL, VERTICAL EXAGGERATION xlO

600 700 780

3001

275

140 160 180 200 220 240 

HORIZONTAL DISTANCE, IN METERS

DATUM IS SEA LEVEL, NO VERTICAL EXAGGERATION

260 280

Figure 3. Gray-scale images of computer-processed 80-MHz ground penetrating radar data along west-to-east traverses, row 5B, at the 

Princeton, Minnesota, Management Systems Evaluation Area. (Location is shown in fig. 1.)
ROW 16

USGS Well MC-11 

Lithology Log Summary

black top soil

brown medium sand

cobbles

298

brown medium to coarse sand

light brown coarse sand

dark brown coarse sand

brown coarse sand with fines

brown medium sand

288

3456789 10

HORIZONTAL DISTANCE, IN METERS 

DATUM IS SEA LEVEL, NO VERTICAL EXAGGERATION

Figure 4. Summary of lithologic log for USGS well MC-11 near the start of ground penetrating radar traverse row 

16 at the Princeton, Minnesota, Management Systems Evaluation Area. (Locations are shown in fig. 1.)

depending on antenna frequency and spacing of 

traverses.

At all scales, hydraulic-flow patterns are observed 

to be three dimensional and controlled by hydrogeologic heterogeneity. Uncertainties caused by the spatial 

variability of hydraulic properties, such as unsaturated 

hydraulic conductivity, affect the ability to model and 

forecast flow and transport of solutes in porous media. 

GPR can provide the additional information required to 

extend hydrologic information obtained at test holes

and wells to much larger areas. In some cases, GPR 

data can directly provide hydrologic information such 

as moisture content (Duke, 1990; Sutinen and others, 

1992) or transmissivity (by repeated measurements 

across time to track the position of moving fluids or 

wetting fronts; Sander and others, 1992).

GPR measurements record signals that are interpreted as reflections from boundaries of hydrogeologic 

units with contrasting electrical properties (relative 

dielectric permittivity and electrical conductivity)

587
controlled by water content, bulk density, lithology, 

and grain-size distribution. Because the signals are 

reflected by contrasts between layers, the processed 

images give good structural and geometric information, whereas information about absolute properties 

must be deduced by modeling individual signals 

(Duke, 1990; Powers and others, 1992). The relation 

between GPR measurements and hydraulic properties 

is complex. However, an empirical relation between 

relative dielectric permittivity and volumetric water 

content for glacial materials can be determined 

(Sutinen and others, 1992). An infiltration test has been 

planned so that the GPR system can be used repetitively over an area to determine the progressive 

movement of the wetting front and thereby infer 

vertical hydraulic conductivity (see Olhoeft, 1985; 

Sander and others, 1992).

The GPR system may prove extremely helpful to 

hydrologists in characterizing a new site. However, in 

an agricultural setting, the GPR system cannot directly 

detect very low concentrations of pesticides or fertilizers in the subsurface. If the site is relatively free of clay 

minerals, GPR data can be useful in (1) determining 

depth and topography of bedrock or other formation 

(such as till); (2) measuring variations in the water 

table (Shih and others, 1986); (3) detecting the effects 

of well draw-down tests; (4) identifying perched water 

tables; (5) determining the lateral variations of lithologies (extent and thickness) near test wells and between 

wells; (6) measuring the type and thickness of soil horizons with the high-frequency antennas (Collins, 1992; 

Doolittle and Asmussen, 1992); and (7) detecting the 

presence of preferential flow paths (Donohue and others, 1992; Sander and others, 1992).

GPR-data collection and processing is relatively 

fast and cost effective in comparison to other methods 

that might collect near-continuous data, such as closespaced drilling or trenching. The GPR system also has 

the advantage of being noninvasive. For each of the 

780-m-long GPR traverses, approximately 15 minutes 

were needed to collect the GPR data using a pickup 

truck and three people. The 80 to 110 topographic 

stations along a traverse took about 1 hour to survey 

using an electronic total station connected to a pocket 

computer. Processing GPR data in a computer is not 

necessary in all applications, as some problems do not 

require any enhancement or geometric correction of the 

data. A 780-m traverse of GPR data required 1 to 2 

hours to process (using software developed by the 

authors) and produce an image (stored on disk as an

Encapsulated PostScript (EPS) file) such as those 

shown in the figures.

CONCLUSIONS

Understanding the hydrogeology of a site is 

integral to determining the potential for ground water 

contamination by surface-applied agricultural chemicals. Analysis of the geometrically corrected GPR 

images shows that reflector continuity, amplitude, configuration, and spatial frequency reveal details of the 

subsurface unavailable by point-source measurements 

provided by drilling. Lithologic structure, depositional 

processes, moisture content, stratification patterns, and 

grain-size distributions can all be interpreted from GPR 

data using antennas of appropriate frequency and 

images displayed at appropriate scale. Ground penetrating radar offers a fast, cost-effective, high-resolution method for extending information acquired at test 

wells and for presenting a comprehensive view of 

subsurface structure. With the addition of computer 

processing, GPR data can also produce information 

about hydrologic properties. GPR does not work at all 

sites, but for areas with very little or no clay near the 

surface, it should be considered as a preliminary study 

tool and as a geophysical technique to provide better 

understanding of geologic and hydrologic information.

REFERENCES

Beres, M., Jr., and Haeni, P.P., 1991, Application of groundpenetrating-radar methods in hydrogeologic studies: 

Ground Water, v. 29, p. 375-386.

Collins, M.E., 1992, Soil Taxonomy A useful guide for 

the application of ground penetrating radar, in Hanninen, P., and Autio, S., eds., Fourth international conference on ground penetrating radar, June 8-13, 1992: 

Rovaniemi, Finland, p. 125-132.

Delin, G.N., and Landon, M.K., 1993, Effects of focused 

recharge on the transport of agricultural chemicals at 

the Princeton, Minnesota Management Systems Evaluation Area, 1991-1992: U.S. Geological Survey OpenFile Report 93-42, 8 p.

Delin, G.N., Landon, M.K., Anderson, J.L.. and Dowdy, 

R.H., 1992, Hydrologic research at the Princeton, 

Minnesota Management Systems Evaluation Area: 

U.S. Geological Survey Open-File Report 92-107, 2 p.

588
Donohue, S.V., Cheng, X-X., and Kung, K-J. Samuel, 1992, 

Improving solute sampling protocol in sandy soils by 

using ground penetrating radar, in Lucius, I.E., Olhoeft, 

G.R., and Duke, S.K., eds., Third international conference on ground penetrating radar, abstracts of the technical meeting, Lakewood, Colorado, USA, May 14-18, 

1992: U.S. Geological Survey Open-File Report 

90-414, p. 20-21.

Doolittle, J.A., and Asmussen, L.E., 1992, Ten years of 

applications of ground penetrating radar by the United 

States Department of Agriculture, in Hanninen, P., and 

Autio, S., eds., Fourth international conference on 

ground penetrating radar, Rovaniemi, Finland, June 8- 

13, 1992: p. 139-147.

Duke, S., 1990, Calibration of ground penetrating radar and 

calculation of attenuation and dielectric permittivity 

versus depth: Colorado School of Mines, Master of science thesis, Department of Geophysics, Golden, Colo.: 

236 p.

Lucius, J.E., and Olhoeft, G.R., 1993, Statistical analysis of 

field-scale subsurface heterogeneity at the Princeton, 

Minnesota, Management Systems Evaluation Area 

using ground penetrating radar, //; Proceedings of the 

agricultural research to protect water quality conference, February 21-24, 1993, Minneapolis, Minn., sponsored by the Soil and Water Conservation Society: 

p. 355-358.

Olhoeft, G.R., 1985, Ground-penetrating radar images of 

the unsaturated zone: EOS Transactions American 

Geophysical Union, v. 66, p. 881.

Powers, M.H., Duke, S.K., Huffman, A.C., III, and Olhoeft, 

G.R., 1992, GPRMODEL One-dimensional full 

waveform forward modeling of ground penetrating 

radar data: U.S. Geological Survey Open-File Report 

92-532, 22 p., and one diskette.

Sander, K.A., Olhoeft, G.R., and Lucius, J.E, 1992, Surface 

and borehole radar monitoring of a DNAPL spill in 3D 

versus frequency, look angle and time, in Bell, R.S., 

ed., Proceedings of the symposium on the application 

of geophysics to engineering and environmental problems. April 26-29, 1992, Oakbrook, 111.: Society of 

Engineering and Mineral Exploration Geophysicists, 

v. 2, p. 455-469.

Sangree, J.B,. and Widmier, J.M., 1979, Interpretation of 

depositional facies from seismic data: Geophysics, 

v.44,p. 131-160.

Shih, S.F., Doolittle, J.A., Myhre, D.L., and Schellcntrager, 

G.W., 1986, Using radar for groundwater investigation: 

Journal of Irrigation and Drainage Engineering, v. 112, 

p. 110-118.

Sutinen, R., Hanninen, P., Cromwell, R., and Hyvonen, E., 

1992, GPR and dielectric classification of glacial material, in Hanninen, P., and Autio, S., eds., 1992, Fourth 

international conference on ground penetrating radar. 

June 8-13, 1992, Rovaniemi, Finland: p. 133-138.

589
590
Geochemistry of Nitrogen in a Farmed Watershed Near

Princeton, Minnesota

By Richard B. Wanty1 , Michele L. Tuttle1 , Matthew K. Landon2, 

Geoffrey N. Delin2 , and J.K. Bohlke3

Abstract

A broad interagency research program is 

being conducted near Princeton, Minnesota, to 

investigate possible advantages of various modifications to existing farming techniques with the 

goal of decreasing contamination of ground-water 

and surface-water supplies by agricultural chemicals. Most of the program is focused on a 

160-acre test farm, with the intention that the 

results from the test farm be applied to other 

farms in the area. To facilitate extending the 

results to a broader region, a study was conducted 

to investigate regional effects of agricultural 

activity on nitrogen in ground waters and surface 

waters of the Battle Brook watershed. Shallow 

ground water has received the greatest inputs 

of nitrate-nitrogen. Concentrations of nitratenitrogen (nitrate-N) exceeding 10 milligrams per 

liter (mg/L) as N are common near the water 

table. Nitrate-N concentrations decreased rapidly 

with depth (deeper than about 8 to 10 meters) in 

the saturated zone throughout the watershed, 

possibly because of mixing with uncontaminated 

waters or because of denitrification. Extensive 

wetlands in the area are fed by the discharge 

of shallow ground water, yet nitrate-N was not 

detected in any wetland waters. Biological activity in the wetland and its sediments is capable of 

decreasing nitrate-N concentrations. Monthly to

'U.S. Geological Survey, M.S. 916, Denver Federal Center, 

Denver, CO 80225

2 U.S. Geological Survey. 2280 Woodale Dr., Mounds View, 

MN 55112

3 U.S. Geological Survey, M.S. 431 National Center, Reston, 

VA 22092

bimonthly sampling of a piezometer near the 

water table at the Princeton test farm during 

July, 1991 to August, 1992, shows that nitrate-N 

concentrations varied between 13 and 25 mg/L in 

response to seasonal recharge events. The magnitude of the variation decreased with depth; about 

2 meters below the water table, nitrate-N concentrations ranged from 12 to 17 mg/L.

INTRODUCTION

The Management Systems Evaluation Area 

(MSEA) program is a multiscale interagency program 

with participation by the U.S. Geological Survey, the 

U.S. Department of Agriculture, the Cooperative State 

Research Service, and the U.S. Environmental Protection Agency. The northern Midwest of the United 

States was selected for this program because this area 

receives approximately 60 percent of fertilizers and 

pesticides used in the United States annually (Burkhart 

and others, 1990). Extensive nitrate-N contamination 

has resulted from the continuous broad application of 

fertilizers in this region (see Power and Schepers, 

1989). One of the five MSEA sites in this region is 

located near Princeton, Minnesota (fig. 1). The goal of 

the Princeton MSEA program is to evaluate the effects 

of best-management practices on ground-water quality 

through a variety of research projects being conducted 

at scales ranging from laboratory to small watersheds 

(about 90 square kilometers). The modified farming 

system being evaluated involves irrigated ridge-tillage 

in a corn and soybean rotation, whereas the prevailing 

farming system in the surrounding Anoka Sand Plain 

involves irrigated conventional tillage in continuous 

corn.

The Princeton MSEA program is primarily 

designed for intensive study at the field scale. The

591
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EXPLANATION OF

SAMPLING SITES 

o Water samples from wells

less than ten mdeep 

o Water samples from wells

greater than ten m deep 

 Surface-water samples 

D Outlets of drain tiles 

O Multilevel piezometers

Princeton

Management

Systems Evaluation

Area

10 KILOMETERS

5 MILES

Figure 1. Location of the Battle Brook watershed near Princeton, Minnesota.

purpose of this study was to investigate the ground and 

surface-water chemistry in the Battle Brook watershed 

in order to link the results of the effects of farming 

systems at the field scale to the watershed scale. This 

report presents preliminary interpretations based on 

data collected during June, 1991-August, 1992, at the 

Princeton MSEA and in the surrounding watershed. 

Battle Brook drains a low-relief, moderately 

wooded area of about 90 km2 (fig. 1). The maximum

elevation in the watershed is about 390 m above mean 

sea level and the mean elevation of Elk Lake, at the 

downstream end of the field area, is 346 m. Because 

the Battle Brook watershed is bounded by low-relief 

land-surface features, the surface-water divide shown 

in figure 1 is not likely to correspond to a groundwater divide. Regional ground-water flow in the 

southern half of the Battle Brook watershed is from 

approximately north to south in the surficial aquifer

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0 5 10 15 20 25 30

NITRATE AS NITROGEN, IN

MILLIGRAMS PER LITER

Figure 2. Nitrate-nitrogen concentrations in ground-water 

and surface-water collected in the Battle Brook watershed 

near Princeton, Minnesota, June-July 1991.

(Lindholm, 1980). Farmland within the watershed is 

planted in com, wheat, and soybeans. Battle Brook is 

bounded by narrow riparian wetlands throughout most 

of its length. Battle Brook flows through a large wetland preserve, the Kunkel State Wildlife Management 

Area (fig. 1), in the middle of the study area. On the 

basis of hydraulic data near the Princeton MSEA, 

shallow ground water discharges to Battle Brook and 

the associated wetlands (Delin and others, 1992).

Possible sources of nitrate-N in ground water 

include leaching of soil organic matter, excess fertilizers, and animal wastes (Komor and Anderson, 1993). 

Nitrate-N concentrations may be decreased by denitrification by anaerobic bacteria or by dissimilatory 

nitrate reduction; both processes use nitrate as an 

electron acceptor in the process of metabolizing 

organic matter (see Greyson, 1990; Komor and Anderson, 1993). The former reaction produces nitrogen 

gas, the latter ammonium.

STUDY METHODS

To assess the chemistry of surface water in the 

Battle Brook watershed, each flowing reach of the 

stream and each tributary was sampled (fig. 1). 

Surface-water samples also were collected from 

reaches of Battle Brook, irrigation ditches, and Elk 

Lake. Shallow (less than 10 m) and deep (greater than 

10 m) wells were sampled to characterize nitrogen 

concentrations in ground water as a function of depth 

and to assess the nature and extent of ground-water 

contamination in the area. The shallow wells are all 

driven sand points, with the exception of the multilevel piezometers at the MSEA. The deeper wells all 

were drilled into aquifers underlying the near-surface 

aquifer. Each multilevel piezometer consists of six 

0.6-cm diameter stainless-steel tubes housed in a 

5.1 -cm diameter PVC casing; each stainless-steel tube 

has a 3-cm-long screened interval (port), which is 

external to the PVC casing. The sampling ports were 

installed at 0.5-m intervals. Sample port 2 was just 

below the water table when the multilevel piezometer 

was installed. Lowering of the water table in late 1991 

has left sample port 3 as the highest sample port in the 

saturated zone. These samples were collected within a 

3-week period in June-July 1991.

Field analyses were performed for nitrate-N, 

nitrite-N, and ammonium-N. Detailed descriptions of 

the field analytical procedures are given in Wanty and 

others (1993). All other analyses were performed in

593
the laboratory on samples collected and preserved in 

the field. Samples analyzed for cations were filtered 

through 0.45 ^im silver-metal membrane filters and 

acidified to a pH less than 1 with concentrated hydrochloric acid. Cations were analyzed by inductively 

coupled plasma atomic emission spectroscopy (Lichte 

and others, 1987). Samples analyzed for anions were 

filtered through the same silver-metal membrane filters and chilled. The silver-metal filters kill any 

microorganisms in the sample. Anions were analyzed 

by ion chromatography with a mixed bicarbonate/carbonate eluent (Fishman and Pyen, 1979). Eluent and 

column conditions were optimized for maximum peak 

separation so individual concentrations of fluoride, 

chloride, nitrite, nitrate, phosphate, and sulfate could 

be resolved.

RESULTS AND DISCUSSION

All surface-water samples, except one, 

contained less than 1 mg/L NO3~-N (as N; fig. 2); most 

surface-water samples contained less than 0.02 mg/L 

NO3 "-N. One sample from an irrigation ditch had a 

NO3 "-N concentration of 1.2 mg/L NO3 " and most 

likely represents water pumped from an irrigation well 

shortly before the sample was collected. The temperature of this sample (15 °C) was significantly lower 

than any other surface-water sample (20 to 31 °C), 

which is consistent with a ground-water source.

Water samples from the shallow wells had 

nitrate-N concentrations ranging up to 25 mg/L. 

Water samples from deeper wells (greater than 10 m) 

generally had nitrate-N concentrations less than 

0.3 mg/L. Thus, only shallow ground water is appreciably affected by agricultural practices. The degree 

to which the shallow ground water is contaminated 

varies considerably in the watershed; the greatest 

nitrate-N concentrations were detected in wells completed within or adjacent to farms (for example, multilevel piezometers in fig. 1), or in densely populated 

areas such as near Elk Lake (fig. 1). The lowest 

nitrate-N concentrations were in water from wells 

adjacent to wooded areas or in low-density residential 

areas away from farmed areas. Dissolved-oxygen 

concentrations were low (less than 1 mg/L) throughout 

the field area (Wanty and others, 1993). Lukashev and 

Onoshko (1989) found qualitatively similar results for 

nitrate-N sources and sinks in a region that includes 

industrial and agricultural activity as well as 

undeveloped land.

Wetlands occupy a large percentage of the 

surface area of the Battle Brook watershed (fig. 1). 

Water in Battle Brook is supplied by short-term 

surface runoff, or by discharge of shallow ground 

water. Numerous springs are found along the edge of 

the wetland area. Although the shallowest ground 

water generally had the greatest nitrate-N concentrations in areas outside the wetlands, none of the 

surface-water samples collected from the Battle Brook 

wetlands had detectable concentrations of nitrate-N. 

Nitrate-N may be taken up by plants in the wetland 

and by bacteria within the wetland sediments. 

Concentrations of nitrate-N in water from springs 

discharging at the edge of the wetland contained 1 to 

2 mg/L nitrate-N, and 0.2 to 0.4 mg/L ammonium-N. 

Wetland pore waters, collected from water that filled 

shallow holes dug in the wetland, did not contain 

nitrate-N but did contain greater than 5 mg/L of 

ammonium -N. A series of experiments was conducted using two columns collected in the wetland; 

each column had a 15-cm diameter and was approximately 1 meter high. The columns contained about 

one-third each of wetland sediment, water and head 

space. The water column in both was spiked with 

nitrate and sulfate. Over the course of about two 

weeks, nitrate concentrations approached zero, while 

sulfate concentrations remained nearly unchanged. 

This result demonstrates the capability of the wetlands 

to take up nitrate-N. Therefore, wetlands may play an 

important role in attenuating anthropogenic nitrate-N 

concentrations. Continuing research is being conducted to determine the rate and extent of nitrate 

removal by wetlands as well as the seasonal variability 

of nitrogen concentrations in waters in and adjacent to 

the wetlands.

Samples were collected from three multilevel 

piezometers at the MSEA during July, 1991 - August, 

1992 to evaluate temporal changes in the groundwater nitrate-N concentrations. Concentrations of 

nitrate-N in one of the multilevel piezometers are 

shown in figure 3. Nitrate-N concentrations in the 

upper 1 m of the saturated zone (ports 2, 3, and 4) 

varied greatly over the year, in contrast to the deepest 

port (port 6). Increases in nitrate-N concentrations in 

the upper meter of the saturated zone occurred during 

the growing seasons in 1991 and 1992. As shown for 

port 4, nitrate-N concentrations in ground water are 

lowest during the winter months. The variations in 

nitrate-N in the upper 1 m of the saturated zone occur 

because recharging water has greater concentrations of

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with the interpretation that fertilizer is the main source 

of nitrate-N in the ground water at the MSEA. Such a 

source is consistent with the transient inputs of nitrateN derived from excess fertilizer by recharging ground 

water throughout the growing season.

CONCLUSIONS

Infiltration of precipitation or irrigation in the 

Princeton, Minnesota study area leaches nitrate-N 

from the soil and carries it to the shallow ground-water 

regime. The greatest concentrations of nitrate-N in 

water from shallow wells were detected during the 

growing season (April through August of 1991 and 

1992) following application of fertilizers. The greatest and most variable nitrate-N concentrations were 

observed in the shallowest wells. Nitrate-N concentrations in water sampled from drain tiles and multilevel piezometers (fig. 2), for example, commonly 

exceed 20 mg/L. Deeper within the saturated zone, 

nitrate-N concentrations are typically less than 

0.3 mg/L.

Surface-water samples from the Battle Brook 

watershed generally contained less than 1 mg/L of 

nitrate-N. If the surface water is derived largely from 

discharge of shallow ground water, then most of the 

nitrate-N must have been removed from the shallow 

ground water by the time it discharges to the surface. 

Nitrate-N could be consumed by biota in the wetland, 

leading to the negligible nitrate-N concentrations 

observed in water in Battle Brook and the adjacent 

wetlands. Comparatively low concentrations of 

nitrate-N in ground-water samples from depths greater 

than 2 m below the water table likely result from 

microbial reduction of nitrate-N or from mixing with 

ground water that has relatively low concentrations of 

nitrate-N. Drain tiles, which remove water from lowlying fields and discharge it directly into the wetland, 

may be carrying significant amounts of dissolved 

nitrate-N into the surface-water regime, but that 

nitrate-N is rapidly and completely attenuated in the 

wetland environment.

REFERENCES

Burkhart, M.R., Onstad, C.A., and Bubenzcr, G.D., 1990, 

Research on agrichemicals in water resources: EOS, 

Transactions of the American Geophysical Union, 

v. 71, no. 29, p. 980-988.

Delin, G.N., Landon, M.K., Anderson, J.L., and Dowdy, 

R.H., 1992, Water Fact Sheet: Hydrologic research at 

the Princeton, Minnesota, Management Systems 

Evaluation Area: U.S. Geological Survey Open-File 

Report 92-107, 2 p.

Delin, G.N., and Landon, M.K., 1993, Effects of focused 

recharge on the transport of agricultural chemicals at 

the Princeton, Minnesota Management Systems 

Evaluation Area, 1991-92: U.S. Geological Survey 

Open-File Report 93-42, 8 p.

Delin, G.N., and Landon, M.K., 1996, Effects of topography 

on the transport of agricultural chemicals near Princeton, Minnesota, 1992, in Morganwalp, D.W., and 

Aronson, D.A., cds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting , Colorado Springs, Colorado, September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Fishman, M.J., and Pycn, G., 1979, Determination of

selected anions in water by ion chromatography: U.S. 

Geological Survey Water-Resources Investigations 

Report 79-101.

Grcyson, J., 1990, Carbon, Nitrogen, and Sulfur Pollutants 

and Their Determination in Air and Water: New York, 

Marcel Dckker, Inc., 338 p.

Komor, S.C., and Anderson, H.W., Jr., 1993, Nitrogen 

isotopes as indicators of nitrate sources in Minnesota 

sand-plain aquifers: Ground Water, v. 31, no. 2, 

p. 260-270.

Lichte, F.E., Golightly, D.W., and Lamothe, P.J., 1987, 

Inductively coupled plasma-atomic emission spcctrometry, in Baedecker, P.A.,, U. S. Geological Survey 

Bulletin 1770: p. B1-B10.

Lindholin, G.F., 1980, Ground-water appraisal of sand 

plains in Benton, Shcrburnc, Stearns, and Wright 

counties, central Minnesota: U.S. Geological Survey 

Water-Resources Investigations Open-File Report 

80-1285, 103 p.

Lukashev, K.I., and Onoshko, M.P., 1989, Human impact 

upon nitrogen geochemistry in landscapes of 

Byelorussia, USSR: Applied Geochemistry, v. 4, 

p. 437-444.

Power, J.F., and Schcpers, J.S., 1989, Nitrate contamination 

of groundwater in North America: Agriculture, 

Ecosystems and Environment, v. 26, p. 165-187.

Wanty, R.B., Tuttle, M.L., Bohlke, J.K., and Briggs, PH., 

1993, Chemical analyses of ground and surface water 

from Battle Brook drainage basin, near Princeton, 

Minnesota: U.S. Geological Survey Open-File Report 

94-129, 26 p.

596
Chemical Loads of Nitrate, Atrazine, and Metolachlor in 

Walnut Creek Watershed near Ames, Iowa, 1991-92

By RJ. Soenksen1 , J.L. Hatfield2 , andD.J. Schmitz2

Abstract

Streamflow, stormflow, and tileflow were 

measured and sampled at numerous sites in the 

Walnut Creek watershed near Ames, Iowa, to 

determine the chemical discharges, loads, and 

yields of nitrate as nitrogen (nitrate-N) and the 

herbicides atrazine and metolachlor. From April 

1991 to September 1992, nitrate-N yield was 58.4 

kg/ha (kilograms per hectare), and atrazine and 

metolachlor yields were 4.5 g/ha (grams per 

hectare) and 7.3 g/ha, respectively, in streamflow 

from a 2,540-ha (hectare) subwatershed. Large 

chemical discharges and increases in cumulative 

loads were closely related to stormflow. Chemical 

loads were larger during the 1991 growing 

season, when soils were wet and stormflow 

occurred shortly after application of chemicals, 

compared to loads during the 1992 growing 

season when stormflow did not occur until several 

months after application of chemicals. Nitrate-N 

was transported during periods of base flow 

throughout the year; there were many periods 

when the herbicides were not detected.

For July-August 1992, a comparison of loads 

and yields in streamflow from the subwatershed 

was made to those in stormflow and tileflow from a 

360-ha small basin within the subwatershed. The 

comparison indicates that most of the chemical 

load in streamflow was transported by tileflow, and 

little was transported by stormflow, even during 

stormflow. There were four periods of stormflow 

during July-August 1992, but there were none in

'U.S. Geological Survey, Iowa City, IA 52244. 

2 Agricultural Research Service, Ames, IA 5001

the preceding 2 months. Nitrate-N yields were 

8.0 kg/ha in streamflow, 0.06 kg/ha in stormflow, 

and 7.9 kg/ha in tileflow. Atrazine yields were 

0.42 g/ha in streamflow, 0.010 g/ha in stormflow, 

and 0.50 g/ha in tileflow. Metolachlor yields were 

0.42 g/ha in streamflow, 0.007 g/ha in stormflow, 

and 0.35 g/ha in tileflow. Comparison of daily 

yields, one during stormflow and one during base 

flow, with those from a field tile also indicate 

tileflow as the primary transporter of agricultural 

chemicals for this period.

INTRODUCTION

The presence of agricultural chemicals in the 

surface- and ground-water resources in many parts of 

the midwestern United States is a growing concern for 

the public and many government officials. As part of 

the Federal Water Quality Initiative (Onstad, Burkart, 

and Bubenzer, 1991), the Management Systems 

Evaluation Area (MSEA) program was established to 

evaluate the effects of current farming practices on 

water quality and to develop new practices that are 

both environmentally and economically sound. Five 

MSEA projects were selected to represent some of the 

principal hydrogeologic settings of the region. The 

Walnut Creek watershed near Ames, Iowa, is one of 

four primary research sites of the Iowa MSEA project. 

The U.S Geological Survey (USGS), through their 

Toxic Substances Hydrology Program, is cooperating 

with the Agricultural Research Service (ARS) and the 

Cooperative State Research Service of the U.S. 

Department of Agriculture in this research. Walnut 

Creek is also the site of the Environmental Protection 

Agency's Midwest Agricultural Surface/Subsurface 

Transport and Effects Research (MASTER) project to 

evaluate the ecological and toxilogical effects of 

agricultural chemicals.

597
Because a number of extensively used 

agricultural chemicals are soluble in water, their 

environmental fate needs to be related to the major 

water-transport processes (Wauchope, 1978; Freeze 

and Cherry, 1979, p. 413). Gaining knowledge of the 

transport processes that affect chemicals and the 

conditions that control transport is an important step 

toward understanding the effects of existing 

management practices. This knowledge will aid in the 

selection of existing and the design of new 

management practices to minimize chemical transport 

from the point of application.

This paper documents the nitrate as nitrogen 

(nitrate-N), atrazine, and metolachlor transported by 

streamflow from a subwatershed of Walnut Creek 

from April 1991 to September 1992. A comparative 

analysis of the amounts transported during JulyAugust 1992 to the amounts transported by stormflow 

and tileflow from two smaller basins within the 

subwatershed also is made. Chemical discharges, 

loads, and yields (load per unit area) are presented 

along with flow hydrographs and rainfall amounts.

WALNUT CREEK WATERSHED

Walnut Creek watershed is located in and 

representative of much of Midwest Land Resource 

Region 103 (U.S. Department of Agriculture-Soil 

Conservation Service, 1981), which covers much of 

north-central Iowa and southern Minnesota. The 

topography is nearly level with numerous potholes in 

the upper one-third of the watershed, nearly level or 

gently rolling in most of the other uplands, and steeper 

near the streams in the lower part of the watershed. 

Surficial deposits of till overlie carbonate bedrock. Till 

thickness is 60 to 90 m on the uplands, and total relief 

is about 60 m. Soils are in the Clarion-NicolletWebster association. The Clarion soils are well 

drained and moderately permeable on the uplands, 

whereas the adjacent Nicollet soils are poorly drained 

with moderate permeability. Natural drainage is poor 

in the upper part of the watershed, requiring the use of 

subsurface drainage tiles many with surface inlets 

from potholes and drainage ditches. Surface drainage 

areas are difficult to determine because of the nearlylevel topography and subtle drainage features. In some 

cases, tile lines transport subsurface flow to an 

adjacent surface drainage basin. Large parts of some 

drainage areas, especially in the upper part of the 

watershed, may not contribute to stormflow but may

contribute to tileflow via surface inlets and 

subsequently to streamflow.

Above where Walnut Creek flows across the 

South Skunk River alluvial plain, the watershed has a 

drainage area of about 5,100 ha that was divided into 

subwatersheds of about 1,300 to 2,500 ha (fig 1). The 

northwestern and southwestern subwatersheds are 

almost completely in row-crop agriculture and have 

large percentages of tile-drained land. The eastern 

subwatershed is steeper, with woodlands near some of 

the streams and a smaller percentage of tile-drained 

land. Study interests also included small basins of 

about 250 to 500 ha and field basins of about 5 to 

25 ha. The locations of all USGS streamflow and 

stormflow sites and two ARS tileflow sites are shown 

in figure 1; sites also are listed in table 1 along with the 

study scale, surface drainage area, and types of flow 

evaluated. Of interest for this paper are sites 310, 220 

and 110. The tileflow from site 220 contains stormflow 

from surface inlets to the tile, whereas site 110 has no 

surface inlet from the pothole field it drains. Rainfall 

data was collected at site 710. Study sites became fully 

operational as follows: site 310, streamflow, April 

1991; site 220, stormflow, August 1991; site 220, 

tileflow, July 1991; site 110, tileflow, December 1991; 

site 710, rainfall, May 1991.

STUDY APPROACH

Discharge at streamflow and stormflow study 

sites was computed from recorded water-stage 

measurements and theoretical or measured stagedischarge relations (Rantz and others, 1982). 

Automatic stage measurements were made by float or 

balanced-beam manometer. Discharge at tileflow 

study sites was computed in real time by "flow 

meters" using recorded stage and a theoretical weir 

rating, or recorded stage and velocity measurements. 

Stage was measured with a pressure transducer, and 

velocity was measured with an electromagnetic sensor. 

Rainfall data were collected with a tipping-bucket rain 

gage. Data were recorded on electronic data collectors.

Water-quality samples were collected manually, 

on a weekly basis, and by automatic samplers. At 

streamflow and stormflow sites the samplers were 

triggered by the data loggers using real-time stage data 

and programmed decision aids. Tileflow samplers 

were triggered by timers in the samplers. Samples 

were analyzed for concentrations of nitrate-N, 

atrazine, metolachlor, alachlor, and metribuzin at the

598
93*42'30> 

42"

r" 

* ' n Northwestern 

subwatcrshcd

Iowa

Walnut Creek 

watershed

Boundary of 

Walnut Creek 

watershed Southwestern 

subwatershed 3 MILES

Base from U.S. Geological Survey dglal data. 124.000.1075 

Universal Transvene Mercator projection 

Zone 15

3 KILOMETERS

EXPLANATION

TOWN

STREAM DRAINAGE

STUDY SITES

320A Streamliow site and she number

 FIELD BOUNDARY

220

T Stormllow she and site number

 SMALL BASIN BOUNDARY 110A Tilellow she and site number

710.

 SUBWATERSHED 

BOUNDARY

Figure 1. Location of study sites in Walnut Creek watershed.

ARS's National Soil Tilth Laboratory in Ames, Iowa 

using the methods described by Pfeiffer and 

Steinheimer (1992). Nitrate-N concentrations were 

determined using a colormetric method. The 

quantification limit was 1 mg/L. Pesticide analytes 

were extracted by use of a Cjg solid-phase extraction 

technique and analyzed with a gas chromatograph/ 

mass spectrometer in a selective ion monitoring mode. 

For these analytes the detection limit in water was 

0.2

Rain gage and site number

Chemical discharges were computed from 

instantaneous values of concentration and the 

corresponding values of discharge. Discharge was 

computed for 5- and 15-minute intervals, depending 

on the site. Instantaneous concentration values were

determined from time-concentration curves developed 

from the discrete sample concentrations. Generally, 

this involved linear interpolation between sample 

points, except just prior to stormflow periods where 

concentrations were kept at base-flow levels. 

Herbicide samples with concentrations below the 

detection limit were assigned a value of 0.1 ng/L, onehalf the detection limit. Outliers were eliminated on 

the basis of data trends and comparison with the flow 

record. Chemical loads were computed by multiplying 

the chemical discharge by the appropriate time interval 

and appropriate unit conversion, and then summing 

the unit values of load as necessary.

Streamflow and stormflow sites were installed 

by USGS and ARS personnel, and tileflow and rainfall

599
Table 1. Study sites in Walnut Creek watershed, iowa

Site 

number

(«g. D

110

140

220

310

320

330

Site name

Pothole field

Walnut Creek tributary near Kelley

South Branch Ditch of Walnut 

Creek at Kelley

Walnut Creek at Kelley

Walnut Creek near Kelley

Walnut Creek near Cambridge

Study scale

Field basin

Field basin

Small basin

Subwatershed

Subwatershed

Watershed

Surface 

. drainage 

area 

(hectares)

8.9

13

360

2,540

3,760

5,100

Type of flow 

evaluated

Tileflow

Streamflow

Stormflow 

Tileflow1

Streamflow

Streamflow

Streamflow

Includes stormflow from surface inlets.

sites were installed by ARS personnel. Routine site 

operation and data collection were done by the ARS, 

with quality control provided by the USGS at 

Streamflow and stormflow sites. Streamflow, 

stormflow, and chemical discharge and load data were 

computed by the USGS. Tileflow and rainfall data 

were computed by the ARS.

RESULTS

The automatic sampling systems were effective 

in collecting a sufficient number of samples for 

analyses and subsequent development of timeconcentration curves. More than 430 samples were 

collected at site 310 from April 1991 to September 

1992, and many of those were collected during 

stormflow when concentrations were usually changing 

rapidly. Concentrations of nitrate-N in Streamflow 

generally ranged from 10 to 20 mg/L during base flow, 

but concentrations decreased substantially, by as much 

as about 75 percent, during stormflow. Concentrations 

of atrazine and metolachlor generally were less than 

about 3.0 |Ag/L or less than the detection limit of 

0.2 ng/L during base flow, but were as large 29 and 

80 ng/L during stormflow in May 1991. The largest 

concentrations generally occurred on the rising limbs 

of the flow hydrographs or near the peaks. On the 

receding limbs of the hydrographs, concentrations 

decreased rapidly at first and then more slowly to very 

small levels.

Subwatershed Streamflow - Site 310

Sample analyses from April 1991 to September 

1992 were used to determine daily loads of nitrate-N, 

atrazine, and metolachlor in a 2,540-ha Subwatershed 

of Walnut Creek (fig. 2). A cumulative-load curve also 

is shown for each chemical along with the Streamflow 

hydrograph. The cumulative loads and yields (not 

shown) for the entire period were 150,000 kg and 

58 kg/ha for nitrate-N, 11,000 g and 4.5 g/ha for 

atrazine, and 19,000 g and 7.3 g/ha for metolachlor. 

For July-August 1992, the cumulative loads and yields 

were 20,000 kg and 8.0 kg/ha for nitrate-N, and 1,100 g 

and 0.42 g/ha for both herbicides. All chemical loads 

increased during stormflow; the herbicides almost 

exclusively so, but cumulative nitrate-N loads also 

increased during periods of base flow, even during the 

winter months. Most of the atrazine and metolachlor 

loads were transported during four different periods of 

stormflow from May 15 to June 4 of 1991 shortly after 

these chemicals were applied. Soils were very wet 

prior to this period. In 1992, by contrast, stormflow 

was less and it occurred later in the growing season 

resulting in decreased transport of herbicides.

Small-Basin Stormflow and Tileflow - Site 220

Data for July-August 1992 were examined for 

a 360-ha small basin (site 220) to help determine 

the sources of chemical loads at site 310. Hourly 

values of chemical discharge and stormflow, and 

chemical discharge and tileflow are shown in figures 3 

and 4. Daily precipitation also is shown. Cumulative 

nitrate-N loads and yields (not shown) were 22 kg

600
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to
and 0.06 kg/ha in stormflow (fig. 3), and 2,800 kg and 

7.9 kg/ha in tileflow (fig. 4). Cumulative loads and 

yields of atrazine were 3.5 g and 0.010 g/ha in 

stormflow, and 180 g and 0.50 g/ha in tileflow. 

Cumulative loads and yields of metolachlor were 2.6 g 

and 0.007 g/ha in stormflow, and 130 g and 0.35 g/ha 

in tileflow. For each chemical the yield from tileflow at 

site 220 was similar to that from streamflow at site 

310. The extremely small chemical yield in stormflow 

by comparison is expected because stormflow was so 

much less than tileflow. However, the chemical 

discharge rate in the stormflow was also much less 

than that in the tileflow. Together, these results would 

seem to indicate that, for the prevailing conditions, 

stormflow was not a major transport mechanism for 

any of these chemicals and that most of the chemical 

loads were transported by subsurface flow to the tile 

lines after leaching into the soil. This is reasonable 

because there were more than 75 mm of rainfall in 

July before any stormflow occurred, as shown in 

figures 3 and 4. If the stormflow at site 220 is 

representative of the flow into the surface inlets, then 

the effect of the inlets on the tileflow at site 220 was 

probably small. It is not known how significant 

stormflow was in transporting the much larger loads of 

atrazine and metolachlor in 1991 when stormflow 

occurred closer to the time following chemical 

application in the subwatershed.

Field-Basin Tileflow - Site 110

To investigate further the possible effects of 

surface inlets on the tileflow at site 220, some 

comparisons were made with the tileflow for a 8.9-ha 

field basin (site 110), which has no surface inlets. 

Yields of nitrate-N and atrazine are shown for each

of the sites for two days during July-August 1992 

(table 2). Data for July 16 is for a period of stormflow, 

the first in 2 months, and data for August 5 is for a 

period of base flow. Except for stormflow site 220, 

yields for all sites for both chemicals are very similar. 

On the basis of this comparison, it would seem that 

during July-August 1992 subsurface flow was the 

primary transport mechanism for each of the 

chemicals studied at site 220, even during periods of 

stormflow.

SUMMARY

Streamflow, stormflow, and tileflow were 

measured and sampled at several sites in the Walnut 

Creek watershed near Ames, Iowa, to determine the 

chemical discharges, loads, and yields of nitrate-N, 

atrazine, and metolachlor. The total load and yield of 

nitrate-N in streamflow from a 2,540-ha subwatershed 

(site 310) of Walnut Creek from April 1991 to 

September 1992 was 150,000 kg and 58 kg/ha. Loads 

and yields of selected herbicides was 11,000 g and 

4.5 g/ha for atrazine, and 19,000 g and 7.3 g/ha for 

metolachlor. Large chemical loads and increases in 

loads are closely related to periods of stormflow. 

Cumulative chemical loads were greater during the 

1991 growing season, when soil conditions were wet 

and stormflow occurred shortly after application of 

chemicals, compared to the cumulative loads of 

chemicals during the 1992 growing season when 

stormflow did not occur until 2 months after 

application of chemicals. This was especially true for 

the herbicides. Nitrate-N is transported in significant 

quantities during periods of base flow at all times of 

the year.

Table 2. Daily nitrate-as-nitrogen and atrazine yields at selected study sites in the Walnut Creek 

watershed, Iowa, July 16 and August 15, 1992

Site number (fig. 1) and

Nitrate-as-nitrogen yield 

(kilograms per hectare)

Atrazine yield 

(grams per hectare)

310-Streamflow

220-Stormflow

220-Tileflow

HO-Tileflow

July 16

0.51

0.003

0.49

0.51

August 5

0.11

0

0.11

0.09

July 16

0.066

0.0031

0.050

0.087

August 5

0.002

0

0.003

0.003

'July 15.

604
For July-August 1992, loads and yields in 

streamflow from the same subwatershed were 

20,000 kg and 8.0 kg/ha for nitrate-N and 1,100 g and 

0.42 g/ha for both atrazine and metolachlor. Loads 

and yields in stormflow from a 360-ha small basin 

(site 220) within the subwatershed were only 22 kg 

and 0.06 kg/ha for nitrate-N, 3.5 g and 0.010 g/ha for 

atrazine, and 2.6 g and 0.007 g/ha for metolachlor. 

Concurrent loads and yields in tileflow for the same 

small basin were 2,800 kg and 7.9 kg/ha for nitrate-N, 

180 g and 0.50 g/ha for atrazine, and 130 g and 0.35 g/ha 

for metolachlor. On the basis of yields for this period 

of flow following 2 dry months, little of the chemical 

loads from the subwatershed appeared to have come 

from stormflow, whereas most of the loads appeared to 

have'come from tileflow.

The tile lines draining this small basin do have a 

number of surface inlets, but the chemical loads from 

these inlets should be similar to those measured in the 

stormflow at the small basin outlet and should, 

therefore, have had little effect on the chemical loads 

and yields from the tile lines. Daily yields for July 16 

(stormflow) and August 5 (base flow) compared to 

those from a 8.9-ha field-basin tile (site 110) with no 

surface inlets also indicates that tileflow is the primary 

transporter of agricultural chemicals for this period.

Nitrate-N and atrazine yields were nearly identical for 

streamflow, tileflow with surface inlets, and tileflow 

without surface inlets, whereas yields for stormflow 

were much smaller.

REFERENCES CITED

Freeze, R. A., and Cherry, J.A., 1979, Goundwater: 

Englewood Cliffs, N.J., Prentice-Hall, Inc., 604 p.

Onstad, C.A., Burkart, M.R., and Bubenzer, G.D., 1991, 

Agricultural research to improve water quality: 

Journal of Soil and Water Conservation, v. 46, p. 184- 

188.

Pfeiffer, R.L., and Steinheimer, T.R., 1992, Quality

assurance - quality control manual of the National Soil 

Tilth Laboratory: NSTL Document 92-1, 41 p.

Rantz, S.E., and others, 1982, Measurement and

computation of streamflow: U.S. Geological Survey 

Water-Supply Paper 2175, 631 p.

U.S. Department of Agriculture-Soil Conservation Service, 

1981, Land resource regions and major land resource 

areas of the United States: Washington D.C., 

Agricultural Handbook 296, 156 p.

Wauchope, R.D., 1978, The pesticide content of surface 

water draining from agricultural fields A review: 

Journal of Environmental Quality, v. 7, no. 4, p. 459- 

472.

605
606
Atrazine Transport Through Preferential Pathways

at the Missouri MSEA

By James A. Tindall 1 and William K. Vencill2

Abstract

Results of experiments at the Missouri 

Management System Evaluation Area (MSEA) 

indicate that atrazine moved rapidly through the 

soil, probably due to the presence of a large 

number of cracks and macropores. Data also 

show that concentrations of atrazine exceeding 

0.50 ^ig ml/ 1 were observed with depth (45- 

135 cm) after several months following heavy 

recharge events. It appears that macropores were 

a major factor in atrazine transport at the site. 

Agriculture practices designed to reduce surface 

cracking (which will reduce access to 

macropores) and chemical input could play a 

major role in lessening contamination of ground 

water in the Midwest.

INTRODUCTION

Research and monitoring studies have shown 

widespread detection of various herbicides in shallow 

ground water in the midwestern United States as well 

as other locations (Hallberg, 1989). Both atrazine and 

alachlor, common postemergent herbicides used on 

corn, soybeans, and other field crops, have been 

detected in numerous states (U.S. Environmental 

Protection Agency, 1987). Residues of these two 

herbicides are commonly found in the Midwest and in 

California. Atrazine and alachlor concentrations as 

high as 10 fig L' 1 and 16 jig L" 1 respectively, have 

been reported in ground water in Iowa (Wintersteen 

and Hertzler, 1987). The processes of sorption, 

volatilization, degradation, persistence, mobility, and

'U.S. Geological Survey, NRP, Denver, Colo..

2 Department of Agronomy, University of Georgia, Athens, Ga.

plant uptake, as well as the presence of preferential 

flowpaths interact over time to determine the fate and 

transport of herbicides. Environmental factors, such as 

evapotranspiration, precipitation, and temperature, 

also affect herbicide transport (Jury, 1986). There is 

convincing evidence that the presence of preferential 

pathways promotes rapid, deep movement of 

chemicals within agricultural soils (White, 1985; 

Edwards and others, 1990; Luxmoore and others, 

1990).

Because non-point source (NPS) pollution has 

become a major water-quality concern in the United 

States, especially in the midcontinent area, five 

Management System Evaluation Areas (MSEA) have 

been established. These MSEA sites are located in 

Minnesota, Missouri, Iowa, Ohio, and Nebraska and 

have become part of the well known "midcontinent 

herbicide initiative." Research at these sites is being 

conducted through a collaborative effort with various 

universities, the U.S. Department of AgricultureAgricultural Research Service, U.S. Environmental 

Protection Agency (USEPA), and U.S. Geological 

Survey. Because contamination of ground water by 

the major herbicides atrazine, alachlor, cyanazine, 

metolachlor, and metribuzin, as well as nitrate 

contamination from fertilizers have been found in 

these States, research at each MSEA site is focusing 

on the transport and transformation processes of these 

chemicals. Although it is well known that 

contamination by these compounds enters surface 

water and ground water through various pathways, the 

mechanisms by which they are transported are poorly 

understood. The research for this report was begun to 

determine how atrazine moves through the soil matrix, 

and the role of macropores in that movement through 

the clay soils at the Missouri MSEA.

607
MATERIALS AND EXPERIMENTAL 

METHODS

Six undisturbed experimental plots with a 1 

m2 surface area (two sets of three each) were 

instrumented in a fine montmorillonitic, mesic 

Udollic Ochraqualf Mollic albaqualf, silty loam soil 

at the Missouri MSEA near Centralia in Boone 

County, Missouri, on May 11, 1991 (fig. 1). Three 

plots (plots 1 to 3) were 1.35 m deep, and the 

remaining three (plots 4 to 6) were 1.05 m deep. 

Zero-tension pan lysimeters were installed at 1.35 

m in plots 1 to 3 and at 1.05 m in plots 4 to 6. 

Additionally, each plot was planted with soybeans 

(Glycine max L.) and instrumented with suction 

lysimeters and tensiometers at 15-cm increments. 

The tensiometers were automated with transducers 

attached to a data logger. Neutron-access tubes

were installed in each plot and were read at 15-cm 

increments to determine moisture content. The

^

zero-tension pan lysimeters (1 m ) was installed at 

the bottom of each plot without disturbing the soil 

above the pans. To accomplish this, the plots were 

excavated on the eastern and western sides and 

lined with 4-mil plastic to prevent lateral ground 

water flow through the plots. Plots were encased 

with a raised frame (8-cm high) to prevent any 

surface runoff (fig. 2). After installation, the plots 

were allowed to equilibrate for 1 year before data 

collection. During this waiting period, all suction 

and pan lysimeters were purged monthly and were 

sampled prior to herbicide application in May 1992. 

Initial atrazine concentrations were below 

detectable limits (0.05 ng/mL).

94'6

40' 30' -

CENTRALIA:

MISSOURI 

MANAGEMENT

SYSTEM

EVALUATION

AREA

COLUMBIA

MISSOURI

36' 48' 

225

Figure 1 . Site-location map.

mile 

inch

608
Neutron Access Tubes 

Sampling Tub" ^

Lysimeters (1 for each Tensiometer)

^Transducer 

Connections

^Tensiometers

Lysimeter (1m2 )

Sampling. 

Bucket

Figure 2. Diagram of plot instrumentation (not to scale).

The herbicide atrazine was applied on May 18, 

1 992 at a rate of 1.4 kg ha" 1 . Six hours after 

application, an infiltrometer was used to apply 5 cm of 

water to each plot (1.5 cm hr" 1 ). Water samples were 

extracted from suction and pan lysimeters monthly 

(June through November) and analyzed for atrazine on 

a high-precision liquid chromatograph (hplc) with a 

25-mm by 4.6- mm column and an 80:20 volume-pervolume methanol-to-water ratio (adjusted to pH 7.4 

with 0.05 M ammonium acetate). Wavelength was set 

at 222 nm with a retention time of 9.5 minutes and a 

flow rate of 8 mL minute" 1 . All samples were 

maintained at 0°C until analysis and filtered through a 

0.45-m syringe filter prior to analysis. Atrazine 

concentrations were determined from standards and a 

linear regression curve (r2 = 0.991). Standards as well 

as blank samples were inserted at frequent (blind) 

intervals to insure quality control. Soil samples also 

were extracted and analyzed by hplc. A 20-g sample 

of soil was cored from each depth increment, and 

atrazine was extracted by adding 42 mL of 90 percent 

methanol and 10 percent hplc-grade water, shaking for 

2 hours, and centrifuging. The supernatant was 

collected, filtered, and analyzed.

EXPERIMENTAL RESULTS AND 

DISCUSSION

The zero-tension pan lysimeters were designed to 

intercept all water flowing vertically through each plot 

so that total flux (volume basis) could be determined. 

Concentrations of atrazine at depths of 60 to 135 cm 

(plot 1; fig. 3A) and 90 to 105 cm (plot 6; fig. 3B) 1 

month after application are greater than would be 

expected if flow and chemical transport were through 

the soil matrix alone, because atrazine would normally 

be diffused within the soil matrix, allowing for little if 

any transport below 15- to 30-cm depths. Plots I and 6 

(figs. 3A and 3B) were selected for illustration because 

they are representative of each of the six plots. In plots 

4-6, the pan lysimeter was just below a claypan that 

exhibited severe cracking with slight decreases in water 

content (claypan cracks were observed on separate, 

adjacent plots). Cracking that was evident on the 

surface was likely caused by the development of 

cleavage planes at points of high-water content. As the 

soil dried, the cleavage planes separated, causing large 

cracks. Several surface cracks were 2 cm wide by about 

20 m long; these cracks did not follow cultivation 

planes. Cracking of the claypan was probably the cause 

of the rapid transport of atrazine through the soil profile.

609
Sample Date

 - June 16 

A-A July 18

 - August 16

 -* September 21 

T-T October 19 

o-o November 22

Sample Date

 - June 16 

A-A July 18

 - August 16

 « September 21 

Y-» October 19 

c-e November 22

B

0.2 0.4 0.6 0.8 1 

Atrazine Concentration, 

In Micrograms Per Milliter

Figure 3. (A) Atrazine concentrations in ground water from 

suction and pan lysimeters (plot 1).; and (B) Atrazine 

concentrations in ground water from suction and pan 

lysimeters (plot 6).

A companion laboratory study on undisturbed, 

unsaturated soil cores (28 cm diameter by 40 cm 

height) that were extracted by the method of Tindall 

and others (1992) and analyzed by the method of 

Green and others (1990), snowed that the claypan, 

which is 40 to 80 cm below land surface, had 

anomalously high saturated hydraulic conductivity 

(Ks) values greater than 30 cm d" 1 . After staining the 

cores in the laboratory study with crimson-red dye, 

large numbers of macropores were found in the 

claypan cores compared to cores collected from above 

and below the claypan, where Ks averaged about 

10 cm d . Because of high clay content, Ks values for 

claypan sediments are generally in the range of 3 to 4 

orders of magnitude less than 30 cm d" 1 . Thus, the 

high Ks values are likely a result of the macroporosity 

of the claypan cleavage planes. Unsaturated hydraulic 

conductivities of the laboratory cores were in the same 

range as those which had been measured in the field 

plots (10"' to 10"3 cm d" 1 ) at the same relative water 

content (0 « 0.23). Therefore, rapid transport of 

atrazine probably occurred under saturated conditions 

during times of large recharge events and, as the 

atrazine moved from the soil surface through the top 

15-cm, it was able to move quickly through the 

claypan to the water table at a depth of about 1.5 m 

below land surface. The high Ks values for the 

claypan probably account for the high peak 

concentrations of atrazine collected in the zero-tension 

pan lysimeters.

Generally, atrazine concentrations decreased 

with depth; however, large concentration peaks were 

observed at various depths and times within all plots. 

These peaks confirm that atrazine was being 

transported to a depth below 15 cm relatively 

quickly. Although the concentrations were lower with 

depth than those at the time of application, however, 

considerable amounts of atrazine were still being 

transported within the soil profile, as evidenced by the 

high peaks in August 16 samples (figs. 3A and 3B).

During the 6 month sampling period, 52-cm of 

rainfall was recorded. At the end of July 1992, the 

plots received about 15 cm rainfall just prior to the 

August 16 sampling. The atrazine concentrations in 

the 60- to 120-cm depth interval (fig. 3A) and 90- to 

105-cm depth interval at plot 6 (fig. 3B) in August 

were much higher than atrazine concentrations near 

land surface. Prior to the November 22 sampling, 

about 7 cm of rainfall was recorded and similar, but 

smaller increases in atrazine concentration with depth 

occurred at the 90- and 135-cm depth interval at plot 1 

(fig. 3 A) and at the 90-cm and 105-cm depth interval 

at plot 6 (fig. 3B). High atrazine concentrations 

observed as late as August and November may reflect 

entrapment and partial desorption of atrazine from soil 

within the macropore system where it remained in 

immobile pockets until it was later flushed by 

recharging water.

Compared to concentrations in the water 

samples (fig. 4) extracted from the suction and pan 

lysimeters, the concentrations in the soil samples are 

greater near the surface but decrease with depth, as 

would be expected for bulk soil. At the surface (fig. 

4), atrazine concentrations collected in cores in June 

are similar to applied concentrations, but soil samples 

collected in November indicate that large rainfall 

events caused fluctuations in atrazine concentrations 

in the soil matrix similar to fluctuations in the 

lysimeter extracts. The same trends are quite 

noticeable in soil and extract samples.

The difference in atrazine concentration in soil 

and extract samples at the 15-cm and other depths 

results from the fact that the suction lysimeters collect 

water that moves relatively quickly through the larger 

pores. As water and chemicals flow into the soil 

matrix, flow is retarded by the small size of the pores, 

which, because of capillary tension, restrict the flow of 

water to large conducting pores; because of these 

zones of restricted flow, some of the atrazine may be 

entrapped while additional amounts are adsorbed onto 

soil particles. Consequently, atrazine concentration in

610
Sample Date

 - June 16 

A-A July 18

 - August 16

 - September 21 

T-T October 19 

e e November 22

200 400 600

Atrazine Concentration, 

In Micrograms Per Kilogram

Figure 4. Atrazine concentrations in soil-core samples 

(plotl).

the matrix is higher than can be sampled by the 

lysimeters, and each recharge event probably flushed 

the atrazine into the larger pores where it became 

much more mobile and was sampled later by suction 

lysimeters. With continual recharge and subsequent 

flushing during the growing season, atrazine 

concentrations in water from the suction lysimeters 

will gradually decrease to less than detectable limits, 

while concentrations will remain high within the soil 

matrix; accordingly, sail samples would represent the 

amount of atrazine stored in the soil. The results 

indicate that a portion of this soil stored atrazine may 

become available for transport and be flushed during 

large rainfall events.

SUMMARY AND CONCLUSIONS

Macropores are a major factor in transport of 

atrazine at the Missouri field site, where most of the 

macropores present are likely the result of cleavage 

plane separation. It appears that atrazine is not 

irreversibly adsorbed to the soils at this site and, 

during early stages of the growing season (within 1 to 

2 months of application), it may subsequently leach 

after periods of heavy rainfall. Because large cracks 

and pores within the soil profile are at atmospheric 

potential, saturated or near saturated conditions are 

required for water to infiltrate the land surface in these 

areas. Consequently, a management system such as

minimum tillage and normal tillage operations (this 

would not include no-till), that reduces surface 

cracking, thereby preventing chemical inputs from 

reaching macropores, could play a major role in 

lessening ground water contamination in this area. 

Also, the timing and frequency of application of the 

fertilizer or pesticide would be incorporated into a 

management strategy. Accordingly, management of 

current pesticide application practices may play a 

more important role than previously believed and, a 

mass-balance study of atrazine using 14C labeling 

would help quantify total amounts that may be 

reaching ground water.

REFERENCES

Edwards, W.M., Shipitalo, J.J., Owens, L.B., and Norton, 

L.D., 1990, Effect of Lumbricus terrestris L. burrows 

on hydrology of continuous no-till corn fields: 

Gcodcrma v. 46 p. 73-84.

Green, R.E., Ahuja, L.R., and Chong, S.K., 1990, Hydraulic 

conductivity, diffusivity and sorptivity of unsaturated 

soils-Field Methods:(Arnold Klute Ed.) Methods of 

soil analysis: Part 1-Physical and mineralogical 

methods, 2nd Edition, Monograph 9, Madison, Wise., 

American Society of Agronomy, p. 771-798.

Hallberg, G.R., 1989, Agricultural Ecosystems and the 

Environment, v. 26 p. 299-367.

Jury, W.A., 1986, Chemical movement through soil: Vadose 

Zone Modeling of Organic Pollutants. S.C. Hern and 

S.M. Melancon, eds., Chelsea, Michigan, Lewis 

Publishers, Inc.,p. 135-158.

Luxmoore, R.J., Jardine, P.M., Wilson, G.V., Jones, J.R., 

and Zelazny, L.W., 1990, Physical and chemical 

controls of preferred path flow through a forested 

hillslope. Geoderma v. 46, p. 139-154.

Tindall, J.A., Hemmen, K., and Dowd, J.F., 1992, An 

improved method for field extraction and laboratory 

analysis of large, intact soil cores: Journal of 

Environmental Quality, v. 21, p. 259-263.

U.S. Environmental Protection Agency, 1987, Agricultural 

chemicals in ground water Proposed pesticide 

strategy: Washington, D.C., Office of Pesticides and 

Toxic Substances, 150 p.

White, R.E., 1985, The influence of macropores on the 

transport of dissolved and suspended matter through 

soil. In Advances in Soil Science. B.A. Stewart, ed., 

New York, Springer-Verlag, v. 3, p. 95-120.

Wintersteen, W., and R. Hertzler. 1987. Pesticides used in 

Iowa, Crop production in 1985: Ames, Iowa., Iowa 

Cooperative Extension Service, p. 1-18.

611
612
U.S. Geological Survey 

Toxic Substances Hydrology ProgramProceedings of the Technical Meeting, 

Colorado Springs, Colorado, 

September 20-24, 1993

By David W. Morganwalp and David A. Aronson

U.S. GEOLOGICAL SURVEY

WATER-RESOURCES INVESTIGATIONS REPORT 94-4015

Volume 2 of 2

Tallahassec, Florida 

1996
U.S. DEPARTMENT OF THE INTERIOR 

BRUCE BABBITT, Secretary

U.S. GEOLOGICAL SURVEY 

Gordon P. Eaton, Director

For additional information write to:

Assistant Coordinator

Toxic Substances Hydrology Program

U.S. Geological Survey

412 National Center

Reston, Virginia 22092

Copies of this report may be 

purchased from:

U.S. Geological Survey

Earth Science Information Center

Open-File Reports Section

Federal Center, MS 517

Box 25286

Denver, CO 80225
PREFACE

The Toxic Substances Hydrology (Toxics) Program of the U.S. Geological Survey (USGS) provides 

earth-science information needed to prevent or mitigate contamination of the Nation's ground-water and 

surface-water resources. The Toxics Program's objective is to understand the movement and fate of toxic 

substances in the Nation's ground water and surface 

water. The program's research activities include 

( i) neict and laboratory investigation of the physical, 

chemical, and biological processes that control the 

transport and fate of contaminants in ground water and 

surface water; (2) the development of new analytical 

methods and sampling techniques to measure the concentrations of toxic substances in water; and (3) the 

investigation of the relations among human activities, 

environmental characteristics, and regional patterns of 

water quality. Most of the research activities take 

place at contaminated sites. The major research 

objective is not characterization of contaminated sites, 

but is obtaining knowledge about contaminant-transport processes transferable to other contaminated sites. 

For example, the knowledge obtained at the research 

sites could be used to predict contaminant transport at 

other sites, improve methods for remediating contaminated ground water, design water-quality assessments, 

and identify potential causes of contamination.

Recent accomplishments of the Toxics Program 

have been presented at a series of technical meetings 

held in Tucson, Arizona (1984), Cape Cod, Massachusetts (1985), Denver, Colorado (1987), Pensacola, 

Florida (1987), Phoenix, Arizona (1988), and 

Monterey, California (1991). This report documents 

the oral and poster presentations given at the most 

recent meeting, held in Colorado Springs, Colorado, 

in September 1993. The sections in this report correspond to the topical organization of the meeting's oral 

sessions. Each session focused on research at a specific contaminated site or on a topical subject. Papers 

associated with poster sessions were integrated into 

the appropriate section. The lead paper in most of the 

sections gives an overview of the research conducted 

at the site discussed.

Section A addresses the biodegradation of gasoline in ground water that resulted from a leaking 

underground storage tank in Galloway Township, 

New Jersey. The papers describe methods to determine in situ biodegradation rates of gasoline on the 

water table, the geochemistry of gasoline constituents

dissolved in ground water, and air-flow modeling of 

gas transport in the unsaturated zone.

Section B describes fluid flow and solute transport in fractured rock at the Mirror Lake Site in New 

Hampshire, where methods to characterize contaminated fractured rock are being developed. The methods are interdisciplinary and draw on techniques from 

hydrology, geology, geochemistry, and geophysics. 

Specific examples of papers are (1) discussions on the 

use of well logs to help design tracer tests and (2) comparisons of subsurface fracture mapping with electromagnetic and seismic tomography and with borehole 

acoustic televiewer and borehole flowmeter logs.

Section C contains papers on the migration of 

sewage-contaminated ground water on Cape Cod, 

Massachusetts, and on the transport and fate of metals, 

microbes, and organic chemicals in ground water. The 

papers describe the use of a large array of multilevel 

samplers to conduct large- and small-scale (hundreds 

of meters to meters) tracer tests designed to study a 

variety of transport processes in ground water. Examples of topics of additional papers in this section 

include: phosphate contamination in ground water, the 

role of protozoa in an organically contaminated aquifer, and numerical simulation of contaminant transport.

Section D is on the biodegradation and transport 

of chlorinated solvents, such as trichloroethylene 

(TCE), in ground water at Picatinny Arsenal, New 

Jersey. The papers describe the desorption of TCE 

from aquifer sediments, the microbial degradation and 

transport of TCE in ground water, the volatilization 

and transport of TCE in the unsaturated zone, and the 

surfactant-enhanced remediation of ground water contaminated with TCE.

Section E presents the development of methods 

to analyze and collect water samples, and to investigate transport processes. These papers discuss, for 

example, the development of a new method to collect 

sediment cores without loss of sample, new analytical 

methods to measure the concentrations of pesticides 

and pesticide metabolites in water, and the use of 

nitrogen isotopes to study the subsurface transport of 

nitrogen.

Section F contains papers on the environmental 

fate of agricultural chemicals in the Midwest. The 

papers describe the occurrence, distribution, and transport of agricultural chemicals in ground water, 

streams, large rivers, reservoirs, and precipitation at
regional to test plot scales. Examples of specific topics in this section are papers on the relation of nitrate 

concentrations in surface water to land use, the persistence of herbicides in reservoirs, and the results of a 

reconnaissance of pesticides in near-surface aquifers. 

The final eight papers in this section present results 

from the Management Systems Evaluation Areas 

(MSEA) program, which is part of an interagency 

research program to evaluate the effects of agricultural-management (farming) systems on water quality.

Section G contains papers on the natural biodegradation and transport of crude oil in the subsurface at Bemidji, Minnesota. The papers describe 

crude-oil contamination that resulted from a pipeline 

failure. Because the time the contamination occurred 

and the amount of crude oil in the subsurface are 

known, this site has provided USGS researchers and 

others a unique opportunity to study contaminant 

hydrology; information on the source of contamination usually is not available for most sites of groundwater contamination. Examples of specific topics in 

this section are papers on the use of simulation to 

study field-scale solute transport and biodegradation, 

and on small-scale chemical heterogeneities in the 

contaminated aquifer.

Section H focuses on the fate of anthropogenic 

contaminants in the San Francisco Bay Estuary and 

tributaries in California. The papers describe the 

transport of pesticides from the San Joaquin and 

Sacramento Rivers, California, through the Sacramento-San Joaquin Delta, and into the Bay. In addition, the papers in this section describe the fate of 

organochlorine compounds, the results of a survey of 

molecular marker compounds, and the distribution 

and flux of sulfides in the Bay.

Section I contains papers on the transport and 

fate of toxic metals in acidic mine drainage in the 

headwaters of the Upper Arkansas River in Colorado. Results of tracer and pH-modification experiments in St Kevin Gulch, Leadville, Colorado~a 

small mountain watershed-are discussed. These 

experiments were designed to isolate hydraulic transport processes so that geochemical processes can be 

studied.

Section J addresses a variety of topics relating 

to the hydrology of toxic substances in ground water 

and surface water. The papers are grouped by topic, 

and discuss, for example, the fate of organic compounds in surface water, the fate of hydrocarbons in 

the subsurface, the microbial degradation of explosive

wastes, and the measurement of unsaturated-zone 

properties.

Section K reports on the transport of agricultural chemicals in ground water at an agricultural 

research plot near Plains, Georgia, where USGS and 

U.S. Department of Agriculture researchers collaborated on a tracer test involving bromide, atrazine, and 

other agricultural chemicals beneath a plot of irrigated 

corn. The papers discuss, for example, the twodimensional distribution of a bromide tracer in the 

unsaturated zone, and the spatial variability of soil 

properties affecting the transport of agricultural chemicals.

Section L contains papers on nonpoint sources 

of ground-water contamination. These papers, which 

report results of studies on Long Island, New York, 

and the New Jersey coastal plain, describe the relation 

between land use and ground-water contamination.

Section M presents papers on acidic groundwater and surface-water contamination from copper 

mining near Final Creek, Arizona. The papers 

discuss, for example, the use of ground-water age dating to confirm estimates of traveltimes, the interaction 

of the contaminant plume with Final Creek, and the 

relations among algal species and chemical gradients 

in Final Creek.

Compiling and editing a proceedings of this size 

is a complex and difficult task. The task was made 

easier by the work and dedication of David Aronson, 

who died on April 7, 1996. David coedited two proceedings of technical meetings for the Toxic Substances Hydrology Program, and received awards for 

his work. This proceedings is a testament to his many 

contributions to the U.S. Geological Survey.

David W. Morganwalp

Acting Coordinator

Toxic Substances Hydrology Program

fv
CONTENTS 

Volume 1

Page 

SECTION A Gasoline in Ground Water in Galloway Township, New Jersey

Overview of Research on the Transport, Microbial Degradation, and Remediation of Hydrocarbons at a 

Subsurface Gasoline-Spill Site in Galloway Township, New Jersey

By Arthur L. Baehr and Jeffrey M. Fischer.............................................................................................................. 1

Hydrogeology, Contaminant Distribution, and Biodegradation Processes at a Gasoline-Spill Research Site 

in Galloway Township, New Jersey

By Jeffrey M. Fischer, Nicholas P. Smith, and Arthur L. Baehr.............................................................................. 7

Biogeochemical Fate of Organic Acid Metabolites Produced During the Microbial Degradation of 

Hydrocarbons in Ground Water

By Isabelle M. Cozzarelli and Mary Jo Baedecker.................................................................................................. 17

Method for Estimating the Rate of Microbial Degradation of Hydrocarbons Associated with Bioventing 

ul a Gaoolinc-Spill Site in Galloway Township, New Jersey

By Matthew A. Lahvis, Arthur L. Baehr, and Jeffrey M. Fischer............................................................................ 27

Determination of Vapor-Phase Diffusion Coefficients for Unsaturated-Zone Sediments at a 

Gasoline-Spill Site in Galloway Township, New Jersey

By Jeffrey M. Fischer, Ronald J. Baker, Matthew A. Lahvis, and Arthur L. Baehr................................................ 35

Use of Column Studies and a Reactive Transport Model to Measure Biodegradation Rates of Hydrocarbons

By Ronald J. Baker and Arthur L. Baehr................................................................................................................ 43

Method for Monitoring Biodegradation of Hydrocarbons in Unsaturated Porous Media

By John G. Nolan, Ronald J. Baker, and Arthur L. Baehr ...................................................................................... 49

Optimizing the Design of Vapor-Extraction Remediation Systems for Removal of Organic 

Contaminants From the Unsaturated Zone

By Claire Welty, Craig J. Joss, and Arthur L. Baehr............................................................................................... 55

A Multispecies Transport Model to Simulate Vapor Extraction and Bioventing Remediation of Unsaturated 

Zones Contaminated with Gasoline

By C.J. Joss and A.L. Baehr..................................................................................................................................... 61

A Comparison of Two Analytical Techniques For Identifying and Measuring the Concentrations of 

Volatile Hydrocarbons in Ground Water

By Curtis S. Phinney and Isabelle M. Cozzarelli..................................................................................................... 65

SECTION B Research on Flow and Transport in Fractured Rock at Mirror Lake, New Hampshire

Overview of Research on Use of Hydrologic, Geophysical, and Geochemical Methods to

Characterize Flow and Chemical Transport in Fractured Rock at the Mirror Lake Site, New Hampshire

By Alien M. Shapiro and Paul A. Hsieh .................................................................................................................. 71

Characterizing Bedrock Fractures in Outcrop for Studies of Ground-Water Hydrology: An Example 

from Mirror Lake, Grafton County, New Hampshire

By Christopher C. Barton......................................................................................................................................... 81

Use of a Borehole Color Video Camera to Identify Lithologies, Fractures, and Borehole Conditions in 

Bedrock Wells in the Mirror Lake Area, Grafton County, New Hampshire

By Carole D.Johnson.............................................................................................................................................. 89

naciuie Detection in Crystalline Rocks, Mirror Lake Area, Grafton County, New Hampshire

By P.P. Haeni, John W. Lane, Jr., C.C. Barton, and David A. Lieblich.................................................................. 95

Use of Well Logs to Prepare the Way for Packer Strings and Tracer Tests: Lessons from the 

Mirror Lake Study

By Frederick L. Paillet............................................................................................................................................. 103

Velocity, Attenuation, and Dispersion Electromagnetic Tomography in Fractured Rock

By David L. Wright, Gary R. Olhoeft, and Thomas P. Grover................................................................................ 111
Electromagnetic and Seismic Tomography Compared to Borehole Acoustic Televiewer and Flowmeter Logs 

for Subsurface Fracture Mapping at the Mirror Lake Site, New Hampshire

By David L. Wright, Gary R. Olhoeft, Paul A. Hsieh, Ernest L. Majer, Frederick L. Paillet, and

John W. Lane, Jr................................................................................................................................................... 119

Hydraulic Characteristics of Fractured Bedrock Underlying the FSE Well Field at the Mirror Lake Site, 

Grafton County, New Hampshire

By Paul A. Hsieh and Alien M. Shapiro .................................................................................................................. 127

A New Method of Performing Controlled Injection of Traced Fluid in Fractured Crystalline Rock

By Alien M. Shapiro and Paul A. Hsieh.................................................................................................................. 131

Field Instrumentation for Multilevel Monitoring of Hydraulic Head in Fractured Bedrock at the Mirror 

Lake Site, Grafton County, New Hampshire

By Paul A. Hsieh, Richard L. Perkins, and Donald O. Rosenberry......................................................................... 137

Factors Affecting Recharge to Crystalline Rock in the Mirror Lake Area, Grafton County, New Hampshire

By Philip T. Harte and Thomas C. Winter............................................................................................................... 141

Concentrations of Chlorofluorocarbons and Other Gases in Ground Water at Mirror Lake, New Hampshire

By Eurybiades Busenberg and L. Niel Plummer...................................................................................................... 151

Helium Isotope Analysis and Tritium-Helium Age Dating in the Mirror Lake Basin, Grafton County, 

New Hampshire

By S. Drenkard, T. Torgersen, R. Weppernig, K. Parley, P. Schlosser, R.L. Michel, A.M. Shapiro and

W.W. Wood ....................................................................................................................................................^ 159

Observational, Experimental and Inferred Evidence for Solute Diffusion in Fractured Granite Aquifers: 

Examples from the Mirror Lake Watershed, Grafton County, New Hampshire

By Warren W. Wood, Alien M. Shapiro, Paul A. Hsieh, and Terry B. Councell.................................................... 167

A Conceptual Model for Estimating Regional Ground-Water Velocity in Bedrock of the Mirror Lake Area, 

Grafton County, New Hampshire

By Alien M. Shapiro, Warren W. Wood, Eurybiades Busenberg, Stefan Drenkard, L. Niel Plummer,

Thomas Torgersen, and Peter Schlosser.............................................................................................................. 171

SECTION C Contaminant Transport Research on Cape Cod, Massachusetts

Overview of Research at the Cape Cod Site: Field and Laboratory Studies of Physical, Chemical, and 

Microbiological Processes Affecting Transport in a Sewage-Contaminated Aquifer

By Denis R. LeBlanc................................................................................................................................................ 179

The Transport of Inorganic Contaminants in a Sewage Plume in the Cape Cod Aquifer, Massachusetts

By Brigid A. Rea, Douglas B. Kent, Linda C. D. Anderson, James A. Davis, and Denis R. LeBlanc.................... 191

Potential Long-Term Effects of Phosphate Contamination in a Sand and Gravel Aquifer, 

Cape Cod, Massachusetts

By Kenneth G. Stollenwerk...................................................................................................................................... 199

Background Aqueous Chemistry and Effects of Carbon Dioxide Variations in Recharge in Shallow 

Ground Water in a Glacial Outwash Aquifer, Cape Cod, Massachusetts

By Roger W.Lee...........................................................................................^ 207

Heavy-Metal Transport in a Sand and Gravel Aquifer with Variable Chemical Conditions, Cape Cod, 

Massachusetts

By James A. Davis, Douglas B. Kent, Jennifer A. Coston, andKathrynM. Hess.................................................. 219

Evaluation of Geochemical Indicators of Metal Adsorption in a Sand and Gravel Aquifer, Cape Cod, 

Massachusetts

By Christopher C. Fuller, Jennifer A. Coston, James A. Davis, and Eleanor Dixon............................................... 227

Spatial Variability of Metal-Ion Adsorption and Hydraulic Conductivity in a Sand and Gravel Aquifer, 

Cape Cod, Massachusetts

By Kathryn M. Hess, James A. Davis, Christopher C. Fuller, and Jennifer A. Coston........................................... 235

Small-Scale Tracer Tests Applied to the Measurement of in situ Denitrification Rates in a SewageContaminated Aquifer, Cape Cod, Massachusetts

By Myron H. Brooks, Richard L. Smith, and Stephen P. Garabedian..................................................................... 243

vi
Overview of Research on the Distribution and Role of Protozoa in an Organically Contaminated Aquifer at 

Cape Cod, Massachusetts

By Nancy E. Kinner and Ronald W. Harvey............................................................................................................ 249

Transport of Indigenous Protozoa in a Sandy Aquifer, Cape Cod, Massachusetts

By Ronald W. Harvey, Nancy E. Kinner, Dan MacDonald, Amoret L. Bunn, and David W. Metge. .................... 253

Use of Static Column Experiments to Identify Factors Affecting Bacterial Attachment in Contaminated 

Aquifer Sediments From Cape Cod, Massachusetts

By David W. Metge, Ronald W. Harvey, George R. Aiken, and Larry B. Barber, II ............................................. 259

An Experiment to Quantify Temporal Variability of Water Samples Obtained from Screened Wells, 

Cape Cod, Massachusetts

By Thomas E. Reilly and Denis R. LeBlanc.. .......................................................................................................... 265

Preliminary Type-Curve Analysis of an Aquifer Test in an Unconfined Sand and Gravel Aquifer, 

Cape Cod, Massachusetts

By Allan F. Moench, Denis R. LeBlanc, and Stephen P. Garabedian ..................................................................... 273

Numerical Simulation of Downward Movement of Solutes During a Natural-Gradient Tracer Test in Sand and 

Gravel, Cape Cod, Massachusetts

By Denis R. LeBlanc and Michael A. Celia............................................................................................................. 283

Effects of Changes in Aquifer Properties on Simulated Fluid-Particle Pathlines, Cape Cod, Massachusetts

By John P. Masterson and Donald A. Walter........................................................................................................... 289

SECTION D Chlorinated Solvents in Ground Water at Picatinny Arsenal, New Jersey

Overview of Research Activities on the Transport and Fate of Chlorinated Solvents in Ground Water at 

Picatinny Arsenal, New Jersey, 1991-93

By Thomas E. Imbrigiotta and Mary Martin............................................................................................................ 297

Simulation of Transport, Desorption, Volatilization, and Microbial Degradation of Trichloroethylene in 

Ground Water at Picatinny Arsenal, New Jersey

By Mary Martin........................................... ............................................................................................................. 307

Simulation of Trichloroethylene Volatilization in the Unsaturated Zone during a Field Infiltration Experiment at 

Picatinny Arsenal, New Jersey

By H. Jean Cho, James A. Smith, and Peter R. Jaffe ............................................................................................... 315

Cometabolic Biotransformation of Trichloroethylene and cis- 1 ,2-Dichloroethylene in Unsaturated-Zone 

Soil at Picatinny Arsenal, New Jersey

By Theodore A. Ehlke and Thomas E. Imbrigiotta.................................................................................................. 321

Desorption of Trichloroethylene From Aquifer Sediments at Picatinny Arsenal, New Jersey

By David Roller, Thomas E. Imbrigiotta, Arthur L. Baehr, and James A. Smith ................................................... 329

Effect of Air Drying on Solvent Extraction of Trichloroethylene-Contaminated Soil, Picatinny Arsenal, 

New Jersey

By David Roller, Thomas E. Imbrigiotta, and James A. Smith............................................................................... 339

In Situ Biotransformation of Trichloroethylene and cis- 1 ,2-dichloroethylene at Picatinny Arsenal, 

New Jersey

By Theodore A. Ehlke, Barbara H. Wilson, John T. Wilson, and Thomas E. Imbrigiotta...................................... 347

Surfactant-Enhanced Remediation of Ground Water at Picatinny Arsenal, New Jersey

By James J. Deitsch and James A. Smith................................................................................................................. 355

SECTION E Methods Development

g Borehole Flow to Quantify Aquifer Cross-Contamination by Volatile Organic Compounds 

By Ronald A. Sloto.................................................................^ 361

A Sample-Freezing Drive Shoe for a Wireline-Piston Core Sampler

By Fred Murphy and W.N. Herkelrath..................................................................................................................... 369

Determination of Surface Area of Soil Components by Sorption of Nitrogen and Ethylene Glycol Monoethyl 

Ether Vapors and Evaluation of Adsorption, Partition and Cation Solvation Processes

By Gary T. Chiou and David W. Rutherford............................................................................................................ 375

vii
Isolation of Lipophilic Organic Contaminants Along the Upper Mississippi River Using a 

Semipermeable Membrane Device

By Geoffrey S. Ellis and Colleen E. Rostad............................................................................................................. 381

Measurement of Concentrations of Alachlor and Its Ethanesulfonic Acid Metabolite in Water by Solid-Phase 

Extraction and Enzyme-Linked Immunosorbent Assay

By Diana S. Aga, E. Michael Thurman, and Michael L. Pomes.............................................................................. 385

Isocratic Separation of Alachlor Ethanesulfonic Acid, Alachlor Oxoacetic Acid, and Hydroxyatrazine by 

Reversed-Phase Liquid Chromatography

By Michael L. Pomes, Douglas F. Holub, Diana S. Aga, and E.M. Thurman......................................................... 393

Use of 15N to Trace Movement of Nitrogen Fertilizer at a Field Plot

By Dale W. Blevins, Donald H. Wilkison, Steven R. Silva, and Brian P. Kelly..................................................... 401

SECTION F Occurrence, Distribution, and Transport of Agricultural Chemicals in the Midwest

Occurrence, Deposition, and Long Range Transport of Herbicides in Precipitation in the Midwestern and 

Northeastern United States

By Donald A. Goolsby, E. Michael Thurman, Michael L. Pomes, Michael Meyer, and William A. Battaglin...... 411

Atrazine Transport and Degradation in a Pristine Watershed: The Fate of Atrazine Deposited by Precipitation

By Aron E. Cromwell and E. Michael Thurman...................................................................................................... 423

Annual Use and Transport of Agricultural Chemicals in the Mississippi River, 1991-92

By William A. Battaglin, Donald A. Goolsby, and Richard H. Coupe.................................................................... 429

Relation of Nitrate Concentrations in Surface Water to Land Use in the Upper-Midwestern United States, 

1989-90 

By David K. Mueller, Barbara C. Ruddy, and William A. Battaglin....................................................................... 441

Reconnaissance Data for Selected Herbicides and Two Atrazine Metabolites in Surface Water of the Midwestern 

United States: Chemical Analysis by Immunoassay and Gas Chromatography/Mass Spectrometry

By Elisabeth A. Scribner, E. Michael Thurman, and Donald A. Goolsby............................................................... 449

Chemistry, Degradation, and Transport of Triazine Herbicide Metabolites in Surface Water

By E.M. Thurman, M.S. Mills, and M.T. Meyer..................................................................................................... 455

The Transport and Degradation of Cyanazine Metabolites in Surface Water of the Midwestern United States

By M.T. Meyer and E.M. Thurman.......................................................................................................................... 463

Transport and Concentrations of Selected Herbicides and Nitrate in the Sangamon River, Illinois, 

April 1991 -March 1992

ByG.P. Johnson and R.H. Coupe ............................................................................................................................ 471

Temporal Distributions of Herbicide and Nitrite-Nitrate Concentrations and Loads in the West Fork of the 

Big Blue River, Nebraska, 1990 and 1991-92

By A.D. Druliner, D.A.Goolsby, and M. Meyer...................................................................................................... 479

Persistence of Herbicides in Selected Reservoirs in the Midwestern United States: Some Preliminary Results 

By Donald A. Goolsby, William A. Battaglin, James D. Fallon, Diana S. Aga, Dona W. Koplin, and

E. Michael Thurman............................................................................................................................................. 487

Determining the Relative Age, Transport, and Three-Dimensional Distribution of Atrazine in a Reservoir 

Using Immunoassay

By James D. Fallon and E.M. Thurman................................................................................................................... 499

Pesticides in Near-Surface Aquifers: Results of the Midcontinental United States Ground-Water 

Reconnaissance, 1991-92

By Dana W. Koplin, Donald A. Goolsby, Dana S. Aga, Jana L. Iverson, and E. Michael Thurman...................... 505

Predicting Nitrate-Nitrogen and Atrazine Contamination in the High Plains Aquifer in Nebraska

ByA.D. Druliner and T.S. McGrath........................................................................................................................ 515

Comparison of Simulated and Observed Movement of Bank-Storage Water Adjacent to the Cedar River, Iowa

By Paul J. Squillace and David W. Pollock............................................................................................................. 525

Fate of Alachlor, Atrazine, and Bromide in a Corn Plot near Topeka, Kansas

By D.A. Eckhardt, R.J. Wagenet, E.M. Thurman, and P.L. Barnes......................................................................... 535

viii
Effects of Farming Systems on Ground-Water Quality at the Princeton, Minnesota, Management 

Systems Evaluation Area, 1991

By Matthew K. Landon, Geoffrey N. Delin, John A. Lamb, Robert H. Dowdy, and James L. Anderson.............. 543

Effects of Topography on the Transport of Agricultural Chemicals near Princeton, Minnesota, 1992

By Geoffrey N. Delin and Matthew K. Landon....................................................................................................... 555

Spatial Variability of Unsaturated-Zone Properties in Relation to Topography in a Sand-Plain Setting near 

Princeton, Minnesota

By G.N. Delin, M.K. Landon, R.W. Healy, and H.W. Olsen.................................................................................. 565

Nitrogen Uptake and Soil Water Variability Across a Sand Plain Landscape

By J.A. Lamb, M.D. Tomer, J.L. Anderson, and R.H. Dowdy................................................................................ 575

Geophysical Investigations of Heterogeneity and Scale at the Princeton, Minnesota, Management 

Systems Evaluation Area

By Jeffrey E. Lucius and Gary R. Olhoeft............................................................................................................... 581

Geochemistry of Nitrogen in a Farmed Watershed Near Princeton, Minnesota

By Richard B. Wanty, Michele L. Tuttle, Matthew K. Landon, Geoffrey N. Delin, and J.K. Bohlke.................... 591

Chemical Loads of Nitrate, Atrazine, and Metolachlor in Walnut Creek Watershed near Ames, Iowa, 1991-92

By P.J. Soenksen, J.L. Hatfield, and D.J. Schmitz................................................................................................... 597

Atrazine Transport Through Preferential Pathways at the Missouri MSEA

By James A. Tindall and William K. Vencill........................................................................................................... 607

Volume 2

SECTION G Crude Oil in the Subsurface Environment at Bemidji, Minnesota

Evolution of the Contaminant Plume in an Aquifer Contaminated with Crude Oil, Bemidji, Minnesota

By Mary Jo Baedecker, Isabelle M. Cozzarelli, Philip C. Bennett, Robert P. Eganhouse, and Marc F. Hult......... 613

Use of Simulation to Study Field-Scale Solute Transport and Biodegradation at the Bemidji, Minnesota, 

Crude-Oil Spill Site

By H.I. Essaid, M.J. Baedecker, and I.M. Cozzarelli............................................................................................... 621

Source Mass Balance Calculated from Changes in Composition of Spilled Crude Oil in the Subsurface near 

Bemidji, Minnesota

By Matthew K. Landon and Marc F. Hult................................................................................................................ 631

Studies of the Molecular Size of Dissolved Organic Carbon Fractions Downgradient from the Oil Body at 

Bemidji, Minnesota

By George Aiken...................................................................................................................................................... 641

Small-Scale Chemical Heterogeneities in a Crude-Oil-Contaminated Aquifer, Bemidji, Minnesota

By Isabelle M. Cozzarelli, Mary Jo Baedecker, George Aiken, and Curtis Phinney............................................... 647

Use of Carbon and Hydrogen Stable Isotopes to Investigate the Production and Fate of Methane at a 

Toxic Waste Site, Bemidji, Minnesota

By Kinga Revesz, Tyler Coplen, Mary J. Baedecker, and Marc Hult..................................................................... 653

Sorption of Polycyclic Aromatic Hydrocarbons to Aquifer Materials Containing a Low-Organic Carbon 

Content from Bemidji, Minnesota

By Joseph J. Piatt, Steven J. Eisenriech, Debera A. Backhus, and Paul D. Capel................................................... 663

Investigations of Microbial Weathering of Minerals

By Philip C. Bennett and Franz K. Hiebert.............................................................................................................. 671

Multiphase-Flow Modeling of the Bemidji, Minnesota Crude-Oil-Spill Site with Geostatistical Simulation of 

Hydraulic Properties

By L.A. Dillard, H.I. Essaid, and W.N. Herkelrath.................................................................................................. 679

ix
SECTION H Anthropogenic Contaminants in the San Francisco Bay Estuary and Tributaries, 

California

Challenges in Detecting Contaminant Effects on an Estuarine Ecosystem Affected by Many Different 

Disturbances: San Francisco Bay

By Samuel N. Luoma, and Frederic H. Nichols....................................................................................................... 689

Hydrodynamics of California's Sacramento-San Joaquin Delta, and the Collection of Flow Data for the 

Tracking of Pesticides

By Richard N. Oltmann............................................................................................................................................ 695

Diazinon Concentrations and Transport in the Sacramento River and San Francisco Bay, California, 

February 1993

By Kathryn M. Kuivila and Dwight D. Copeland.................................................................................................... 699

Distribution and Possible Biological Effects of Diazinon in the San Joaquin River and the Sacramento-San 

Joaquin Delta, California, February 1993

By Kathryn M. Kuivila, Robert C. Sheipline, and Christopher G. Foe.................................................................... 705

Concentrations of Dissolved Rice Pesticides in the Colusa Basin Drain and Sacramento River, 

California, 1990-92

By Kathryn L. Crepeau, Kathryn M. Kuivila, and Joseph L. Domagalski.............................................................. 711

Method for Estimating the Activity of Carbofuran-Degrading Microorganisms in Environmental Samples

By William J. Meyers and Kathryn M. Kuivila....................................................................................................... 719

Distributions of Organochlorine Compounds in Sediments and Livers of Striped Bass (Morone Saxatilis) in 

San Francisco Bay-Delta Estuary

By Wilfred E. Pereira, Frances D. Hostettler, John R. Cashman and Richard S. Nishioka..................................... 725

A Survey of Molecular Marker Compounds in Sediments of San Francisco Bay, California

By Frances D. Hostettler, John B. Rapp, Wilfred E. Pereira, and Keith A. Kvenvolden........................................ 735

Distributions and Benthic Flux of Dissolved Sulfides in the Toxic Water Column of San Francisco Bay, California

By James S. Kuwabara, Yvonne R. Hunter, and Cecily C.Y. Chang...................................................................... 747

SECTION I Metals in the Headwaters of the Upper Arkansas River, Colorado

Past and Present Research on Metal Transport In St. Kevin Gulch, Colorado

ByBriant A. Kimball............................................................................................................................................... 753

Iron and Zinc Budgets in Surface Water for a Natural Wetland Affected by Acidic Mine Drainage, 

St. Kevin Gulch, Lake County, Colorado

By Kathcrinc Walton-Day........................................................................................................................................ 759

Tracer Injections in Small Streams Why and How We Do Them

By Gary W. Zellweger............................................................................................................................................. 765

Processes Controlling Dissolved Copper Concentrations during an Instream pH-Modification Experiment

By Kathleens. Smith............................................................................................................................................... 769

An Equilibrium-Based Simulation Model for Reactive Solute Transport in Small Streams

By Robert L. Runkel, Kenneth E. Bencala and Robert E. Broshears....................................................................... 775

Simulation of Reactive Transport During a pH Modification Experiment in a Mountain Stream Affected by Acid 

Mine Drainage

By Robert E. Broshears, Briant A. Kimball and Robert L. Runkel......................................................................... 781

Difficulties with Intermediate-Scale Experiments for Studies of Iron Chemistry in Streams Affected by Acidic 

Mine Drainage

By Diane M. McKnight, Kenneth E. Bencala, Richard A. Hamish, and Robert L. Runkel.................................... 789

Metal Concentrations in Sedges in a Wetland Receiving Acidic Mine Drainage from St. Kevin Gulch, 

Leadville, Colorado

By B.M. Erickson, P.H. Briggs, and T.R. Peacock.................................................................................................. 797

Uraniferous Waters of Southeastern Colorado: A Function of Geology, Climate and Land Use

By Robert A. Zielinski and Sigrid Asher-Bolinder.................................................................................................. 805
SECTION J Research on the Fate and Transport of Contaminants in Ground Water and Surface 

Water

Dissolved Organic Compounds in the Mississippi River from Minneapolis-St. Paul, Minnesota 

to New Orleans, Louisiana

By Larry B. Barber, II, Jerry A. Leenheer, Wilfred E. Pereira, and Mary A. Noriega............................................ 813

Linear Alkylbenzene Sulfonate in the Mississippi River

By Charles F. Tabor, Jr., and Larry B. Barber, II..................................................................................................... 821

Organic Carbon Associated With Suspended Silt and Colloid From the Mississippi River and Some of 

Its Tributaries, July 1991-May 1992

By Colleen E. Rostad, Stephanie G. Monsterleet, LaDonna M. Bishop, and Geoffrey S. Ellis.............................. 825

Accumulation of Organochlorine Compounds in Hydrilla Verticillata Relative to Sediments in the Tidal 

Potomac River

By Jessica A. Hopple and Gregory D. Foster........................................................................................................... 831

Pathways of Methanogenic Biodegradation of Creosote-Derived Aromatic Compounds

By E. Michael Godsy, Donald F. Goerlitz, and Dunja Grbic'-Galic'....................................................................... 835

Modeling Steady-State Methanogenic Degradation of Phenols in Ground Water at Pensacola, Florida

By Barbara A. Bekins, E. Michael Godsy, and Donald F. Goerlitz........................................................................ 843

Modeling Breakthrough of Nitrogen Heterocyclic Compounds in Laboratory Columns Containing 

Creosote-Contaminated Aquifer Material

By Ean Warren and E. Michael Godsy.................................................................................................................... 851

Bioremediation of Petroleum and Metal Contamination with Dissimilatory Metal Reduction

ByDerekR. Lovley....................................................................................................................................^ 859

Effects of Pb and Terminal-Electron-Accepting Processes on Organic Acid Concentrations in Contaminated 

Aquifer Sediments

By Paul M. Bradley, Francis H. Chapelle, and Don A. Vroblesky.......................................................................... 865

Shifting of Terminal-Electron-Accepting Processes in a Petroleum-Hydrocarbon-Contaminated Aquifer

By Don A. Vroblesky and Francis H. Chapelle....................................................................................................... 873

Preliminary Assessment of Jet Fuel Contamination of an Atoll Aquifer

By Gordon W. Tribble, Roger W. Lee, Duane S. Wydoski, Jill D. Torikai, and Charles D. Hunt, Jr..................... 881

Microbial Degradation of Nitrotoluenes in Surface Soils and Aquifer Sediments, Weldon Spring, Missouri

By Francis H. Chapelle and Paul M. Bradley.......................................................................................................... 889

Depletion of Nitrogen-Bearing Explosives Wastes in a Shallow Ground-water Plume near Hawthorne, Nevada

By A.S. Van Denburgh, D.F. Goerlitz, and E.M. Godsy......................................................................................... 895

Reaction of Arsenic Minerals and Gold Ores with Alkaline Cyanide Solutions

By Walter H. Ficklin, Geoffrey S. Plumlee, and John B. McHugh ......................................................................... 905

Anaerobic Growth of Bacterial Strain Ses3 with Selenate as the Electron Acceptor

By Jodi Switzer Blum, Charles W. Culbertson, and Ronald S. Oremland............................................................... 911

Chemical, Isotopic, and Microbiological Evidence for Denitrification During Transport of Domestic Wastewater 

Through a Thick Unsaturated Zone in the Mojave Desert

By Roy A. Schroeder, Peter Martin, and J.K. Bohlke.............................................................................................. 917

Potential Anhydrite Precipitation Associated With Deep Injection of Ground-Water Brine from 

Paradox Valley, Colorado

By Yousif K. Kharaka, William C. Evans, Gil Ambats, and James J. Thordsen..................................................... 927

Apparent Solute Dispersion in an Alluvial/Aquifer Stream System

By J. Jeffrey Starn.................................................................................................................................................... 935

Measurements of Hydraulic Conductivity and Capillary Pressure under Unsaturated Conditions in a Laboratory 

Triaxial System

By Harold W. Olsen................................................................................................................................................. 941

Unsaturated Zone Properties at a Hazardous-Waste-Disposal Site at the Idaho National Engineering 

Laboratory

By Stephanie Shakofsky and John R. Nimmo.......................................................................................................... 949

XI
SECTION K Agricultural Chemicals in the Unsaturated Zone and Ground Water Near Plains, Georgia

Fate and Transport of Atrazine at the Plains, Georgia, Ground Water Study Site

By R. A. Leonard, L.R. Marti, D.W. Hicks, and J.B. McConnell............................................................................ 955

Unsaturated Zone Transport at the Plains, Georgia Field Site

By D.D. Bosch, R.A. Leonard, and R.K. Hubbard.................................................................................................. 965

Two-Dimensional Distribution of a Bromide Tracer in the Unsaturated Zone at the Plains, Georgia, 

Research Site

By D.W. Hicks, J.B. McConnell, H.H. Persinger, J.D. Scholz, and R.K. Hubbard................................................. 971

Spatial Variability of Soil Properties Affecting Chemical Transport at the Plains, Georgia, Research Site

By R. K. Hubbard, D. D. Bosch, R. A. Leonard, and D. W. Hicks ........................................................................ 983

SECTION L Nonpoint-Source Contamination Investigations

Modeling Statistical Relations Among Shallow Ground-Water Quality, Human Activities, Land Use, 

and Thickness of the Unsaturated Zone on Long Island, New York

By Paul E. Stackelberg and David A.V. Eckhardt................................................................................................... 993

Evaluating the Effect of Land Use and Sampling Depth on Ground-Water Quality, Long Island, New York 

By Paul E. Stackelberg.....................................................^

Use of Flowpath Simulation to Determine Contributing Areas and Traveltimes of Nonpoint-Source 

Ground-Water Contamination, Gloucester County, New Jersey

By Anthony S. Navoy...............................................................................................................................................1011

Vulnerability of Water from Public Supply Wells to Contamination by Pesticides, Potomac-Raritan-Magothy 

Aquifer System, New Jersey Coastal Plain

By Eric F. Vowinkel................................................................................................................................................. 1021

A Review of Methods for Assessing the Potential for Regional Ground-Water Contamination

By David R. Seller................................................................................................................................................... 1^

SECTION M Ground-Water and Surface-Water Contamination From Copper Mining Near 

Final Creek, Arizona

Hydrologic and Geochemical Factors affecting Metal Contaminant Transport in Final Creek Basin near 

Globe, Arizona

By James G. Brown and Judson W. Harvey ............................................................................................................ 1035

Dissolved Gas and Chlorofluorocarbon Content of Ground Waters in the Pinal Creek Basin, Arizona

Bv Pierre D. Glynn and Eurybiades Busenberg....................................................................................................... 1043

Unsaturated Zone Diffusion of Carbon Dioxide and Oxygen in the Pinal Creek Basin, Arizona

By Pierre D. Glynn and Eurybiades Busenberg....................................................................................................... 1055

Interactions between Shallow Ground Water and Surface Water that Affect Metal Transport in 

Pinal Creek, Arizona

By Judson W. Harvey, Christopher C. Fuller, and Brian J. Wagner....................................................................... 1065

Association of Selected Metals with Colloidal and Suspended Paniculate Material in Shallow Ground Water 

and Surface Water at Pinal Creek, Arizona

By Judson W. Harvey and Christopher C. Fuller..................................................................................................... 1073

Solute-Transport Parameter Estimation for an Injection Experiment at Pinal Creek, Arizona

By Brian J. Wagner and Judson W. Harvey............................................................................................................. 1081

Alteration of Alluvium by Acidic Ground Water Resulting from Copper Mining at Pinal Creek, Arizona

By Carol J. Lind and Kenneth G. Stollenwerk......................................................................................................... 1089

Chemical Processes in Manganese Oxide and Carbonate Precipitation in Pinal Creek, Arizona

By John D. Hem and Carol J. Lind .......................................................................................................................... 1095

Experimental Studies of Trace-Metal Partitioning in Perennial Reaches of Pinal Creek, Arizona

By Christopher C. Fuller ..........................................................................................................................................1101

Relations Among Algal Species and Chemical Gradients in Pinal Creek, near Globe, Arizona

By Patrice H. Spindler and Milton R. Sommerfeld.................................................................................................. 1107

xii
Evolution of the Contaminant Plume in an Aquifer 

Contaminated with Crude Oil, Bemidji, Minnesota

By Mary Jo Baedecker1 , Isabelle M. Cozzarelli 1 , Philip C. Bennett2 , Robert P. Eganhouse1 ,

and Marc F. Hult3

Abstract

A long-term investigation of the geochemistry of a contaminant plume in an aquifer 

where crude oil floats on the water table indicates 

that the size of the plume of dissolved constituents has changed little over a 8-year period. Even 

though the geochemical reactions have changed 

over time and part of the plume has become 

anoxic, the biodegradation of dissolved hydrocarbons under oxic and anoxic conditions has 

prevented the hydrocarbons, in concentrations 

above Federal maximum contaminant levels, 

from moving more than 137 meters downgradient 

from the oil body. Another factor that helps contain the plume is the presence of silt layers of low 

hydraulic conductivity near the oil and the top of 

the saturated zone. The findings of this study support the conclusions that significant concentrations of petroleum-type hydrocarbons can be 

attenuated or removed from aquifers by natural 

hydrologic and biogeochemical processes.

INTRODUCTION AND OVERVIEW OF 

PROJECT

An investigation of the effects of a crude-oil 

spill on an aquifer was undertaken near Bemidji, Minnesota, as part of the Toxic Substances Hydrology 

Program of the U.S. Geological Survey (USGS). An 

underground pipeline carrying crude oil ruptured in

1979, and the oil sprayed over the land surface 

(referred to as the spray zone in fig. 1) . After partial

'U.S. Geological Survey, Reston, VA. 

"University of Texas, Department of Geological 

Sciences, Austin, TX.

3 Macalester College, Department of Geology, St. Paul, MM.

spray area

A - A' Line/section, 

see Fig. 2

e Ground water 

sampling sites

533 Well numbers 

0 45 Feet

15 Meters

Figure 1. Map of Crude-oil site near Bemidji, Minnesota, 

showing the location of the pipeline that ruptured, 

approximate location of the oil body, area over which oil 

was sprayed (spray zone), and locations of wells.

613
removal of the product, about 410 m3 of crude oil 

remained at the site (Hult, 1984). Some of the crude 

oil reached the ground water and is floating on the 

water table and a contaminant plume has developed in 

the aquifer downgradient from the oil body. A preliminary investigation and site characterization that began 

in 1983 was expanded in 1985 to examine (1) the fate 

of the crude oil, (2) development of the contaminant 

plume, (3) factors that affect the transport of chemical 

constituents, and (4) the effect of biogeochemical processes on aquifer solids. The scope of the work was 

expanded again in 1989 to investigate (5) the movement of the oil body, (6) partitioning at the oil/water 

interface, and (7) solute transport.

A summary of the major topics that have been 

investigated at the site is listed in table 1. The references are not inclusive and additional information can 

be found in publications referred to in the table and in 

USGS reports (see Mallard and Aronson, eds., 1991, 

for the most recent compilation of papers on this 

work). Many of the studies mentioned above are continuing and investigations are being expanded to 

include colloidal transport, sorption characteristics of 

the sediment, and microbial activity in the aquifer.

This paper describes the evolution of the contaminant plume and discusses the importance of assessing 

natural processes for removing petroleum-type hydroTable 1. Topics of investigation and selected references for research at the Bemidji, Minnesota site from 1984-93

[Publications marked * are in this volume]

Topic of research Selected references

Site description and geology

Ground penetrating radar

Configuration and movement of the oil body

Multi-phase flow modeling

Source mass balance 

Alkane mass transfer 

Unsaturated zone gases

Hydrogeology and inorganic geochemistry

Geochemistry of silica 

Formation of iron-bearing minerals 

Organic geochemistry 

Biogeochemical processes 

Solute transport and biodegradation 

Methane production and consumption 

Small-scale chemical heterogeneities 

Low-molecular-weight organic acids 

Dissolved organic carbon 

Non-volatile organic acids

Sorption on aquifer solids

Toluene oxidation by iron-reducing bacteria

Microbial weathering of silicates

Aerobic degradation__________

Hult, 1984; Franzi, 1988 

Olehoeft, 1991 

Hess and others 1992 

Essaid and others, 1993 

Dillard and others, 1994 * 

Landon and Hult, 1994 * 

Pfannkuch and others, 1988 

Hult and Grabbe, 1988; 

Baehr and Hult, 1991 

Siegel and others, 1986; 

Bennett and others, 1993 

Bennett, 1991

Baedecker and others, 1992 

Eganhouse and others, 1993 

Baedecker and others, 1993 

Essaid and others, 1994 * 

Revesz and others, 1994 * 

Cozzarelli and others, 1994 * 

Cozzarelli and others, 1990 

Aiken, 1994* 

Aiken and Thorn, 1989; 

Thorn and Aiken, 1989 

Piatt and others, 1994* 

Lovley and others, 1989 

Hiebert and Bennett, 1992 

Bennett and Hiebert, 1994 * 

Chang and others, 1988

614
carbons from ground water. The geochemistry of the 

plume has been examined in detail for a 8-year period to 

document the changes in the aquifer caused by to the 

alterations and transport of hydrocarbons.

DESCRIPTION OF SITE AND METHODS

The study site is located in the Bagley outwash 

plain Bemidji, Minnesota. The surficial aquifer is 

about 20 m thick and consists of sand and gravel outwash overlying till. Thin, discontinuous silt layers are 

interbedded with sand near the water table (Franzi, 

1988). The water table, in the part of the aquifer studied in this report, is 6 to 10m below land surface. 

Only the upper 5 to 8 m of the aquifer has been 

affected by crude oil. Locally ground water flows to 

the northeast and discharges to a small lake. The average linear flow velocity of water in the aquifer is 0.1 to 

0.5 m/d, but velocities as low as 0.05 m/d were measured in the fine-grained silty material (Bennett and 

others, 1993). The aquifer is mostly quartz with 30 

percent feldspar, 5 percent carbonate, and less than 5 

percent clay minerals (Bennett and others, 1993).

Piezometers were installed by hollow-stem 

auger and the wells were screened in depth intervals of 

0.15, 0.61 or 1.52 m. The screens were stainless steel 

and the casings were polyvinyl chloride. Water samples were collected with submersible pumps for analyses of inorganic constituents and dissolved organic 

carbon and with a Teflon bailer for analyses of organic 

constituents. Methods of chemical analyses presented 

in this paper are found in Eganhouse and others (1993) 

for hydrocarbons; in Baedecker and Cozzarelli (1992) 

for dissolved oxygen (DO), ferrous iron (Fe2+), volatile dissolved organic carbon (VDOC), methane 

(CH4) and carbon isotopes; and in Baedecker and others (1993) for other analytical methods and field sampling methods.

DEVELOPMENT OF THE CONTAMINANT 

PLUME

Plume development depends primarily on aquifer transmissivity, the extent of volatilization and solubilization of the contaminants, the amount of sorption, 

and the degradability of the dissolved constituents. 

The main components of crude oil at the Bemidji site 

are normal, alicyclic, and aromatic hydrocarbons. The 

most soluble aromatic hydrocarbons, benzene and the 

alkylbenzenes, are known to degrade by microbial

processes (Atlas, 1984). When the plume developed 

shortly after the spill, concentrations of oxygen 

present in the aquifer were sufficient to oxidize the 

hydrocarbons by aerobic microbial processes. After 

oxygen was depleted, the major processes that oxidized hydrocarbons in the anoxic zone were iron and 

manganese reduction and methanogenesis (Baedecker 

and others, 1989, 1993). It was demonstrated that 

these processes are linked with hydrocarbon oxidation 

in laboratory experiments using pure cultures for iron 

reduction (Lovely and others, 1989) and mixed cultures for methanogenesis (Grbic-Galic and Vogel, 

1987). Aerobic oxidation of hydrocarbons continues 

to be a major process at the edges of the anoxic zone 

where hydrocarbons come in contact with oxygenated 

ground water.

The plume at the Bemidji site can be defined by 

the distributions of VDOC; specific hydrocarbons 

(Eganhouse and others, 1993); geochemical indicators 

such as pH and DO (Baedecker and others, 1993); and 

inorganic solutes such as calcium (Bennett and others, 

1993). The distributions of DO, VDOC, and Fe2+ in 

1987 and 1992 are shown in fig. 2. Oxygenated 

ground water moved upward in the middle of the 

anoxic zone in 1992. Concentrations of DO ranged 

from 0.05 to 2.4 mg/L in a 4-m-thick zone below the 

top of the saturated zone that extended 70 m downgradient from the anoxic zone. These concentrations 

were less than background DO concentrations of 8 

mg/L. The distributions of VDOC were similar in 

1987 and 1992 as shown in fig. 2, but the zone having 

VDOC concentrations greater than 10 mg/L has 

become smaller, the plume has become less contaminated in the middle, and it is sinking at the leading 

edge. A major change from 1987 to 1992 was 

observed for the leading edge of the 10 mg/L contour 

of the Fe plume which has moved downgradient at a 

rate of about 6 m/yr.

The change over 8 years (1984-1992) in the 

chemical composition of the plume near the oil body is 

shown in fig. 3 for Mn2+, Fe2+, CH4, and the 5 I3C of 

the total inorganic carbon. The data are for a sampling 

point in the plume at the downgradient edge of the oil 

(location 533 on fig. 1), and the screened interval is 

about 1 m thick at the water table. These results indicate that Mn(IV) reduction preceded Fe(III) reduction 

and methanogensis (Baedecker and others, 1993). 

The change in measurements of the total inorganic 

carbon toward heavier numbers over time is due to the

615
Dissolved 

Oxygen

Meters 

425

420

425

420

Meters 

425

. A .

water table

A1 I

1987

- B 1992

Volatile Organic 

Carbon

Ferrous Iron

EXPLANATION

Screened interval 

of sampling sites

) Line of equal con10 centraDon, in milli- 

( grams per liter

Altitude is relative 

to sea level

Meters

-100 

Datum is Sea Level

50 100 150

Vertical Scale is Exaggerated

Distance from Center of Oil Body, in Meters

Fiqure 2. Hydrologic sections (A-A1 in fig. 1) of the contaminant plume showing the anoxic zone, and distribution of volatile 

dissolved organic carbon and ferrous iron in 1987 and 1992. The water table is 6 to 10 meters below land surface.

616
<

d

10 c

1 r

Q 0.1 r

DChZ 

LLJ 

O

O 0-01

O

0.001

Ill

I I I I I I I I I

1984 1986 1988 1990 1992

Figure 3. Concentrations of dissolved ferrous iron, manganese, and methane, and 6 13C measurements for total inorganic 

carbon for the years 1984-92 at location 533 (see fig. 1). Modified from Baedecker and others, 1933.

formation of methane. Concentrations of Mn 

increased to 0.12 mM and then decreased to 0.01 mM, 

whereas Fe concentrations increased by a factor of 30 

to 0.92 mM, and concentrations of CH4 increased 

from the detection limit (0.006 mM) to 1 mM (fig. 3). 

Concentrations of both Fe and CH4 have decreased 

slightly in recent years. The data indicate that Mn(IV) 

reduction has become a less important reaction over 

time near the oil body. The data suggest that both 

Fe(III) reduction and methanogensis continue to be 

major reactions in the anoxic plume.

Concentrations of total benzene and alkylbenzenes in ground water were 10.4 to 18.4 mg/L, respectively, at three sampling locations within 1 meter of 

the oil body (1990 and 1992 data). Concentrations of 

hydrocarbons in the ground water varied over time and 

spatially near the oil body, probably because small oil 

stringers were associated with the sediment near the 

source. Oil stringers on a scale of millimeter to a few 

centimeters were observed in sediment cores obtained 

downgradient from the oil body. Also, ground water 

pumped near the oil body had an oily sheen even 

where a separate fluid phase was not encountered.

At location 518 (fig. 1), 56 m downgradient 

from the middle of the oil body, concentrations of benzene and alkylbenzenes were more consistent over a 

5-year period (fig. 4) than they were at locations close

to the oil body. Benzene was the major hydrocarbon 

present, and its concentration decreased over time, 

whereas concentrations of the other hydrocarbons 

were low but increased over time. Concentrations of 

toluene were equal to or less than 0.007 mg/L and 

concentrations of the xylenes, the C3-benzenes and the 

C4-benzenes were equal to or less than 0.35 mg/L. At 

locations beyond 137 m the oil body, concentrations of 

hydrocarbons were less than Federal drinking water 

standards.

DISCUSSION AND CONCLUSIONS

Although the geochemical processes have 

changed over time, the plume has not migrated as far 

as predicted considering the ground water flow velocities and sorption constants for these compounds 

(Baedecker and others, 1993). A factor that may 

affect migration of contaminanted ground water is the 

presence of discontinuous silt layers (Franzi, 1988; 

Baehr and Hult, 1991) that have low hydraulic conductivity. Thus, the flow velocities in these zones may 

be low and retard the movement of fluids. However, 

the primary reason that the anoxic plume has not farther and that only trace concentrations of hydrocarbons are found in ground water farther than 137 m

617
DC 

111

D

cc 111

Q. 

O)

DC 

CD

<

CC

111 

o

o o

1987

1990

1992

Bz = Benzene

Tol = Toluene

Xyl = o-, m-, p-Xylene

C3-Bz = Alkylbenzenes with 9 carbons

C4-Bz = Alkylbenzenes with 10 carbons

Tol Xyl CS-Bz C4-Bz

Figure 4. Concentrations of benzene and alkylbenzenes at location 518 (see fig. 1) for the years 1987,1990, and 1992.

downgradient from the plume is that the hydrocarbons 

have biodegraded under oxic and anoxic conditions. 

migratedThe rate of removal of organic contaminants 

by natural biodegradative processes and the factors 

that affect these rates are important considerations in 

making decisions concerning cleanup of contaminated 

ground water. Biodegradation of petroleum-derived 

hydrocarbons in aerobic and sub-oxic environments is 

generally considered a more efficient attenuation 

mechanism than is biodegradation in anoxic environments. However, biodegradation in anaerobic environments also may remove significant amounts of 

hydrocarbons from ground water. Estimated half-lives 

of benzene, toluene, and the xylenes in methanogenic 

field sites were 0.5 to 3.8 years at five sites (Barker 

and Wilson, 1992). Natural processes may account for 

the removal of significant quantities of petroleum 

hydrocarbons in the subsurface. Additional work 

needs to be done to determine ground-water flow 

velocities and rates of intrinsic degradation on small 

scales in contaminated aquifers to evaluate natural 

processes as part of long-term remedial action programs.

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Toxic Substances Hydrology Program Proceedings of 

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Baedecker, M.J. and Cozzarelli, I.M., 1992, The determination and fate of unstable constituents in contaminated 

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Groundwater Quality and Analysis at Hazardous Waste 

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618
Baedecker, M.J., Cozzarelli, I.M., Evans, J.R., and Hearn, 

P.P., 1992, Authigenic mineral formation in aquifers 

rich in organic material, in Kharaka, Y.K., and Maest, 

A.S., eds., Proceeding of the 7th International Symposium on Water-Rock Interaction WRI-7, Park City, 

Utah, July 9-23, 1992: Balkema, Rotterdam, p. 257- 

261.

Baedecker, M.J., Cozzarelli, I.M., Eganhouse, R.P., Siegel, 

D.I., and Bennett, P.C., 1993, Crude oil in a shallow 

sand and gravel aquifer, III Biogeochemical reactions and mass balance modeling in anoxic groundwater: Applied Geochemistry, v. 8, p. 569-586.

Baedecker, M.J., Siegel. D.I., Bennett, P.C., and Cozzarelli, 

I.M., 1989, The fate and effects of crude oil in a shallow aquifer, I The distribution of chemical species 

and geochemical facies, in Mallard, G.E., and Ragone, 

S.E., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting. Phoenix, Arizona, September 26-30, 1988: 

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Water Resources Research, v. 27, no. 18, p. 2605-2617.

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attenuation of aromatic hydrocarbons under anaerobic 

conditions, Subsurface Restoration Conference 

Proceedings of the third international conference on 

ground water quality research, Dallas, Texas, June 21- 

24, 1992: p. 57-58.

Bennett, P.C., 1991, Quartz dissolution in organic-rich

aqueous systems: Geochimica et Cosmochimica Acta, 

v. 55, p. 1781-1797.

Bennett, P.C., and Hiebert, F.K., 1996, Microbial control of 

silicate weathering in a petroleum contaminated aquifer, in Morganwalp, D.W., and Aronson, D.A., eds., 

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Program Proceedings of the technical meeting, Colorado Springs, Colorado, September 20-24, 1993: U.S. 

Geological Survey Water-Resources Investigations 

Report 94-4015.

Bennett, P.C., Siegel, D.I., Baedecker, M.J. and Hult, M.F., 

1993, Crude oil in a shallow sand and gravel aquifer, 

I Hydrogeology and inorganic geochemistry: 

Applied Geochemistry, v. 8, p. 529-549.

Chang, F.H., Noben, N.N., Brand, Danny, and Hult, M.F., 

1988. Microbial degradation of crude oil and some 

model hydrocarbons, in Ragone, S.E., ed., U.S. Geological Survey Toxics Substances Hydrology Program 

on Toxic Waste Ground-Water Contamination 

Proceedings of the second technical meeting, Cape 

Cod, Massachusetts, October 21-25, 1985: U.S. Geological Survey Open-File Report 86-481, p. C33-C42.

Cozzarelli, I.M., Baedecker, M.J., Aiken, G.R., and Phinney, Curtis, 1996, Small-scale heterogeneities of biogeochemical reactions in a crude-oil contaminated 

aquifer, Bemidji, Minnesota, in Morganwalp, D.W., 

and Aronson, D.A., eds., U.S. Geological Survey

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of the technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

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1990, Transformation of monoaromatic hydrocarbons 

to organic acids in anoxic groundwater environment: 

Environmental Geology and Water Sciences, v. 16, no. 

2, p. 135-141.

Dillard, L.A., Essaid, H.I., and Herkelrath, W.N., 1996, 

Multiphase-flow modeling at the Bemidji, Minnesota 

crude-oil spill site using geostatistical simulation of 

soil hydraulic properties, in Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water- Resources Investigations Report 94-4015.

Eganhouse, R.P., Baedecker, M.J., Cozzarelli, I.M., Aiken, 

G.R., Thorn, K.A., and Dorsey, T.F., 1993, Crude oil in 

a shallow sand and gravel aquifer, II Organic 

geochemistry: Applied Geochemistry, v. 8, p. 551- 

567.

Essaid, H.I., Baedecker, M.J., and Cozzarelli, I.M., 1996, 

Use of simulation to study field-scale solute transport 

and biodegradation at the Bemidji, Minnesota crudeoil site, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water- Resources 

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Essaid, H.I., Herkelrath, W.N., and Hess, K.M., 1993, Simulation of fluid distributions observed at a crude-oil 

spill site incorporating hysteresis, oil entrapment, and 

spatial variability of hydraulic properties: Water 

Resources Research, v. 29, no. 6, p. 1753-1770.

Franzi, D.A., 1988, Sedimentary and post-depositional processes related to aquifer properties at the Bemidji 

Research Site, North-Central Minnesota, in Franks, 

B.J., ed., U.S. Geological Survey Program on Toxic 

Waste Ground-Water Contamination Proceedings 

of the third technical meeting, Pensacola, Florida, 

March 23-27, 1987: U.S. Geological Survey Water 

Resources Investigations Report 88-4220, p C7-C8.

Grbic-Galic, D., and Vogel, T.M., 1987, Transformation of 

toluene and benzene by mixed methanogenic cultures: 

Applied Environmental Microbiology, v. 53, p. 254- 

260.

Hess, K.M., Herkelrath, W.N., and Essaid, H.I., 1992,

Determination of subsurface fluid contents at a crudeoil spill site: Journal of Contaminant Hydrology, v. 10, 

p. 75-96.

619
Hiebert, F.K., and Bennett, P.C., 1992, Microbial control of 

silicate weathering in organic-rich ground water: 

Science, v. 258, p. 278-281.

Hult, M.F., 1984, Ground-water contamination by crude oil 

at the Bemidji, Minnesota Research Site an introduction, in Hult, M.F., ed., Ground-Water Contamination 

by Crude Oil at the Bemidji, Minnesota Research Site, 

U.S. Geological Survey Toxic Waste Ground-Water 

Contamination Study Proceedings of the technical 

meeting, Tucson, Arizona, March 20-22, 1984: U.S. 

Geological Survey Water-Resources Investigations 

Report 84-4188, chap. A, p. 1-15.

Hult, M.F., and Grabbe, R.R., 1988, Distribution of gases 

and hydrocarbons vapors in the unsaturated zone, in 

Ragone, S.E., ed., U.S. Geological Survey Program on 

Toxic Waste Ground-Water Contamination-Proceedings of the technical meeting, Cape Cod, Massachusetts, October 21-25, 1985: U.S. Geological Survey 

Open-File Report 86-481, p. C21-C26.

Landon, M.K., and Hult, M.F., 1996, Source mass balance 

calculated from changes in composition of spilled 

crude oil in the subsurface near Bemidji, Minnesota, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, Colorado 

Springs, Colorado, September 20-24, 1993: U.S. 

Geological Survey Water- Resources Investigations 

Report 94-4015.

Lovley, D.R., Baedecker, M.J., Lonergan, D.J., Cozzarelli, 

I.M., Phillips, E.J.P., and Siegel, D.I., 1989, Oxidation 

of aromatic contaminants coupled to microbial iron 

reduction: Nature, v. 339, p. 297-299.

Mallard, G.E., and Aronson, D.A., 1991, U.S. Geological 

Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, Monterey, California, March 11-15, 1991: U. S. Geological Survey 

Water Resources Investigations Report 91-4034, 730 p.

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subsurface heterogeneities with ground penetrating 

radar at Bemidji, Minnesota, in Mallard, G.E., and 

Aronson, D.A., eds,. U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water Resources Investigations Report 91-4034, p. 650-653.

Pfannkuch, H.O., Nourse, S.M., and Hult, M.F., 1988, Mass 

transfer at the alkane-water interface in laboratory columns of porous media, in Ragone, S.E., ed., U.S. Geological Survey Program on Toxic Waste GroundWater Contamination Proceedings of the technical 

meeting, Cape Cod, Massachusetts, October 21-25, 

1985: U.S. Geological Survey Open-File Report 

86-481, p. C29-C32.

Piatt, J.J., Eisenreich, S.J., Backhus, D.A., and Capel, P.D., 

1996, Sorption of polycyclic aromatic hydrocarbons to 

aquifer solids with low organic-carbon from Bemidji, 

Minnesota, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Revesz, K.M., Coplen, T.B., and Baedecker, M.J., and Hult, 

M.F., 1996, Methane production and consumption 

monitored by stable hydrogen and carbon isotope 

ratios at Bemidji, Minnesota, in Morganwalp, D.W., 

and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings 

of the technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Siegel, D.I., Baedecker, M.J., and Bennett, P.C., 1986, The 

effects of petroleum degradation on inorganic waterrock interactions, in Proceedings of the Fifth International Symposium on Water-Rock Interaction, Reykjavik, Iceland, August 8-12, 1986: International 

Association of Geochemistry and Cosmochemistry, 

p. 524-527.

Thorn, K.A., and Aiken, G.R., 1989, Characterization of 

nonvolatile organic acids resulting from the biodegradation of crude oil by nuclear magnetic resonance 

spectrometry, in Mallard, G.E., and Ragone, S.E., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, Phoenix, Arizona, September 26-30, 1988: U.S. Geological 

Survey Water-Resources Investigations Report 

88-4220, p. 41-51.

620
Use of Simulation to Study Field-Scale Solute Transport 

and Biodegradation at the Bemidji, Minnesota,

Crude-Oil Spill Site

SyH.I. Essaid1 , M.J. Baedecker2, anc/I.M. Cozzarelli2

Abstract

A two-dimensional, multispecies solutetransport model that includes aerobic and 

anaerobic degradation processes was developed 

and applied to the Bemidji, Minnesota, crude-oil 

spill site. The model was used as a tool to study 

the field-scale solute-transport and degradation 

processes. The simulations included the transport 

of degradable and refractory dissolved organic 

carbon, dissolved oxygen, and methane, as well as 

aerobic and anaerobic (methanogenic) 

biodegradation. Despite the considerable 

uncertainty in the model parameter estimates, 

results of simulations reproduced the general 

features of the observed ground-water plume. In 

addition to the kinetics of the biodegradation 

processes, important factors that affect the 

distribution of the solutes in the field are the 

recharge influx and the degree of dispersion and 

mixing in the ground water system. Recharge 

influx causes flow lines in the aquifer to be 

deflected downward from the water table causing 

downward movement of the plume. 

Biodegradation results in a narrow plume with 

sharp gradients at the margins. Spatial variability 

of hydraulic conductivity increases the amount of 

carbon degraded by aerobic and anaerobic 

processes.

'U.S. Geological Survey, Menlo Park, CA 

2U.S. Geological Survey, Reston, VA

INTRODUCTION

Numerous researchers working at the Bemidji, 

Minn. crude-oil spill site (fig. 1) have documented 

evidence for microbial degradation of petroleum 

hydrocarbons in the field. Chang and others (1991) 

studied the kinetics of aerobic crude-oil 

biodegradation in laboratory microcosms using 

materials from the site. Methanogenesis has been 

observed in the unsaturated (Hult and others, 1991) 

and saturated zones (Bennett and others, 1993; 

Eganhouse and others, 1993; Baedecker and others, 

1993). Cozzarelli and others (1990) identified organic 

acids produced by oxidation of monoaromatic 

hydrocarbons by anaerobic degradation processes, and 

Lovley and others (1989) showed that Fe(III)-reducing 

organisms can oxidize an aromatic hydrocarbon 

anaerobically. Hiebert and Bennett (1992) used in-situ 

microcosms to study the accelerated chemical 

weathering of silicates by surface-adhering bacteria.

As a result of these studies, five geochemical 

zones in ground water (fig. 2) have been identified at 

the crude-oil spill site (Baedecker and others, 1989; 

Baedecker and others, 1993; and Bennett and others, 

1993). Zone 1 consists of oxygenated uncontaminated 

native ground water. Zone 2, which is below the area 

where the land surface was sprayed by oil following 

the pipeline rupture, is characterized by reduced 

oxygen concentrations and the presence of refractory 

high molecular-weight hydrocarbons. Zone 3, beneath 

and immediately down gradient from the separatephase oil lens, consists of an anoxic plume of groundwater that contains high concentrations of 

hydrocarbons and methane. Zone 4 is the transition 

zone from anoxic conditions to fully oxygenated 

conditions, and concentrations of hydrocarbons decrease 

rapidly as a result of aerobic degradation processes.

621
95'5'40" 95"5'4"

47'34'32"

47-34'14'

Well Location

Area of pipeline 

break

Trace of simulated 

section

Railroad 

Direction of track

ground-water flow

SCALE 

50 100 150 Mctcn

0 400 Feet 

Figure 1. Map of the Bemidji research site, Minnesota, showing the trace of the simulated vertical section.

Zone 5 consists of oxygenated water downgradient of 

the contamination plume that contains slightly 

elevated concentrations of dissolved constituents. 

Long-term monitoring of the plume since 1984 has 

shown that, near the water table, the concentration of 

total dissolved organic carbon (TDOC) and dissolved 

oxygen (DO) downgradient from the oil lens has 

remained relatively stable with time. In the anoxic 

zone (Zone 3) concentrations of reduced manganese 

(Mn) and iron (Fe), and methane have increased with 

time, indicating a sequence of Mn reduction followed 

by Fe reduction and methanogenesis.

In this study, a two-dimensional, multispecies 

solute-transport model that incorporates 

biodegradation was developed and applied to the 

ground-water system at the Bemidji, Minn. crude-oil 

spill site. The model was used as a tool to study the 

subsurface solute-transport and degradation processes. 

The goal of this study is to use the model to quantify 

the field-scale degradation processes and to identify 

the important factors affecting the distribution of 

solute species in the field. This report describes the 

preliminary results of the solute-transport and 

biodegradation modeling at the Bemidji, Minn. site.

622
THE NUMERICAL MODEL

A numerical model was developed to simulate 

the aerobic and anaerobic degradation processes that 

have been observed in the contaminated ground-water 

plume at the spill site. The model was used as a tool to 

study the physical, chemical, and microbial processes 

affecting the transport of dissolved species. Steep 

concentration gradients observed in the field data 

required that the model be capable of handling sharp 

fronts without introducing numerical dispersion. For 

this reason, the U.S. Geological Survey's Method of 

Characteristics transport model (Konikow and 

Bredehoeft, 1978, Goode and Konikow, 1989) was 

expanded to handle multiple solutes and to include 

biodegradation terms. The approach of Kindred and 

Celia (1989) was used to represent the biodegradation 

terms in the transport equation as outlined below.

The two-dimensional transport equation solved 

in the model for each solute species includes linear 

sorption and biodegradation assuming a uniform 

porosity and fluid density as follows:

dC. .

n _ / _ j. O

idt bdx. -l^ Jkd -V.-^-C. + Jdx. i 

J

of zero order, and when C; is small, v" » 

first order.

V"C

is of

As outlined by Kindred and Celia (1989) the 

biodegradation terms in equation 1 can be represented 

as follows. If degradation is substrate limited, the first 

order degradation term is given by

(3)

where Xn is the biomass concentration of microbial 

population w, Ih and Ic are the biomass and competitive

Q inhibition factors, respectively, given by L = 1 + -^

ki

where Qj is the concentration of the inhibiting substance /, and kj is the inhibition constant for that substance. The zero-order degradation term is given by:

(4)

(1)

where C, is the concentration of the ith solute, Rf is the 

retardation factor for the /th solute, t is time, b is the 

width of the vertical cross-section through the aquifer, 

Dfi is the dispersion tensor, jty and x^ are the spatial 

coordinates, Vj is the average linear velocity of the 

fluid, Wis the source fluid flux, £ is the porosity, CV is 

the concentration of the ith solute in the source fluid, 

K(l is the first-order biodegradation reaction term, and 

KL ' is the zero-order biodegradation reaction term.

Michaelis-Menton uptake kinetics is used to 

represent the degradation process, whereby

where Y", is the cell-yield coefficient of the limiting 

substrate /, Y^. is the cell-yield coefficient of substrate 

/, KI is the total uptake of the limiting substrate, Jpi is 

the i-p reaction ratio, Kp is the total amount of parent 

compound p metabolized, and k"n is the death rate or 

maintenance constant. The terms on the right hand 

side of equation 4 represent, respectively, the zeroorder uptake of substrate /, the production of substrate 

/ by metabolism of a parent compound, and release of 

substrate / because of cell death.

The rate of biomass growth is given by:

v, =

A-+C,

(2)

where v" is the specific uptake rate of solute / by 

microbial population n, V"m is the asymptotic maximum specific uptake rate, and tCh is the half-saturation constant. The total uptake of solute / is simply 

equal to the specific uptake rate multiplied by the biomass of population n. As can be observed by examining equation 2, when C, is large, v"» V"n is

The model solves the ground-water-flow 

equation and then uses the velocities obtained from the 

flow solution to compute the movement of particles 

through the system. A separate set of particles is used 

for each solute in order to allow for different 

retardation factors. The degradation terms Kj and Kc 

are calculated explicitly at the end of each time step. 

The amount of biomass is then calculated using the 

integral of equation 5.

623
RECHARGE ZONES

c/>

EXPLANATION 

X mldpolnl of woll

screen 

Datum is sea level Direction of Ground-Water Flow

416

100 200 

DISTANCE, IN METERS

300 400

Figure 2. The simulated section showing ground-water zones 1 through 5 of the contaminated ground-water plume and 

simulated recharge zones A through E. (Modified from Baedecker and others, 1989. fig. A-6.).

For one-dimensional test cases, the results of 

this model compared favorably to results of the model 

of Kindred and Celia (1989).

SIMULATION OF THE OBSERVED PLUME

A two-dimensional cross-section was designed 

to simulate the observed contaminated ground water 

plume at the Bemidji, Minn. site. The section is 

approximately parallel to the direction of groundwater flow. Ground-water samples from numerous 

wells along this section have been analyzed over time 

(Baedecker and others, 1993; Bennett and others, 

1993; Eganhouse and others, 1993). The locations of 

the midpoints of the screened intervals of the wells 

that have been sampled are shown in figure 2. Water 

table wells were installed with a 1.52-m screen length 

and deeper sampling wells with 0.15 or 0.61-m screen 

lengths.

A vertical cross-section of unit width 

approximating the sampling transect was simulated 

using the transport model for the period from the time 

of the spill in 1979 until September 1990. Steady-state 

flow, no sorption, and isothermal conditions were 

assumed; the decrease in aquifer thickness caused by 

the change in water-table elevation was neglected. 

Four solutes and two microbial populations were 

modeled. Total dissolved organic carbon (TDOC) was 

split into two fractions degradable dissolved organic 

carbon (DDOC) and refractory dissolved organic 

carbon (RDOC). The remaining two solutes modeled 

were dissolved oxygen (DO) and methane. To 

represent the aerobic and anaerobic degradation

processes, aerobic and methanogenic populations of 

bacteria were included in the simulations. Iron and 

manganese reductions were not included because of 

the complexity of incorporating the rock-water 

interactions of dissolution and precipitation in the 

transport model.

For simulation purposes, the system was 

represented by an initially clean aquifer with 

background dissolved organic carbon concentrations 

and fully oxygenated water. Following the oil spill, 

DDOC and RDOC dissolved and entered the aquifer 

with recharge water. In the spray zone, we assumed 

that the major fraction of carbon entering the aquifer 

was refractory, whereas in the immediate zone of the 

oil body, we assumed most carbon was degradable. 

Hult and others (1991) have shown that 

biodegradation above the water table depletes the 

oxygen resulting in an anoxic unsaturated zone in the 

vicinity of the oil. Therefore, it was assumed that there 

was no oxygen in the recharge water near the oil body, 

and that oxygen was partially depleted in the recharge 

water in the spray zone. Competitive inhibition was 

used to represent the suppression of methanogenesis 

by oxygen. In this manner, as oxygen in the ground 

water is consumed and an anoxic zone develops, the 

methanogens begin to flourish resulting in increased 

methane production.

The simulated section is shown in figure 2. The 

oil present in the pore space within the oil body makes 

it difficult for water to flow through this zone. The 

magnitude of reduction of water flow is a complex 

function of the oil distribution. As a first approximation of this effect, the hydraulic conductivity and

624
recharge rate were reduced to 25 percent of the aquifer 

values in the zone of the oil body. Constant heads of 

424.0 m and 422.6 m were specified at the upgradient 

and downgradient lateral boundaries, respectively. The 

bottom boundary condition was no flow, and a 

constant rate of recharge was specified across the top 

of the aquifer. The estimated values of initial 

concentrations and recharge water concentrations for 

each solute are given in table 1. The parameter values 

used in the simulation are given in table 2. There is 

considerable uncertainty in the values of the model 

parameters. Estimates of permeability (Essaid and 

others, 1991; Dillard and others, 1996) and 

measurements of velocity (White, 1991) have been 

made at the site; however, the values range over 

several orders of magnitude. No measurements or 

estimates are available for the transport and 

biodegradation parameters under the natural field 

conditions. Therefore, reasonable estimates of these 

values were used in the simulations.

Table 1. Initial and recharge water concentration in 

milligrams per liter

[Recharge /ones A through Fi are shown in figure 2; DDOC, degradablc 

dissolved organic carbon; RDOC, refractory dissolved organic carbon; DO, 

dissolved oxygen]

Recharge Zone

Solute lni

DDOC

RDOC

DO

Methane

itial Cone

0.0

2.0

9.0

.0

en- A

0.0

2.0

9.0

.0

B

10.0

20.0

3.0

.0

c

100.0

30.0

.0

.0

D

0.0

2.0

.0

.0

E

0.0

2.0

9.0

.0

RESULTS AND DISCUSSION

Observed and simulated profiles of TDOC, DO, 

and methane near the water table are plotted in 

figure 3. Measurements in wells near the water table 

and immediately below the oil body were used for 

comparison. The screen length of most wells is 1.52 

m; approximately half of the screen is below the water 

table. To approximate the measured values, the 

simulated values near the water table were calculated 

by averaging the concentrations of the first 0.8 m 

below the top of the simulated section.

The concentration profile of TDOC is shown in 

fig. 3a. The observed points show a concentration 

distribution at the water table that is relatively stable 

with time. The simulation has captured this feature as 

can be seen by the similarity between the simulated 

1986 and 1990 concentration profiles. There is an 

increase in TDOC concentration in the upgradient

spray zone, followed by a rapid increase in TDOC 

concentration in the zone of the oil body.

Table 2. Parameters used in the transport model simulation

[Vm, maximum specific uptake rate; Kh, half-saturation constant; Y, cell 

yield coefficient; m, meters; m/s, meters per second; m/d, meters per day; 

m/yr, meters per year; /d, per day; mg/L, milligrams per liter; L/mg, liters 

per milligram]

Flow and Transport Parameters: 

X-direction grid spacing 

Y-direction grid spacing 

Porosity

Hydraulic conductivity. 

Average linear velocity 

Water table gradient 

Recharge

Longitudinal dispersivity 

Transverse dispersivity

Biodegradation Parameters: 

Aerobic degradation of DOC: 

Vm

Kh 

Y

Aerobic consumption of DO: 

Vm

Kh 

Y

2m

0.1 m

0.38

I x I0-4 m/s

0.07 m/d

0.0035

O.I26 m/yr

0.1 m

0.001 m

1.1 /d 

0.09 mg/L 

0.21 L/mg

0.99 /d 

O.I4 mg/L 

0.08 L/mg

Anaerobic degradation of DOC (mcthanogencsis): 

Vm 0.25 /d 

K h 0.06 mg/L 

Y 0.06 L/mg 

Biomass inhibition constant (//,) 1.0 mg/L 

Methanogenesis inhibition 0.005 mg/L

constant (/c .)

Initial aerobic biomass (X1 ) 0.001 mg/L 

Initial anaerobic biomass (X ) 0.001 mg/L 

Aerobic maintenance constant (* /) 0.005 /d 

Anaerobic maintenance constant (kn ~) 0.011 /d

Downgradient from the oil body, the TDOC concentration decreases gradually to the background 

concentration. This decrease is a result of microbial 

and physical processes. First, there is the anaerobic 

degradation of DDOC within the anoxic zone near the 

oil body and aerobic degradation of DDOC as it 

moves away from the oil body and encounters oxygenated recharge water (fig. 3b). However, even the concentration of refractory DOC (fig. 3c) decreases to 

background levels and achieves a steady-state profile 

near the water table. This is a result of the physical 

processes of displacement and mixing of flowing 

ground water with recharge water.

625
8

o

80

60

40

20

TDOC

100 200 300

DISTANCE, IN METERS

400

I 80

_i

DC 

01 

O.

CO

60

40

20

UJ

o

§ o

o

DDOC

100 200 300 

DISTANCE. IN METERS

400

80

60

40

20

o

1

DO METHANE

10

co 8

< 

cc

O o

~i D

-. 4

o

100 200 300 

DISTANCE. IN METERS

400

o

30

20

Q 10

I

100 200 300 

DISTANCE. IN METERS

400

E

1986 Sim

1990 Sim

1986 Obs

Figure 3. Graphs showing the observed and simulated concentrations near the water table for (a) total dissolved organic carbon 

dissolved organic carbon (DDOC), (c) refractory dissolved organic carbon (RDOC), (d) dissolved oxygen (DO), and (e) methane. Th

the simulated section in meters.
The dissolved oxygen concentration profile 

(fig. 3d) decreases in the spray zone because of the 

decrease in DO concentration in recharge water 

(assumed to occur because of degradation of 

hydrocarbons in the unsaturated zone) and the 

consumption of oxygen by degradation in the ground 

water. Near the oil body and immediately 

downgradient from it, the DO of the recharge water is 

assumed to have been completely consumed in the 

unsaturated zone. An anoxic zone develops in this 

area. However, DO begins to increase as oxygenated 

recharge water enters the system farther downgradient.

Methane is produced in the anoxic zone that 

develops in the immediate vicinity and downgradient 

of the oil body. The methane peak is displaced 

downgradient from the center of the oil body because 

of the input of methane-free ground water from the 

upgradient area. The decline in methane concentration 

at a distance of 230 m (fig. 3e) is a result of the 

upwelling of oxygenated water caused by the upward 

bending of flow lines around the less water-permeable 

oil body as the oil body decreases in thickness. The 

simulated profiles show a marked increase in methane 

production from 1986 to 1990 as population of 

methanogens increases. This increase in 

methanogenesis results in a slight decrease in DDOC 

concentrations from 1986 to 1990 (fig. 3b).

In order to examine the effect of degradation on 

DDOC in the aquifer, two-dimensional distributions of 

DDOC for three different simulations are shown in 

figure 4. In the first simulation, there is no degradation 

(fig. 4a); in the second simulation (described above) 

degradation occurs (fig. 4b); and in the third 

simulation degradation occurs and the hydraulic 

conductivity distribution is heterogeneous (fig. 4c). 

The distribution of DDOC for the case with no 

degradation (fig.4a) reflects the physical processes of 

dispersion, diversion of flow around the oil body 

because of its reduced water hydraulic conductivity, 

and the depression of the plume beneath the water 

table because of the deflection of flow lines by 

incoming recharge water.

In the second simulation, the anaerobic and 

aerobic degradation processes result in a contaminant 

plume that is narrower than the plume in the first 

simulation and whose concentration gradients are 

comparatively sharp at the edges (fig. 4b). In this 

simulation, 46 percent of the total DDOC mass 

entering the aquifer is degraded 14 percent by 

anaerobic degradation and 32 percent by aerobic

degradation. Previous work has shown that the 

hydraulic properties of the aquifer are spatially 

variable (Essaid and others, 1991; Essaid and others, 

1993; Dillard and others, 1996). To make the 

simulation more realistic, a heterogeneous hydraulic 

conductivity distribution was created using the 

methods of Dillard and others (1996) and was used in 

the transport model. Because of the complex flow 

field, an irregularly shaped plume develops (fig. 4c). 

The average linear velocities of ground water (Darcy 

velocity divided by porosity) in this simulation range 

from 3.0 x 10' 10 to 1.6 m/d, with a mean of 0.13 m/d. 

The variability in flow paths and flow velocities results 

in increased mixing and dispersion of ground water. 

This, in turn, results in increased biodegradation. Of 

the total DDOC mass entering the aquifer, 60 percent 

is degraded 21 percent by anaerobic degradation and 

39 percent by aerobic degradation.

The simulations presented in this paper represent 

a highly simplified representation of the true field 

conditions and neglect Fe and Mn reduction. Also, the 

parameters used in the simulations are highly uncertain. 

Nevertheless, the results do reproduce the general 

features of the observed contaminated ground water 

plume. In addition to the kinetics of the biodegradation 

processes, important factors that affect the magnitude of 

degradation and the distribution of the solutes in the 

field are the recharge influx and the degree of dispersion 

and mixing in the ground water system.

SUMMARY

A two-dimensional, multispecies solutetransport model that includes aerobic and anaerobic 

degradation processes was developed and applied to 

the Bemidji, Minn., crude-oil-spill site. The model 

was used as a tool to study field-scale solute transport 

and degradation processes. Simulation results indicate 

that, in addition to the kinetics of the biodegradation 

processes, important factors that affect the distribution 

of the solutes in the field are the recharge influx and 

the degree of dispersion and mixing in the groundwater system. Recharge influx deflects flow lines in 

the aquifer downward from the water table causing 

downward movement of the plume. Biodegradation 

produces a narrow plume with sharp gradients at the 

margins. Spatial variability of hydraulic conductivity 

increases the amount of carbon degraded by aerobic 

and anaerobic processes.

627
RECHARGE ZONES

< 416

c/D 424

DC 

LJJ

LJJ

LJJ 

Q

ID

420

< 416

50

EXPLANATION 

Y. midpoint of well screen 

zone boundary 

1 __ isoconcentration line 

Datum is sea level

I

100 200 

DISTANCE, IN METERS

Direction of Ground-Wate

300 

Figure 4. Sections showing the simulated distribution of degradable dissolved organic carbon with (a) no biodegradation, (b) biod

and a heterogeneous distribution of hydraulic conductivity.
REFERENCES

Baedecker, M.J., Cozzarelli, I.M., Siegel, D.I., Bennett, 

P.C., and Eganhouse, R.P., 1993, Crude oil in a shallow 

sand and gravel aquifer, 3 Biogeochemical Reactions 

and mass balance modeling in anoxic ground water: 

Applied Geochemistry, v. 8, p. 569-586.

Baedecker, M.J., Siegel, D.I., Bennett, P., and Cozzarelli, 

I.M., 1989, The fate and effects of crude oil in a 

shallow aquifer, I The distribution of chemical 

species and geochemical facies, in Mallard, G.E., and 

Ragone, S.E., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

Technical Meeting, Phoenix, Arizona, September 26- 

30, 1988, Water-Resources Investigations Report 88- 

4220, p. 13-20.

Bennett, P.C., Siegel, D.I., Baedecker, M.J., and Hult, M.F., 

1993, Crude oil in a shallow sand and gravel aquifer,

I Hydrogeology and inorganic geochemistry: 

Applied Geochemistry, v. 8, p. 529-549.

Chang, F.H., Wang, H., Denzin, B., and Buller, J., 1991, 

Kinetics of crude-oil biodegradation by bacteria 

indigenous to sediment and ground water, in Mallard, 

G.E., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the Technical Meeting, Monterey, California, March

II-15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 646-649. 

Cozzarelli, I.M., Eganhouse, R.P., and Baedecker, M.J., 

1990, Transformation of monoaromatic hydrocarbons 

to organic acids in anoxic groundwater environment: 

Environmental Geology and Water Science, v. 16, 

p. 135-141.

Dillard, L.A., Essaid, H.I., and Herkelrath, W.N., 1996, 

Multiphase flow modeling at the Bemidji, Minnesota, 

crude-oil spill site using geostatistical simulation of 

soil hydraulic properties, in Morganwalp, D.W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Eganhouse, R.P., Baedecker, M.J., Cozzarelli, I.M., Aiken, 

G.R., Thorn, K.A., and Dorsey, T.F., 1993, Crude oil in 

a shallow sand and gravel aquifer, 2 Organic 

geochemistry: Applied Geochemistry, v. 8, p. 551-567.

Essaid, H.I., Herkelrath, W.N., Hess, K.M., 1991, Air, oil, 

and water distributions at a crude-oil spill site, 

Bemidji, Minnesota, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 614-620.

Essaid, H.I., Herkelrath, W.N., Hess, K.M., 1993,

Simulation of fluid distributions observed at a crudeoil spill site incorporating hysteresis, oil entrapment, 

and spatial variability of hydraulic properties: Water 

Resources Research, v. 29, no. 6, p. 1753-1770.

Goode, D.J., and Konikow, L.F., 1989, Modification of a 

method-of-characteristics solute-transport model to 

incorporate decay and equilibrium-controlled sorption 

or ion exchange: U.S. Geological Survey WaterResources Investigations Report 89-4030, 65 p.

Hiebert, F.K., and Bennett, PC., 1992, Microbial control of 

silicate weathering in organic-rich ground water: 

Science, v. 258, p. 278-281.

Hult, M.F., Landon, M.K., and Pfannkuch, H.O., 1991, 

Field validation of conceptual models of mobilization 

and transport of volatile petroleum derivatives in the 

unsaturated zone near Bemidji, Minnesota, in Mallard, 

G.E., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the Technical Meeting, Monterey, California, March 

11-15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 621-626.

Kindred, J.S., and Celia, M.A., 1989, Contaminant transport 

and biodegradation, 2 Conceptual model and test 

simulations: Water Resources Research, v. 26, no. 6, 

p. 1149-1160.

Konikow, L.F., and Bredehoeft, J.D., 1978, Computer 

model of two-dimensional solute transport and 

dispersion in ground water: U.S. Geological Survey 

Techniques of Water-Resources Investigations, book 7, 

chap. C2, 90 p.

Lovley, D.R., Baedecker, M.J., Lonergan, D.J., Cozzarelli, 

I.M., Phillips, E.J.P., and Siegel, D.I., 1989, Oxidation 

of aromatic contaminants coupled to microbial iron 

reduction: Nature, v. 339, p. 297-299.

White, D.L., 1991, Point dilution method determination of 

groundwater velocities at the Bemidji research site: 

Unpublished M.A. thesis, University of Texas, Austin, 

Texas, 145 p.

629
630
Source Mass Balance Calculated from Changes in

Composition of Spilled Crude Oil in the

Subsurface near Bemidji, Minnesota

By Matthew K. Landon 1 and Marc F. Hult 1

Abstract

The transport and fate of crude oil that spilled 

from a break in a buried pipeline near Bemidji, 

Minnesota, on August 20, 1979, has been monitored by the U.S. Geological Survey since 1983. 

The oil percolated to the water table in a glacial 

outwash aquifer. The oil has moved about 30 

meters as a separate fluid phase in the aquifer and 

is preferentially losing soluble and volatile compounds through dissolution into ground water and 

vaporization into the unsaturated zone. As part of 

the research effort at the site, the evolution of the 

composition and physical properties of the oil in 

the subsurface was studied. The primary objectives were to (1) quantify changes in the composition and physical properties of the oil at the site, 

(2) compare field results with oil alteration under 

controlled laboratory conditions that simulated the 

field environment, and (3) estimate mass balances 

for the total oil and for individual compounds 

mobilized during natural alteration of the oil 

source.

Analysis of oil collected from monitoring 

wells screened in the oil bodies during 1988 and 

1989 indicated that areal differences in the physical properties of the oil have developed because of 

spatially and temporally variable rates of alteration of the originally uniform oil. The composition of the oil fraction lighter (with a lower carbon 

number) than dodecane (a normal alkane with 12 

carbon atoms) of field and laboratory aged samples of oil was determined using gas chromatography/mass spectroscopy (GC/MS).

U.S. Geological Survey, Mounds View, Minn.

The oil has selectively lost aromatic compounds 

and aliphatic compounds (normal, branched, and 

cyclic alkanes) more volatile than nonane (normal 

alkane with nine carbons). About 50 detectable 

compounds have been depleted in the fraction of 

the oil that was analyzed. Maximum depletions of 

the total mass of C3 (compounds with three carbons), C4, C5, C6, C7, C8, C9, and CIO compounds were 100, 100, 95, 81, 66, 48, 49, and 25 

percent, respectively. Depletion of the total mass 

of aromatic compounds ranged from 30 to 84 percent and exceeded depletion of the total mass of 

aliphatic compounds, which ranged from 0 to 73 

percent. The greater depletion of aromatic compounds may indicate that dissolution of aromatic 

compounds into water initially occurs at greater 

rates than losses of aliphatic compounds by volatilization. Because of the much greater abundance 

of aliphatic compounds in the oil, however, more 

than 70 percent of the mass loss from the oil likely 

occurs as a result of volatilization. Losses of C6 

and C7 aliphatic compounds account for a majority of the total loss of mass from the oil. Total mass 

loss from the oil calculated from changes in oil 

composition agrees within 0 to 3.8 percent 

(median of 1.6 percent) to mass loss calculated 

from changes in specific gravity, refractive index, 

and kinematic viscosity, indicating that any of 

these properties can be used for mass-balance 

calculations. Minimum estimates of total mass 

loss from the oil based on changes in composition 

and physical properties range from 0 to 11 percent 

(average is 3.8 percent) since the spill occurred in 

1979. The results of mass loss calculations provide 

an independent estimate of the spatially variable 

source term for transport models of volatile and 

soluble hydrocarbons emanating from the oil 

source in the subsurface.

631
INTRODUCTION

The composition and physical properties of spilled 

crude oil in a shallow subsurface environment change 

because of vaporization of volatile organic compounds 

into the unsarurated zone, dissolution of soluble 

organic compounds into ground water, and soil moisture, and biodegradation. The potency of a body of 

spilled petroleum as a source of environmental contamination, therefore, can change with time because of 

natural alteration processes.

Research to quantify changes in the composition 

and physical properties of spilled crude oil in the 

shallow subsurface in an undisturbed setting was 

 conducted at a site near Bemidji Minn. (fig. 1). A 

description of the site was given by Hull (1984). This 

study is part of an interdisciplinary, long-term research 

effort at the Bemidji site to improve understanding of 

the mobilization, transport, and fate of petroleum contaminants in the shallow subsurface (Hull, 1991). The 

objectives of this study are to (1) quantify changes in 

composition and physical properties of the oil at the 

site, (2) compare field results with oil alteration under 

controlled laboratory conditions that simulate the field 

environment, and (3) estimate mass balances for the 

total oil and for individual compounds mobilized 

during natural alteration of the oil source. This paper

presents measurements of oil composition and calculations of mass loss from the oil on the basis of changes 

in composition. Measurements of physical properties 

and calculations of mass loss from the oil in the field 

using laboratory relations between changes in physical 

properties and mass loss were presented in Landon and 

Hull (1991).

METHODOLOGY

Oil was collected from 14 wells during July 1988 

and from 17 wells during June 1989. Each well was 

completed with a 5-cm inside-diameter (i.d.) 1.5-mlong stainless-steel screen. The samples were collected 

with a Teflon bailer, placed in airtight glass containers 

to prevent loss of volatile constituents, and stored at 

approximately 5 °C until analysis. The oil samples that 

were collected represented vertically integrated composites of the oil in the vicinity of the well screen over 

the length of the screen. Analysis of oil samples collected 24 hrs after initial sampling indicated that the oil 

collected from the wells had the same physical properties as oil that seeped in from the surrounding aquifer 

material; the oil samples collected from the wells were, 

therefore, representative of oil in the subsurface.

44*

ft!*

309

EXPLANATION

- Well sampled and 

identifier

SCALE

50 100 METERS!

100 200 300 FEET

Pipeline 

break

Approximate 

extent of 

oil bodies

Figure 1 . Location of Bemidji, Minn., research site and areal map of monitoring wells sampled for crude oil in July 1988 

and June 1989.

632
The kinematic viscosity, specific gravity, and 

refractive index of oil samples collected in the field 

were measured at 7.5 °C-the approximate mean 

annual ground-water temperature at the Bemidji site 

(Landon and Hult, 1991). Split samples of the oil collected in the field were allowed to volatilize inside a 

chamber at 7.5 °C to simulate oil alteration by vaporization under subsurface conditions and to determine 

relations among mass loss and changes in physical 

properties (Landon and Hult, 1991).

Gas chromatography/mass spectroscopy (GC/MS) 

was used to analyze the composition of the oil fraction 

lighter (with a lower carbon number) than dodecane 

(a normal alkane with 12 carbon atoms) in 13 samples 

collected in the field in 1989, and in three sets of samples from the laboratory aging experiments. Because 

the oil fraction lighter than dodecane contains the 

hydrocarbons that have the greatest vapor pressures 

and aqueous solubilities, this fraction is most likely to 

undergo changes in composition and loss of mass. 

Measurements of loss of mass from this lightest fraction of the oil are likely to account for the large majority of the total loss of mass from the oil, because the 

heavier fractions of the crude oil (with larger carbon 

numbers than dodecane) contain hydrocarbons with 

smaller vapor pressures and solubilities.

One part of oil was dissolved in 33 parts dodecane. 

The oil-dodecane mixtures were thoroughly mixed, 

and 1-mL samples were extracted and injected into a 

GC attached to a mass selective detector. A 60-m 

length of a DB-1,0.325-mm i.d., fused-silica capillary 

column with a 5.0-|um film thickness was used in the 

GC. Helium was used as the carrier gas with a flow rate 

of 1 mL/min at 40 °C. The oven temperature of the GC 

was held at 40 °C for 5 minutes after injection and then 

the temperature was raised at a rate of 2 °C per minute 

from 40 to 200 °C from 5 to 85 min after injection.

Standards were analyzed to establish the retention 

times and retention order of major compounds and to 

compare the mass spectral characteristics of known 

compounds to sample peaks to facilitate compound 

identification. The relation between peak area and concentration was assumed to be constant for each hydrocarbon so that peak area was used as a measure of 

compound abundance. Areas of individual peaks were 

normalized to the decane (normal alkane with 10 carbons) peak to compare the relative concentrations of 

individual compounds in different oil samples. Decane 

was assumed to have the same abundance in all oil 

samples. This assumption was based on the fact that all

crude oil at the site originally was uniform in composition and that decane has a sufficiently low vapor pressure and solubility not to be volatilized or dissolved.

The percentage of the initial mass (in the original 

unaltered oil) of each individual compound lost from 

altered oil samples was calculated by comparing the 

normalized abundance of compounds in the altered oil 

to the normalized abundance in the least altered oil 

from the research site. The least altered oil (from well 

604a, fig. 1) was used as a proxy for the original oil 

because there are no adequately preserved samples of 

the original oil. The large abundance of volatile compounds in the least altered oil, and the fact that four of 

the least altered samples have very similar compositions and physical properties, indicate that these samples may closely approximate the composition and 

physical properties of the original oil. Values of percentage mass loss for individual compounds or the total 

oil determined are minimum estimates of mass loss, 

because it was assumed that the least altered oil has 

the same composition and physical properties as the 

original oil.

Boiling-point distillations of five crude-oil samples 

were performed using the Engler method (ASTM D86) 

to estimate the percentage of the total mass of the oil 

composed of the lighter-than-dodecane fraction of the 

oil that was analyzed by GC/MS.

CHANGES IN OIL COMPOSITION

The 57 peaks detected in the chromatogram of the 

oil sample from well 604a are the major peaks and are 

present in nearly all oil samples (fig. 2). Of the total 

integrated area under the chromatogram peaks, 74.08 

percent was attributable to aliphatic compounds (30.37 

percent to normal alkanes, 25.91 percent to cyclic 

alkanes, and 17.80 percent to branched alkanes), 13.87 

percent was attributable to aromatic compounds, and 

12.05 percent was attributable to unidentified compounds (table 1). The median carbon number was 7 

(C7). Fifty percent of the mass in the analyzed fraction 

had a carbon number of C6 to C8, 25 percent of the 

mass had a carbon number of C3 to C5, and 25 percent 

of the mass had a carbon number of C9 to C11.

Selective loss of aromatic compounds and aliphatic 

compounds with molecular weights less than nonane 

has occurred in altered samples of oil from the Bemidji 

site (fig. 2, table 1). Approximately 50 individual compounds have been depleted in the analyzed fraction. 

Maximum losses of the total mass of compounds with

633
LJJ 

Q_

Q 

LJJ

LJJ 

CC

O

LJJ

LJJ

O

cc

LJJ 

Q_

1

_ A

i^

1 1 1 1 1 1 1 1

C74

N6 N7

N5

B5

N4

B4 

N3

1 1 | 1

B61

1

B62

|

C(

C(

1

51

A6

P

|

32

372

O/

C71

II

C72

71

II

! 

N8

B81

A

C82

C81

C76 

C75

, III

"

C83

| |

1 1 1 1 1

Peak labels correspond

to ident 

table 1 .

tier labels in

h

N9

A°5

A81

Aft

1

ill

ll II

N10

A'9's

ll,

Nil

A1.01 A102

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100

90

80

z 

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LJJ "

cc

LJJ Q60

40

30

20

10

0 5 10 15 20 25 30 35 40 45 50 55 60 65 70 75 80 85

RETENTION TIME, IN MINUTES

_ R

D

B4

IV 3

N4 1

1

5 N 1 ' ' At 5

B

1

31 

N6 

B62

|

C 51

.1111

PRO ... .................. .^-- --- -

L C72 

P?2\N7

"Wll

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C 73 

. .....C74......

<r

1

C76

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C81 

C82

J

Peak labels correspond 

to identifier labels in 

table 1.

A82

B81

C83 

C 8

AS

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. A102

A1

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01

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0 5 10 15 20 25 30 35 40 45 50 55 60 65 70 75 80 85

RETENTION TIME, IN MINUTES

Figure 2 Composition of oil samples collected from the Bemidji, Minn., research site in 1989: (A) bar graph representation 

of a total ion chromatogram of the oil fraction lighter than dodecane for the least altered sample (from well 604a) and (B) 

percentage depletion of compounds in most altered oil (from well 309). [Peak identifier key: N = Normal alkanes, B - 

Branched alkanes, C = Cyclic alkanes, A = Aromatics; number following letter indicates number of carbons in molecule 

(N7=normal alkane with 7 carbons); where more than one compound with the same class and number of carbons is 

present, compounds are numbered sequentially.]

634
Table 1. Compounds present and distribution of mass in least altered oil collected from the Bemidji, Minn., research site in 

1989 (from well 604a) and minimum, median, and maximum depletions (as percent of initial) of compounds for set of 12 oil 

samples collected at the site in 1989

[N = Normal alkanes, B = Branched alkanes, C = Cyclic alkanes, A = Aromatics; number following letter indicates number of carbons in 

molecule (N7=normal alkane with 7 carbons); where more than one compound with the same class and number of carbons is present, 

compounds are numbered sequentially]

Peak 

Compound name identifier

Normal alkanes

Propane

Butane

Pcnianc

Hexanc

Heptane

Octane

Nonanc

Dccane

Undccanc

Sum:

Branched alkanes

Isobutane

Isopcnianc

2- mcth ly pcnianc

3-mclhylpcnUinc

2-mcthylhcxanc

3-methylhexanc

2-mclhylhcptanc

Sum:

Cyclic alkanes

Melhylcycpcnlane

Cyclohcxanc

Cis- 1 ,3-dimcthylcycIopcntanc

Trans- 1 ,3-di mclhylcyclopcntanc

Trans-l,2-dimcthylcyclopcnianc

Multiple cyclic alkanes

Multiple cyclic alkanes

Elhylcyclopcntanc

l-trans-2-cis-4-lrimcihylcyclopcntanc

l-trans-2-cis-3-lrimclhylcyclopcnlanc

Trans-1 ,2-dimcthylcyclohcxanc

Sum:

Aromatics

Benzene

Toluene

Ethylbcnzcne

M & P-xylcnc

O-xylcnc

Isopropylbcnzenc

N-propylbcnzene

Alkylbcnzene (unidentified)

Alkylbcnzcnc (unidentified)

1,3,5-trimclhylbcnzene

Alkylbcnzcnc (unidentified)

l-mcthyl-4-isopropylbcnzcnc

N-bulylbcnzenc

Sum:

N3

N4

N5

N6

N7

N8

N9

N10

Nil

B4

B5

B61

B62

B71

B72

B81

C61

C62

C71

C72

C73

C74

C75

C76

C81

C82

C83

A6

A7

A8I

A82

A83

A91

A92

A93

A94

A95

A96

A101

A102

Retention 

time, in 

minutes

7.216

10.758

17.691

27.935

39.665

51.382

62.453

72.696

82.240

9.180

15.589

24.555

26.176

36.355

36.954

47.524

31.574

35.782

38.272

38.632

38.966

43.294

43.947

44.383

45.218

46.093

49.915

34.497

47.173

58.210

59.118

61.893

65.234

68.615

69.029

69.278

69.752

70.847

75.591

78.708

Normalized 

peak area 

(to dccanc)

0.07

.49

1.86

2.20

2.23

1.92

1.03

1.00

.31

0.13

1.26

1 .65

.90

.27

1.11

1.19

1.37

1.57

.67

.48

1.05

2.34

.10

.22

.41

.48

.80

0.87

.87

.63

.84

.46

.25

.16

.20

.10

.27

.09

.19

.14

Percentage of 

area of total 

resolved peaks

0.19

1.34

5.08

6.00

6.09

5.26

2.82

2.73

0.85

30.37

0.35

3.44

4.52

2.47

.75

3.03

3.25

17.80

3.75

4.30

1.82

1.31

2.86

6.41

.26

.59

1.11

1.32

2.18

25.91

2.37

2.37

1.73

2.30

1.25

.69

.45

.56

.27

.75

.25

.51

.39

13.87

Carbon 

number

3

4

5

6

7

8

9

10

11

4

5

6

6

7

7

8

6

6

7

7

7

7

7

7

8

8

8

6

7

8

8

8

9

9

9

9

9

9

10

10

Percentage of compound lost

Minimum

0.00

.00

.00

.00

.00

.00

.00

.00

.00

33.01

.12

.00

.(X)

.00

.00

.00

0.00

.00

.00

.00

.00

.00

.00

.00

.00

.00

.00

39.76

12.17

20.97

35.46

27.04

77.39

20.80

.00

.00

71.05

.00

10.10

59.63

Median

58.69

60.52

49.75

32.72

30.44

.00

.00

.00

.00

100.00

47.50

28.31

26.86

25.66

30.33

9.48

29.47

32.47

25.65

25.28

26.93

21.87

34.36

16.29

10.35

10.99

6.94

81.33

51.93

38.95

50.07

40.07

100.00

100.00

.00

72.81

100.00

.00

36.16

100.00

Maximum

100.00

100.00

94.07

82.11

64.99

33.91

10.31

.00

.00

100.00

94.65

84.83

80.14

62.11

64.93

46.23

76.36

68.76

64.77

64.40

61.78

54.70

100.00

100.(X)

50.16

49.07

42.80

100.00

100.00

54.89

68.09

62.31

100.00

100.00

47.60

100.00

100.00

100.00

100.00

100.00

635
the following carbon numbers were measured: C3 (100 

percent); C4 (100 percent); C5 (95 percent); C6 (81 

percent); C7 (66 percent); C8 (48 percent); C9 (49 percent); and CIO (25 percent). Median losses, by carbon 

number, for the set of oil samples generally were onethird to two-thirds of the maximum numbers. Because 

aliphatic constituents have large Henry's law coefficients, these compounds preferentially partition to the 

vapor phase from the oil; most aromatic constituents 

have small Henry's law coefficients and will preferentially partition into the aqueous phase (Hult and others 

1991). Benzene and toluene (C6 and C7 aromatic 

compounds), will readily partition to both the vapor 

and aqueous phases from the oil (Hult and others, 

1991). Losses of C3-C8 compounds occurred primarily 

by volatilization of aliphatic compounds. Losses of 

C9-C10 compounds were primarily by dissolution into 

ground water and soil moisture of aromatic constituents. Benzene and toluene likely were primarily volatilized; losses by dissolution must have also occurred 

because losses of benzene and toluene exceeded losses 

of aliphatic compounds with similar molecular weights 

and vapor pressures (fig. 2). Methane (Cl) and ethane 

(C2), if initially present in the oil, were not detected in 

the oil samples and likely were lost from the oil prior to 

or during the sample collection and storage process 

because of their high volatility. Because some loss of 

C3 and C4 compounds may have occurred during sample collection and storage, abundances and losses of 

these compounds may be underestimated. The total 

mass of C1-C4 compounds in the original oil likely was 

relatively small compared to the total mass of the oil; 

losses of these compounds, therefore, has a negligible 

affect on the total mass balance of the oil. Maximum 

percent loss of the total mass in different structural 

classes of hydrocarbons were normal alkanes (57 percent), branched alkanes (72 percent), cyclic alkanes 

(62 percent), and aromatics (84 percent). Percent 

depletions of aromatic compounds, as a class, generally 

were higher than for aliphatic compounds (table 1). 

Minimum losses of the total mass of aliphatic compounds, as a class, were 0 percent; minimum losses 

of the total mass of aromatic compounds, as a class, 

were 30 percent. These larger depletions of aromatic 

compounds in otherwise only slightly altered oil samples indicates that dissolution of soluble compounds 

into water may initially occur at greater rates than 

volatilization.

CALCULATION OF MASS LOSS

Areal differences in the physical properties of the 

oil at the site have developed because of variable 

spatial and temporal alteration rates of the originally 

uniform oil (Landon and Hult, 1991). Changes in the 

physical properties of oil during laboratory aging 

experiments closely matched alteration trends in physical properties measured in the field. The percentage of 

the initial mass lost from oil collected in the field was 

calculated using measured changes in the values of 

physical properties of altered oil samples relative to the 

least altered oil and relations between percent mass 

loss and changes in physical properties of oil measured 

in the laboratory. Linear relations between changes in 

specific gravity and percent mass loss (fig. 3) and 

changes in refractive index and percent mass loss exist 

over the entire range of observed values of physical 

properties for Bemidji crude oil. The relation between 

kinematic viscosity and mass loss is nonlinear but systematic. Changes in physical properties of the oil indicate that the rate of mass loss from the original oil 

ranged from 0 to 1.25 percent per year (average of 0.5 

percent per year) and total mass loss ranged from 0 to 

11 percent (average is 3.8 percent) since the 1979 spill 

(Landon and Hult, 1991).

The percentage of total mass lost from altered oil 

also can be calculated using changes in the composition of the oil. These calculations were performed to 

compare the values of total mass loss determined on the 

basis of compositional changes to those calculated 

using changes in physical properties and to quantitatively determine the compounds and classes of compounds that account for most of the mass loss from the 

oil. Because the methods of calculating total mass loss 

based on composition and on physical properties were 

independent of each other, comparison of results of 

the two methods is a test of the reliability of the 

calculations.

Total-mass loss was calculated from changes in oil 

composition by summing the percentage losses of each 

individual compound weighted by the abundance of 

each compound in the least altered oil. Abundance of 

each compound (as a percentage of the total mass of the 

oil) was determined from the initial percentage of the 

mass of the resolved peaks for each compound (table 1) 

and the percentage of the mass of the oil in the analyzed 

fraction. Results of the boiling-point distillations indicated that approximately 32 percent of the total mass of 

the oil had a boiling point less than undecane (normal 

alkane with 11 carbons), the compound with the highest

636
0.000 0.005 0.010 0.015 0.020 0.025 0.030

CHANGE IN SPECIFIC GRAVITY 

FROM INITIAL FIELD VALUE

lie

rJO

O!

.O

la!

COW

O

£ 1

Q.

B

12345678

PERCENT MASS LOSS CALCULATED FROM 

CHANGE IN COMPOSITION

Figure 3. (A) Relation of change in specific gravity from initial 

field value to percent mass loss for oil subjected to volatilization 

under controlled laboratory aging experiments. (B) Percentage 

mass loss from oil of original spill composition estimated from 

changes in composition and specific gravity of oil collected from 

the Bemidji, Minn., research site in 1989.

molecular weight and boiling point in the analyzed 

fraction. Of this 32 percent, a median of 40 percent of 

the mass of the analyzed fraction was comprised of 

individually resolved chromatogram peaks whose 

areas were reported in the GC/MS analyses. The other 

60 percent of the mass was composed of the unresolved 

compound mixture (UCM), a complex mixture of 

many hydrocarbons with individual concentrations 

below the detection limit but which, when taken as a 

whole, comprise a substantial percentage of the mass of 

the oil. The resolved peaks within the analyzed fraction 

of the oil, therefore, comprise about 12.8 percent (32 

percent multiplied by 40 percent) of the mass of the oil.

Values of mass loss calculated from changes in oil 

composition deviate from those calculated using 

changes in specific gravity by 0.1 to 3.8 percent, with a 

median deviation of 1.6 percent (fig. 3). For two highly 

altered samples, mass loss calculated from changes in 

physical properties greatly exceeded losses calculated 

from compositional changes. Mass loss determined 

from compositional changes can underestimate the 

actual mass loss from highly altered samples because 

losses from the UCM in the sub-dodecane fraction and 

from fractions of the oil heavier than dodecane were 

not measured. In slightly and moderately altered samples, values of mass loss from compositional changes 

closely approximated or exceeded those determined 

from changes in physical properties. Physical-property 

measurements are bulk measurements that, in some 

cases, might not be sufficiently precise to detect small 

changes in physical properties. Comparisons of percent 

mass loss calculated from changes in composition of 

oil samples to percent mass loss measured during the 

laboratory aging experiments indicated that mass loss 

from changes in composition were within 1 to 2 percent 

of the measured value. Generally, close agreement was 

found between amounts of mass loss calculated independently on the basis of changes in kinematic viscosity, specific gravity, refractive index, and composition. 

It appears that changes in any of these parameters can 

be used in constructing a whole oil mass balance.

Losses of aliphatic compounds account for about 

80 percent of the mass lost from the highly altered oil 

(maximum, fig. 4). Losses of aromatic compounds 

account for the other 20 percent of the mass of the 

mobilized hydrocarbons. Losses of aliphatic compounds account for approximately 70 percent of the 

mass lost from the median oil (fig. 4). Because some 

removal of aromatic compounds (benzene and toluene) 

likely occurred as a result of volatilization, it appears

637
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co 

O

<co £co 1.5

Occ tu

tu 

o 

cc tu 

a.

Maximum

Median

Minimum

NORMAL ALKANES BRANCHED ALKANES CYCLIC ALKANES

HYDROCARBON CLASS

AROMATICS

CO

O

2.5

2.0

P I*

H-Z

1.0

O

£ 0.5 

DL

B

Maximum 

Median 

ill Minimum

C3 C5 C6 C7 C8 

CARBON NUMBER

C9 C10 C11

Figure 4. (A) Percentage of total oil lost associated with losses of hydrocarbons in each class: minimum, median, 

and maximum losses for the set of 12 oil samples collected from the Bemidji, Minn., research site in 1989. 

(B) Percentage of total oil lost associated with losses of hydrocarbons classified by carbon number; minimum, 

median, and maximum losses for the set of 12 oil samples collected from the Bemidji, Minn., research site in 1989.

638
that more than 80 percent and 70 percent of the loss of 

mass from highly altered and median oil samples, 

respectively, occurs because of volatilization. For 

slightly altered oil samples (minimum, fig. 4) loss of 

aromatic compounds accounts for nearly all the loss of 

mass from the oil.

The largest proportion of the total amount of mass 

loss is associated with losses of C6 and C7 compounds 

(fig. 4). Although greater percentage losses of C3, C4, 

and C5 compounds were due to the greater volatility of 

these compounds, C6 and C7 compounds were present 

in greater abundance in the oil than were C3, C4, and 

C5 compounds, and this greater initial abundance 

accounts for the greater total loss amount associated 

with C6 and C7 compounds.

DISCUSSION

Variations in the physical properties and composition of the oil observed in the field have developed 

from alteration of an originally uniform source. The 

variations in the composition and physical properties of 

different oil samples, therefore, represent oil at different stages of evolution from the original oil spilled at 

the site.

Most of the loss of mass from the oil occurs as a 

result of volatilization. The effects of biodegradation 

are poorly known but are likely to be small relative to 

the effects of volatilization and dissolution. Although 

percentage losses of aromatic compounds exceed percentage losses of aliphatic compounds, aromatic compounds in the analyzed fraction comprise less than 2 

percent of the total oil; losses of aromatic compounds, 

therefore, are not of major significance in the total mass 

balance for the oil. The distribution of the aromatics is 

significant from an environmental perspective because 

the aromatic compounds make up the vast majority of 

the contaminant mass dissolved in ground water 

(Baedecker and others, 1989; Hult and others, 1991).

Essaid and others (1991) found that the oil saturation in aquifer material in the oil bodies at the Bemidji 

site was highly variable. Even in aquifer material in the 

middle of the oil bodies no more than 50 percent of the 

pore space was saturated with oil. The remaining pore 

space is filled with water and (or) air. There are likely 

spatial variations in the composition and physical properties of oil on a very fine scale. Such microscale variations may be related to the degree of contact of oil 

droplets with air and (or) water phases and to the rate at 

which volatile hydrocarbons in the gas phase or soluble

hydrocarbons in the aqueous phase can move away 

from the oil. These microscale variations in oil properties and composition will not be detected by sampling 

oil from monitoring wells. The samples collected from 

monitoring wells represent vertically integrated average samples of the oil in the vicinity of the monitoring 

well. These bulk average samples, however, are still 

highly useful for studying the magnitude of the 

changes in oil characteristics and for calculating bulk 

or average rates of mass loss from the oil bodies for different areas of the oil pools.

The values of mass loss for the total oil and individual compounds in the oil provide an independent 

estimate of the spatially variable source term for transport models of volatile and soluble hydrocarbons emanating from the oil source in the subsurface at the 

Bemidji site. The methods used to quantify these loss 

rates can be used at other subsurface crude oil spill 

sites.

REFERENCES

Baedecker, M.J., Siegel, D.I., Gennett, P., and Cozzarelli, 

I.M., 1989, The fate and effects of crude oil in a shallow aquifer, I The distribution of chemical species 

and geochemical facies, in Mallard, G.E., and Ragone, 

S.E., eds, U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Phoenix, Arizona, September 26-30, 1988: 

U.S. Geological Survey Water-Resources Investigations Report 88-4220, p. 13-20.

Essaid, H.I., Herkelrath, W.N., and Hess, K.M., 1991, Air, 

oil, and water distributions at a crude-oil spill site, 

Bemidji, Minnesota, in Mallard, G.E., and Aronson, 

D.A., eds, U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 614-620.

Hult, M.F., 1991, Overview of research on contamination of 

the subsurface by crude oil at the Bemidji, Minnesota, 

toxic substances research site, in Mallard, G.E., and 

Aronson, D.A., eds, U.S. Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 611-613.

Hult, M.F., Landon, M.K., and Pfannkuch, H.O., 1991, 

Field validation of conceptual models of mobilization 

and transport of volatile petroleum derivatives in the 

unsaturated zone near Bemidji, Minnesota, in Mallard, 

G.E., and Aronson, D.A., eds, U.S. Geological Survey

639
Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Montercy, California, March 11- 

15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 621-626.

Hult, M.F, 1984, Ground-water contamination by crude oil 

at the Bemidji, Minnesota, research site An introduction, in Hult, M.F., ed., U.S. Geological Survey Toxic 

Substances Hydrology Program Waste technical meeting, Tucson, Arizona: U.S. Geological Survey WaterResources Investigations Report 84-4188, p. 1-15.

Landon, M.K., and Hult, M.F., 1991, Evolution of physical 

properties and composition of a crude oil spill, in 

Mallard, G.E., and Aronson, D. A., eds, U.S. Geological 

Survey Toxic Substances Hydrology ProgramProceedings of the technical meeting, Monterey, 

California, March 11-15, 1991: U.S. Geological Survey 

Water-Resources Investigations Report 91 -4034, 

p. 641-645.

640
Studies of the Molecular Size of Dissolved Organic 

Carbon Fractions Downgradient from the Oil Body at

Bemidji, Minnesota

By George Aiken1

Abstract

Preliminary results of equilibrium dialysis 

measurements on whole-water samples and 

isolates of the hydrophobic acid fraction of the 

dissolved organic carbon (DOC) in ground water 

at the Bemidji, Minn., toxic-waste study site are 

presented. Membranes with 2,000 (2K)-dalton 

and 12,000 (12K)-dalton cutoffs were used. 

Approximately 10 percent of the DOC in the 

whole-water sample from well 530 B was 

excluded from the 2K-dalton-cutoff membrane. 

Similar results were obtained with aqueous 

solutions of the hydrophobic acid fraction isolated 

from water from well 530 B. These results 

differed markedly from those obtained with 

aqueous solutions of the hydrophobic organic acid 

fraction from water from well 310, a backgroundwater-quality well, where approximately 30 

percent of the DOC was excluded from the 2Kdalton membrane. Measurements made on 

whole-water samples from two other wells in the 

contaminated plume indicate that the amount of 

organic matter excluded from the 2K-dalton 

membrane increases with decreasing distance to 

the oil body, possibly because of organicmatter/iron or organic- matter/organic-matter 

interactions. The results suggest that colloidsized organic constituents are generally absent 

from the ground water but that interactions of 

DOC with other dissolved constituents could

'U.S. Geological Survey, WRD, 3215 Marine Street, Boulder, 

Colorado 80303

result in an increase in molecular size of DOC 

fractions.

INTRODUCTION

The U. S. Geological Survey (USGS) study site 

near Bemidji, Minnesota (fig. 1), is an excellent site 

for studying interactions controlling organiccompound transport in ground water. In August 1979, 

a pipeline carrying crude oil burst; the spilled oil 

contaminated a shallow sand and gravel aquifer. 

Microbiological degradation of crude oil at the water 

table has resulted in the development of an extensive 

plume of dissolved organic carbon (DOC) that has 

migrated downgradient from the oil body. This DOC 

is a heterogeneous mixture of organic compounds, 

most of which are nonvolatile, yellow-colored organic 

acids that are intermediates in the microbiological 

degradation of the crude oil (Aiken and Thorn, 1989). 

In addition, the plume contains distinct redox zones 

(Baedecker and others, 1989). Measurable 

concentrations of hydrophobic contaminants, such as 

phenanthrene, downgradient of the oil body are strong 

evidence for facilitated transport at this site (Aiken 

and others, 1992), and the well-defined redox 

conditions and high carbon content in the mobile and 

stationary phases make this a desirable site for 

studying the processes that control solute migration. 

The interaction of hydrophobic organic species with 

colloids is a significant, albeit poorly understood, 

mechanism for transport of hydrophobic organic 

species in aquatic systems. Although the potential 

significance of colloids in affecting transport has been 

recognized (McCarthy and Zachara, 1989; Capel and 

Eisenreich, 1989), a number of problems associated 

with the sampling and analysis of colloids have

641
Water Well and Number

Figure 1. Location map of the Bemidji research site, Minnesota. Open circles show the approximate locations of selected 

wells at the site.

hampered research in this area. Preliminary equilibrium dialysis experiments indicate that the molecular 

sizes of aquatic humic substances from a number of 

environments are larger than expected on the basis of 

analyses of molecular weight of organic-matter isolates. Equilibrium dialysis provides information about 

molecular size rather than molecular weight and has 

been used successfully to study interactions of hydrophobic organic species with larger humic substances 

(McCarthy and Jimenez, 1985). In this paper, prelimi642
nary results of equilibrium dialysis measurements for 

the purpose of estimating the molecular size of organic 

matter at the Bemidji site are presented. The method 

used was designed to minimize concentration shifts 

an important consideration when studying concentration-dependent phenomena.

SITE DESCRIPTION

The sand and gravel aquifer underlies a pitted 

and dissected outwash plain; the aquifer is underlain, 

at a depth of about 20 m, by low-permeability till 

(Hult, 1987). Aquifer material consists of well-sorted 

medium quartz sand containing about 6 percent 

carbonate minerals, 30 percent feldspars, less than 5 

percent clay minerals, and less than 2 percent organic 

carbon (Baedecker and others, 1989). After sitecleanup efforts were completed, approximately 

400,000 L of crude oil remained in the aquifer and 

overlying unsaturated zone. The resulting 2- to 4mthick lens of crude oil floats on the water table about 8 

m below land surface.

In the 14 years since the spill, petroleum, as a 

separate fluid, has moved about 30 m downgradient, 

organic constituents in the ground water have moved 

at least 200 m downgradient, and vapors in the 

unsaturated zone have moved horizontally about 100 

m from the spill. The petroleum source is becoming 

increasingly viscous, dense, and depleted in volatile 

organic compounds as a result of selective dissolution 

and volatilization. Biological degradation of oil 

components has created an expanding anoxic zone 

around, immediately under, and downgradient from 

the spill, and has resulted in the development of 

distinct redox zones within the plume downgradient 

from the oil body. Microbiological degradation of 

crude oil at the water table also has resulted in an 

extensive plume of DOC that is migrating from the oil 

body.

METHODS

Samples were pumped from select wells along 

the plume axis by means of a Keck pump and 

collected in prebaked, 1-L pesticide bottles. DOC 

concentrations were determined by use of an 

Oceanographic International Model 700 carbon 

analyzer. The hydrophobic-acid isolates were 

obtained by extraction using XAD resins by the 

method of Aiken and Thorn (1989). Elemental

analyses were performed by the methods reviewed 

by Huffman and Stuber (1985), and molecular 

weights were determined by vapor-pressure 

osmometry by the method of Aiken and Malcolm 

(1987).

Molecular size-distribution measurements of 

the DOC in whole-water samples and solutions of 

isolated organic acids from a contaminated well 

(530 B) and from an uncontaminated well (310) 

were made by an equilibrium dialysis technique 

using Spectrum Spectra/por membranes with 2Kand 12K- dalton molecular weight cutoffs. In 

addition, molecular size-distribution measurements 

of the DOC in solutions of potassium acid phthalate 

and 4-[4-(2-carboxybenzoyl)phenyl]butyric acid 

also were made. In all cases, the membranes were 

first soaked in, and then rinsed with, distilled water 

to minimize the potential for sample contamination 

by organic materials leached from the membrane. 

After rinsing, membranes were filled with 50 to 70 

mL of distilled water and placed in 1-L widemouth 

sediment jars containing approximately 950 mL of 

the solution of interest. DOC concentrations in the 

solutions both inside and outside the dialysis bags 

were measured. The ratio of the DOC concentration 

inside the dialysis bag to the DOC concentration 

outside the bag (Cj/C0) was determined for a given 

solution at a given time to determine the rate and 

degree of equilibration. All reagents were reagent 

grade.

RESULTS AND DISCUSSION

The results of equilibrium dialysis 

measurements on contaminated ground water from 

well 530 B are given in figure 2. Well 530 B is 

located in a region of the plume that is a transition 

zone (Zone 4, according to Baedecker and others, 

1989) between the anoxic water near the oil body 

and the oxygenated, low-DOC waters 

downgradient. The water in well 530 B had low 

dissolved oxygen concentrations (= 0.3 mg/L, 

Baedecker and others, 1989), and a DOC 

concentration of 17.1 mg C/L. After 48 hours 

equilibration time, C/C0 equaled 0.98 in the case of 

the 12K-dalton membrane and C/C0 equaled 0.89 

in the case of the 2K-dalton membrane. Neither 

sample attained equilibrium after 48 hours, as 

defined by a flattening of the curve.

643
Well 530 B, 2K-dalton membrane 

Well 530 B, 12K-dalton membrane 

Well 530 B HPOA, 2K-dalton membrane 

Well 530 B HPOA, 12K-dalton membrane

20 30 40 50 60 

TIME, IN HOURS

10 20 30 40 50 60 70 80 90 100 

TIME, IN HOURS

Figure 2. Dialysis equilibration curves for whole-water 

samples and the hydrophobic acid (HPOA) fraction of the 

DOC isolated from water from well 530 B using 2K-dalton 

and 12K-dalton cutoff membranes. [Ci/Co is the ratio of the 

dissolved organic carbon (DOC) concentration inside the 

dialysis bag to the concentration outside the bag.]

The results of molecular size-distribution 

measurements of the DOC in whole-water samples 

from well 530 B were compared with results obtained 

using aqueous solutions of standard compounds and 

isolates of the hydrophobic acid fraction of the DOC 

in water from well 530 B and in uncontaminated 

ground water (fig. 3). The elemental compositions 

and molecular weight data for the isolates and the 

model compounds are given in table 1. The model 

compounds were chosen because of the similarity 

between these compounds and the isolate from well 

530 B with respect to polarity and molecular weight. 

Potassium hydrogen phthalate, which had the lowest 

molecular weight of the compounds studied, 

approached equilibrium most rapidly. The 

hydrophobic acid isolate from well 310 had the 

highest number average molecular weight (650 

daltons) and was the slowest to approach equilibrium. 

The 530 B hydrophobic acid fraction (335 daltons) 

and the 4-[4- (2-carboxybenzoyl)phenyl]butyric acid 

(CPBA, 313.33 daltons) approached equilibrium at an 

intermediate rate.

Figure 3. Dialysis equilibration curves for the whole-water 

sample from well 530 B, the hydrophobic acid (HPOA) 

fraction isolates in water from wells 530 B (530 B HPOA) 

and 310 (310 HPOA), potassium hydrogen phthalate (KHP), 

and 4-[4-(2- carboxybenzoyl)phenyl]butyric acid (CPBA) 

using the 2K-dalton cutoff membrane. Ci/Co is the ratio of 

the dissolved organic carbon (DOC) concentration inside the 

dialysis bag to the DOC concentration outside the bag.

A difference between the standards and the 

whole-water and isolate samples is that Cj/C0 equaled 

1.05 for both of the standards within 72 hours, 

whereas Cj/C0 was less than 1 for the other samples in 

the same time (fig. 3). The standards are 

homogeneous solutions containing a single organic 

constituent. The DOC in the whole-water sample and 

the organic matter contained in the hydrophobic acid 

isolates, on the other hand, is a complex mixture of 

high- to low- molecular-weight species. The 

molecular-weight data given for the isolate samples 

are number-average molecular weights, whereas the 

data for the standards represent true molecular weights 

for these species. It is likely that Cj/C0 will not 

approach 1 if the 2K-dalton membranes for the wholewater and isolate samples are used because of the 

presence of high- molecular-weight compounds in 

these samples.

644
Table 1. Ash-free elemental compositions and molecular-weight data for the hydrophobic-acid fractions isolated from water 

from wells 530 B and 310, and the corresponding data for phthalic acid and 4-[4- (2-carboxybenzoyl)phenyl]butyric acid.

[Elemental data reported as weight percent]

Sample 

(percent)

Hydrophobic-acid fraction from well 530 B

Hydrophobic-acid fraction from well 310

Phthalic acid

4-[4-(2-carboxybenzoyl)phenyl]butyric acid

Element

Carbon

61.9

58.4

57.8

69.2

Hydrogen

5.7

5.5

3.6

5.1

Oxygen

30.7

34.9

38.6

25.6

Nitrogen 

(daltons

0.3

0.7

0.0

0.0

Molecular 

weight

335

650

166.12

312.33

The results of molecular size-distribution 

measurements of the DOC in whole-water samples 

from well 530 B were similar to those obtained using 

aqueous solutions of the hydrophobic acid fraction 

isolated from well 530 B. After 72 hours, Cj/C0 

equaled 0.92 for the hydrophobic acid fraction sample, 

indicating that approximately 8 per cent of this sample 

was excluded by the membrane (fig. 3). In the original 

sample, 95 percent of the DOC was accounted for in 

the solutions inside and outside the membrane, indicating the possibility for sorption of some material on 

the membrane surface. In addition, complete equilibration across the membrane was not attained with a 

Cj/C0 of 0.98 using the 12K-dalton membrane (fig. 2). 

These results indicate the presence of organic components with molecular sizes larger than indicated by the 

number-average molecular weight of the sample. The 

difference in results obtained using the 2K-dalton and 

12K-dalton membranes also suggests that organic constituents with larger molecular sizes, considered to be 

material excluded from the 12K-dalton membrane, 

comprise only a small percentage of the DOC in 

ground water at well 530 B. It is possible that the 

large molecular-size material in the water sample from 

well 530 B is derived from the naturally occurring 

organic matter in the ground water that contributes to 

the DOC at well 530 B. A relatively low Cj/C0 of 0.7 

was obtained using the 2K-dalton membrane for the 

fulvic acid sample isolated from well 310 (fig. 3), indicating that approximately 30 percent of the naturally 

occurring background DOC was excluded (93 percent 

of the DOC was accounted for in this sample in the 

solutions inside and outside of the dialysis membrane).

Finally, molecular size measurements of the 

DOC in whole-water samples from two other wells 

within the contaminated plume were made using the 

2K- dalton membrane Well 532 B is located in the 

anoxic part of the plume (Zone 5, Baedecker and others,

1989) a zone characterized by the presence of 

dissolved iron and relatively large DOC concentrations. 

Well 515 B is located in the oxygenated waters 

downgradient from well 530 B (Zone 3, Baedecker and 

others, 1989) that contain little dissolved iron and 

decreased amounts of DOC. A Cj/C0 of 0.86 was 

determined after 48 hours with the 2K-dalton 

membrane for the sample from well 532 B. This Cj/C0 

ratio was the lowest ratio determined in this study; it 

increased downgradient to well 515 B where it was 0.9. 

These data suggest the possibility that either organic 

matter/iron interactions, or organic-matter/organicmatter interactions have produced organic- compound 

assemblages whose molecular size and weight increase 

with decreasing distance to the oil body.

REFERENCES

Aiken, G., Capel, P., Furlong, E., Hult, M. and Thorn, K., 

1992, Mechanisms controlling the fate and transport of 

organic chemicals in groundwater, in Mallard, G. E., 

and Aronson, D. A., eds.,U. S. Geological Survey 

Toxic Substances Hydrology Program Proceedings 

of the Technical Meeting, Monterey, California, 

March 11-15, 1991: U.S. Geological Survey WaterResources Investigations Report 91-4034, p. 633-637.

Aiken, G.R., and Malcolm, R.L., 1987, Molecular weight of 

aquatic fulvic acids by vapor pressure osmometry: 

Geochimica et Cosmochimica Acta, v. 51, p. 2177- 

2184.

Aiken, G.R., and Thorn, K.A., 1989, Variability in the 

chemistry of nonvolatile organic acids downgradient 

from the oil body at Bemidji, Minnesota, in Mallard, 

G.E. and Ragone, S. E., eds, U.S. Geological Survey 

Toxic Substances Hydrology Program-Proceedings 

of the Technical Meeting, Phoenix, Arizona, 

September 26-30, 1988: U. S. Geological Survey, 

Water-Resources Investigations Report 88-4220, 

p. 35-40.

645
Baedecker, M.J., Siegel, D.I., Bennett, P., and Cozzarelli, I. 

M., 1989, The fate and effects of crude oil in a shallow 

aquifer, I The distribution of chemical species and 

geochemical facies, in Mallard, G.E. and Ragone, S.E., 

eds, U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Phoenix, Arizona, September, 26-30, 1988: 

U. S. Geological Survey, Water-Resources 

Investigations Report 88-4220, p. 13-20.

Capel, P., and Eisenreich, S., 1989, Sorption of

organochlorines by lake sediment porewater colloids, 

in Suffet, I. H., and MacCarthy, P., eds., Advances in 

Chemistry Series 219, Aquatic Humic Substances: 

Influence on Fate and Transport of Pollutants: 

Washington, D.C., American Chemical Society, 

p. 185-207.

Hult, M.F., 1987, Movement and fate of crude oil 

contaminants in the subsurface environment at 

Bemidji, Minnesota, in Franks, B. J., ed., U. S. 

Geological Survey Program on Toxic Waste-GroundWater Contamination: Proceedings of the third 

technical meeting, Pensacola, Florida, March 23-27, 

1987: U. S. Geological Survey Open-File Report 87- 

109, p. C3-C5.

Huffman, E.W.D. and Stuber, H.A., 1985, Analytical 

methodology for elemental analysis of humic 

substances, in Aiken, G.R., McKnight, D.M., 

Wershaw, R.L. and MacCarthy, P., eds., Humic 

Substances in Soil, Sediment and Water: New York, 

John Wiley & Sons, p. 433-455.

McCarthy, J.F., and Jimenez, B.D., 1985, Interactions 

between polycyclic aromatic hydrocarbons and 

dissolved humic material: Binding and dissociation: 

Environmental Science and Technology, v. 19, 

p. 1072-1076.

McCarthy, J.F., and Zachara, J.M., 1989, Subsurface

transport of contaminants: Environmental Science and 

Technology, v. 23, p. 496-502.

646
Small-Scale Chemical Heterogeneities in a Crude-Oil 

Contaminated Aquifer, Bemidji, Minnesota

By Isabelle M. Cozzarelli 1 , Mary Jo Baedecker1 , George Aiken2 ,

anc/Curtis Phinney1

Abstract

A study of the spatial heterogeneity of 

biogeochemical reactions was undertaken in an 

aquifer affected by a crude-oil spill near Bemidji, 

Minnesota. An anoxic plume extended 90 meters 

downgradient from the crude-oil body in 1992, 

and contained elevated concentrations of 

dissolved organic carbon, methane, ferrous iron, 

and manganese. Two 1.5-meter-long continuous 

cores were collected from two locations within the 

anoxic plume to determine small-scale chemical 

heterogeneities in the ground water. Chemistry of 

pore water in a core collected from near the edge of 

the crude-oil body indicates that oxygenated water 

does not mix with the anoxic plume. The effects of 

mixing oxygenated ground water with the anoxic 

plume is reflected in the chemistry of pore water in 

a core collected 25 meters farther downgradient; 

mixing resulted in significant small-scale chemical 

heterogeneity in the concentrations of organic and 

inorganic chemical constituents.

INTRODUCTION

The geochemical character of a shallow sand and 

gravel aquifer near Bemidji, Minn., has been altered by 

the release of crude oil from a broken pipeline in 1979. 

Previous investigations at this site documented microbial 

degradation of monoaromatic hydrocarbons (Eganhouse and others, 1987, 1993; Cozzarelli and others, 

1990) and formation of redox zones (Baedecker and 

others, 1989, 1993; Bennett and others, 1993) downgradient from oil that had accumulated at the water 

table. The degradation of dissolved organic compounds

'U.S. Geological Survey, Reston, Va. 

2 U.S. Geological Survey, Boulder, Colo.

occurs by aerobic as well as anaerobic microbial 

degradation processes, resulting in the reduction of 

manganese and iron and the generation of methane.

Steep chemical gradients have been observed by 

other investigators in the study of contaminated aquifers (Nicholson and others, 1983; Back and Baedecker, 

1989; Cozzarelli and others, 1991; Smith and others, 

1991). These gradients can usually be explained by 

redox boundaries caused by a thermodynamic progression of degradative processes (Baedecker and Back, 

1979). Resolution of water-quality problems requires 

characterization of these chemical heterogeneities in an 

attempt to understand the factors that control changes 

in water quality. Novel techniques commonly are 

required to sample such environments at spatial and 

temporal scales appropriate to the reactions being studied. This paper describes small-scale chemical heterogeneities in the ground water downgradient from the 

crude-oil body at the Bemidji research site. The spatial 

variability of biogeochemical processes and the resulting effects on the fate of degradable hydrocarbons in 

the plume are discussed.

DESCRIPTION OF THE FIELD SITE

The surficial aquifer is an outwash plain in northcentral Minnesota, near Bemidji, (lat. 47°34'25"; long. 

95°5'25"). The aquifer was contaminated by crude oil 

when a pipeline burst in 1979 and sprayed oil onto the 

land surface (Hult, 1984). Oil reached the water table at 

a depth of 5 to 7 m below land surface and formed an 

oil body approximately l-m thick that floats on the 

water table (fig. 1). The aquifer sediment is heterogeneous and consists of medium- to poorly sorted, fineto medium-grained sand near the oil body (D.A. Franzi, 

State University of New York, written commun., 

1989). Estimates of ground-water velocities near the 

water table range from 0.05 m/d in fine-grained sediment to 0.5 m/d in coarse-grained sediment (Bennett 

and others, 1993).

647
EXPLANATION 

A A' Line of section

Oil Ground water sampling 

/ site and number spray 

area x Core

0 100 200 300 FEET

50 100 METERS

Figure 1. The study site at Bemidji, Minn., showing locations of watertable well, cores, the oil body, and oil pipeline. (A-A' is the sampling 

transect in figure 2.)

STUDY METHODS 

Field Procedures

Ground-water samples were collected from watertable wells and deeper wells along a transect in the 

direction of ground-water flow (fig. I). Ground-water 

samples for analysis of unstable constituents were 

collected with a submersible pump, and ground-water

samples for analysis of organic compounds were 

collected with a Teflon bailer by the methods of Baedecker and others (1993) and Eganhouse and others 

(1993).

The 1.5-m-long continuous cores were collected 

from two locations (A and B, fig. 2) downgradient from 

the crude-oil body using a modified version of a fixedpiston, wireline-core barrel (Murphy and others, 1996).

648
The coring method involved drilling with a hollowstem auger to the top of the desired coring interval, followed by driving of the core barrel beneath the augers 

to the desired depth. The coring method uses a piston 

that creates a suction on saturated sediment within the 

core barrel as the barrel is driven into the sediment. The 

bottom of the core is frozen in situ to prevent loss of 

fluids out the core bottom. The 47-mm diameter cores 

were collected in clear polycarbonate liners. The cores 

were removed from the core barrel and immediately 

sealed with plugs at the top and bottom after retrieval. 

The top plug was connected to N2 gas which was at a 

pressure of less than 30 kPa during sampling of the 

core.

Water from the saturated cores was collected at 

16-cm intervals beginning at the top of the core. The 

core was drained at each interval before the next interval 

was sampled. The pore water was sampled by drilling

a 4-mm-diameter hole in the polycarbonate liner and 

then inserting a leur-tip syringe snugly into the hole. 

Approximately 10 to 15 mL of pore water was 

collected at each interval.

Water samples were collected for methane (CH4 ) 

and monoaromatic hydrocarbon analyses without 

filtration. Samples for dissolved organic carbon (DOC) 

analysis were filtered through 0.45-um silver filters. 

Samples collected for ferrous iron (Fe2+) analysis were 

filtered through 0.1 -um Nuclepore filters. Water samples collected for measurement of CH4 concentrations 

were collected in Glasspak syringes. The water sample 

was transferred from the syringe into 25-mL serum 

bottles containing enough mercuric chloride (HgCl2 ) to 

produce an Hg concentration of 0.2 mM. Ground water 

for determination of volatile hydrocarbons was collected in 1.8-mL glass vials containing HgCl2 .

A

METERS 

435

430

425 h

420

415

s s / S

Oil body A B Water table

- ,-- * f-v

^- - -^ ' x i^ ^ o

^ - - - O ,. - <i- i ^

Direction of

ground-water flow

VERTICAL EXAGGERATION x 10 

DATUM IS SEA LEVEL

0 100 200 300 FEET

0 50 100 METERS 

EXPLANATION

I Screened interval of sampling sites 

/ Core

-- 0.5 Line of equal O2 concentration 

Interval, in milligrams per liter, is variable

Figure 2. Cross section along transect A-A' of the Bemidji, Minn., aquifer 

showing location of water table and deeper wells, cores, and oil body, and the 

concentrations of dissolved oxygen for 1992.

649
Analytical Procedures

Samples collected for Fe2+ analysis were fixed with 

reagents in the field; Fe2+ concentrations were determined colorimelrically by a modified bipyridine 

method (Baedecker and Cozzarelli, 1992). Dissolved 

methane concentrations were measured by headspace 

analysis and gas chromalography (Baedecker and 

Cozzarelli, 1992). DOC concentrations were measured 

by the persulfate wet-oxidation technique using a carbon analyzer. Concentrations of volatile monoaromatic 

hydrocarbons were measured by purge-and-trap capillary gas chromatography using a DB-5 bonded-phase 

fused-silica capillary column as described by Phinney 

and Cozzarelli (1996). The hydrocarbons were analyzed by gas chromatography with ion-trap mass-spectrometric detection.

GROUND-WATER CHEMISTRY AND 

SPATIAL HETEROGENEITIES

An anoxic zone, approximately 1 '/2- to 3-in thick, 

extended 90 m downgradicnt from the oil body in 1992 

(fig. 2). Elevated concentrations of reduced chemical 

species, CH4 (29.7 mg/L), Fe 2+ (57.6 mg/L), and Mn2+ 

(2.7 mg/L) were found in this anoxic zone but were at 

concentrations less than 0.1 mg/L in the background 

water. The highest concentration of DOC, 35 mg/L, 

was found in ground water beneath the oil body. The 

primary processes controlling the oxidation of hydrocarbons in the anoxic zone were identified as microbially mediated iron and manganese reduction and 

methanogenesis. The steepest gradients in chemical 

constituents occurred along the edges of the anoxic 

plume where recharge water and ground water upgradient from the oil body mix with the plume (fig. 2).

At 20 m downgradient from the edge of the 

crude-oil body, the concentrations of chemical constituents in the pore water from core A indicate that oxygenated water does not infiltrate into the anoxic plume 

(figs. 3 A,B,C) as illustrated by the high concentrations 

of DOC, Fe2+, and hydrocarbons (total monoaromatic 

hydrocarbons) that are present less than 10 cm below 

the water table. The shallowest pore-water sample was 

collected at a depth of 8.8 m within the capillary fringe, 

10 cm above the water table. Oxygen in infiltrating 

water is consumed in the unsaturated zone because of 

the presence of hydrocarbons (2.7 mg/L) (fig. 3B). The 

absence of oxygen results in a stable, small-scale 

homogeneity in the geochemical zones in ground water

at this location. This lack of heterogeneity is further 

indicated by the constant molar ratios of several 

monoaromatic hydrocarbons relative to 1,2,3,4-tetramethylbenzene (1,2,3,4-TMBz) throughout the vertical 

profile (fig. 3C). 1,2,3,4-TMBz is the most stable of the 

soluble oil constituents and is the best available conservative tracer of the plume (Eganhouse and others, 

1993). On the basis of the constant ratios, the conclusion is that no preferential microbial degradation has 

occurred near the water table.

The effects of mixing of the anoxic plume with 

infiltrating oxygenated ground water is observed in the 

pore-water geochemistry of core B, which was collected near the edge of the anoxic plume, 45 m downgradient from the edge of the crude-oil body. 

Significant small-scale chemical heterogeneity indicates that, at the downgradient edge of the anoxic 

plume, the biogeochemical reactions are changed as 

infiltrating oxygenated water mixes with the anoxic 

plume (figs. 3 D,E,F). The DOC concentrations greatly 

increase from 5.6 mg/L near the water table to more 

than 17 mg/L at less than 1 in below the water table. 

Concentrations of Fe2+ increased from 2.3 mg/L at the 

water table to 38 mg/L at about a meter below the water 

table. The monoaromatic hydrocarbons represent only 

4 percent of the DOC within the first meter below the 

water table. Analyses of pore water in core B from 

greater than 1 m below the water table, in contrast, 

indicate that 14 percent of the DOC is composed of 

monoaromatic hydrocarbons. The remainder of the 

DOC is most likely composed of nonvolatile compounds, such as hydrophobic and hydrophilic acids that 

are degradation products of the oil (Eganhouse and 

others, 1993). The ratios of individual monoaromatic 

hydrocarbons relative to 1,2,3,4-TMBz indicate that 

hydrocarbons that are persistent within the anoxic 

plume (Eganhouse and others, 1987; Cozzarelli and 

others, 1990) are degrading by aerobic processes in a 

thin zone (<l-m thick) near the water table (fig. 3F).

SUMMARY

The results of this study underscore the need to 

determine small-scale geochemical changes so that the 

biogeochemical processes occurring in organic-rich 

subsurface environments can be evaluated. Pore-water 

samples from two cores collected downgradient from a 

subsurface oil body indicate that oxygenated water 

does not mix with the anoxic plume 20 m from the edge 

of the oil body. The chemistry of the water at this loca650
CO

S

W

CO

W 

PQ

CO

W

U

D

CO

Q

W 

PQ

8.8

9.0

9.2

9.4

9.6

9.8

10.0

10.2

METHANE (CH4) CONCENTRATION,

IN MILLIGRAMS PER LITER 

0 4 8 12 16 20

HYDROCARBON CONCENTRATION 

IN MILLIGRAMS PER LITER

S 10-4

s.^

0 5 10 15 20 25 30 35 40

DISSOLVED ORGANIC CARBON, 

IN MILLIGRAMS PER LITER

D METHANE (CH^ CONCENTRATION,

IN MILLIGRAMS PER LITER 

0 2.5 5 7.5 10 

8.8

9.0

9.2

9.4

9.6

9.8

10.0

10.2

10.4

10 15 2C

8.8

9.0

9.2

9.4

9.6

9.8

10.0

10.2

10.4

10

0 20 40 60 80 100 

FERROUS IRON, IN MILLIGRAMS PER LITER

HYDROCARBON CONCENTRATION 

IN MILLIGRAMS PER LITER

c

8.8 

9.0

9.2 

9.4

9.6

9.8

10.0

10.2

>n A

RELATIVE CONCENTRATION OF BENZENE 

0 50 100 150 200 250

I \ » .

( Propylbenzene / ^ Benz / A V"

Ethylbenzene/ \ .

f r

/ T/

\ /

-

i

ene

DISSOLVED ORGANIC CARBON. 

IN MILLIGRAMS PER LITER

0 10 20 30 40 

FERROUS IRON, IN MILLIGRAMS PER LITER

9 2 4 6 8 10 12

RELATIVE CONCENTRATION OF 

ETHYLBENZENE AND PROPYLBENZENE

RELATIVE CONCENTRATION OF BENZENE 

0 25 50 75 100 125

8.8

9.0

9.2

9.4

9.6

9.8

10.0

10.2

10.4

10

RELATIVE CONCENTRATION OF 

ETHYLBENZENEAND PROPYLBENZENE

Figure 3. Vertical cross sections showing concentrations of:

(A) Methane (CH4) and dissolved organic carbon (DOC);

(B) Total monoaromatic hydrocarbons (HC) and ferrous iron (Fe2+); and

(C) Benzene, ethylbenzene, and propylbenzene relative to 1,2,3,4-tetramethylbenzene in pore water from core A as a function of 

depth below land surface; and the concentrations of

(D) Methane (CH4) and dissolved organic carbon (DOC);

(E) Total monoaromatic hydrocarbons (HC) and ferrous iron (Fe2+); and

(F) Benzene, ethylbenzene, and propylbenzene relative to 1,2,3,4-tetramethylbenzene in pore water from core B as a function of

depth below land surface.

The water table was at 8.9 m below land surface at the location of core A, and at 9.1 m below land surface at the location of core B.

651
tion is characterized by high Fe2+ and CH4 concentrations and DOC, a large traction of which is composed 

of monoaromatic hydrocarbons. In contrast, mixing 

with infiltrating oxygenated water at a site 20 m farther 

downgradient has produced spatial heterogeneities in 

the composition of chemical constituents and their concentrations. Aerobic degradation of hydrocarbons in a 

thin zone near the water table results in their depletion 

relative to a more conservative organic compound, 

1,2,3,4-TMBz. The DOC in this area of the aquifer is 

composed of mostly nonvolatile hydrocarbons.

REFERENCES

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of Hydrology, 106, p. 1-28.

Bacdcckcr, M.J., and Back, W., 1979, Modern marine 

sediments as a natural analog to the chemically stressed 

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Groundwatcr quality and analysis at hazardous waste 

sites: N.Y., Marcel Dckkcr, p. 25-461. 

Bacdcckcr, M.J., Cozzarelli, I.M., Eganhousc, R.P., Siegcl, 

D.I., and Bcnnctt, P.C., 1993, Crude oil in a shallow 

sand and gravel aquifer, 3 Biogcochcmical reactions 

and mass balance modeling in anoxic groundwatcr: 

Applied Geochemistry, v. 8, p. 569-586.

Bacdcckcr, M.J., Siegcl, D.I., Bcnnctt, P.C., and Cozzarelli, I.M., 1989, The fate and effects of crude oil in a 

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1993, Crude oil in a shallow aquifer, I Aquifer characterization and liydrogcochcmical controls on inorganic solutes: Applied Geochemistry, v. 8, p. 529-549.

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Gcochcmical gradients in shallow ground water caused 

by the microbial degradation of hydrocarbons at Galloway Township, New Jersey, in Mallard, G.E., and 

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technical meeting, Montcrcy, California, March 11-15, 

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Cozzarelli, I.M., Eganhousc, R., and Bacdcckcr, M.J., 

1990, Transformation of monoaromatic hydrocarbons 

to organic acids in anoxic ground water environment: 

Environmental Geology and Water Sciences, v. 16, 

p. 135-141.

Eganhouse, R.P., Baedcckcr, M.J., Cozzarclli, I.M., Aikcn, 

G.R., Thorn, K.A., and Dorscy, T.F., 1993, Crude oil in 

a shallow sand and gravel aquifer, 2 Organic 

geochemistry: Applied Geochemistry, v. 8, p. 551-567.

Eganhouse, R.P., Dorsey, T.F., Phinney, C.S., Baedecker, 

M.J., and Cozzarclli, I.M., 1987, Fate of monoaromatic 

hydrocarbons in an oil-contaminated aquifer: Evidence 

for the importance of microbial activity, in Geological 

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Hult, M.F., 1984, Ground-water contamination by crude oil 

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Report 94-4015.

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Migration of contaminants in groundwatcr at a landfill: 

A case study, 6 Hydrogcochcmistry: Journal of 

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two analytical techniques for identifying and measuring 

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Importance of closely spaced vertical sampling in 

delineating chemical and microbiological gradients in 

groundwatcr .studies: Journal of Contaminant 

Hydrology, v. 7, p. 285-300.

652
Use of Carbon and Hydrogen Stable Isotopes to

Investigate the Production and Fate of Methane at a

Toxic Waste Site, Bemidji, Minnesota

By Kinga Revesz 1 , Tyler Coplen 1 , Mary J. Baedecker1 , and Marc Hult2

Abstract

Stable carbon and hydrogen isotope ratios 

of dissolved methane and aqueous carbon dioxide 

in the saturated zone of a crude oil spill near 

Bemidji, Minn., support the concept of methane 

production by acetate fermentation with 

concomitant increase in bicarbonate concentration 

as opposed to carbon dioxide reduction (and 

bicarbonate consumption). Oxidation of dissolved 

methane along the lateral flow path seems to be 

minimal because no measurable change in 

isotopic composition of methane occurs with 

distance from the oil body. As methane from the 

ground water diffuses upward through a 5- to 

7-meter-thick unsaturated zone, it is partially 

oxidized to carbon dioxide, increasing the 5 C of 

the remaining methane, decreasing the 5 13C of 

the carbon dioxide, and increasing the pressure of 

carbon dioxide in the unsaturated zone.

No increase of concentrations of 

atmospheric methane was detected at ground 

level directly above the oil plume; concentrations 

were identical to those of a background 

atmospheric sample.

INTRODUCTION

Although oil spills are becoming increasingly 

common, few geochemical studies have investigated 

in detail the production and fate of methane (CH4), 

one of the products of biodegradation of oil. We 

investigated CH4 geochemistry at Bemidji, Minn.

(fig. 1), where crude oil from a spill floats on the water 

table in a shallow sand and gravel aquifer. Some 

components of the crude oil are degraded to CH4 and 

other products in the ground water by anaerobic 

microbes. The CH4 produced in the ground water 

diffuses upward into the unsaturated zone, and is 

transported horizontally by the ground-water flow. To 

identify the reactions generating CH4 in the saturated 

zone at Bemidji, we analyzed CH4, carbon dioxide 

(CC>2), and water for their isotopic compositions.

At Bemidji, substantial amounts of volatile 

petroleum hydrocarbons are transported from the 

water table through the unsaturated zone as vapor, 

which subsequently dissipates to the atmosphere or is 

biodegraded. An increase in oxygen concentration 

and a decrease in CC>2 concentration toward land 

surface in the unsaturated zone indicate that volatile 

hydrocarbons are being oxidized to CC>2 (Hult and 

Grabbe, 1986). Recent gas-chromatographic 

measurements show that the concentrations of C3-C8 

alkanes and CH4 decrease rapidly toward land surface 

(Hult and others, 1991), although the decrease in CH4 

is less than that of the alkanes. Data from Hult and 

others (1991) indicate that the transport of 

hydrocarbons in the unsaturated zone is through 

diffusion coupled with oxidation. To provide 

information on the sources and sinks of CH4 in the 

unsaturated zone, we measured the concentration of 

CH4 and C(>> , the hydrogen and carbon isotopic 

composition of CH4, and the carbon isotopic 

composition of C(>>. 3

'U.S. Geological Survey, 431 National Center, Reston, VA 

22092.

2 U.S. Geological Survey, 2280 Woodale Drive, Mounds 

View, MN 55112.

3 We express hydrogen and carbon isotope ratios as delta values given in per mil (parts per thousand, or %o) relative to a working standard VPDB (Vienna PeeDee Belemnite) for 8 I3C values 

on a scale normalized such that NBS19 carbonate is +1.95%o (Hut, 

1987) and VSMOW (Vienna Standard Mean Ocean Water) for 6D 

values on a scale normalized such that 8D of SLAP (Standard 

Light Antarctic Precipitation)'is -428 %o relative to VSMOW. 

(Coplen, 1988.)

653
o>

2

EXPLANATION

Direction of ground-water 

flow

spray

, A A ' Line of section

417 (around water sampling 

site and number

100 200 300 FEET

I i

METERS 

435 1

430 

425 

420 

415

 ATMOSPHERE SAMPLING SITE 

WATER TABLE

DIRECTION OF 

GROUND-WATER FLOW

OIL BODY

VERTICAL EXAGGERATION x 10 

DATUM IS SEA LEVEL

EXPLANATION

i SCREENED INTERVAL 

OF SAMPLING SITES

100 200 

50 

Figure 1. Bemidji research site, Bemidji, 

arrow shows approximate direction of grou

hydrogeologic section through the oil plum

Cozzarelli, 1991, fig. 2).
THEORETICAL BACKGROUND

Anoxic aqueous environments are commonly 

characterized by the formation of biogenic CH4 

(Whiticar, 1986). Two biogenic CH4 production 

pathways are commonly recognized: the reduction of 

CC>2 by hydrogen and the fermentation of acetate. 

Acetate fermentation refers collectively to 

methanogenesis that involves the transfer of a methyl 

group from any substrate. These two pathways can be 

distinguished on the basis of stable carbon and 

hydrogen isotope ratios.

The carbon isotope fractionation factor between 

coexisting CO2 and CH4 can be expressed as

a

( 13C/ 12C) CO,

CO2 -CH4

(1)

where <xCO2-CH4 * s 8enera^y 1 -05 to 1.09 for reduction 

of CO2 and 1.04 to 1.06 for fermentation reactions 

(Whiticar, 1986). The carbon fractionation factors of 

the two production pathways overlap. In some cases, 

carbon isotope data are not sufficient to identify the 

reaction pathway, as was the case at Bemidji. Therefore, determination of the hydrogen isotope ratios is 

useful because the isotopic difference between the two 

CH4 -production pathways is large. The formation 

water is the only source of the hydrogen in the CH4 

formed by reduction of CO2 . A linear relation 

between 5D in CH4 and 5D in water for the reduction 

pathway in per mil is (Nakai and others, 1974; 

Schoell, 1980)

6DCH = 6DHO -160± 10 . (2)

For acetate fermentation processes, 25 percent of the 

hydrogen in the CH4 is derived from water and the 

remaining 75 percent is transferred from methyl 

carbon acetate (Pine and Barker, 1956; Daniels and 

others, 1980). The relation between 6DCH4 and 

8DH o f°r tne reaction in per mil is

= 0.258DH0-321 (3)

Methanogenesis in freshwater usually represents a 

mixture of acetate fermentation and CO2 reduction 

pathways. Many investigators have concluded that 

fermentation of acetate is the source of about

70 percent of the CH4 in freshwater (Jeris and 

McCarty, 1965; Balyaev and others 1975; Cappenberg 

andJongejan, 1978; Mountfort and Asher, 1978). The 

production of CH4 by CO2 reduction is thought to be 

responsible for the remaining 30 percent.

Microbial anaerobic oxidation of CH4 can also 

be identified by isotopic abundance measurements, 

which indicated that the carbon isotope fractionation 

factor ranges from 1.002 to 1.014, a range that is 

smaller than that associated with methanogenesis by 

CO2 reduction or by acetate fermentation.

EXPERIMENTAL PROCEDURES

A vacuum preparation system and sampling 

technique were used to determine the 6D and 6 I3C 

values of CH4. The major experimental steps involved 

(1) sample collection of CH4 from ground water, the 

unsaturated zone, and air, (2) determination of CH4 

and CO2 concentrations by use of gas 

chromatography, (3) quantitative separation and 

conversion of CH4 into CO2 and H2, and (4) isotoperatio analysis of CO2 and H2 by mass spectrometry.

Sample collection of CH4 from ground water, 

modified from the method of Oremland and others 

(1987), involved combining about 50 mL of N2 with 

450 mL of water sample and shaking vigorously for 

equilibration. The headspace gas mixture was entirely 

removed by a syringe and it was stored by injecting it 

into a stoppered, crimped serum bottle (60 mL) that 

was previously filled with deionized water. The water 

was partly replaced by inverting the bottle and 

inserting an 18-gauge needle through the stopper. This 

procedure allowed recovery of the entire sample 

without contamination or dilution by air. Furthermore, 

this segregation of the extracted gases eliminated any 

possibility of postsampling microbial alteration.

Gas samples from the unsaturated zone were 

collected from a sampling point 1 m above the center 

of the oil plume. Samples were collected through 

10-cm-long, 6.4-mm-o.d. stainless steel screens 

attached to 3.2-mm-i.d. stainless steel tubing 

permanently installed 0.6, 1.5, 2.4, 4, and 5 m below 

land surface. Sampling tubes were placed in 

10-cm-diameter auger holes, and the holes were 

backfilled with native sand and bentonite (Hult and 

Grabbe, 1986). Samples were collected in 

preevacuated 250-mL gas-sample tubes by use of a 

peristaltic pump by a modified method of Wood and 

Petraitis (1984). The gas samplers have a stopcock at

655
each end and an opening with a septum. Aliquots of 

10 to 20 mL were extracted for sample preparation by 

inserting a syringe through the septum.

Air samples 1 m in volume were collected from 

5 cm above the ground surface directly above the oil 

plume and about 50 m away from the oil plume (for a 

background sample). The air samples were 

pressurized in an aluminum cylinder by use of an oilfree air compressor (Wahlen and others, 1989).

All samples were analyzed for CH4 

concentration and, when necessary, for CO2 

concentration by means of a gas chromatograph with a 

flame-ionization detector. Gas chromatographic 

samples were no larger than 1 fjL for soil gas samples 

and 10 mL for air samples to ensure that the isotopic 

composition of samples was not altered on the basis of 

mass balance considerations.

Sample preparation for analysis of 8 C and 8D 

in CH4 consisted of (1) separation of the CH4 from a 

large volume of air by passing the sample slowly over 

charcoal at -190°C (Wahlen and others, 1989), which 

absorbs the CH4, (2) purification of CH4 by gas 

chromatography (Revesz and Coplen, 1990), (3) 

quantitative oxidation of CH4 to CO2 and H2O by use 

of a copper furnace at 450°C and a copper oxide 

furnace at 800°C, and (4) decomposition of water to 

H2 by passage over U at 800°C (Bigeleisen and others, 

1952). Air samples with a volume fraction of 1 to 3 

ppm CH4 required all four steps. Gas samples with a 

volume fraction of 5 to 20 percent CH4 however, did 

not require preenrichment (step 1) before gaschromatographic separation. Gas samples consisting 

of 100 percent CH4 or containing only inert gases and 

no other hydrocarbons beside CH4 were combusted 

directly. The 1-a standard deviations of the system if 

all steps are used are 0.13 %o for 8 C and 1.55 %o for 

5D values, nearly as small as those obtained for 

carbonate or water samples by many laboratories.

RESULTS AND DISCUSSION

CH4 Production in Ground Water

In this study, the concentration of CH4 and the 

stable-isotope ratios for CH4, CO2 , and water were 

measured along the ground-water flow path. The 

results are summarized in table I and figure 2. Along 

a flow path at an altitude of around 422 m above sea 

level, CH4 concentrations decrease with distance from 

the oil plume, whereas the 8D and 8 C ratios for the 

CH4 are relatively constant (fig. 2 and table 1). The

8 C ratio for dissolved inorganic carbon (DIG) 

decreases by about 30 %o and HCO^ concentrations 

decrease between the center of the plume and 100 m 

downgradient (fig. 2). Baedecker and others (1989, 

1993) noted that with time, CH4 concentrations 

increased sharply and the S 13C of DIC became 

progressively heavier. The dramatic increase in CH4 

production in 1987 is typically observed in freshwater 

environments when Fe(III) is no longer available and 

the S I3C of DIC increases (Lovley and others, 1989). 

Since 1987, there has been no major change in CH4 

concentration or its S 13C ratio in the ground water.

In order to identify the pathway for CH4 

genesis, one can look at the relations among SDH O' 

SDCH 8 I3CCH , 8 I3CCO , and their isotopic 

fractionation factors (Oremland and others, 1987). As 

shown in the plot of 8 13C for CH4 as a function of SD 

of CH4 (fig. 3A), the CH4 is produced by 

fermentation. A plot 8D of CH4 and 8D for coexisting 

formation water (modified from Whiticar, 1986) 

shows that the Bemidji data fall near the line 

indicating fermentation (fig. 3B). A plot S 13C of 

coexisting aqueous CO2 as a function of 8 C of CH4 

shows that CH4 in Bemidji is on the borderline (close 

to the middle line) between the fermentation and the 

CO2 reduction pathways (fig. 3C). From a plot of 

hydrogen-isotope fractionation factors against the 

carbon-isotope fractionation factors (fig. 3D), one can 

see that CH4 seems to be produced by fermentation at 

Bemidji. This is the characteristic CH4 production 

pathway in freshwater environments and this 

observation is in good agreement with the study of 

Cozzarelli and others (1990; 1993 in press), who 

detected acetate in the ground water at Bemidji.

Furthermore, production of CH4 by the acetate 

fermentation pathway is accompanied by production of 

CO2 , which can react with calcium carbonate to 

produce bicarbonate. Thus, the high values of 

bicarbonate that are observed at Bemidji (Baedecker 

and others, 1989) are inconsistent with acetate 

fermentation. 14C studies (Blair and others, 1987) 

indicate that the methyl group of acetic acid is 

converted to CH4. Because the methyl groups of acetic 

acid are depleted in 13C relative to total acetic acid 

carbon, CH4 produced by acetate fermentation is 

significantly depleted in 13C relative to acetate. 

Consequently, inorganic carbon is enriched in C and 

its value can exceed 0 %o as shown for the Bemidji site 

in fig. 2.

656
Table 1. Geochemistry of water samples collected from selected wells, Bemidji, Minn., 1990

[m, meter; %o, per mil; mg/L, milligram per liter; 8D and SO are expressed relative to Vienna SMOW water; 6 

belemnite; a, isotopic fractionation factor]

Well

B015

B026

B420

B421A

B5IO

B5I5A

B516

B518

B530A

B531

B532A

B533A

B707A

B417F

B515B

B530B

B530C

B532B

B532C

B532D

B533B

B533C

B707B

B707C

B801B

Distance8 

(m)

-

--

21

0

105

137

110

57

91

67

47

36

-70

--

137

91

91

47

47

47

36

36

-70

-70

--

Altitude 

(m)

-

-

423.0

421.0

423.0

422.5

-

422.5

422.5

422.2

423.5

422.5

423.5

~

420.0

421.5

420.2

421.2

421.0

417.0

421.5

420.5

421.0

418.0

--

513C, 

CO2

(% )

+6.45

-16.80

-9.15

+ 12.95

-12.20

-13.35

-17.70

-7.85

-14.45

-

-.05

+2.10

-17.75

-15.40

-13.05

-6.30

-13.15

-.05

-15.00

-12.85

-1.00

-14.30

-12.60

-12.35

-13.00

CH4

Concen- §D §13c 

tration 

(mg/L) (/oo) (/6o)

20.3 -359.0 -54.85

.9

-359.0 -60.6

20.6 -352.5 -52. 1 6

2.9

..

..

1 3.2 -347.5 -57.21

6.5

9.15

12.1 -334.0 -54.82

16.7 -338.0 -54.08

.001

.08

-

10.5

.41

11.5

.01

.0023

9.15

0

.001

.0008

.93

C is expressed relative to Vienna Peedee

H2O

SD

(%o)

-83.1*

-

-83.1*

-88.5

-80.5

-82.0

~

-83.1*

-82.5

-83.0

-78.5

-

-86.0

--

-

-77.5

-84.0

-83.0

-84.0

-

-80.5

-

-86.0

-86.0

-84.0

5180 

(% )

-

-

-

-12.IO

-1 1. 40

-II.90

-

~

-H.65

-II. 50

-II. 15

--

-12.00

~

--

-11.10

-12.00

-11.65

-11.90

-

-11.40

-

-12.00

-12.00

-11.95

Fractionation Factor

u a 

CO2-CH4 H2O-CH4

1 .06 1 .43

_.

1.05

1 .07

..

-

1 .05 1 .40

..

..

1.06 1.38

1.06 1.38

-

-

..

..

~

-

~

..

-

-

..

-

*Estimate of value.

aDist. is lateral distance to wells and negative values are upgradient.

The decrease in CH4 in the ground water 

downgradient from the oil plume (fig. 2) can be 

explained by dilution by native ground water and by 

the loss of CH4 to the unsaturated zone because 5 I3C 

and 5D of CH4 change little with distance along the 

flow path. Had oxidation been an important 

contributor to the decrease in CH4, the 5 13C values of 

CH4 would have increased along the flow path 

because oxidized CH4 is depleted in I3C by as much 

as 14%o (Whiticar, 1986) and the resulting DIG would 

presumably be depleted in C.

CH4 Consumption in the Unsaturated Zone

CH4 concentration and isotopic data for the 

unsaturated zone are summarized in table 2 and

figures 4 and 5. CH4 concentrations are 

approximately constant within 1 meter of the water 

table and then decrease gradually to the surface 

(fig. 4). CC>2 concentrations are constant from the 

water table to within 1 m of the surface. Depth is 

plotted against 5 13C of CH4 and CC>2 in figure 5. 

8 13C values of CH4 are inversely correlated with 

CH4 concentrations and indicate that the decrease in 

CH4 concentration toward the surface is caused not 

only by diffusion but also by a mechanism that 

preferentially consumes 12CH4 .

One would expect little change in the carbon 

isotopic composition of CH4 if CH4 concentration 

decreases solely by diffusion. However, one might 

expect a change in carbon-isotope composition of

657
J 10

i 10

n! 5

-. o

5 -5

fe -10

& -15

»o -9n

-

-

" 

m

" " 

 

1 1 1 1 1 1

/&u 

j 700

i 650

- 600

n

8 55°

1 500

_, -400 

1 -380

Z -360

5* -340

o .320

3

-300

strui - 0

a.

z

V -55

X

o

o -50

* -45

22

20

16

c3 16

I* 8

0 6

4

2

n

I

-

. 

1 1 1 1 1 ' 1 '

-

 .

_

1 1 1 1 1 1

 *

' 

*

-

1 1 1 I 1 1

.

- .

"" * 

 ~ *

"" 1 1 1 1 1 1

0 20 40 60 80 100 120 140

LATERAL DISTANCE FROM OIL PLUME, IN METERS

Figure 2. Relation of geochemical data and lateral distance 

from the center of oil plume at the Bemidji research site.

and CH4 with depth if CH4 is beeing oxidized

I ^

to CO2 . Oxidation explains (1) the increasing 8 C 

of CH4 with decreasing CH4 concentration towards 

the surface, (2) the decrease in the 13C content of 

CO2 towards the surface (from 1 to 3 m below land 

surface, fig. 5), and (3) the stable CO2 concentration 

between the water table and a depth of about 1 m 

(the CO2 production of CH4 oxidation compensates 

for the decrease in CO2 concentration caused by 

diffusion).

The fractionation factor between CO2 and CH4 

ranges from 1.008 to 1.028, values that are in good 

agreement with the values of 1.005 to 1.03 obtained by 

Whiticar and Faber (1986) for similar field areas 

where CH4 oxidation occurs.

CH4 Contribution to the Atmosphere

No difference in CH4 concentration in 

atmospheric samples was found between a background 

sample and a sample collected just above the 

contaminated area; both samples had a CH4 content of 

1.79 ppm by volume. A typical 8 C for atmospheric 

CH4 is -46%o; -33%o was measured for the CH4 gas 

from the unsaturated zone. The addition of 1 part CH4 

from the unsaturated zone to 1 part atmospheric CH4 

would give an atmospheric concentration of 3.58 ppm 

by volume and a 8 13C of -39%o. Thus, these results 

indicate that carbon-isotope ratios probably cannot be 

used at Bemidji to quantify transport of CH4 from the 

unsaturated zone to the atmosphere.

CONCLUSIONS

A dramatic increase in CH4 production was 

observed in the saturated zone above a crude oil plume 

near Bemidji, Minn. Findings from our study include 

the following:

In the Saturated Zone:

1. Carbon and hydrogen isotopic ratios of CH4 

support CH4 production by acetate fermentation (as opposed to CH4 production by CO2 

reduction).

2. Carbon-isotope fractionation between CH4 

and dissolved inorganic carbon also supports 

the acetate fermentation production pathway.

3. Oxidation of dissolved CH4 along the 100-m 

long flow path investigated seems to be minimal because there is no measurable change in 

the isotopic composition of CH4 with distance; therefore, the decreases of CH4 concentration along the flow path is caused by 

dilution and the loss of CH4 to the unsaturated 

zone.

In summary, the CH4 production by acetate 

fermentation in Bemidji is causing an increase in 

concentration and 8 13C of dissolved inorganic carbon, 

apparently because of the intermolecular isotopic 

inhomogeneity of acetate.

658
_l

5

oc

LJJ 

0.

Z

til*

X

»-

tll

5

u.

O

Oo

*<<>

-100

-88

-75

-62

-50

-38

-25

-12

Cti2 REDUCTION

o

cP

0 °fc°

V

~

v ^_ «

^ o " "o° o o o 0

V^^ o O O O^ ^^S^Q

^^o ° o5L oTRANSITION ^^ °

° B B^^^O^ ° C>

^^H PTTI^^ O O

o ,_

o

Tyy^"M^

- m DATA FROM BEMIDJI 

0 DATA FROM WHITICAR (1986)

-400 -300 -200 -100

til

til

X

SD OF METHANE, IN PERMIL

-150

^ -200

cc 

tii 

a.

z

H -250

-300

S -350

-400

o

ERMENTATION

 DATA FROM BEMIDJI ° 

O DATA FROM WHITICAR (1986)

-100 -50 50

20

10

O 

O 

u. 

O 

O

-10

-20

-30

-40

-50

CO0 REDUCTION

-100 

813C OF MET

z

o

o

»-

^

GC

til

Q_

o

w

z 

til

8

o:

o

i

tf

1 .0

1.5

1.4

1.3

1.2

U 

i i i 

 

0 

FERMENT

O Q O 0 9fiO _

' 1 

.

r> ,-

_ oQ

0 o°b °

° 

0 0

1 1 1 

1.03 1.04 1.05 1.06 

5D OF WATER, IN PERMIL «C- CARBON ISOT

Figure 3. Distinction between acetate fermentation and C02 reduction methanogenesis by use of isotopes, based on Whiticar's (19

function of 5D of CH4. The Bemidji data fall into the area indicating that CH, is produced by fermentation. (B) 5D of CH4 as a func

Bemidji data fall on the fermentation line. (C) 513C of C02 as a function of 5^0 of CH4. The CH4 production in Bemidji is on the bord

production pathways. <xc is the carbon isotope fractionation factor. (D) Hydrogen isotope fractionation factor versus carbon isotop

Bemidji indicates production of CH4 by the acetate fermentation pathway.
Table 2. Geochemical data for the unsaturated zone at Bemidji, Minn., 1991

[m, meter; vol. %, volume fraction in percent; %o, per mil; 8 I3C is expressed relative to Vienna Peedee belemnite; 

8D is expressed relative to Vienna SMOW; a, isotopic tractionatipn factor]

Altitude 

(m)

430.5

429.5

428.5

427.5

426.5

425.5

Well

301.1

301.2

301.3

301.4

301.5

301.7

CH4 

(vol. %)

0

1.95

5.14

9.30

11.05

11.46

CO2 

(vol. %)

4.45

11.72

11.27

12.05

12.12

12.26

5"<2r

~

-33.35

-38.25

-42.05

-41.85

-42.05

 & 

-

-292.5

-323.5

-317.5

-337.0

-323.5

513CC02

-24.50

-25.55

-24.65

-18.95

-16.20

-14.95

a 

CO2-CH4

-

1.008

1.015

1.024

1.027

1.028

430

429

,- 428

427

426

B Methane

+ Carbon Dioxide

I

2 4 6 8 10 12 

VOLUME FRACTION. IN PERCENT

14

430 -

-35 -30 -25 -20 

813C. IN PERMIL

Figure 4. Change in concentration of CH4 and CO2 in the 

unsaturated zone at Bemidji test site as a function of depth 

below the surface.

In the Unsaturated Zone:

1. CH4 is apparently oxidized between 427.5 m 

and the surface, as indicated by the increase in 

5 13C values and the gradual decrease in CH4 

concentration with decreasing depth.

2. The CO2 concentration as a function of depth 

curve is approximately constant below a depth 

of 1 m, in accordance with oxidation of CH4 to 

C02.

3. The 5 13C of CO2 increases with depth in accordance with oxidation of CH4 that is depleted in 

13C. Between the altitude of 425.5 and 427.5 m,

Figure 5. 513C change of CO2 and CH4 in the unsaturated 

zone at the Bemidji test site as a function of depth below the 

surface.

the CH4 and CO2 is probably a mixture of CH4 

production and consumption. The shapes of the 

CH4 and CO2 distributions in the upper 3 m are 

consistent with CH4 consumption (producing a 

concave upward curve) and CO2 production (elevated CO2 concentration at shallow depth).

In the Atmosphere:

No difference in CH4 concentration in 

atmospheric samples was found between a background 

sample and a sample collected just above the 

contaminated area. Thus, isotope data cannot be used at 

this test site to quantify discharge of CH4 from the 

unsaturated zone to the atmosphere.

660
SELECTED REFERENCES

Balyaev, S.S., Finkelstein, Z.I., and Ivanov, M.V., 1975, 

Intensity of bacterial methane formation in ooze 

deposits of certain lakes: Microbiology, v. 44, 

p. 272-275.

Baedecker, M.J., Siegel, D.I., Bennett, Philip, and

Cozzarelli, I.M., 1989, The fate and effects of crude 

oil in a shallow aquifer, I The distribution of 

chemical species and geochemical facies: U.S. 

Geological Survey Water-Resources Investigations 

Report 88-4220, p. 13-20.

Baedecker, M.J., and Cozzarelli, I.M., 1991, Geochemical 

modeling of organic degradation reactions in an 

aquifer contaminated with crude oil: U.S. Geological 

Survey Water-Resources Investigations 

Report 91-4034, p. 627-632.

Baedecker, M.J., Cozzarelli, I.M., Eganhouse, R.P., 1993, 

Cruide oil in a shallow sand and gravel aquifer, III 

Biochemical reactions and mass balance modeling in 

anoxic ground water: Applied Geochemistry, v. 8, 

p. 569-586.

Bigeleisen, J., Perlman, M.L., and Prosser, H.C., 1952, 

Conversion of hydrogenic materials to hydrogen for 

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662
Sorption of Polycyclic Aromatic Hydrocarbons to Aquifer 

Materials Containing a Low-Organic Carbon Content

from Bemidji, Minnesota

By Joseph J. Piatt1 , Steven J. Eisenriech2 , Debera A. Backhus3 , and Paul D. Capel4

Abstract

Three polycyclic aromatic hydrocarbons, 

naphthalene, phenanthrene, and pyrene were 

sorbed to a single sample of aquifer material containing a low-organic carbon content at two different temperatures, 5 °C and 25 °C, using batch 

and column experimental techniques. Controls on 

batch and column techniques were incorporated 

to ensure that the experimental vessels were 

organic-carbon free, biological and photolytic 

activity was absent, and solution supernates were 

colloid-free. The sorptive uptake curves indicated 

a two-step approach to equilibrium. The data 

were modelled using a two-site sorption kinetic 

model. Times to equilibrium were at least 40 

hours. Small increases in equilibrium distribution 

coefficients were observed with decreased temperature which resulted in enthalpy of sorption 

values which were consistent with a partitioning 

mechanism. Small decreases in rate constants 

were occurred as temperature decreased. The 

values of the solids/water distribution coefficients 

for the batch experiments were consistently 4 to 

60 times larger than those obtained from the 

column experiments, probably because of differences in experimental boundary conditions or to 

creation of fine particles in the batch experiments

Dept. of Soil and Water Science, Univ. of Arizona, Tucson, 

AZ

Gray Freshwater Biological Institute, Univ. of Minnesota, 

Navarre, MN

School of Public and Environmental Affairs, Indiana Univ., 

Bloomington, IN

4U.S. Geological Survey, Navarre, MN

that would increase the surface area available for 

sorption. Although the relative effects of organiccarbon phase sorption versus mineral phase 

sorption were not isolated directly, the magnitude 

of the observed distribution coefficients and rate 

constants are consistent with sorption controlled 

by the organic-carbon content of this aquifer 

material.

INTRODUCTION

Contamination of ground water by toxic organic 

chemicals has become a national problem (Hult, 1984; 

CheremisinorT, 1987). One group of organic chemicals, the polycyclic aromatic hydrocarbons (PAH's), 

are of particular concern because many are suspected 

mutagens, teratogens, and carcinogens (Abdul and 

Gibson, 1986). Thus, the processes that affect the 

transport and fate of these contaminants in the subsurface must be understood in order to assess exposure 

risks to humans or other living species and develop 

efficient and cost-effective remediation strategies.

Sorption is an important process affecting the 

transport and fate of hydrophobic organic chemicals in 

aquifer systems. Sorption is a term used to describe 

the movement of a chemical species between the aqueous and solid phases (Weber 1972; Voice and Weber, 

1983). The magnitude of sorption is defined by a 

distribution coefficient (Kd), which is the ratio of the 

concentration of species in the sorbed phase to the 

concentration of species in the aqueous phase. If a 

chemical sorbs to a stationary sorbent, such as the 

porous media of an aquifer, its transport will be 

retarded. If a chemical sorbs to a mobile sorbent 

phase, such as dissolved organic matter or mobile 

colloids, its transport will be enhanced as compared to

663
the stationary sorbed phase (McCarthy and Zachara, 

1989; Aiken and others, 1991). The major factors 

affecting sorption include the solute's hydrophobicity 

and the sorbent's organic carbon content, mineral composition, and surface area (Voice and Weber, 1983; 

Chiou 1989; Hassett and Banwart, 1989; Harmon and 

others, 1989; Karickhoff, 1984; Karickhoff, 1981).

Aquifer materials commonly have fractional 

organic carbon (foe; mass of organic carbon per mass 

of total dry solids) contents of less than 0.001. There 

has been little research done on the sorption of PAH's 

to low foe solids, but it is known that the linear freeenergy relations developed for predicting sorption to 

sorbents with a foe greater than 0.1 do not work as 

well for low foe sorbents (Schwarzenbach and 

Westall, 1981; Stauffer and others, 1989; Piwoni and 

Banerjee, 1989).

Even less well understood are the kinetics of 

hydrophobic organic chemical sorption to mineral surfaces. Most studies usually allowed 24 to 48 hours for 

sorptive equilibrium (Piwoni and Banerjee, 1989). 

Sorptive equilibrium to mineral oxides has been 

reported to be as short as a few hours (Backhus, 1990; 

Perlinger and others, 1993). But a recent long-term 

equilibration study concluded that sorptive equilibrium can take weeks or even years depending on the 

solute (Ball and Roberts, 199la). The time-dependent 

sorption data show an initial rapid uptake followed 

by a much slower extended uptake of the chemical 

(Karickhoff, 1980). The presence of organic matter 

and mineral surface, however, confounds efforts to 

define a sorptive mechanism. Researchers generally 

agree that sorption is rate-limited, but they disagree as 

to the cause. Two major hypotheses exist: the first 

hypothesis is that rate-limited sorption is due to 

retarded intraparticle diffusion (Ball and Roberts, 

199Ib; Harmon and others, 1989). The second 

hypothesis is that rate-limited sorption is due to 

intraorganic matter diffusion (Brusseau and others, 

1991; Szecsody and Bales, 1989).

Some solute-transport models assume that 

local equilibrium exists between the solute and the 

sorbent ~ that is, that sorptive interactions are fast 

compared to advective/dispersive transport (Lapidimus and Amundson, 1952; Gaber and others, 1992). 

But these equilibrium models do not accurately predict 

solute movement in transport studies (Brusseau and 

Rao, 1989; Pignatello, 1989; Harmon, 1989). Some 

other, more robust, models deal with nonideal transport by incorporating pseudo-first-order rate constants

to describe rate-limited physical, transport-related, or 

chemical-sorption-related processes. Physical 

processes tend to be rate limited owing to diffusional 

mass transfer of solutes between mobile and immobile 

flow regimes (van Genuchten and Wierenga, 1976a, 

1976b). On the field-scale, physical nonequilibrium 

probably dominates nonideal solute transport induced 

by spatial variations in hydraulic conductivity and distribution coefficients, soil aggregation, and fractures 

and fissures (Brusseau and Rao, 1989).

Sorption-related nonequilibrium is also important. For hydrophobic organic chemicals, sorption to 

solids has been shown to be a two-step process: a fast 

initial solute uptake taking minutes to hours followed 

by a slower rate of uptake taking days to weeks 

(Karickhoff, 1984; Ball and Roberts, 199la). For these 

chemicals, sorption nonequilibrium is thought to be 

due to intrasorbent diffusion, and the rate limitations 

are due to intraorganic-matter diffusion and (or) 

retarded intraparticle diffusion.

This paper summarizes the results of batch and 

column sorption experiments, which are used as complimentary techniques to determine equilibrium distribution coefficients and sorptive/desorptive rate 

constants for experiments conducted at 5 °C and 

25 °C. The distribution coefficients obtained from 

these experiments could be used in ground-water contaminant transport models to predict and interpret the 

transport and behavior of PAH's in the subsurface.

MATERIALS AND METHODS

Materials

Naphthalene (greater than 99 percent pure), 

phenanthrene (greater than 99.5 percent pure) and 

pyrene (99 percent pure) were chosen as the PAH 

compounds because of their ubiquity at contamination 

sites and range of physical-chemical properties 

(table 1). Sodium bicarbonate (NaHCO3 ) was used 

as a background electrolyte.

The sorbent used was an uncontaminated 

aquifer sample taken from near Bemidji, Minnesota, at 

a site of a crude-oil spill that occurred in 1979 and that 

is currently under study by the U.S. Geological Survey 

as part of its Toxic Substances Hydrology Program 

(Hull, 1984). The ground water has a pH 7.7 and an 

alkalinity of about 230 meq/L (Bennett and Siegel, 

1987). The bulk aquifer material can be described as a 

moderately sorted, medium-grained sand containing 

57 per cent quartz, 30 percent feldspar minerals,

664
Table 1 - Physical and chemical properties of three polycyclic aromatic hydrocarbons 

at 25°C and measured sorptive distribution coefficients to aquifer material 

at 5°C and 25°C

[°C, degrees Celsius, Pa, Pascals; crrrVmole, cubic centimeters per mole, mg/L, milligrams 

per liter; mUg, milliliters per gram]

Naphthalene Phenanthrene Pyrene

Molecular weight 

Vapor Pressure (Pa) 

Molar volume (cm3/mole) 

Water solubility (mg/L)

Distribution coefficients (mUg) 

batch - 5°C 

batch - 25°C 

column - 5°C 

column - 25°C

128.2

36.8

148

31

178.2 202.3

.113 .0119

199 214

1.10 .132

2.3

1.6

.036

.038

3.0 

2.7 

.75 

.61

13

8.3

3.2

2.1

Desorption rate constants (hour1 )

batch - 5°C

batch - 25°C

column - 5°C

column - 25°C

.14

.057

1.2

1.0

.037

.095

.45

.57

.030

.067

.18

.25

6 percent carbonate minerals, and less than 5 percent 

clay minerals (Aiken and others, 1991). The bulk 

material was wet sieved to exclude grain diameters 

less than 62 (im and greater than 2,000 (im to avoid 

phase-separation problems in the experiments.

The bulk aquifer material had a fractional 

organic carbon (foe) of 0.00025 ± 0.00003, the sieved 

aquifer material had a foe content of 0.00019 ± 

0.00008. Sieving the aquifer material had no measurable effect on the foe of the soil. The specific surface 

area of the soil was measured by a nitrogen three-point 

surface area analysis. Bulk aquifer material had a 

specific surface area of 1.37 m2/g. Sieved aquifer 

material had a specific surface area of 1.08 m /g. 

Thus, the portion of the aquifer material removed by 

sieving accounted for about 20 percent of the bulk 

soil's total specific surface area.

Batch Experiments

Batch experiments were done in 35 mL glass 

centrifuge tubes. Aqueous PAH solutions were 

prepared in 0.001 M NaHCO3 electrolyte solution at 

greater than 50 percent of their aqueous solubility by

adding a small amount of a PAH in methanol stock 

solution. The amount of methanol added from the 

stock solutions was always less than 0.05 percent by 

volume. The same amount of methanol was added to 

quantitation standards and to control and experimental 

blanks.

Each experiment consisted of duplicate sets of 

four tubes. Each set had a control tube that contained 

only electrolyte solution and the aquifer material. 

Each set also had three experimental tubes that 

contained PAH solution and the aquifer material. The 

amounts of aquifer material and solution were determined gravimetrically through sequential weighing of 

the tube and cap; tube, cap and soil; and tube, cap, soil, 

and PAH or electrolyte solution. The mass of sorbent 

was chosen to ensure that there would be at least a 

50 percent difference between initial and final aqueous 

concentrations at equilibrium. The masses of soil 

added to the tubes were 20 g, 15 g, and 10 g for naphthalene, phenanthrene, and pyrene, respectively. All 

centrifuge tubes were filled to the top with solution so 

that the solutes could not volatilize into a headspace. 

The tubes were wrapped with aluminum foil to 

prevent photodegradation.

665
After filling, the tubes were placed on a wrist 

shaker and shaken for a predetermined length of time. 

Times were chosen to maximize the resolution of the 

sorptive uptake curve. Several "early" time points 

(0.3, 1, 3, and 9 hours) were chosen in the first few 

hours of the experiment to coincide with the often 

reported fast sorption step. After the first several 

hours, the duration between time points was increased 

considerably (27, 54, and 100 hours) to accommodate 

a "slow" sorptive approach to equilibrium.

Following shaking, tubes were centrifuged at 

approximately 630 to 875 RCF (relative centrifugal 

force) for about 25 minutes. Part of the supernate was 

removed via a glass syringe and transferred to 10-mL 

centrifuge tubes. The tubes were filled to the top and 

capped with solvent-rinsed aluminum foil-lined caps. 

These smaller tubes were then centrifuged at approximately 1330 to 1680 RCF for about 45 minutes to 

remove any traces of fine particles. The supernate 

from these tubes was analyzed for PAH concentration 

by fluorescence spectroscopy.

The consistency and constancy of bulk-phase 

characteristics was monitored throughout the experiments by measuring the pH and specific conductance 

of the supernate solutions after solute analysis. It was 

reasoned that if all tubes in a given experiment had 

consistent pH and specific conductance readings, then 

all tubes were exposed to the same conditions. In 

addition, the turbidity of the supernate solutions was 

measured to detect the presence of colloids.

Column Experiments

A column system was designed for continuous 

flow-through detection. The system consisted of a 

glass column, stainless-steel plumbing, highperformance liquid-chromatography (HPLC) pumps, 

glass solution reservoirs, and flow-through fluorescence and specific conductance detectors. The glass 

column (28 cm in length and 2.54 cm in diameter) was 

modeled after those used by Lion and others (1990), 

except it was fabricated from only glass and stainless 

steel. The ends of the column were slightly rounded to 

avoid a dead volume. One end of the column formed a 

cap with a ground-glass base which was mounted on 

the main column section, which also had a groundglass base. Stainless-steel frits, 0.5-um nominal pore 

size, were seated at the column inlet and outlet. At the 

column outlet, a small plug of glass wool was placed 

between the solids and the frit to prevent the frit from

becoming clogged with clay-size particles. The 

column was wet-packed with aquifer material and 

flushed with 0.001 M NaHCO3 electrolyte solution for 

approximately 2 days.

HPLC pumps delivered aqueous solutions at 

0.8 mL/min, which corresponded to porewater velocities of about 27 cm/hr. This pore-water velocity was 

chosen because it was as close to estimated actual field 

pore-water velocities as the pumps could deliver accurately. The general range for other studies is 0.1 to 

100 cm/hr. One pump was used to deliver the electrolyte solution and a second pump was used to deliver 

the PAH solutions. A stainless-steel six-port valve 

facilitated switching between the two pumps. Three 

4-L reservoirs, which were closed to the atmosphere, 

held the electrolyte, "PAH", and sodium chloride 

(NaCl) solutions.

After the column was stabilized with electrolyte 

solution and the solute solution was equilibrated with 

the pump parts, a solute front was started to begin formation of the break-through curve (BTC). NaCl was 

used a tracer to define the hydrodynamic characteristics of the column. Flow rates were checked several 

times during the course of the experiments to ensure 

constancy. The column effluent was directed through 

transfer lines to flow through detectors. NaCl was 

analyzed using a specific conductance meter and the 

PAH solutions were analyzed using a fluorescence 

detector (fig. 1). The influent concentration was 

monitored at the beginning and end of the solute pulse 

for constancy. The specific conductance, pH, and 

turbidity of effluent concentrations were monitored, as 

they were for the batch experiments.

RESULTS AND DISCUSSION

Bulk-solution-phase pH and specific conductance measurements were constant, indicating that 

sorption and not proton exchange or mineral dissolution was the only process occurring. In the tracer 

column experiments, transport-related nonideality in 

the BTC was not observed because the system was 

homogenous. Thus, sorption nonequilibrium was the 

only process that caused asymmetric BTC for this 

system.

Batch and column data sets showed that Kj 

values increased with a decrease in temperature for a 

given compound and increased Kj values for more 

hydrophobic chemicals at a constant temperature. 

These observations were a direct result of the

666
Naphthalene 

Phenanthrene 

T Pyrene 

 Modelled Fits

10 20 30 40 50 

PORE VOLUMES

EXPLANATION

60 70 80

First Second

run run

+ O

 a

T V

Naphthalene

Phenanthrene

Pyrene

Figure 1. Breakthrough curves for duplicate experiments of naphthalene, phenanthrene, and pyrene at 5 °C for the column 

studies. Duplicate experimental runs for the same compound produced identical breakthrough curves. The relative aqueous 

concentration is the measured concentration divided by the influent concentration. The lines are mathematically modeled 

break-through curves.

increased activity coefficient for the PAH's as the temperature was decreased. The overall increase in Kd 

values with a 20 °C decrease in temperature, however, 

was only about 1.4 to 1.6 times. Independent analysis 

of the contributions of organic-carbon and mineral 

phases to total sorption indicated that organic carbon 

was the major contributor to total sorption, even 

though the Bemidji aquifer material had a low fraction 

of organic carbon (0.00019). In addition, the thermodynamic data, calculated from the observations at two 

temperatures, indicated small enthalpy of sorption 

values (van der Waal's forces) consistent with partitioning of the PAH's into the aquifer material's organic

carbon. Entropy values greater than measured in 

these experiments might suggest the formation of 

hydrogen-bonds indicating mineral-phase contributions to sorption due to solute/sorbent specific interactions. The batch Kd values were consistently larger 

than the column Kd values, either because (1) modelling the column data entails normalizing the aqueous 

solute concentration to the total volume of the system, 

whereas calculation of batch-derived Kd values normalizes the aqueous concentration to the volume of 

solution only, or (2) the residence time of the PAH's in 

the column experiments did not allow true sorptive 

equilibrium to be obtained.

667
z 

o

H

H

UJ 

O

o 

o

CO

o

tu

LU

F

LU 

DC

1.0

0.8

0.6

0.4

i= 0.2 -

0.0

I I

O Naphthalene 25°C

a Phenonthrene 25°C

 Phenanthrene 5°C

 Naphthalene 5°C - 

A Pyrene 25°C 

A Pyrene 5°C

20 40 60 80 100

TIME, IN HOURS

EXPLANATION 

5'C 25'C

 O Naphthalene

 D Phenanthrene 

T V Pyrene

Figure 2. Decrease in relative aqueous concentration over time for naphthalene, phenanthrene, and pyrene for the batch 

studies. The relative aqueous concentration is the measured concentration divided by the initial concentration. Vertical bars 

represent the error associated with the aqueous concentration.

The kinetic data from the batch studies showed 

a characteristic two-step approach to equilibrium: 

initial fast decrease in solution concentration occurring over 30 minutes to 2 hours, followed by a slow, 

extended decrease in aqueous concentration (fig. 2). 

Sorptive equilibrium was attained after about 40 hours 

for naphthalene and phenanthrene and longer than 100 

hours for pyrene. The data was mathematically 

described using the two-site sorption kinetic model,

which is based on the assumption that a fraction of the 

total sorptive sites will be rate-limited.

The desorption-rate constants decreased by a 

factor of 1.5 to 2.5 as the temperature decreased for 

batch and column experiments, except for the batch 

experiments with naphthalene (table 1). At a given 

temperature, the rate constants decreased as the hydrophobicity of the chemical increased. The retarded 

intraparticle-diffusion model predicted the magnitude

668
of this effect because of increased Kd values and 

decreased aqueous diffusion. The intraorganic-matter 

diffusion model also predicted a decrease in desorption rate constants with a decrease in temperature as 

the result of a decrease in the polymer diffusion coefficients.

SUMMARY

Although the relative effects of organic-carbon 

sorption compared to mineral-phase sorption were not 

directly separated, it appears that the organic-carbon 

content, although very small, controls the rate and 

extent of sorption because the bonding forces are 

weak, the rate constants were much smaller than those 

reported for site-specific sorption, and the organiccarbon phase was responsible for most of the sorption.

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669
Perlinger, J.A., Eisenreich, S.J. and Capel, P.D., 1993,

Application of headspace analysis to the study of sorption of hydrophobic organic chemicals to alumina: 

Environmental Science and Technology, v. 27, 

p. 928-937.

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K., eds., Reactions and movement of organic chemicals in soils, SSSA Special Publication Number 22, 

p. 45-56.

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organic solvents from aqueous solution onto subsurface soils: Journal of Contaminant Hydrology, v. 4, 

p. 163-179.

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nonpolar organic compounds from surface waters to 

groundwater. Laboratory sorption studies: Environmental Science and Technology, v. 15, p. 1360-1367.

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Sorption of nonpolar organic chemicals on lowcarbon-content aquifer material: Environmental 

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low-molecular weight hydrophobic organic compounds on surface-modified silica: Journal of Contaminant Hydrology, v. 4, p. 181-203.

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670
Investigations of Microbial Weathering of Minerals

By Philip C. Bennett and Franz K. Hiebert 1

Abstract

Mineral weathering in an oil-contaminated 

aquifer near Bemidji, Minnesota, was investigated 

using in situ methods. Small microcosms 

containing clean, freshly fractured, mineral 

fragments were suspended for ~1 year in wells 

penetrating the aquifer, and the mineral surfaces 

were then examined for evidence of microbial 

colonization and mineral diagenesis. After 

reaction, feldspar and quartz surfaces were lightly 

colonized with a variety of microfauna. Feldspars 

were deeply etched in the vicinity of attached 

microbes, whereas an unidentified clay precipitate 

was found on uncolonized surfaces of the same 

fragment. Quartz was lightly etched, with a 

generally roughened surface and scattered 

triangular etch pits, whereas most of the calcite 

showed evidence of carbonate precipitation. 

These results, combined with 10 years of 

geochemical data collected at the site, illustrate a 

complex geochemical system where microbial 

processes drive multiple reactions that are 

reflected in the mineral-surface diagenetic 

textures.

INTRODUCTION

In 1979, a petroleum pipeline rupture near 

Bemidji, Minnesota, contaminated an unconfined sand 

and gravel aquifer (Hull, 1984). The hydrology and 

geochemistry of the site has been intensively studied 

for the past decade (Baedecker and others, 1993; 

Bennett and others, 1993; Eganhouse and others, 

1993). Although a wide variety of hydrologic and

'Department of Geological Sciences, The University of 

Texas, Austin, TX

geochemical processes has been documented, a 

recurrent finding is that microbes affect oil fate, 

ground-water geochemistry, and mineral diagenesis. 

Reported here is direct evidence for mineral 

weathering associated with the presence of oil 

degrading native microbes.

Study Area Physiography and Hydrogeology

The study site is a shallow sand and gravel 

aquifer near the small city of Bemidji, in the Bagley 

outwash plain (fig. 1). The aquifer material is typical 

for the region, consisting of outwash and morainal 

sands and gravels, with interbedded lacustrine silts 

(Franzi, 1988). The ground water in this 

heterogeneous and largely permeable aquifer flows 

generally northeastward through the site at a rate of 

0.05 to 1.0 m/d (White, 1991). Hydraulic conductivity 

ranges from 0.02 to 5 cm/s. The lacustrine silt 

deposits near the water table have the lowest conductivities. Locally, the aquifer is recharged by surface 

infiltration and discharges to a small lake northeast of 

the site (Bilir, 1992).

Ground-Water Geochemistry

Details of the inorganic ground-water chemistry 

have been summarized in Bennett and others (1993) 

and Baedecker and others (1993). Background water 

is a dilute Ca-Mg-HCO3 type, with minor 

concentrations of silica and trace concentrations of Na, 

K, and Sr. In the contaminated area, the water gains 

Ca, Mg, dissolved inorganic and organic carbon (DIC 

and DOC), silica, and iron by the aerobic and 

anaerobic degradation of oil followed by reaction with 

silicate and carbonate minerals. Downgradient 

oxygen is reintroduced, aerobic oil oxidation 

continues, and the water gradually returns to nearly 

background conditions (Baedecker and others, 1993).

671
 431 . Line of equal elevation 

Contour interval 1 meter

Microcosm Location '" Monitoring well with well number 

Railroad Staff

Lakehead Pipeline Company *". Piezometer nest with well number 

Site road

Figure 1 . Map of study-site area showing geologic and 

microcosm locations. (Modified from Bilir, 1992.)

Changes in chemical characteristics along the 

contaminated ground-water flow path suggest that 

several fundamental diagenetic reactions occur. In 

areas of aerobic degradation, increasing Ca, Mg, and 

DIG concentrations suggest that carbonates are 

dissolving, whereas increasing Fe2+ concentrations in 

anaerobic areas suggest that iron minerals are 

undergoing reductive dissolution (Baedecker and 

others, 1993) coupled with carbonate precipitation to 

buffer pH (Bennett and others, 1993). An increase in 

dissolved silica (from 20 to >60 mg/L) under the oil 

suggests that silicate minerals also are dissolving. 

Indirect evidence of quartz and feldspar dissolution

was obtained by SEM examination of native aquifer 

sands (Bennett and Siegel, 1987), but these studies 

assumed that the precontamination grain surfaces were 

pristine and not previously altered by other 

geochemical processes. The goal of the study reported 

here is to characterize mineral-water reactions by 

examining surface alteration of pristine mineral 

fragments.

METHODOLOGY

Our in-situ microcosms consist of clean mineral 

fragments within a specified ground-water flow path in 

order to characterize macroscopic and microscopic 

surface alteration during a known reaction period. We 

are examining a range of mineral types, including 

quartz, microcline, albite, anorthoclase, calcite, 

dolomite, biotite, and muscovite. Mineral specimens 

were crushed to a size range of 5 to 20 mm, and the 

selected fragments were cleaned by brief sonication in 

distilled water and were air-dried. Multiple 20 g 

monomineralic and mixed-mineral splits were 

prepared for use in field reaction, field background, 

and laboratory control experiments, as well as for 

permanent archive. In the most recent experiment, 

each mineral fragment was also scribed with a unique 

identification code, and weighed to the nearest 10 p,g, 

and the total surface area and microporosity of a 

complete microcosm fragment population were 

determined by standard krypton and argon-sorption 

methods, respectively.

After preparation and characterization, mineral 

fragments were placed in porous 15 mL polyethylene 

containers, and four to eight filled containers were 

assembled in a polyethylene mesh and sterilized. The 

sterile microcosms were suspended in the screened 

part of the aquifer by nylon monofilament, and the 

well bore was purged with nitrogen and sealed for a 

period of 10 to 14 months (fig. 2). Laboratory control 

experiments were divided into an archival fraction and 

a fraction that was immersed in distilled water. After 

recovery the microcosms were preserved in 2 percent 

gluteraldehyde at 2 °C until analysis; similar 

procedures using reserved control experiments did not 

result in detectable surface-texture artifacts.

In the laboratory, selected fragments were 

prepared for scanning electron microscope 

examination by dehydration and critical-point drying 

before stub-mounting and gold-coating. The attached 

microbial population was examined for density,

672
METERS 

440-,

430-

420-

410-

310

Land Surface

y 4p <a go loo METEI

1 80 loO'ZJO'a* FEET

Oround Water Row

Sand and 

gravel aquifer

Base of Aquifer

DATUM IS SEA LEVEL VERTICAL EXAGGERATION x4

EXPLANATION

310

Well and identifier, showing 

center of screened interval Location of microcosm

. Approximate location of oil pool

Figure 2. Hydrologic section showing microcosm placement in a well bore.

attachment, and morphology, and the mineral surface 

was examined for evidence of biogeochemical 

precipitation and dissolution. A subset of mineral 

fragments was sonicated to remove adherring material 

and was similarly examined for weathering features. 

The geochemistry of the silicate surfaces was 

examined by SEM with energy dispersive analysis of 

X-rays and by sequential leaching techniques (Berndt, 

1987). In particular, organic carbon coatings, 

carbonates, and iron precipitates were examined. The 

current experimental series will also characterize mass 

loss and gain caused by dissolution and precipitation 

reactions, and changes in microporosity.

Quartz

The quartz fragments from the control 

experiments consist of clean fractured surfaces with 

fine ridges and steps (fig. 3a). No evidence of 

chemical weathering or remnant inclusions is visible. 

After reaction in the contaminated aquifer, many of the 

surfaces had been colonized by a variety of bacteria 

seen as individual organisms rather than as biofilms. 

Much of the visible quartz surface is subtly roughened 

by extremely fine pitting or more extensively etched in 

some areas to produce distinctive triangular etch pits 

(fig. 3b).

RESULTS TO DATE

Three years of data have been collected from a 

total of 24 field-reaction microcosms. Mineral 

surfaces were characterized for microbial 

colonization, microfauna morphology and attachment, 

surface chemistry, and diagenetic textures. 

Preliminary leaching experiments have also been 

completed to characterize the chemistry of the surface 

precipitates. The results to date for three mineral 

groups are summarized here.

Feldspars

The control feldspar surfaces are clean and 

unweathered, with evidence of linear steps along ; 

cleavage and very small ellipsoidal inclusion pits 

(fig. 4a). As with quartz, some of the reacted feldspar 

surfaces were colonized, while other surfaces 

remained uncolonized. The colonized surfaces show 

evidence of intense chemical weathering, with deep 

prismatic etch pits oriented along inferred cleavage 

planes (fig. 4b). The pits are several microns wide.

673
and over 1 u,m deep, with irregular edges. Little 

evidence of clay mineral precipitation is seen around 

the etched areas except for thin strands of an unknown 

amorphous material (fig. 4b). Uncolonized surfaces 

on the same grain, however, are unetched, and support 

extensive aluminous clay precipitates (fig. 4c).

Calcite.

Calcite fragments used in the microcosm 

experiments were prepared in a manner similar to the 

silicates, except that the duration of the water rinses 

was much shorter. The clean rhombohedral crystal 

fragments exhibit flat surfaces with regular 100 to

Figure 3. Scanning electron microscope micrographs of 

quartz fragments, (a) Clean quartz fragment from control 

group; (b) etched quartz fragment from the field reaction 

group, showing roughened surface and triangular etch pits.

Figure 4. Scanning electron microscope micrographs of 

microcline fragments, (a) Clean microcline surface from 

control group showing cleavage and remnant inclusion 

features; (b) etched feldspar surface from field- reaction 

group; (c) unetched surface from field-reaction group 

showing precipitated kaolins.

674
500 nm high steps where cleavage planes have been 

fractured during crushing (fig. 5a). After reaction for 

~1 year, only very sparse microbial colonization was 

found, and the amount of colonization was much less 

than that on the silicate surfaces. Where bacterial 

colonization was observed, etching of the mineral 

surface was evident, and was characterized by deep 

irregular prismatic dissolution pits along cleavage 

planes (fig. 5b).

The most common surface alteration of calcite was 

precipitation of calcite overgrowths. Growth of 500- nmthick tabular calcite on the previously clean surfaces was 

recorded on almost all crystal faces examined. A 

hypothesized crystal-growth progression begins with 

100- to 500-n-high individual spikes extending upward 

from a flat background surface (fig. 5c). The spikes form 

straight rows and rhombohedral enclosures uniformly 

1000 nm across. The open enclosures of calcite spikes 

fill to form isolated plateaus 500 nm high that apparently 

coalesce into sheets of smooth calcite, thus adding a new 

layer to the original calcite grain (fig. 5c).

DISCUSSION

Previous research into silicate and carbonate 

weathering at the Bemidji site relied on geochemical 

characterization of ground water and examination of 

native sand grains collected from contaminated and 

pristine locations (Bennett and others, 1993; Baedecker 

and others, 1993; Bennett and Siegel, 1987). This 

approach suggested that silicates are weathering, but the 

rate and extent of dissolution of a specific silicate type is 

obscured by lack of knowledge of the starting condition 

of the mineral surface. Similarly, the carbon massbalance calculations suggest that a carbonate mineral 

may be precipitating in the anoxic region under the oil 

(see, for example, Baedecker and others, 1993), but the 

composition of the mineral and the rate of precipitation 

were unclear. In both cases, the role of microbes was 

unknown and initially was thought to be primarily as net 

producers of carbonic and organic acids released directly 

into the bulk ground water.

The experimental results described here suggest 

that microbes in this aquifer colonize fresh mineral 

surfaces, resulting in a non-uniform pattern of attached 

biologic material. In the vicinity of attached microbes, 

feldspar and quartz dissolve, even in waters greatly 

supersaturated with respect to quartz. We hypothesize 

that attached microbes on the surface create a 

microreaction zone, where extremely high

Figure 5. Scanning electron microscope micrographs of 

calcite fragments, (a) Clean calcite surface from control 

group showing 500 nm cleavage steps; (b) calcite surface 

from field-reaction group showing etch- pitting associated 

with attached bacteria; (c) precipitated calcite overgrowth 

showing inferred progression of growth from 1 to 5: 1. 

initial calcite spike; 2. linear group of spikes forming 

enclosure; 3. fully formed enclosure consisting of individual 

spikes; 4. filled enclosure forming 500-nm-high plateau; 5. 

coalesced plateaus forming a nascent 500-nm-high step.

675
concentrations of extracellular organic ligands 

accumulate (Hiebert and Bennett, 1992). In this small 

area, silicate weathering is greatly enhanced, 

increasing both the dissolution rate due to ligandpromoted mechanisms and the solubility due to 

solution complexes. The rate of feldspar dissolution, 

inferred from an estimate of mineral volume removed, 

is many times faster than that predicted from inorganic 

experimental systems (Bennett and others, 1991; 

Hiebert and Bennett, 1992).

Uncolonized silicate surfaces, in contrast, are 

unweathered or show evidence of clay precipitation. 

Away from the microbial microreaction zone, the 

reactive ligands and chelated products of dissolution 

mix with the bulk aquifer water, possibly shifting 

equilibria from that established at the dissolving 

surface. This water may be greatly supersaturated 

with respect to kaolin-group minerals, thus 

precipitating aluminum as a kaolinite or halloysite and 

leaving excess silica in solution. This hypothesis is 

supported by the chemical composition of the ground 

water, where silica concentration increases greatly, but 

aluminum concentration never exceeds 2 \\M (Bennett 

and others, 1993). We are currently investigating the 

possibility that microbes preferentially colonize the 

upstream side of a fragment, where microfauna will 

first attach and where nutrients are supplied by flowing 

ground water.

In the carbonate system, the microbial use of 

iron oxide as an electron acceptor for carbon 

metabolism may shift the carbonate geochemistry 

toward calcite supersaturation by increasing pH and 

producing bicarbonate (Bennett and others, 1993). 

Calcite precipitates, forming the distinctive 

overgrowth morphologies seen on the cleaved calcite 

fragments. The form of the carbonate precipitate and 

results of ED AX analysis of the individual spikes 

suggest that the calcite is nearly pure, with no 

detectable iron or magnesium. We are presently 

investigating the controls on the geometry of calcite 

overgrowth spikes, and overgrowth step height.

CONCLUSIONS

This ongoing study of microbial processes in an 

oil-contaminated aquifer suggests that native 

microfauna have a significant effect on mineralweathering mechanisms and rate. Where microbes 

attach to silicate surfaces, microenvironments 

produced by metabolic processes greatly enhance 

silicate dissolution. The rate and extent of dissolution

cannot be evaluated from the bulk chemical 

composition of the water, because this is a dilute 

reflection of the environment immediately adjacent to 

a metabolizing microbe. The nature of this 

microenvironment and the types and concentrations of 

reactive ligands are expected to be affected by the 

biogeochemical environment, the nature of the carbon 

substrate, and the availability of nutrients. In the 

nutrient- poor, anoxic, and carbon-rich environment of 

the Bemidji aquifer, the microbial microenvironment 

apparently is very aggressive toward silicate 

minerals. The carbonate system, in contrast, is 

apparently perturbed by the overall biogeochemical 

reactions that consume the carbon in petroleum, 

resulting in calcite precipitation.

The in situ microcosm approach allows for the 

direct examination of the form and rate of both silicate 

dissolution and calcite precipitation in lowtemperature systems using well-characterized starting 

surfaces. By combining surface-texture examination 

with long-term geochemical characterization, rockwater reactions can be demonstrated directly .

REFERENCES

Baedecker, M.J., Cozzarelli, I.M., Bennett P.C., and Siegel, 

D.I., 1993, The fate of crude oil in a sand and gravel 

aquifer. III Geochemical processes: Applied 

Geochemistry, v. 8.

Bennett, P.C., Siegel, D.I., Baedecker M.J., Cozzarelli, I.M., 

and Hull, M., 1993, The fate of crude oil in a sand and 

gravel aquifer, I Inorganic geochemistry: Applied 

Geochemistry, v. 8.

Bennett, P., and Siegel, D.I., 1987, Increased solubility of 

quartz in water due to complexation by dissolved 

organic compounds: Nature, v. 326, p. 684-687.

Berndt, M.P., 1987, Metal partitioning in a sand and gravel 

aquifer contaminated by crude oil: Syracuse, N.Y.. 

Syracuse University, M.S. Thesis, 84 p.

Bilir, S. II, 1992, Hydrogeologic characterization of a

glacial aquifer contaminated by crude oil near Bemidji, 

Minnesota: Austin, Tex., The University of Texas at 

Austin, 140 p.

Eganhouse, R.P., Baedecker, M.J., Cozzarelli, I.M., Aiken, 

G.R., Thorn, K.A., and Dorsey, T.F., Crude oil in a 

shallow sand and gravel aquifer, II Organic 

geochemistry: Applied Geochemistry, v. 8.

676
Franzi, D.A., 1988, Sedimentary and post-depositional 

processes related to aquifer properties at the Bemidji 

Research Site, North-Central Minnesota: in Franks, 

B.J., ed., U.S. Geological Survey Program on Toxic 

Waste Ground-Water Contamination Proceedings of 

the third technical meeting, Pensacola, Florida, March 

23-27, 1987, U.S. Geological Survey Water-Resources 

Investigative Report. 88-4220, p. C7-C8.

Hiebert, F.K., and Bennett, P.C., 1992, Microbial control of 

silicate weathering in organic-rich ground water: 

Science, v. 258, p. 278-281.

Hult, M.F., 1984, Ground-water contamination by crude oil 

at the Bemidji, Minnesota Research Site an introduction: in Ground-Water Contamination by Crude 

Oil at the Bemidji, Minnesota, Research Site: U.S. 

Geological Survey Toxic Waste Ground-Water 

Contamination Study Proceedings of the technical 

meeting, Tucson, Ariz., March 20-22, 1984, U.S. 

Geological Survey Water-Resources Investigative 

Report. 84-4188, p. 1-15.

White, D.L., 1991, Point Dilution Method Determination of 

Ground Water Velocities at the Bemidji Research 

Site: Austin, Tex., The University of Texas at Austin, 

M.S. Thesis, 142 p.

677
678
Multiphase-Flow Modeling of the Bemidji, Minnesota 

Crude-Oil-Spill Site with Geostatistical Simulation of

Hydraulic Properties

By LA. Dillard 1 ' 2 , H.I. Essaid2 , andW.N. Herkelrath2

Abstract

This paper describes a study aimed at 

improving the understanding of multiphase flow 

in a heterogeneous aquifer by using data collected 

at the north pool sub-region of a crude-oil-spill 

site near Bemidji, Minnesota. Data collected 

include grain-size distributions, oil and water saturations , and porosities of samples collected at 

the north pool site. These data are used to estimate permeability (k), retention curves, and the 

geometric mean and standard deviation of the 

sampled grain-size distributions. The k estimates 

are statistically distributed in a bimodal lognormal fashion, whereby the two k population 

distributions correspond to the two predominant 

aquifer lithologies at the north pool: a coarse 

glacial-outwash deposit and interbedded fine-silt 

lenses. A dual geostatistical approach involving 

sequential indicator and Gaussian simulation is 

used to characterize the bimodal aquifer heterogeneity present at the north pool by creating one 

realization of a three-dimensional grid of log(k) 

values conditioned on the original data.

Field oil-saturation data indicate that the 

spatial distribution of oil in the subsurface has 

been affected by the bimodal aquifer heterogeneity present at the north pool. The shape of

Department of Applied Earth Sciences, Stanford University, 

Stanford, CA, 94305-2225.

2U.S. Geological Survey, 435 Middlefield Rd., Menlo Park, 

CA, 94025.

Fluid saturation refers to the volume of the fluid phase relative to the volume of pores in the sample.

the subsurface oil-saturation distribution is highly 

irregular; more specifically, the center of the oil 

lens is actually depressed below the water table in 

some locations, and oil saturations almost as high 

as 0.3 occur in the unsaturated zone. A crosssectional numerical model is used to simulate the 

flow of oil and water at the north pool, assuming 

the air phase is at atmospheric pressure. A slice 

of the three-dimensional grid of log(k) values produced through geostatistical simulation is used as 

input in the multiphase model to investigate the 

effect of aquifer heterogeneity on multiphase 

flow. Flow simulations using the bimodal representation of heterogeneity produce an oilsaturation distribution that is similar to that found 

in the field. Other factors affecting multiphase 

flow, such as hysteresis and three-phase oil relative permeability, also are investigated. The 

effects of hysteresis are not as significant when 

heterogeneity is adequately characterized.

INTRODUCTION

Contamination of ground water by nonaqueousphase liquids (NAPL's) is a problem that has become 

increasingly prevalent in hydrogeology. Such NAPL's 

can range from synthetic organic compounds such as 

trichloroethylene (TCE), to natural and refined hydrocarbons such as crude oil. A limited amount of work 

has focused on the effects of spatial variability on the 

movement of NAPL's in the subsurface. Previous 

work includes one- and two-dimensional laboratory 

experiments on the distribution and movement of oil 

(Schwille, 1984, 1988; Schiegg and McBride, 1987; 

Kueper and others, 1989), and numerical simulations

679
of two-phase flow (Kueper and Frind, 199 la, b). Laboratory experiments are unrealistic representations of 

field-scale processes. Furthermore, because NAPL 

contamination problems often involve infiltration of 

NAPL and water into the unsaturated zone, numerical 

simulations should include the three-phase flow of air, 

water, and NAPL. Petroleum engineers have shown 

that reservoir spatial variability can have a significant 

effect on multiphase flow (Haldorsen and Chang, 

1986; Lasseter and others, 1986). However, because 

hydrogeologists work at a more shallow and local 

scale, small-scale heterogeneities that can be overlooked at the reservoir scale need to be incorporated 

into the hydrogeologist's model.

The extensive fluid saturation and sediment data 

set acquired at the Bemidji crude-oil-spill site 

prompted an investigation of the effect of aquifer heterogeneity on three-phase flow in the field. A portion 

of these data was used in previous studies of the effect 

of heterogeneity on multiphase flow (Essaid and others, 1993; Essaid and Hess, 1993; Essaid and others, 

1991). However, these previous studies used data 

from the south pool sub-region of the Bemidji field 

site where the aquifer does not display the complex 

bimodal heterogeneity present at the location of the 

present study, the north pool. Consequently, the geostatistical simulation techniques used in the previous 

south pool studies to characterize aquifer heterogeneity 

do not apply to the more complex heterogeneity analyzed in the present north pool study.

The purpose of the north pool study is to explore 

and challenge the results of the previous south pool 

study. The current north pool study offers a more 

detailed examination of the effects of aquifer heterogeneity on multiphase flow, and it explores the significance of other factors affecting the migration of oil, 

water, and air in the subsurface. This paper briefly 

describes the contamination history of the entire 

Bemidji site and the north pool data used in the current 

study. A numerical aquifer that represents the bimodal 

heterogeneity present at the north pool is generated 

using geostatistical simulation. This numerical aquifer 

is used in multiphase flow simulations to ascertain the 

effect of heterogeneity on the flow of water and oil in 

the presence of air in the subsurface. Other factors 

affecting multiphase flow in a heterogeneous aquifer, 

such as hysteresis and three-phase oil relative permeability, are also investigated.

SITE DESCRIPTION

In August 1979, a submerged pipeline ruptured 

and spilled crude oil at a site near Bemidji, Minnesota. 

The topography of the Bemidji site consists of mostly 

flat terrain with gently undulating hills; most of the 

area is underlain by a coarse glacial outwash deposited 

during the most recent glaciation. Because the crude 

oil was under pressure when it spilled, much of the oil 

was sprayed over the area. The crude oil eventually 

settled, either by flowing from the point of rupture or 

by being sprayed, in two topographic depressions, 

termed the north and south pools. Despite the proximity of the north and south pools at the site, the local 

geology at each is quite different. The south pool is 

underlain almost exclusively by coarse glacial outwash, whereas the north pool is underlain by glacial 

outwash with interbedded silt lenses (Franzi, 1988). 

Previous studies focused on data from the south pool; 

the focus of the present study is on the more heterogeneous north pool.

Since the time of the pipeline rupture, numerous 

boreholes have been drilled at the entire site. For the 

purpose of this study, data from 613 core samples collected from the north pool are analyzed. The cores 

were obtained from 20 boreholes aligned along a 90- 

m-long transect approximately parallel to the direction 

of groundwater flow (fig. 1). Some (269) of the core 

samples were collected by the authors and are 7.8 cm 

in length. The remainder (344) of the core samples 

were collected by Dave Franzi and are variable in 

length (Dave Franzi, State University of New York, 

Platsburg, written communs., 1989-93). Grain-size 

distributions were determined for all 613 core samples, and oil and water saturations and porosities were 

determined for the subset of 269 core samples using 

the methods of Hess and others (1992). Following the 

procedure of Krumbein and Monk (1942), permeabilities were estimated for all 613 core samples on the 

basis of the grain-size distribution data. Soil moisture 

retention characteristics were estimated for the subset 

of core samples by the method of Arya and Paris 

(1981), and retention curves were fitted to the estimated points by means of a van Genuchten function 

(van Genuchten, 1980).

Fluid saturation and grain-size distribution data 

suggest that the north pool aquifer heterogeneity has a 

significant effect on the flow of crude oil, water and air 

at the site. Measured oil saturations in the unsaturated

680
95°10'

47-35'

47°30'

direction of 

ground-water flow

0 30 60 rte~i

EXPLANATION

WATER BODIES

mm MUNICIPAL AREAS

 430- TOPOGRAPHIC CONTOUR, IN METERS

A A' NORTH POOL TRANSECT 

000

BOREHOLE USED IN THIS STUDY AND 1DENTIFIE

Figure 1. Location of field site near Bemidji, Minnesota. Inset corresponds to the north pool subregion of the field site.

zone are as high as 0.3 (fig. 2). Silt lenses in the unsaturated zone appear to have impeded oil infiltration and 

trapped this oil. Silt lenses at the water table also 

appear to have an effect on the spatial distribution of 

the floating oil and lateral oil migration. Profiles of 

the following sampled data from borehole 016 illustrate the effect the fine fraction of sediment has on 

multiphase flow (fig. 3): percent grain size less than 1 

mm, oil saturation, and water and total fluid saturation. 

Note that in borehole 016 the peak oil saturation is 

below the water table. Apparently, the oil follows a 

high permeability pathway that is located beneath a 

silt lens.

GEOSTATISTICAL MODELING of 

AQUIFER HETEROGENEITY

Summary Statistics

The statistical distributions of permeability (k), 

van Genuchten retention curve parameters (a and n), 

geometric mean (gmean) and standard deviation of 

the grain-size distribution (gstd), and porosity, are 

analyzed using probability plots and histograms. 

Permeability is the most useful parameter for characterizing aquifer heterogeneity, because it is distributed in a lognormal bimodal fashion with the two

681
435

DATUM IS SEA LEVEL

VERTICAL EXAGGERATION IS X 4.5

60 FEET

EXPLANATION ° 

THICKNESS OF OIL LAYER IN WELLS

,ro 20 METERS

015

OIL SATURATION SAMPLE 

OBSERVATION WELL AND IDENTIFIER

1 QUESTION MARK INDICATES DEPTH TO BOTTOM OF WELL 

? IS UNKNOWN

-0.2- LINE OF EQUAL OIL SATURATION 

CONTOUR INTERVAL IS VARIABLE

Figure 2. Hydrogeologic section of subsurface showing oil saturation distribution as determined through sampling at the north 

pool site. Line of section shown in figure 1.

682
Borehole 016 

Percent Fine Grain Size

Borehole 016 

Oil Saturation

430

420

430

o

420

0 percent 100

Borehole 016

Water and Total Fluid

Saturation

430

£

-o

420

0 saturation 1 0 saturation 1

Figure 3. Profiles of percent fine grain size, oil saturation, and water and total fluid saturation for borehole 016.

population distributions corresponding to the coarse 

glacial outwash and the fine-silt lenses. A cut-off of 

log(k)=-l 1.957 [log(m2)] is used to separate the coarse 

and fine population distributions.

Variogram Analysis

The variogram can be used as a tool by which 

the degree of spatial continuity of a random function 

can be quantified. The following variograms are calculated and modeled in the vertical and horizontal 

directions: variograms that use an indicator transform 

of log(k) to estimate the probability that log(k) will be 

less than or equal to the specified cut-off (indicator 

variograms); and variograms of coarse log(k) that 

have been transformed to fit a normal distribution 

(normal scores variograms). The indicator variograms 

provide a measure of the spatial correlation of the indicator transform of log(k) for the specified cut-off 

(Johnson and Dreiss, 1989). In other words, the indicator variogram provides a measure of the connectivity of the highs and lows, that is, the coarse glacial 

outwash and the fine silt lenses, respectively. On the 

other hand, the variograms of the normal scores transform of coarse log(k) provide a measure of variability 

within the coarse glacial outwash. Variograms are 

modeled by visually fitting a combination of nested

exponential and spherical structures. Such a combination of nested models is needed to capture the rapid 

increase in variance at low separation distances or 

lags, and to represent the relatively slower increase in 

variance to the sill at larger lags. A summary of the 

correlation structure of each variogram is in table 1. 

The models for the indicator and normal scores variograms are used in sequential stochastic simulation.

Sequential Stochastic Simulation

Kriging is a linear estimation procedure through 

which values of a random variable at unsampled 

points are calculated, given measured values of the 

random variable at sampled points. When kriging is 

combined with a multi-Gaussian or nonparametric 

(indicator) estimation approach, it not only provides a 

linear estimator, but it also identifies the posterior conditional cumulative distribution function (ccdf) (Deutsch, 1992). The posterior ccdf is an integral part of the 

sequential simulation process. Furthermore, if a different random path of estimation is taken in each 

sequential simulation, a series of equally-probable yet 

different results can be achieved. The uncertainty that 

can be quantified upon accumulation of such results 

characterizes stochastic simulation. Two such

683
Table 1. Correlation structure used to model indicator variograms of log(k) and variograms of normal scores transform (NSC) 

of coarse fraction of log(k)

[Note: y denotes the semivariogram, exp denotes an exponential model, and Sph denotes 

a spherical model]

Indicator of log(k)

vertical 

horizontal

y(h) = 0.04 + 0.04*exp (2h) + 0.01735*Sph

y(h) = 0.04 + 0.04*exp L + 0.01735*Sph L

NSC coarse fraction of log(k)

vertical

horizontal

y (h) = 0.229 + 0.687*exp (h ) + 0.084*Sph

y(h) = 0.229 + 0.687*exp f ^-} + 0.084*Sph f

geostatistical estimation procedures are sequential 

Gaussian simulation (GSS) and sequential indicator 

simulation (ISS). The correlation structure described 

by the indicator and normal scores variograms are 

used in ISS and GSS respectively to construct a permeability distribution conditioned on the original data. 

Only one realization is generated because the intent of 

the geostatistical estimation procedure is to produce 

not a host of equally probable results, but one realization that demonstrates detailed aquifer heterogeneity 

and can be used in flow simulations.

In geostatistical estimation procedures, if the 

phenomenon under study is a mixture of different 

physical or statistical populations, the geometry of the 

mixture should be modeled and simulated first; then 

the attribute within each population can be simulated 

(Deutsch and Journel, 1992; Journel and Isaaks, 1984). 

The bimodal north pool aquifer heterogeneity is reproduced by using a dual estimation approach to generate 

a three-dimensional grid of log(k) values conditioned 

on the original data. ISS is used to describe the structure of the two log(k) classes representing the fine silt

and coarse sand population distributions. GSS is used 

on the normal scores transform of the coarse fraction 

of the original data to reproduce the spatial variability 

within the coarse outwash deposit. The GSS approach 

is not used on the fine population distribution due to 

the limited amount of conditioning data available representing fine material. The final distribution is 

obtained by substituting GSS values for ISS values at 

each node where the ISS value is greater than the specified log(k) cut-off, indicating the presence of coarsegrained sand deposits. This combined k distribution 

(IGSIM) not only shows the structure of the fine silt 

and coarse sand deposits, but also the variability 

within the coarse sand.

A vertical slice of the three-dimensional simulation result is then taken along an extended 180-m-long 

transect at which the original boreholes are located. 

Examples of one realization each of GSS and IGSIM 

can be found in figures 4b and 4c. These slices of 

GSS and IGSIM of log(k) values are used as input into 

the cross-sectional numerical multiphase flow model.

684
A.

constant oil pressure head 

oil infiltration nodes

Uniform mean 

= -11.28[log(m2)]

oil saturation

1.00

B.

log(k) nog(m2)] 433.0

m

m -10.000

Hydrostatic water-pressure boundary'

-12.400

C.

0.00

oil saturation 

1.00

415.0

DISTANCE, IN METERS

119.0

DISTANCE, IN METERS

Iog(k) [log(mz)] 433.0 

-10.000

415.0

0.00

oil saturation 

1.00

DISTANCE, IN METERS

420.0* 

180.0 29.0

0.00

119.0

DISTANCE, IN METERS

Figure 4. Sections showing permeability distributions (left) and resulting simulated oil saturation distributions (right) for 

conditions of (A) uniform mean permeability field, (B) coarse permeability field generated using sequential Gaussian 

simulation (GSS), and (C) bimodal permeability field generated using joint sequential indicator and Gaussian simulation 

(IGSIM). Datum is sea level and origin for horizontal distance is borehole 707 as shown in figure 1.

685
NUMERICAL MULTIPHASE-FLOW 

MODELING

The objectives of the numerical modeling are 

threefold. The first goal is to investigate the appropriateness of the assumptions used in the model. The 

second objective is to produce a simulated oil saturation spatial distribution similar to the observed oil saturation distribution. For this purpose, the oil 

saturation data are used not as a matching device but 

as a clue to the subsurface spatial structure of the oil 

body. Of course, given the amount of uncertainty in 

model parameters and approximations, the simulated 

oil saturation distribution will never exactly match the 

observed oil saturation data. Therefore, the third goal 

of the study is to conduct a sensitivity analysis of the 

factors most affecting multiphase flow and the resulting oil saturation distribution.

A vertical, two-dimensional, numerical model 

that simulates the flow of oil and water is used (Essaid 

and others, 1993). The model is based on the assumption of atmospheric air pressure and includes spatial 

variability of soil hydraulic properties using Miller 

and Miller scaling (Miller and Miller, 1956) and hysteresis with oil entrapment. The numerical model is 

used to simulate oil movement along the 90-m-long 

north pool transect from the time of the spill in August 

1979 to the approximate time at which core samples 

were collected in June 1990. The initial condition for 

the simulation is a hydrostatic water pressure distribution corresponding to a water table altitude computed 

from water level data collected during sampling from 

wells 604b and 534b (423.67 m and 423.46 m respectively) in June 1990. The boundary conditions are 

labeled in figure 4a. The left, bottom, and right boundaries of the domain are assumed to be constant hydrostatic water-pressure boundaries and the oil pressure 

head is assumed to be 0 m at five infiltration nodes at 

the top boundary until oil infiltration ceases. Oil infiltration ceases when cumulative oil infiltration is equal 

to the actual oil mass of 11513 kg as estimated from 

the measured oil saturation distribution. The oil saturation spatial distributions produced through numerical simulation are visually compared to the observed 

oil saturation data to determine the most effective 

modeling scheme.

The effect of heterogeneity on multiphase flow 

is examined by doing simulations using the following 

permeability distributions: a uniform mean k 

(UNIFK); coarse k generated through GSS; and bimodal k produced through IGSIM (figs. 4a, 4b, and 4c).

The simulation using UNIFK produces an oil saturation distribution reminiscent of the typical vision of a 

subsurface oil lens floating on the water table, where 

the highest oil saturation is located in the middle of the 

oil body. None of the characteristics of the actual oil 

saturation distribution that reflect aquifer heterogeneity, such as the high oil saturations in the vadose zone 

and below the water table, and the irregular oil lens, 

are reproduced. The simulation using a GSS realization produces a generally uniform oil saturation distribution with oil saturations in the unsaturated zone as 

high as 0.1 and some variability in the floating oil lens 

because the heterogeneity within the coarse population 

is included. Compared to the simulations using 

UNIFK and GSS permeability distributions, the simulation using an IGSIM realization creates an oil saturation distribution that is most similar to the actual data. 

The oil lens produced through the IGSIM simulation is 

highly irregular, with fingering and oil saturations 

below the water table as high as 0.9. Oil saturations as 

high as 0.3 are reproduced in the unsaturated zone; 

however, these high values are not as prevalent as in 

the data. Due to discrepancies between observed and 

simulated oil saturation distributions, other factors 

affecting multiphase flow are being investigated in an 

effort to more closely reproduce the observed oil distribution.

Two simulations with hysteresis were done 

using UNIFK and IGSIM k distributions, respectively. 

Hysteresis has a significant effect on the final oil saturation distribution in the simulation using uniform 

mean k, but virtually no effect in the simulation using 

the heterogeneous distribution of k produced through 

IGSIM. Hysteresis, therefore, is important when aquifer heterogeneity is not properly characterized; the 

small scale effects of hysteresis are eclipsed when 

aquifer heterogeneity is adequately represented.

Work in progress involves an investigation of 

the role of relative permeability in multiphase flow. 

Simulations are being conducted using the IGSIM k 

distribution, no hysteresis, recharge, and two different 

relative permeability relations: Parker (Parker, 1987) 

and Stone II (Stone, 1973). Preliminary results demonstrate the sensitivity of multiphase flow to relative 

permeability and reinforce the uncertainty surrounding 

relative permeability relations and their uniqueness to 

every individual scenario. The use of the Bemidji field 

data will lend additional insight on these issues.

686
SUMMARY

Grain-size distribution data and permeability 

estimates from the Bemidji, Minnesota crude-oil-spill 

site indicate the presence of a bimodal heterogeneity; 

the two classes correspond to a coarse glacial-outwash 

sand and interbedded fine-silt lenses. Oil saturation 

data reveal that this heterogeneity has a significant 

effect on fluid flow in the subsurface. Geostatistical 

simulation is conducted in two stages to create a threedimensional grid of permeabilities that represents the 

bimodal heterogeneity. A slice of this grid is used as 

input in a cross-sectional multiphase flow model to 

analyze the effect of heterogeneity on multiphase flow 

at the field scale. Three simulations are conducted 

using increasingly more variable permeability distributions (mean permeability, coarse permeability generated using sequential Gaussian simulation, and 

bimodal permeability generated using dual sequential 

indicator and Gaussian simulation). Results of flow 

simulations using the bimodal permeability distribution generated using dual sequential indicator and 

Gaussian simulation produce oil saturations that most 

closely represent the observed oil saturation distribution. The small scale effects of hysteresis are shadowed by the effect of heterogeneity on the oil 

saturation distribution. The significance of relative 

permeability in the multiphase flow process at the field 

scale is currently being investigated.

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to predict the soil moisture characteristic from particlesize distribution and bulk density data: Soil Science 

Society of America Journal, v. 45, p. 1023-1030.

Deutch, C.V., 1992, Annealing Techniques Applied to 

Reservoir Modeling and the Integration of Geological 

and Engineering (well test) Data: Stanford Ph.D. Dissertation, Stanford University, 306 p.

Deutch, C.V., and Journel, A.G., 1992, GSLIB: Geostatistical Software Library and User's Guide: New York, 

Oxford University Press, 340 p.

Essaid, H.I., Herkelrath, W.N., and Hess, K.M., 1991, Air, 

oil, and water distributions at a crude-oil spill site, 

Bemidji, Minnesota, in Mallard, G.E., and Aronson, 

D.A., eds, U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, Calif, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 614-620.

Essaid, H.I., and Hess, K.M, 1993, Monte Carlo simulations 

of multiphase flow incorporating spatial variability of 

hydraulic properties: Ground Water, v. 31, no. 1, 

p. 123-134.

Essaid, H.I., Herkelrath, W.N., and Hess, K.M., 1993, Simulation of fluid distributions observed at a crude-oil spill 

site incorporating hysteresis, oil entrapment, and 

spatial variability of hydraulic properties: Water 

Resources Research, v. 29, no. 6, p. 1753-1770.

Franzi, D.A., 1988, Surficial and subsurface distribution of 

aquifer sediments at the Bemidji, Minnesota, research 

site, in Ragone, S.E., ed., U.S. Geological Survey program on toxic waste-ground-water contamination, Proceedings of the second technical meeting, Cape Cod, 

Massachusetts, October 21-25, 1985: U.S. Geological 

Survey Open-file Report 86-481, c5-c 10.

Haldorsen, H.E., and Chang, D.M., 1986, Notes on stochastic shales: from outcrop to simulation model, in Lake, 

L.W. and Carroll, H.B., eds., Reservoir characterization: Orlando, Florida, Academic Press, Inc., p. 445- 

485.

Hess, K.M., Herkelrath, W.N., and Essaid, H.I., 1992,

Determination of subsurface fluid contents at a crudeoil spill site: Journal of Contaminant Hydrology, v. 10, 

p. 75-96.

Johnson, N.M., and Dreiss, S.J., 1989, Hydrostratigraphic 

interpretation using indicator geostatistics: Water 

Resources Research, v. 25, no. 12, p. 2501-2510.

Journel, A.G., and Isaaks, E.H., 1984, Conditional indicator 

simulation: application to a Saskatchewan uranium 

deposit: Mathematical Geology, v. 16, no. 7, p. 685- 

718.

Kueper, B.H., Abbot, W., and Farquhar, G., 1989, Experimental observations of multiphase flow in heterogeneous porous media: Journal of Contaminant 

Hydrology, v. 5, p. 83-95.

Kueper, B.H., and Frind, E.O., 1991a, Two-phase flow in 

heterogeneous porous media, 1 Model development: 

Water Resources Research, v. 27, no. 6, p. 1049-1057.

Kueper, B.H. and Frind, E.O., 1991b, Two-phase flow in 

heterogeneous porous media, 2 Model application: 

Water Resources Research, v. 27, no. 6, p. 1059-1070.

Krumbein, W.C., and Monk, G.D., 1942, Permeability as a 

function of the size parameters of unconsolidated sand: 

Transactions of the American Institute of Mining and 

Metallurgical Engineers, v. 151, p. 153-160.

Lasseter, T.J., Waggoner, J.R., and Lake, L.W., 1986, Reservoir heterogeneities and their influence on ultimate 

recovery, in Lake, L.W. and Carroll, H.B., eds., Reservoir characterization: Orlando, Florida, Academic 

Press, Inc., p. 545-559.

Miller, E.E, and Miller, R.D., 1956, Physical theory of capillary flow phenomena: Journal of Applied Physics, 

v. 27, p. 325-332.

687
Parker, J.C., Lenhard, R.J., and Kuppusamy, T., 1987, A 

parametric model for constitutive properties governing 

multiphase flow in porous media: Water Resources 

Research, v. 24, no. 4, p. 618-624.

Schiegg, H.O., and McBride, J.F., 1987, Laboratory setup to 

study two-dimensional multiphase flow in porous 

media, in Proceedings of the NWWA/API Conference 

on Petroleum Hydrocarbons and Organic Chemicals in 

Ground Water: Prevention, Detection and Restoration: 

Dublin, OH, National Water Well Association, p. 371- 

395.

Schwille, F., 1984, Migration of organic fluids immiscible 

with water in the unsaturated zone, in Yaron, B.,

Dagan, G., and Goldshmid, J. (eds.), Pollutants in 

porous media: New York, Springer-Verlag, p. 27-48.

Schwille, F., 1988, Dense chlorinated solvents in porous and 

fractured media: Model Experiments: Michigan, Lewis 

Publishers. 146 p.

Stone. H.L., 1973, Estimation of three-phase relative permeability and residual oil data: The Journal of Canadian Petroleum Technology, October-December, 

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p. 892-898.

688
Challenges in Detecting Contaminant Effects

on an Estuarine Ecosystem Affected by Many

Different Disturbances: San Francisco Bay

By Samuel N. Luoma1 , and Frederic H. Nichols 1

Abstract

The Suisun Bay/Delta at the head of San 

Francisco Bay is an area where complex factors 

converge to affect the success of aquatic species. 

Water diversion and drought have reduced river 

flows to the bay. The historic elimination of 

marshes has permanently removed an important 

source of carbon while river diversion and 

drought have reduced river-borne inputs of 

carbon. The invasion of the bay by an exotic clam 

has affected the abundance of other benthic 

species and has eliminated previously observed 

phytoplankton biomass peaks. Coincidentally, 

populations of important resource species have 

declined. Persistent contamination in Suisun 

Bay/Delta accentuates the adverse effects of the 

above first-order problems. Selenium 

contamination in benthos is sufficient to affect 

upper-trophic-level organisms such as birds and 

fish. Pulse inputs of pesticides occur during high 

flows. Locations in Suisun Bay were identified 

that are chronically contaminated with trace 

organic compounds, cadmium, chromium, and 

vanadium. At these locations resident bivalves 

are in poorer condition and have modified 

reproductive cycles, compared to elsewhere in the 

estuary. Long-term studies are allowing 

identification of the relative importance of 

different factors that disturb biological processes 

in the estuary. Few multi-disciplinary, long-term

i

94025

U.S.Geological Survey, Mail Stop 465, Menlo Park, Calif.

field studies are conducted in environments 

modified by human disturbance. Results from 

Suisun Bay show how continuity in the study of 

cause-and-effect may be the only way to separate 

the influence of the myriad of processes that 

contribute to ecological characteristics in 

estuaries.

INTRODUCTION

The array of human activities that have 

disturbed the San Francisco Bay ecosystem since the 

gold rush of 1849 are well known (Nichols and others, 

1986). However, the effects and relative importance 

of specific activities are not as well understood.

One focus of the initial phase of the San 

Francisco Bay toxics program has been the 

characterization of contamination and its effects in the 

context of the myriad other disturbances that affect the 

Suisun Bay and Delta region. In this paper we discuss 

recent progress in studies of both the natural 

variability of the Suisun Bay/Delta and effects on 

aquatic life of human-induced stresses including 

contaminants.

RIVER FLOW AS A MAJOR 

DETERMINANT OF ECOSYSTEM 

FUNCTIONING

The upper San Francisco Bay is a complex 

estuarine system bounded on its landward side by a 

tidal freshwater Delta, a highly dissected region of 

channels and islands at the mouths of the Sacramento 

and San Joaquin Rivers. The rivers converge at the

689
eastern end of Suisun Bay, an important nursery area 

for estuarine fish populations.

Flow from the Sacramento and San Joaquin 

river system, draining about 40 percent of the land 

area of California, reflects both the pattern of 

precipitation over western North America and the 

operation of the world's largest man-made water 

management system (Peterson and others, 1989). 

More than half of river water flowing toward San 

Francisco Bay is diverted from the drainage basin 

through impoundment of snowmelt in reservoirs and 

pumping of Sacramento and San Joaquin river water 

from the Delta. The diverted water is consumed in 

irrigated agriculture and by urban and industrial uses 

both within and beyond the Bay's watershed (Nichols 

and others, 1986). The result is greatly reduced flows 

into the estuary during spring.

Despite the large-scale water diversion, the 

estuary typically receives large annual winter/spring 

pulses of freshwater runoff. Since 1984, however, 

persistent drought combined with the continuing water 

diversion to virtually eliminate the annual pulses of 

freshwater except for an exceptional flood during the 

spring 1986 (fig. 1).

The reduced flow of freshwater into the estuary 

has contributed to major changes to the Bay's 

ecosystem. As an example, striking declines in the 

abundance of a number of aquatic species were 

observed over the last several decades (Jassby and 

others, 1993a). The mechanisms differ, but reduced 

river flow seems to have affected the success of a 

native species of zooplankton, the local mysid shrimp,

RIVER DISCHARGE

Monthly Averages

1974 1976 1978 1980 1982 1984 1986 1988 1990 1992

Figure 1. Monthly average discharge of the Sacramento 

River at Chipps Island in San Francisco Bay.

striped bass, smelt, flounder, and mollusks that inhabit 

Suisun Bay (Jassby and others, I993a). The 

freshwater/seawater interface, as indicated by the 

geographic location of the 2 0/00 isohaline in bottom 

water, has moved upstream to the detriment of the 

freshwater organisms that inhabit the region. The 

reduced flows and upstream movement of salt also 

have increased the exposure offish populations to the 

diversion pumps in the Delta where huge numbers of 

fish are lost (Jassby and others, 1993a).

Recent studies have demonstrated that reduced 

river flow also has caused a reduction of organic 

carbon inflow into the estuary (Jassby and others, 

I993b). This work demonstrated that San Francisco 

Bay has a low input of organic carbon compared to 

that of many other estuarine systems (Jassby and 

others, 1993b). Poor light penetration into the siltladen waters of the Bay limits in situ carbon 

production by phytoplankton and vascular plants. The 

extensive marsh system that once existed was also 

largely eliminated early in this century, reducing 

carbon inputs from what was once a major source. 

Thus the rivers are now the remaining important 

source of carbon input to Suisun Bay/Delta. As river 

flow was reduced by drought and water diversion, 

carbon input was reduced concomitantly. The 

application of isotopic and molecular tracers has 

revealed that an important source for the organic 

matter now supporting benthic organisms in the upper 

estuary is freshwater algae produced in the rivers and 

Delta (Canuel and others, in preparation). This 

contrasts to South San Francisco Bay where in situ 

production is most important (Jassby and others, 

1993a).

Another important change in the estuarine 

ecosystem since the onset of the drought is the 

disappearance of the summer phytoplankton biomass 

maximum that typically appeared in the Suisun Bay 

region (Alpine and Cloern, 1992). This is likely the 

result of drought-related increases in the abundance of 

estuarine benthic suspension-feeding organisms, 

particularly that of a recently introduced Asian clam.

EFFECTS OF AN EXOTIC SPECIES

Historically, the most pronounced change to the 

benthic biota of San Francisco Bay has come from the 

accidental and intentional introduction of exotic 

species. Large numbers of sediment-dwelling 

invertebrates were first introduced during the latter

690
half of the 19th century with live oysters that were 

shipped by train from East coast estuaries and 

transplanted in the bay (Nichols and others, 1986). 

Exotic invertebrates continue to arrive in the bay in 

ballast water of cargo ships.

A recent arrival from the estuaries of east Asia, 

the clam Potamocorbula amurensis, has been 

remarkable in the rapidity with which it became the 

numerically dominant species throughout the northern 

part of the estuary and in its potential to change foodweb processes (Carlton and others, 1990).

Potamocorbula arrived in San Francisco Bay 

immediately after the major flood of 1986 when most 

of the estuarine benthic invertebrates were eliminated 

from the upper estuary because of their intolerance to 

fresh water. Within months of its first detection in 

October 1986, Potamocorbula became the most 

abundant benthic organism in Suisun Bay, and has 

prevented the bottom dwelling species that once 

dominated during low flow conditions (for example, 

Mya arenaria} from recolonizing the area (Nichols 

and others, 1990) (fig. 2). The combination of its great 

abundance and the efficiency with which it filters food 

particles from the water column has resulted in the 

near elimination of phytoplankton (diatoms and other 

microscopic, single-celled plants that form the base of 

aquatic food chains) from the water of the Suisun Bay 

region (fig. 2; see Alpine and Cloern, 1992). Further, 

there is laboratory evidence that Potamocorbula eats 

the juvenile stages of some zooplankton, possibly

ABUNDANCE OF MYA. POTAMOCORBULA; 

RIVER DISCHARGE; CHLOROPHYLL BIOMASS

4500

1974 1976 1978 1980 1982 1964 1986 1988 1990 1992 

YEAR

Figure 2. Annual discharge of the Sacramento River in 

cubic meters per second (solid line), concentrations of 

chlorophyll a in milligrams per cubic meter (line through 

open squares), and abundance of the clams Mya arenaria 

(solid bars) and Potamocorbula amurensis (striped bars) in 

Suisun Bay expressed as number of individuals per meter 

squared, between the years 1974 and 1992.

contributing to the reductions in zooplankton 

abundance noted during the prolonged drought 

(Obrebski and others, 1992).

The population biology of Potamocorbula has 

been intensively studied in Suisun Bay, San Pablo Bay 

and South Bay in monthly samplings between 1989 

and the present. Genetic studies (electrophoretic 

separation of isozymes) indicated Potamocorbula had 

an unusually high within-population genetic 

variability, contributing to its ability to survive in a 

range of estuarine conditions (Duda, 1993). A study 

of its reproductive activity throughout the Bay, under 

varying salinity, temperature and food regimes, 

demonstrated major differences in reproductive 

seasonality between north and south Bay. The animals 

from Suisun Bay were reproductively active during 

shorter periods than elsewhere. This difference 

coincided with, and thus may be related to, spatial 

differences in food availability and osmotic stress, or 

possibly contaminant stress. Age structure and 

condition index (tissue weight of a bivalve of a given 

shell length) differed among stations in Suisun Bay, 

and between Suisun and other parts of the estuary. 

Clams from the upper end of Suisun Bay were 

generally of longer shell length, had a lower condition 

index and demonstrated less seasonal fluctuation in 

condition index than elsewhere (Brown and Luoma, in 

preparation). Again it is not yet clear whether this is a 

response to reduced food availability or higher 

contaminant exposure.

TRACE ELEMENT CONTAMINATION

It is probable that contamination from 

anthropogenic waste discharges is contributing to 

some of the ecological changes occurring in the upper 

San Francisco Bay estuary, and biological differences 

between Suisun Bay and other regions. Recent results 

of the analysis of sediment cores from San Francisco 

Bay indicate sediments and water have been 

contaminated for decades by potentially toxic trace 

metals (Hornberger and others, in preparation; Van 

Geen and others, in preparation) and trace organic 

contaminants (Pereira and others, in preparation; 

Hostettler and others, in preparation). Industrial and 

urban development coincided with increased 

sedimentation during the last 40 years (Fuller, in 

preparation) and enriched concentrations of Cu, Ag, 

Pb ancj Zn. Sediments contaminated for more than one 

meter in depth are found from Suisun Bay to

691
Richardson Bay, near the Golden Gate. Pesticides 

(predominantly DOT) and petroleum by-products are 

enriched to this depth in Richardson Bay sediments.

The Suisun Bay/Delta was characterized by 

Biggs and others (1989) as one of the estuaries in the 

United States most susceptible to pollution impact 

because the physical characteristics of this system 

limit its natural resilience. River-transported metals, 

and wastes form nine of the largest point source 

dischargers in the Bay area are released directly into 

Suisun Bay (Luoma and others, 1990).

Recent USGS studies illustrate some 

characteristics of modern organic and inorganic 

contamination in Suisun Bay. Water soluble pesticides 

occur periodically in detectable concentrations in 

Suisun Bay when runoff from agricultural fields 

reaches the estuary (Kuivila and others, 1996). These 

pulse inputs from winter storms are superimposed on 

chronic contamination with petroleum by-products 

(Pereira and others, 1992) and trace elements (Luoma 

and others, 1990; Johns and Luoma, 1988; Luoma and 

others, 1992; Brown and Luoma, in preparation).

Selenium is an example of a chronic 

contaminant. Field surveys and mass balance 

calculations (Johns and Luoma, 1988; Cutter, 1989) 

demonstrated that the most important sources of Se are 

industrial discharges located near Carquinez Strait. 

During the low river flows characteristic of the estuary 

since 1986, Se enrichment in water, suspended 

particles and benthic species spread throughout Suisun 

Bay. Substantial Se contamination is found in large 

fish and in diving ducks in Suisun Bay. The 

concentrations of Se in water are well below those that 

cause toxicity in bioassays, but Se concentrations in 

clams (the food of diving ducks and some fish) are at 

the level that causes toxicity to birds and fish in food. 

The combination of studies showed that the 

contamination of the upper trophic level organisms, 

and the concomitant threat to their health, resulted 

from highly efficient transfer of selenite (the form of 

Se discharged by industry in the area) from water, to 

phytoplankton and then to clams via ingestion of 

phytoplankton (fig. 3).

Analysis of Potamocorbula tissues has been 

employed to define chronic trace metal contamination, 

in addition to selenium. The abundance of this clam 

across across a broad range of salinities in the system 

makes it an especially suitable organism for spatially 

comparative studies of contamination. Laboratory 

studies also showed that P. amurensis rapidly 

bioconcentrates metals in its soft tissues from both

 *

Figure 3. An illustration of the concentrations of selenite in 

micrograms per liter, selenium in suspended particles in 

micrograms per gram, selenium in clams in micrograms per 

gram, and the toxicity of selenium in the food of ducks 

typical of Suisun Bay.

food and solution (Brown and Luoma, 1995; Decho 

and Luoma, 1995). Potamocorbula were sampled and 

analyzed monthly at six stations between January, 

1991 and April, 1992, a period of relatively stable 

river flow. Thus, influences of physical and salinity 

variability on assessment of source inputs was 

minimized. A consistent spatial gradient in 

enrichment with Cd, Cr, Ni, and V was observed in the 

clam tissues (fig. 4). Concentrations were highest in 

upper and mid-Suisun Bay and declined through 

Carquinez Straits into San Pablo Bay. All metals 

except silver were more enriched in Suisun Bay than 

in South Bay. The same general gradient in 

contamination was observed at every sampling time 

through the 14 month period. Interestingly, the highest

Chromium in P. amurensis and C. fluminea

i 

Sacramento R.

microgram per gram 

0-2 2-5

Figure 4. Distribution of chromium concentrations in tissues 

of the clams Potamocorbula amurensis and Corbicula 

fluminea in the San Joaquin River, Suisun Bay and San 

Pablo Bay.

692
through the 14 month period. Interestingly, the highest 

concentrations of metals were found in Honker Bay, 

where no immediate point sources of metal input 

occur. These data raise the possibility that Honker 

Bay could be a depositional zone for upper Suisun 

Bay. Hydrodynamic studies will aid an evaluation of 

this hypothesis.

INTERACTION OF CONTAMINANTS WITH 

OTHER DISTURBANCES

The gradient of chronic contamination in Suisun 

Bay coincides with some of the unique biological 

characteristics of the area. In P. amurensis these 

include the gradient of lower condition index in 

Suisun Bay to higher condition index in San Pablo 

Bay, the lack of a seasonal cycle of condition index in 

Suisun Bay, and the spatial differences in the 

seasonality of reproduction in the species. A decline 

in the condition of a bivalve and alterations to 

reproductive capabilities can be symptoms of stress in 

bivalves. Trace elements can cause these responses by 

disrupting the energetic balance of the organism and 

reducing energy available for reproduction and growth 

(Widdows, 1985). On the other hand, food availability 

and other natural or anthropogenic disturbances can 

also cause such changes or interact with contaminants 

in complex ways to produce responses of 

communities, populations and individuals (Luoma and 

Carter, 1991).

Ecologists typically cite natural processes or 

anthropogenically induced changes in habitat 

conditions to explain differences in populations or 

declines in fisheries. Toxicologists often search for 

effects of contaminants where waste discharges from 

human activities impinge on an aquatic environment. 

Sufficiently careful, detailed ecological studies that 

consider contaminants as one of several variables of 

influence are rare. The studies of 1989 - 1993 

identified and specifically characterized ecological and 

biological anomalies in Suisun Bay. These studies 

identified several processes that could cause those 

anomalies, including reduced river flow, lowered 

carbon inputs, changes in water column ecology 

caused by invasion of an exotic species, and persistent 

contamination. Continuity in collection of data that 

describe these processes, including time series that 

extend for decades, have previously provided 

opportunities to distinguish the relative importance of 

different processes in complex estuarine environments

(Jassby and others, 1993a; Nichols and others, 1986). 

This same approach is the best approach to 

distinguishing factors that control ecological and 

biological processes in Suisun Bay.

REFERENCES

Alpine, A.E., and Cloern, J.E., 1992, Trophic interactions 

and direct physical effects control phytoplankton 

biomass and production in an estuary: Limnology and 

Oceanography, v. 37, p. 946-955.

Biggs, R.B., DeMoss, T.B., Carter, M.M., and Beasley, 

E.L., 1989, Susceptibility of U.S. estuaries to 

pollution: Reviews in Aquatic Science, v. 1, p. 189- 

208.

Brown, C.L. and Luoma, S.N., 1995, Use of a euryhaline 

bivalve to assess trace metal contamination in San 

Francisco Bay, I Evaluating the characteristics of 

Potamocorbula amurensis as a biosentinel species: 

Marine Ecology Progress Series, in press.

Canuel, E.A., Cloern, J.E., Ringelberg, D., Guckert, J. and 

Rau, G., 1995, Application of molecular and isotopic 

tracers to understand sources of organic matter and 

tropic relationships in the San Francisco Bay estuary: 

Marine Ecology Progress Series, in press.

Carlton, J.T., Thompson, J.K. Schemel, L.E., and Nichols, 

F.H., 1990, The remarkable invasion of San Francisco 

Bay (California, USA) by the Asian clam 

Potamocorbula amurensis, I Introduction and 

dispersal: Marine Ecology Progress Series, v. 66, p. 

81-94.

Cutter, G.A., 1989, The estuarine behaviour of selenium in 

San Francisco Bay: Estuarine, Coastal and Shelf 

Science, v. 28, p. 13-34.

Decho, A.W. and Luoma, S.N., 1991, Time-courses in the 

retention of food material in the bivalves 

Potamocorbula amurensis and Macoma balthica: 

Significance to the absorption of carbon and 

chromium: Marine Ecology Progress Series, v. 78, p. 

303-314.

Duda, T.F., 1993, The genetic population structure of 

Potamocorbula amurensis in San Francisco Bay: 

Masters Thesis, California State University, San 

Francisco CA, p. 120.

Fuller, C., van Geen, A. and Baskaran, M., 1994, Records of 

contaminant inputs to San Francisco Bay: 3. Sediment 

chronology from 2I ° Pb, l37 Cs, 239- 240Pu and 234Th 

profiles: Ocean Sciences Meeting, American 

Geophysical Union, San Diego, CA, Feb., 1994.

Hornberger, M.I., 1994, Records of contaminant inputs to 

San Francisco Bay: 4. Downcore concentration of Cr, 

Cu, Pb, V, and Zn: Ocean Sciences Meeting, American 

Geophysical Union, San Diego, CA, Feb., 1994.

693
Hostsettler, F.D., Pereira, W.E., Rapp, J.B., and

Kvenvolden, K. A., 1994, Records of contaminant 

inputs to San Francisco Bay, 8 Biomarker profiles in 

sediment cores: Ocean Sciences Meeting, American 

Geophysical Union, San Diego, CA, Feb., 1994.

Jassby, A.D., Kimmerer, W.J., Monismith, S.G., Armor, C., 

Cloern, I.E., Powell, T.M., Schubel, J.R., Vendlinski, 

T.J., 1993a, Isohaline position as a habitat indicator for 

estuarine populations: Ecological Applications.

Jassby, A.D., Cloern, J.E. and Powell, T.M., I993b, Organic 

carbon sources and sinks in San Francisco Bay 

variability induced by river flow: Marine Ecology 

Progress Series, v. 95, p. 39-54.

Johns, C. and Luoma, S.N., 1988, Selenium accumulation in 

benthic bivalves and fine sediments of San Francisco 

Bay, The Sacramento-San Joaquin Delta, and selected 

tirbutaries: Estuarine Coastal and Shelf Science, v. 27, 

p. 381-396.

Kuivila, K.M. and Copeland D.D., 1996, Diazinon

concentrations and transport in the Sacramento River 

and San Francisco Bay, California, February 1993, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Luoma, S.N. and Carter, J.L., 1991, Effects of trace metals 

on aquatic benthos, in Newman, M.C. and Mclntosh, 

A.W., Eds., Metal Ecotoxicology: Concepts and 

Applications, Chelsea, ML, Lewis Publishers, p. 261- 

300.

Luoma, S.N., Dagovitz, R., and Axtmann, E., 1990, 

Temporally intensive study of trace metals in 

sediments and bivalves from a large river-estuarine 

system: Suisun Bay/Delta in San Francisco Bay: 

Science of the Total Environment, v. 97/98, p. 685- 

712.

Luoma, S.N., Johns, C., Fisher, N.S., Steinberg, N.A., 

Oremland, R.S. and Reinfelder, J.R., 1992, 

Determination of selenium bioavailabililty to a benthic 

bivalve from particulate and solute pathways: 

Environmental Science and Technology, v. 26, p. 485- 

491.

Nichols, F.H., Cloern, J.E., Luoma, S.N., and Peterson, 

D.H., 1986, The modification of an estuary: Science, 

v. 231, p. 567-573.

Nichols, F.H., Thompson, J.K., and Schemel, L.E., 1990, 

The remarkable invasion of San Francisco Bay 

(California, USA) by the Asian clam Potamocorbula 

amurensis, II Displacement of a former community: 

Marine Ecology Progress Series, v. 66, p. 95-101.

Obrebski, S., Orsi, J.J., and Kimmerer, W., 1992, Long-term 

trends in zooplankton distribution and abundance in 

the Sacramento-San Joaquin Estuary: Interagency 

Ecological Studies Program Technical Report no. 32, 

42 p.

Pereira, W.E, Hostettler, F.D. and Rapp, J.B., 1992, 

Bioaccumulation of hydrocarbons derived from 

terrestrial and anthropogenic sources in the asian clam, 

Potamocorbula amurensis, in San Francisco Bay 

estuary: Marine Pollution Bulletin, v. 24, p. 103-109.

Pereira, W.E. and Hostettler, F.D., 1994, Records of 

contaminant inputs to San Francisco Bay, 7 

Anthopogenic and biogenic polycyclic aromatic 

hydrocarbons: Ocean Sciences Meeting, American 

Geophysical Union, San Diego, CA, Feb., 1994.

Peterson, D.H., Cayan, D.R., Festa, J.F., Nichols, F.H., 

Walters, R.A., Slack, J.V., Hager, S.E., Schemel, L.E., 

1989, Climate variability in an estuary effects of 

riverflow on San Francisco Bay: American 

Geophysical Union, Geophysical Monograph, v. 55, 

p. 419-442.

vanGeen, A., 1994, Records of contaminant inputs to San 

Francisco Bay: 6. Recent cadmium enrichments in the 

water column from Cd/Ca ratios in foraminifera: 

Ocean Sciences Meeting, American Geophysical 

Union, San Diego, CA, Feb., 1994.

Widdows, J., 1985, Physiological responses to pollution: 

Marine Pollution Bulletin, v. 16, p. 129-133.

694
Hydrodynamics of California's Sacramento-San 

Joaquin Delta, and the Collection of Flow Data for the

Tracking of Pesticides

By Richard N. Oltmann1

Abstract

The Sacramento and San Joaquin Rivers drain 

the Central Valley of California where 10 percent 

of the pesticides used in the United States are 

applied. The two rivers converge at the head of 

San Francisco Bay, forming a complex delta 

consisting of many interconnected sloughs and 

channels. The tracking of pesticides within and 

through the delta is difficult because flows are 

being manipulated constantly and measured-flow 

data are lacking. Ultrasonic velocity meters and an 

acoustic Doppler flow-measuring system are being 

used to collect the flow data necessary to improve 

understanding of the hydrodynamics of the delta 

and to track pesticides entering the delta.

INTRODUCTION

Agriculture in the Central Valley of California 

accounts for 10 percent of the total pesticide usage in 

the United States. The valley's two primary river systems, the Sacramento and San Joaquin Rivers (fig. 1), 

converge to form a large delta at the head of the San 

Francisco Bay estuary. These two rivers, in addition to 

a few smaller rivers, discharge almost half of the entire 

State's runoff into the delta. A study is currently being 

done by the U.S. Geological Survey (USGS) to determine how pesticides are transported and transformed as 

they pass from the valley and through the SacramentoSan Joaquin Delta (Kuivila and Nichols, 1991). The 

hydrodynamics of the delta are dynamic and complex 

and not completely understood, primarily because of 

the lack of measured-flow data. An improved understanding of the hydrodynamics of the delta is needed to 

study the transport of pesticides through the delta.

'U.S. Geological Survey, Sacramento, Calif.

This paper describes the complex hydrodynamics 

of the Sacramento-San Joaquin Delta, how flow data 

are and will be collected, what has been learned from 

the data collected to date, and how the data will be used 

to assist in the study of the transport of pesticides 

within and through the delta.

DESCRIPTION OF SACRAMENTO-SAN 

JOAQUIN DELTA AND THE MANIPULATION 

OF FLOWS

The Sacramento-San Joaquin Delta is about 50 mi 

east of the city of San Francisco, Calif, (fig. 1). The 

delta was originally a tidal marsh that experienced 

floods every winter, very low summer flows, and saltwater intrusion from the bay. The soils of the area are 

rich, and during the mid-1800's to early 1900's levees 

were built to form islands for agriculture. The delta 

presently consists of about 1,100 mi of levees, 700 mi 

of channels, and about 120 flow junctions (fig. l);the 

area is about 1,100 mi2 .

The delta is critical to the State's water distribution 

and supply and is an important region for fish and wildlife resources. During normal and wet years, enough 

precipitation occurs in California to satisfy the statewide demand for water; however, about two-thirds of 

the precipitation occurs north of the delta, and about 

two-thirds of the population and water need is south of 

the delta. Also, the precipitation occurs during the 

autumn and winter, and the need is greatest during the 

hot summer months.

To mitigate these distribution problems, the Federal 

Central Valley Project (CVP) (construction began in 

1930's) and the State Water Project (SWP) (construction 

began in 1950's) were constructed to capture the northern runoff and convey water through the delta to points 

of need south of the delta. Water is exported from the 

southern end of the delta by both projects (fig. 1): the 

CVP delivers water to the San Joaquin Valley primarily 

for agricultural irrigation, and the SWP delivers water to 

the southern part of the State through the 500-mi long

695
122°00'

38°15'

38°00'

37°45'

SUISUN MARSH 

SALINITY CONTROL GATE 

/

,ANA \ '^ u r*vv

I ' - , \ Central Valley Project

/ oVV^P^P'30' Tracy

/ ** ^."""V. '-.

EXPLANATION

i I I I I 

4812 KILOMETERS

-^ Flow routes to export facilities 

0 UVM station

Figure 1. Sacramento-San Joaquin Delta and location of ultrasonic velocity meter (UVM) stations and flow routes to 

export facilities.

696
California aqueduct. The combined pumping capacity of 

the projects is about 11,000 ft3/s with an annual mean of 

about 7,000 ft3/s. In addition, more than 1,800 individual 

irrigation diversions within the delta can divert as much 

as 5,000 ft3/s during the height of the irrigation season; 

annual mean diversion rate is about 1,300 ft3/s. The 

average flow of the Sacramento and San Joaquin Rivers 

is 24,000 and 4,600 ft3/s, respectively.

The diversion of water from and within the delta 

has caused several problems within the estuary, primarily with regard to water quality and fisheries (Miller, 

1993). Because the delta supplies drinking water to 20 

million Californians, maintaining the quality of delta 

water is of paramount importance. The diversions 

greatly reduce the amount of freshwater flowing into 

the bay from the delta, resulting in increased salt concentrations within the delta-a condition detrimental to 

drinking-water and irrigation supplies and to aquatic 

life. The diversions are a possible cause for observed 

reductions in the abundance of several fish species, 

some of which are listed or being considered for listing 

as endangered species. The diversions can affect fish 

populations by entraining eggs and larvae (some diversions are screened) and by causing cross-delta currents 

that can confuse spawning fish.

In order to mitigate problems created by the 

diversions while continuing to export water from the 

delta, various operating strategies and structural modifications to the delta are used by water managers. A 

network of real-time, salinity-monitoring stations is 

operated throughout the delta to monitor water quality. 

If salinities become too high, additional water is 

released from upstream storage, water is redirected, 

and (or) the export rate is reduced to lower the salinity 

at locations where salinity is high. At times, temporary 

rock barriers or tidal flap gates are installed across 

selected channels to prevent the intrusion of brackish 

water, thus improving the water quality. Barriers also 

are sometimes temporarily installed in channels near 

the export facilities to raise the water-surface elevation 

so that intakes to irrigation pumps remain submerged.

Several operating strategies are used in the delta to 

mitigate problems to the fisheries. When salmon and 

striped bass eggs and larvae drift down the Sacramento 

River, the delta cross channel gates near Walnut Grove 

(fig. 1) are closed to allow more of the eggs and larvae 

to move toward the bay and away from the export 

facilities. The CVP constructed the delta cross channel 

to transfer Sacramento River water to the Mokelumne 

River during low- to medium-flow conditions to provide more water to the export facilities in the southern

delta; the delta cross channel gates are closed when the 

flow in the Sacramento River exceeds 25,000 ft3/s. The 

delta cross channel and Georgiana Slough (fig. 1) are 

the primary channels for the conveyance of Sacramento River water to the export facilities, and the sum 

of the two flows is referred to as delta transfer flow. No 

control structures are on Georgiana Slough; however, 

consideration is being given to installing a temporary 

barrier on the slough when eggs and larvae are present 

in the Sacramento River.

A temporary partial rock barrier is installed each 

autumn at the head of Old River (fig. 1) near the San 

Joaquin River during the time when Chinook salmon 

eggs and larvae are adrift down the San Joaquin River. 

This barrier allows a greater percentage of eggs and 

larvae to move north and away from the export facilities. 

The export rate of the CVP and SWP also can be temporarily reduced or terminated to allow the eggs and 

larvae to reach the bay. All the above strategies used to 

control flow complicate the hydrodynamics of the delta 

and increase the difficulty of tracking pesticides and 

interpreting pesticide concentration data.

COLLECTION OF DATA IN THE DELTA 

AND WHAT HAS BEEN LEARNED FROM 

THE DATA

New flow-measuring technology has made it 

possible to measure flow continuously at key sites in 

the delta, thus reducing some of the uncertainties about 

pesticide movement. The USGS is currently collecting 

flow data in the delta using ultrasonic velocity meters 

(UVM) and an acoustic Doppler discharge measuring 

system (ADDMS). A UVM (Laenen, 1985) provides a 

continuous index velocity, which is measured across a 

channel between two transducers that are installed at 

the same depth. The index velocity is correlated to 

mean cross-sectional velocity using measured-flow 

data collected by the ADDMS. Flow then is computed 

by multiplying mean cross-sectional velocity and 

channel cross-sectional area. The ADDMS is a portable system that allows for quick and accurate measurement of discharge as an acoustic Doppler current 

profiler (Simpson and Oltmann, 1992) is traversed 

across a channel. Tidal flow measurements of 600-footwide channels can be made with the ADDMS in about 

3 minutes and without the use of navigational aids.

The USGS has been operating UVM flow stations 

on the Old and Middle Rivers (fig. 1) in the central 

delta since January 1987. Old and Middle Rivers are 

two of the three flow routes by which water is drawn

697
through the delta to the export facilities in the southern 

delta. The third route is through upper Old River and 

Grant Line Canal (fig. 1). Prior to operation of the 

Old and Middle River UVM's, no continuous flowmonitoring stations were operated in the delta.

The flow data collected at the Old and Middle 

River stations have improved the understanding of the 

flow characteristics of the delta. The data show that the 

tides create a dynamic flow pattern with maximum 

tidal flows during both flood and ebb of about 10,000 

ft3/s during low export periods. Daily mean flows at the 

flow-monitoring stations indicate that, when a net positive flow exists at the San Joaquin River at Stockton 

(fig. 1), about 33 percent of the water exported by the 

projects is drawn down Old River, 45 percent drawn 

down Middle River, and the remaining 22 percent 

comes through upper Old River and Grant Line Canal. 

Net reverse flows can exist on the San Joaquin River at 

Stockton when the quantity of water drawn to the 

export facilities through upper Old River and Grant 

Line Canal exceeds the flow of the San Joaquin River 

entering the delta from the south. When a temporary 

partial rock barrier is installed at the head of upper Old 

River each autumn, the flow percentage increases to 40 

percent on the Old River and 52 percent on the Middle 

River and decreases to 8 percent on the upper Old River 

and Grant Line Canal.

Near the end of 1992, two UVM stations were 

installed on the Sacramento River upstream and downstream from Walnut Grove (fig. 1) to provide a measure 

of delta transfer flow. Two additional stations are 

planned for installation during 1993 on the Sacramento 

River downstream from Decker Island and the San 

Joaquin River at Jersey Point (fig. 1). The sum of flows 

at the Decker Island and Jersey Point stations will provide a measure of the flow leaving the delta and entering the bay a quantity commonly referred to as delta 

outflow. The data provided by the Jersey Point station 

also will monitor the net reverse flows of the lower San 

Joaquin River that occur periodically as a result of 

diversions.

In addition to the above new UVM stations, the 

cities of Tracy and Stockton (fig. 1) are planning to 

install UVM stations during 1993 so as to receive 

National Pollution Discharge Elimination System 

permits (Darrel Scott, city of Tracy, Calif, oral commun., 1993; Glen D. Birdzell, city of Stockton, Calif., 

oral commun., 1993). The USGS has been asked to 

calibrate the stations using the ADDMS.

The ADDMS has been used to make flow measurements at 15-minute intervals over two complete tidal

cycles at the western boundary of the delta at Chipps 

Island (fig. 1) during normal low-flow conditions to 

measure delta outflow and to determine the range of the 

tidal flows. The results indicated that the tidal flows 

were about 300,000 ft3/s during both flood and ebb, 

with the daily net flows being about 1 to 2 percent of 

the maximum tidal flows.

Water-level data collected throughout the delta 

show that during strong spring tides, the mean water 

elevation of the delta rises about 1 ft above the preceding neap tide level, which is equivalent to about 50,000 

acre-ft of water. Whether water is being drawn into the 

delta from the bay or from upstream sources is not yet 

clear. Flow data from new and planned UVM stations 

should help define this process.

The data from all the above UVM stations will be 

used to calibrate and validate a hydrodynamic model of 

the delta that the USGS and the California Department 

of Water Resources (DWR) are currently developing. 

DWR has been using a model of the delta for several 

years, but because of the difficulties associated with the 

collection of flow data in tidal zones, they have not 

been able to collect sufficient flow data to calibrate and 

validate the model properly. To supplement the UVM 

data, the ADDMS will be used periodically to collect 

10- to 12-hour time series of flow measurements at critical flow splits in the delta for use in model calibration 

and validation. Once the calibrated and validated flow 

and transport models are available, the model will be 

used to track the movement of dissolved pesticides, as 

well as other constituents, within and through the delta.

REFERENCES CITED

Kuivila, K.M., and Nichols, F.H., 1991, Overview of San 

Francisco Bay estuary toxic contaminants study, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. Geological Survey Toxic Substances Hydrology Program-- 

Proceedings of the technical meeting, Monterey, 

California, March 11-15, 1991: U.S. Geological Survey 

Water-Resources Investigations Report 91 -4034, 

p. 659-663.

Laenen, Antonius, 1985, Acoustic velocity meter systems: 

U.S. Geological Survey Techniques of WaterResources Investigations, book 3, chap. A17, 38 p.

Miller, W.J., 1993, The delta Overview of the Sacramento-San Joaquin Delta: Berkeley, Calif., Consulting 

Engineer Report prepared for California Urban Water 

Agencies, 49 p.

Simpson, M.R., and Oltmann, R.N., 1992, Dischargemeasurement system using an acoustic Doppler current 

profiler with applications to large rivers and estuaries: 

U.S. Geological Survey Open-File Report 91 -487,49 p.

698
Diazinon Concentrations and Transport in the 

Sacramento River and San Francisco Bay, California,

February 1993

ByKathryn M. Kuivila1 anc/Dwight D. Copeland1

Abstract

Previous studies of pesticide concentrations 

in the Sacramento River suggest that narrow 

pulses of diazinon, a major dormant spray 

pesticide, are transported down the river 

following rainfall. A study in February 1993 

followed these anticipated pesticide pulses down 

the Sacramento River and into San Francisco Bay. 

Water samples for pesticide analysis were 

collected daily at four sites in the river and bay. 

Three distinct pulses of diazinon were detected in 

the Sacramento River at Sacramento 1 to 3 days 

after rainfall. The load of diazinon in the 

Sacramento River in February was calculated to 

be 340 kilograms (active ingredient) or 

approximately 0.5 to 1.7 percent of the total 

diazinon applied to orchards in the Sacramento 

Valley in January and February. The diazinon 

pulse detected at Sacramento on February 12 was 

detected the next day at Rio Vista, 2 days later at 

Chipps Island (the boundary between the delta 

and San Francisco Bay), and 3 to 5 days later at 

Martinez (the western edge of Suisun Bay). 

Study results suggest that the transport of 

pesticides from the Sacramento River into San 

Francisco Bay is episodic and highly dependent 

on rainfall and flow conditions.

INTRODUCTION

Because of the diverse and extensive use of 

pesticides in the Sacramento Valley, there is a need for 

an understanding of pesticide inputs to the Sacramento

'U.S. Geological Survey, Sacramento, Calif.

River and the transport of pesticides in the Sacramento 

River into San Francisco Bay. Results of previous 

studies of pesticide concentrations in the Sacramento 

and San Joaquin Rivers in 1991 and 1992 suggest that 

rainfall is a major mechanism for transporting dormant 

spray pesticides from the orchards into the river 

(unpublished data on file in the California District office 

of the U.S. Geological Survey). One major dormant 

spray pesticide is diazinon, an organophosphate 

insecticide with high water solubility (40 mg/L ). 

Diazinon is applied on almond, peach, and prune 

orchards in the Sacramento Valley during January and 

February (California Department of Pesticide 

Regulation, 1990; Worthing and Hance, 1991).

This report describes the results of a February 

1993 study of (1) the temporal variability in diazinon 

concentrations and loads in the Sacramento River, and 

(2) the transport of diazinon down the Sacramento 

River into the northern embayment of San Francisco 

Bay. This study is part of a research effort to assess 

quantitatively the fate and transport of pesticides in the 

river-estuary system.

SAMPLING AND ANALYTICAL METHODS

Four sites were selected along the main flow 

path of the Sacramento River from Sacramento to the 

western boundary of Suisun Bay: Sacramento, Rio 

Vista, Chipps Island, and Martinez (fig. 1). Water 

samples for pesticide analysis were collected daily at 

the four sampling locations using a depth-integrating, 

discharge-weighted sampler (D-77 with a Teflon bottle 

and nozzle) at either one or three verticals, depending 

on the site. Flow at the Sacramento site is 

unidirectional, and sampling studies indicate that the

699
122° 15'

Colusa gaging station is 89.4 miles

upstream from Sacramento 

122°00' \ 45' 30'

39°00' -

38°00' -

0 | 4 | 8 | 12 MILES 

048 12 KILOMETERS

EXPLANATION

A STREAMFLOW GAGING STATION Y PESTICIDE SAMPLING SITE

Figure 1. Study area and location of streamflow-gaging stations and pesticide sampling sites.

700
400

D Rainfall 

B Diazinon

concentration

Discharge

80,000

- 60,000

- 40,000

Q

-* y 

>- en

- 20,000 ^ °

15 20 

FEBRUARY

28

Figure 2. Rainfall at Colusa (shown as bar with amount, in inches per day, above each bar), 

discharge of the Sacramento River at Freeport, and concentrations of diazinon at Sacramento. 

Rainfall is recorded from midnight to midnight; ticks represent noon of indicated day.

composition and concentrations of dissolved 

constituents at a single vertical midchannel is representative of the cross section under most flow 

conditions (data not shown). Samples were collected 

randomly during the day from a bridge. Discharge was 

recorded at Freeport (11 rm downstream from Sacramento) with an ultrasonic velocity meter. Because the 

site at Freeport is affected by the tide, the discharge 

was tidally filtered to calculate a daily mean discharge 

(Godin, 1972, p. 65-66).

In contrast to the Sacramento site, the flow at 

Rio Vista, Chipps Island, and Martinez reverses during 

the tidal cycle. It is logistically infeasible to collect 

sufficient data for load calculations when sampling in 

a tidally affected environment. Flow and solute 

concentrations may not covary over time; therefore, 

extensive sampling during the tidal cycle would be 

necessary to calculate loads. Samples at the tidally 

affected sites were collected during slack after ebbtide. 

This sampling scheme estimated the most seaward 

movement of solutes and created a consistency for 

comparison of daily concentrations.

Pesticides were extracted from filtered 1 -L 

samples onto C8 solid-phase-extraction cartridges and 

analyzed using a gas chromatography/ion-trap 

detector. The reporting limit for diazinon was 15 

ng/L. Details of analytical method and quality 

assurance practices are described by Crepeau and 

others (1994).

RIVERINE PESTICIDE PULSES

Pulses of diazinon were measured following 

rainfall in the Sacramento River in previous years; this 

phenomenon is similar to the spring flush of herbicides 

observed in surface-water runoff in the midwestern 

United States (Thurman and others, 1991). The 

observed riverine pulses of diazinon typically are 

narrow and well-defined; elevated concentrations can 

be measured for only a few days to weeks.

Diazinon was applied primarily at the end of 

January 1993 during 2 weeks of dry weather; a series 

of rainstorms began on February 6. Figure 2 shows 

the relations among rainfall, discharge, and diazinon 

concentrations. Rainfall was 0.51, 0.39, and 1.02 in. 

on February 6, 8, and 9, respectively. A few days after 

the first rainfall, flow at Freeport increased, reaching a 

maximum of 52,800 ft3/s on February 14. Similarly, 

diazinon concentrations increased on February 8 and 

reached a measured maximum of 393 ng/L on 

February 12. It rained again on February 20 (1.14 in.), 

and both discharge and diazinon concentrations soon 

increased. During this second peak, the diazinon 

concentrations were 194 ng/L and 186 ng/L on 

February 21 and 22, respectively. Two days later 

(February 24), the discharge reached a maximum of 

75,400 ft3/s.

701
0 <

< Q

50

40

z a. 30

o ^

 2

3 2

5 g 20

>- -J

10

 Diazinon load 

......... Discharge

80,000

60,000 uT o 

o o

ry LJ 

^ (/)

o: 

o Pn

CO

40,000

 a. 

o

-* o

>- m

20,000 < o

a _

10 15 

FEBRUARY

20 25 28

Figure 3. Calculated daily diazinon load for the Sacramento River and discharge at Freeport, 

February 1993. Ticks represent noon of indicated day.

Traveltimes were estimated by calculating 

velocities for discrete reaches of the river from a 

regression relation between measured instantaneous 

velocity and the logarithmic of instantaneous 

discharge for the cross section (at streamflow-gaging 

stations shown in fig. 1) for the selected time period. 

Most of the orchards are near Colusa, north of 

Sacramento, in a narrow band along the Sacramento 

River or clustered along the Feather River, a tributary 

that enters the Sacramento River 20 rm upstream from 

Sacramento. The estimated traveltime after the 

February 9 rainfall from Colusa down the Sacramento 

River to the city of Sacramento (89.4 rm) was 

1 1/2 days. This traveltime estimate is low because 

the calculation does not include transport time from 

the field to the river and velocities were slower along 

the Feather River than along the Sacramento River. 

The observed lag between rainfall and peak pesticide 

concentrations was 3 days. The lag between the 

second rainstorm on February 20 and the 

corresponding peak diazinon concentration was much 

shorter (1 day). The discharge measured on February 

20 (58,800 ft3 /s) was almost double the discharge 

measured on February 9 (32,000 ft3 /s), indicating 

increased stream velocities, which probably shortened 

traveltimes on February 20. Increased soil-moisture 

content resulting from infiltration of precipitation from 

the February 9 and 20 rainfalls may have shortened 

traveltime from fields to river.

CALCULATION OF DIAZINON LOADS

Diazinon loads for the Sacramento River were 

calculated by multiplying the instantaneous measured 

pesticide concentration by the tidally filtered, daily 

mean discharge (fig. 3). The integrated loads for each 

peak were 163 kg for February 8-16, 129 kg for 

February 19-25, and 30 kg for February 25-28. 

Although the maximum concentration during the 

second peak (194 ng/L) was only half that during the 

first peak (393 ng/L), the load of diazinon in the river 

was similar during the two peaks. In contrast, the 

maximum concentration (79 ng/L) and calculated 

diazinon load (30 kg) were lower during the third peak 

than during the previous two peaks, possibly because 

the amount of diazinon transport from orchards 

decreased.

The total load of diazinon in the Sacramento 

River was calculated and compared with the amount 

of diazinon applied to orchards in the Sacramento 

Valley. During February, the total load was 340 kg. 

The most recent diazinon-use data available is for 

1990 (California Department of Pesticide Regulation, 

1990), when 19,900 kg (active ingredient) was applied 

in January and February of that year. In addition, 

diazinon has replaced an unknown portion of the 

previously used ethyl parathion (now banned) 

(51,300 kg active ingredient). Therefore, assuming a 

maximum 1:1 replacement of ethyl parathion by

702
z 

O

a:

z

O </)

Z 2

O -rf

O a:

-5T O

11

M Z

400

300

200

100

 B Sacramento

 --- Rio Vista

 > Chipps Island

 Martinez

14 

FEBRUARY

19 24

Figure 4. Concentrations of diazinon in the Sacramento River at Sacramento, Rio Vista, 

Chipps Island, and Martinez, February 1993. Ticks represent noon of indicated day.

diazinon, from 19,900 to 71,200 kg of diazinon was 

applied to orchards in the Sacramento Valley in 1993. 

In this context, the calculated load in the Sacramento 

River at Sacramento represents 0.5 to 1.7 percent of 

the amount applied in January and February. These 

results support the findings of Wauchope (1978), who 

reported that pesticide losses in runoff from agricultural fields were up to 0.5 percent of the amounts 

applied, unless heavy rainfall fell within 1 to 2 weeks 

after application.

PESTICIDE TRANSPORT INTO SAN 

FRANCISCO BAY

A main flowpath of water down the Sacramento 

River is along the ship channel past Rio Vista, Chipps 

Island, and Martinez (fig. 1). The first pulse of 

diazinon in the Sacramento River was followed from 

Sacramento through Suisun Bay (the northern 

embayment of San Francisco Bay) (figs. 1 and 4). 

Initially detected at Sacramento, the diazinon 

concentrations reached a maximum value of 393 ng/L 

on February 12 (figs. 3 and 4). This pulse was 

detected at Rio Vista (43 rm downstream from 

Sacramento) on February 13; the maximum 

concentration was 300 ng/L. Two days later (February 

15), the pulse was detected at Chipps Island (16.5 mi

from Rio Vista); the maximum concentration was 200 

ng/L. Finally, diazinon concentrations of 100 and 120 

ng/L were detected on February 18 and 20 at Martinez 

(14.5 rm seaward from Chipps Island). The 

approximate traveltime was 1 day from Sacramento to 

Rio Vista, 2 days from Rio Vista to Chipps Island, and 

3 to 5 days from Chipps Island to Martinez. As the 

pesticide pulse moved seaward, the maximum 

concentration decreased and the pulse dispersed over 

time, in part because of tidal diffusion.

The National Academy of Sciences and 

National Academy of Engineering (1973) 

recommended a guideline of 9 ng/L as a maximum 

diazinon concentration in surface water for the 

protection of freshwater aquatic life; the International 

Joint Commission (1975) suggested a limit of 8 ng/L. 

Currently, the U.S. Environmental Protection Agency 

has not established an aquatic-life water-quality 

criterion for diazinon. All the water samples collected 

at the four study sites in February 1993 contained 

diazinon concentrations above the recommended 

guidelines for protection of aquatic life. In addition, 

other pesticides besides diazinon were present in the 

water samples during these runoff events (data not 

shown), and additive or synergistic effects can 

increase the total toxicity of the water.

703
SUMMARY AND CONCLUSIONS

The results of this study and previous studies 

show that rainfall affects the transport of pesticides 

from orchards into rivers. Each rainstorm in February 

1993 was followed by elevated concentrations of 

diazinon in the Sacramento River at Sacramento. 

Under high-flow conditions, diazinon was transported 

down the Sacramento River and into San Francisco 

Bay in distinct pulses. The concentrations measured in 

the river and bay during this study exceed 

recommended water-quality guidelines for freshwater 

aquatic life.

REFERENCES

California Department of Pesticide Regulation, 1990, 

Pesticide use data Computer tapes available from 

California Department of Pesticide Regulation, 

Sacramento, Calif. 95814.

Crepeau, K.L., Domagalski, S.L., and Kuivila, K.M., 1994, 

Methods of analysis and quality assurance practices of 

the U.S. Geological Survey Organic Laboratory, 

Sacramento, California Determination of pesticides in 

water by solid-phase extraction and capillary-column

gas chromatography/mass spectromctry: U.S. 

Geological Survey Open-File Report 94-362, 17 p.

Godin, G., 1972, The analysis of tides: Toronto, Canada, 

University of Toronto Press, 254 p.

IntemationalJoint Commission, 1975, Diazinon Great 

Lakes Water Quality Board, Water Quality Objectives 

Subcommittee Report, Appendix A, p. 103-106.

National Academy of Sciences and National Academy of 

Engineering, 1973 [1974], Water quality criteria, 1972: 

U.S. Environmental Protection Agency, EPA R3-73- 

033, 594 p.

Thurman, E.M., Goolsby, D.A., Meyer, M.T., and Kolpin, 

D.W., 1991, Herbicides in surface waters of the 

midwestern United States The effect of spring flush: 

Environmental Science and Technology, v. 25, p. 1794- 

1796.

Wauchope, R.D., 1978, The pesticide content of surface 

water draining from agricultural fields a review: 

Journal of Environmental Quality, v. 7, p. 459-472.

Worthing, C.R., and Hance, R.J., 1991, The pesticide 

manual, 9th ed.: Surrey, Great Britain, British Crop 

Protection Council, 1081 p.

704
Distribution and Possible Biological Effects of Diazinon

in the San Joaquin River and the Sacramento-San

Joaquin Delta, California, February 1993

ByKathryn M. Kuivila1 , Robert C. Sheipline1 , and Christopher G. Foe2

Abstract

The distribution and possible biological effects 

of a dormant spray pesticide, diazinon, were 

examined by measuring pesticide concentrations 

and estimating toxicity by means of bioassays at a 

series of sites in the Sacramento-San Joaquin 

Delta. A double peak of diazinon concentration 

was observed in early February 1993 at Vernalis 

on the San Joaquin River after heavy rains; similar 

pulses were detected at Stockton 2 days later. Two 

geographically separate sources of diazinon 

orchards along the western side of the San Joaquin 

River near Patterson and orchards along the 

Merced River could account for the observed 

pattern of diazinon concentrations in early 

February. Distributions of diazinon at central delta 

sites on Old and Middle Rivers were not pulselike; 

instead, the concentrations steadily increased 

during the sampling period. Seven-day bioassays 

indicated that San Joaquin River water at Vernalis 

was acutely toxic to Ceriodaphnia dubia (water 

flea) for the 12 consecutive days (February 8-19) 

with the highest diazinon concentrations. Examination of 96-hours LC50 values (lethal concentration that kills 50 percent of test organisms in 96 

hours) indicates that measured diazinon concentrations could not account for all the observed 

toxicity. Other pesticides present could contribute 

to the toxicity.

'U.S. Geological Survey, Sacramento, Calif. 

2Californis Regional Water Quality Control Board, 

Sacramento, Calif.

INTRODUCTION

The biological effects of dormant spray pesticides 

used on orchards in the San Joaquin River Basin are of 

environmental concern (Foe and Connor, 1991; Foe 

and Sheipline, 1993). Diazinon, an organophosphate 

insecticide, is the primary dormant spray applied in 

January and February on almond, apple, apricot, 

cherry, and peach orchards in the San Joaquin River 

Basin (California Department of Pesticide Regulation, 

1990). Distinct high-concentration pulses of diazinon 

have been detected in the San Joaquin River in January 

and February 1991 and 1992 following rainfall (unpublished data on file in the California District office of the 

U.S. Geological Survey). The results of bioassay surveys indicate that San Joaquin River water containing 

elevated concentrations of dormant spray pesticides, 

including diazinon, is generally toxic to Ceriodaphnia 

dubia (water flea) (Foe and Connor, 1991).

This report summarizes the results of a study on the 

distribution of diazinon concentrations in the San 

Joaquin River and Sacramento-San Joaquin Delta during February 1993 and assesses the potential biological 

effects of the diazinon-laden water. This study is part of 

a research effort to understand the fate and transport of 

pesticides in the river-estuary system.

SAMPLING AND ANALYTICAL METHODS

The sampling approach was to measure pesticides 

and toxicity at four sites along a flow path from Vernalis through Stockton to the export pumps: Vernalis (at 

the mouth of the San Joaquin River), Stockton, Old 

River, and Middle River (fig. 1). Water samples were 

collected at Vernalis twice a day using a depth-integrating, discharge-weighted sampler (D-48) at a single vertical from a bridge. Flow at this site is unidirectional, 

and sampling studies indicated that the composition

705
38°15'r 121°45' 121°00'

38°00' 

37°30' 

048 12 KILOMETERS

EXPLANATION

A Streamflow gaging station 

y Pesticide sampling site

Figure 1. Study area and location of streamflow-gaging stations and pesticide sampling sites.

706
o:

o to

Z 2

o <

o o:

o z

z <

M Z

i s

1,200r-r

900 -

600 -

300 -

8,000

Rainfall

Diazinon 

concentration

Discharge

o

10 15 

FEBRUARY

25 28

Figure 2. Rainfall at Modesto (shown as bar with amount, in inches per day, above each bar), 

and discharge and diazinon concentrations at Vernalis. Rainfall is recorded from midnight to 

midnight; ticks represent noon of indicated day.

and concentrations of dissolved constituents at a single 

vertical is representative of the cross section under 

most flow conditions (data not shown). Discharge was 

also recorded at Vernalis, a streamflow-gaging station.

In contrast to the Vernalis site, flow at the Stockton, 

Old River, and Middle River sites reverses during the 

tidal cycle. Flow and solute concentrations do not 

necessarily covary over time within a tidal regime; 

therefore, extensive sampling during the tidal cycle 

would be necessary to calculate constituent loads. 

Instead, samples were collected at slack after ebbtide at 

Stockton and at slack after floodtide at the Old and Middle River sites (see Oltmann, 1996, for additional details 

of delta hydrodynamics). This sampling scheme was 

designed to estimate the most forward movement of solutes along a flow path through the delta and to enable a 

comparison of daily concentrations in the delta.

One of the water samples collected at Vernalis each 

day during February was split for analysis of dissolved 

pesticides and for bioassays. The other water sample 

from Vernalis and the daily water samples from the 

other three sites also were analyzed for dissolved pesticides. Filtered 1-L samples were extracted using C8 

solid-phase extraction cartridges, and the pesticides 

were analyzed using a gas-chromatograph/ion-trap 

detector. The reporting limit for diazinon was 15 ng/L.

Seven-day C. dubia bioassays (U.S. Environmental Protection Agency, 1989) were done at the Sierra 

Foothill Laboratory. All tests met the U.S. Environmental Protection Agency (USEPA) bioassay acceptance criteria of at least 80-percent control survival. 

A water sample was considered toxic if test organism 

mortality was 30 percent or greater than the associated 

laboratory control mortality.

DIAZINON CONCENTRATIONS IN THE 

SAN JOAQUIN RIVER

In 1993, most of the diazinon was applied to 

orchards during 2 weeks of dry weather at the end 

of January (Elizabeth Deaver, San Joaquin County 

Department of Agriculture, oral commun., 1993). 

Both streamflow and diazinon concentration in the San 

Joaquin River began to increase immediately at Vernalis 

on February 8 in response to the first rainfall on February 

8 and 9 0.59 and 1.29 in. of rain, respectively. Two 

well-defined peaks of diazinon concentration were 

detected; the first had a maximum concentration of 773 

ng/L at 2400 hours on February 8 and the second had a 

maximum concentration of 1,071 ng/L at 1900 hours on 

February 11 (fig. 2). Differences in traveltimes to Vernalis from two geographically separate sources could have 

resulted in the observed double pulse after a single 

rainfall event.

707
Toxicity of San Joaquin River water in January and 

February has been attributed to the presence of dormant-spray orchard pesticides, including diazinon. 

Diazinon is applied to orchards located primarily in 

two areas in the San Joaquin drainage basin-along the 

western side of the San Joaquin River near Patterson 

(26 rm upstream of Vernalis) and along the Stanislaus, 

Tuolumne, and Merced Rivers on the eastern side of 

the San Joaquin River (California Department of Pesticide Regulation, 1990). Analysis of bioassays using C. 

dubia from 1988 through 1990 showed a pattern of 

high invertebrate toxicity in the San Joaquin River 

between the confluence of the Merced and Stanislaus 

Rivers, much lower toxicity in the San Joaquin River 

upstream of the confluence of the Merced River, and no 

recorded toxicity in the Stanislaus River (Foe and Connor, 1991).

Runoff in agricultural drains and creeks on the 

western side of the San Joaquin River responds rapidly 

to rainfall (J.L. Domagalski, U.S. Geological Survey, 

oral commun., 1993) and could be the source of the first 

diazinon peak concentration detected on February 8. 

In contrast, agricultural runoff from orchards located 

along the Merced River has a much longer traveltime 

to Vernalis than runoff from the western drains and 

could be the source of the later diazinon peak concentration detected on February 11.

Traveltime was estimated by calculating velocities 

for discrete reaches of the river from a regression relation between measured instantaneous velocity and the 

logarithm of instantaneous discharge for the cross section (at streamflow-gaging stations shown in fig. 1) for 

the selected time period (Clifton and Gilliom, 1989). 

Because early February was a time of unsteady flow 

and the actual time that runoff containing peak diazinon concentrations entered the river is not known, the 

traveltime estimates are expressed as a range. The 

estimated traveltime along the San Joaquin River from 

Patterson to Vernalis was 19 to 27 hours. In contrast, 

the estimated traveltime from the area of orchards 

around the Merced River (approximately 24 rm 

upstream of the confluence with the San Joaquin River) 

to Vernalis along the Merced and San Joaquin Rivers 

was 55 to 74 hours. The two observed diazinon peaks 

were separated by 2 1/2 days at Vernalis. This difference in arrival time supports the hypothesis of two 

sources of diazinon runoff orchards along the western 

side of the San Joaquin River near Patterson and 

orchards farther upstream along the Merced River.

1,200

900

600

300

o 

z 

<

z 0 

o 400

 Vernalis

 * Stockton

o 

o

o

z

N 

Q

300

200

100

Middle River 

Old River

9 14 19 

FEBRUARY

24 28

Figure 3. Diazinon concentrations at Vernalis and Stockton 

and at Old and Middle Rivers. Ticks represent noon of 

indicated day.

DIAZINON CONCENTRATIONS IN 

THE DELTA

During high-flow conditions, such as occurred 

in January and February 1993, there is positive net 

flow along the San Joaquin River toward Stockton 

(Oltmann, 1996). High concentrations of diazinon, 

similar to those observed at Vernalis (on February 8, 

11, and 19), also were detected at Stockton 2 days later 

(on February 10, 13, and 21) (fig. 3). Because flow 

reverses within the delta, insufficient data are available 

to estimate time of travel from Vernalis to Stockton.

In the central delta, water from the Sacramento, 

Mokelumne, and San Joaquin Rivers mixes in a series 

of complex channels and is subjected to tidal flow 

reversals (Oltmann, 1996). During the same time 

period that diazinon pulses were observed at Vernalis, 

similar diazinon pulses (as high as 393 ng/L) were 

detected in the Sacramento River (Kuivila and Copeland, 1994). Well-defined pesticide pulses were not 

observed at the Old and Middle River sites because of

708
the two riverine sources of diazinon and the hydrodynamic complexity of the delta; instead, the diazinon 

concentrations steadily increased through February 

(fig- 3).

POTENTIAL BIOLOGICAL EFFECTS ON 

THE DELTA

The results of this pesticide study are useful to estimate the possible effects of dormant spray pesticides, 

particularly diazinon, on the ecology of the delta. The 

National Academy of Sciences and National Academy 

of Engineering (1973) has recommended a guideline of 

9 ng/L diazinon as a maximum concentration in surface 

water for protection of aquatic life. The International 

Joint Commission (1975) suggests a similar guideline 

of 8 ng/L diazinon for the Great Lakes. Currently 

(1993), there is no USEPA aquatic-life criterion for 

diazinon.

Bioassays were used to determine the potential biological effects of river water contaminated with dormant 

spray pesticides (U.S. Environmental Protection 

Agency, 1989; 1991). One hundred percent C. dubia 

mortality was observed in water samples collected for 12 

consecutive days (February 8-19) from the San Joaquin 

River at Vernalis. The bioassay mortality corresponded 

with the highest diazinon concentrations (fig. 2 and 

table 1). Conversely, no toxicity was observed in water 

collected before (February 5 and 7) or after (February 

20-25) the peaks of diazinon concentration.

Ambient diazinon concentrations appear sufficiently elevated to contribute to, but not explain all, the 

C. dubia mortality. The laboratory 24- to 96-hour LC50 

values for C. dubia were between 430 and 550 ng/L of 

diazinon. In 7-day tests, concentrations of 120 ng/L of 

diazinon did not result in mortality, but did result in a 

25-percent reduction in reproduction (Foe and 

Sheipline, 1993). In the San Joaquin River, water samples with diazinon concentrations higher than 350 ng/L 

resulted in 100 percent mortality in 48 hours, and samples with diazinon concentrations of 148 to 263 ng/L 

resulted in 100 percent mortality in 7 days (table 1). No 

C. dubia toxic response that is, mortality at or greater 

than 30 percent was measured for diazinon concentrations at or below 84 ng/L. Chlorpyrifos, methidathion, 

and carbaryl were routinely detected in the water samples (table 1), and other pesticides also could have been 

present. Some or all these pesticides could have acted 

together either addictively or synergistically to produce 

the observed toxicity.

Table 1. Vernalis bioassav results and pesticide data

[Bioassay tests were done for 7 days; percentages of mortality 

were based on 10 replicates per sample. The following 

pesticides also were detected in some samples: ca, carbaryl; m, 

methidathion; ch, chlorpyrifos.

Date

i

2-05-93

2-07-93

2-08-93

2-09-93

2-10-93

2-11-93

2-12-93

2-13-93

2-14-93

2-15-93

2-16-93

2-17-93

2-18-93

2-19-93

2-20-93

2-21-93

2-23-93

2-24-93

2-25-93

Percentage 

of mortality

Control

0

0

0

0

0

0

0

0

0

0

0

0

0

0

0

0

0

0

0

Sample

0

0

'100

'100

'100

'100

'100

'100

'100

MOO 2 100

2 100

2 100

'100

0

0

0

20

0

Diazinon 

(ng/L)

73

84

111

586

358

1071

554

396

421

364

263

195

148

350

84

74

79

48

42

Other 

pesticides 

detected

ca,m

ca,m

ca,m,ch

ca,m.ch

ca,m,ch

ca,m,ch

ca,m,ch

ca,m,ch

ca,m,ch

ca,m,ch

m.ch

m.ch

m.ch

m.ch

m

ca

m

m

m

'Mortality occurred within 48 hours. 

2Mortality occurred within 7 days.

The ecological effects of pulses of pesticides on 

aquatic biota in the delta are not known. However, most 

freshwater zooplankton (copepods, rotifers, and cladocerans) in the delta are in decline (Obrebski and others, 

1992); the cause is unknown. The bioassay and pesticide data indicate that dormant spray pesticides enter 

the delta from the San Joaquin River in concentrations 

that can affect select organisms.

SUMMARY

Distinct pesticide pulses, with high concentrations 

of diazinon, were detected at Vernalis on the San 

Joaquin River after rainfall in February 1993. A double 

peak of diazinon observed in early February could be 

from two sources; this hypothesis was supported by 

traveltime estimates. Pesticide pulses were observed at 

Stockton 2 days later than at Vernalis; in contrast, the 

diazinon concentrations at Old and Middle Rivers 

increased steadily over time. San Joaquin River water 

containing the highest diazinon concentrations also 

was the most toxic to C. dubia in bioassay tests. 

Published LC50 values indicate that the observed

709
diazinon concentrations could account for only some of 

the observed toxicity. More extensive chemical and 

toxicological testing needs to be done to ascertain the 

chemicals responsible for causing toxicity, to determine their distribution and fate within the delta, and to 

evaluate their effect on native organisms.

REFERENCES

California Department of Pesticide Regulation, 1990, Pesticide use data: Computer tapes available from California 

Department of Pesticide Regulation, Sacramento, Calif.

Clifton, D.G., and Gilliom, R.J., 1989, Sources and concentrations of dissolved solids and selenium in the San 

Joaquin River and its tributaries, California, October 

1985 to March 1987: U.S. Geological Survey WaterResources Investigations Report 88-4217, 33 p.

Foe, C.G., and Connor, V., 1991, San Joaquin watershed 

bioassay results, 1988-90: Central Valley Regional 

Water Quality Control Board, Sacramento, Calif., Staff 

Report, 79 p.

Foe, C.G., and Sheipline, R.C., 1993, Pesticides in surface 

water from applications on orchards and alfalfa during 

the winter and spring of 1991-92: Central Valley 

Regional Water Quality Control Board, Sacramento, 

Calif, Staff Report, 79 p.

International Joint Commission, 1975, Diazinon: Great 

Lakes Water Quality Board, Water Quality Objectives 

Subcommittee Report, appendix A, p. 103-106.

Kuivila, K.M., and Copeland, D.D., 1996, Diazinon concentrations and transport in the Sacramento River and San 

Francisco Bay, February 1993, in Morganwalp, D.W.,

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

technical meeting, Colorado Springs, Colorado, September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

National Academy of Sciences and National Academy of 

Engineering, 1973 [1974], Water quality criteria, 1972: 

U.S. Environmental Protection Agency, EPA R3-73- 

033, 594 p.

Obrebski, S., Orsi, J.J., and Kimmerer, W., 1992, Long term 

trends in zooplankton distributions and abundance in 

the Sacramento-San Joaquin Estuary: Intcragcncy 

Ecological Studies Program for the Sacramento-San 

Joaquin Estuary, Technical Report 32, 42 p.

Oltmann, R.N., 1996, Hydrodynamics of California's Sacramento-San Joaquin Delta, and the collection of flow 

data for tracking of pesticides, in Morganwalp, D.W., 

and Aronson. D.A., cds, U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

U.S. Environmental Protection Agency, 1989, Short-term 

methods for estimating the chronic toxicity of effluents 

and receiving water to freshwater organisms (2d cd.): 

Cincinnati, Ohio, Environmental Monitoring and 

Support Laboratory, EPA/600/4-89-001, 417 p.

___1991, Methods for measuring the acute toxicity of 

effluents to freshwater and marine organisms (4th ed.): 

Cincinnati, Ohio, Environmental Monitoring and 

Support Laboratory, EPA/600/4-90-027, 293 p.

710
Concentrations of Dissolved Rice Pesticides in the 

Colusa Basin Drain and Sacramento River, California,

1990-92

By Kathryn L. Crepeau 1 , Kathryn M. Kuivila1 , and Joseph L. Domagalski 1

Abstract

The measured concentrations of rice 

pesticides in the Colusa Basin Drain and the 

Sacramento River, California, decreased during 

1990-92. Molinate, carbofuran, and thiobencarb 

are applied to the Sacramento Valley ricefields in 

April, May, and June. These pesticides are of 

concern because of the potential effect of 

discharge water from ricefields on striped bass 

larvae. Concentrations of dissolved pesticides 

from ricefields were measured at the Colusa Basin 

Drain (a major source of ricefield drainage) and 

the Sacramento River at Sacramento, Rio Vista, 

and Chipps Island during May, June, and July 

each year. The highest pesticide concentrations 

were measured at the Colusa Basin Drain with 

progressively decreasing concentrations downstream, principally because of dilution. During 

1990-92, the maximum molinate concentration in 

the Colusa Basin Drain decreased by a factor of 

three each year. The maximum carbofuran 

concentration decreased by a factor of four during 

1990-91. In contrast, the maximum carbofuran 

and thiobencarb concentrations in 1992 remained 

at the 1991 concentrations. The holding time 

before ricefield water could be discharged into the 

Sacramento River increased yearly during 1990-92. 

This allowed further degradation of the pesticides 

and resulted in the decreased concentrations of 

pesticides measured in the Colusa Basin Drain 

water and downstream sites.

U.S. Geological Survey, Sacramento, Calif.

INTRODUCTION

Rice is a major crop grown in the Sacramento 

Valley, California; as many as 500,000 acres are 

harvested each year (California Regional Water 

Control Board, Central Valley Region, 199la). The 

ricefields are flooded with irrigation water from the 

Sacramento River a few days before seeding in April 

or May. Pesticides can be incorporated into the soil 

before flooding or applied by air after flooding and 

seeding. Field water returned to the Sacramento River 

through the Colusa Basin Drain (CBD) (fig. 1) can 

contain rice pesticides. These pesticides are of 

concern because of their potential effect on striped 

bass larvae in the Sacramento River (California 

Regional Water Quality Control Board, 199la). The 

striped bass spawn between early May and mid-June, 

the time of ricefield water discharge.

Several pesticides are applied to rice in the 

Sacramento Valley; however, three pesticides of 

concern for water quality are molinate, carbofuran, 

and thiobencarb. Molinate, a thiocarbamate herbicide 

used on rice to control broad-leaved and grassy weeds, 

is applied usually from the air 7 to 10 days after seeding. The water solubility of molinate is 800 mg/L 

(milligram per liter). In shallow ricefields, molinate is 

lost through volatilization (Soderquist and others, 

1977) because of molinate's high vapor pressure 

[410 mPa (millipascal) at 20 °C (degree Celsius)]. 

In addition, molinate undergoes photolysis to form 

2-keto molinate and 4-keto molinate. The quantity of 

molinate [1,492,300 Ib (pound)] used on rice in the 

Sacramento Valley was more than 15 times greater 

than the quantity of carbofuran (88,240 Ib) or thiobencarb (95,830 Ib) during 1990 (California Department 

of Pesticide Regulation, 1990). Carbofuran, a 

carbamate insecticide used to control the rice water

711
122*00'

38'45' 

I

I Knights I Landing/ " FEATHER RIVER

T°

38°00' 

4 6 12 KILOMETERS

EXPLANATION 

Pesticide sampling site

Figure 1. Pesticide sampling sites in the Sacramento Valley, California.

712
weevil, is incorporated into the soil as a granular 

formulation prior to planting or applied by air after 

flooding. Carbofuran's water solubility and vapor 

pressure are 700 mg/L (25 °C) and 2.7 mPa (33 °C), 

respectively. In irrigation water, the half-life ofcarbofuran ranges from 18 to 25 days (Nicosia, 1989) with 

the primary degradation processes being abiotic 

hydrolysis (Sharom and others, 1980) and microbial 

degradation (Chaudry and Ali, 1988). Thiobencarb, a 

thiocarbamate herbicide used to control grassy weeds, 

particularly barnyard grass, is applied by air when rice 

is at the two-leaf stage (California State Water 

Resources Control Board, 1984). Thiobencarb's water 

solubility and vapor pressure are 30 mg/L (20 °C) and 

2 mPa (20 °C). The primary loss of thiobencarb is 

through photolysis (Crosby, 1983).

This paper presents the results of sampling done 

in the Colusa Basin Drain and Sacramento River during May, June, and July of 1990, 1991, and 1992 and 

discusses the effects of degradation and dilution processes on the concentration of rice pesticides.

METHODS

The CBD and the Sacramento River sites were 

sampled each year during May, June, and July. Water 

was collected using a depth-integrating sampler, such 

as a D-77 (Edwards and Glysson, 1988, p. 13-14). The 

sample volume ranged from 100 mL (milliliter) to 4 L 

(liter), depending on the site and estimated pesticide 

concentration.

The samples were filtered through a 0.7 (o,m 

(micrometer) glass-fiber filter prior to extraction using 

C-8 solid-phase extraction cartridges. The water was 

pumped through the cartridge at a rate of 20 mL/min. 

Terbuthylazine was added to the filtered sample to 

check the extraction efficiency of the cartridge. The 

mean accuracy (percent recovery) and the relative 

standard deviation for molinate, carbofuran, and 

thiobencarb are 70.1± 22.2, 71.6± 35.1, and 56.2± 

17.2. The cartridges were dried by positive pressure 

of high-purity carbon dioxide to remove all the interstitial water. Dried cartridges were eluted with hexane 

ether mixture (1:1 ratio) and evaporated to 200 |o,L. 

The samples were analyzed using a gas chromatograph with an ion trap detector. The detection limits 

for the 1990 analyses were 0.100 (o,g/L for molinate, 

0.050 |j,g/L for carbofuran, and 0.025 |o,g/L for 

thiobencarb. The analytical sensitivity was increased

in 1991 (0.030 ^ig/L for molinate and 0.015 ^ig/L for 

carbofuran).

RESULTS AND DISCUSSION

The maximum pesticide concentrations at the 

CBD and Sacramento River sites decreased each year 

during 1990-92. The molinate, carbofuran, and 

thiobencarb concentrations at CBD are shown in 

figure 2. The maximum concentration of molinate 

decreased from 51 ^ig/L in 1990 to 14.9 ^ig/L in 1991 

and to 4.10 jag/L in 1992, a decrease of about a factor 

of three each year. The average percent recovery and 

standard deviation for molinate are 95.8± 8.2 (Daniel 

Killingsworth, Zeneca Ag Products, written commun., 

1990). For carbofuran, the maximum concentration 

measured in 1990 was 1.1 |o,g/L (California Regional 

Water Quality Control Board, Central Valley Region, 

1991b). This data set had a higher carbofuran detection limit (0.50 (J,g/L); thus, carbofuran was not 

detected in most of the 1990 samples. The average 

percent recovery and relative standard deviation for 

carbofuran was 93± 14 (Smith, 1990). The maximum 

carbofuran concentrations were similar for 1991 and 

1992 (0.308 |.ig/L and 0.229 |.ig/L, respectively). 

Overall, the maximum carbofuran concentration 

decreased by a factor of four from 1990 through 1991, 

but changed slightly from 1991 through 1992. With 

the higher detection limit in 1990 (0.10 ^ig/L), 

thiobencarb was not detected in any of the samples at 

CBD (California Regional Water Quality Control 

Board, Central Valley Region, 1991b). The average 

percent recovery and standard deviation for thiobencarb was 97.7± 11.8 (Daniel Killingsworth, Zeneca Ag 

Products, written commun., 1990). The maximum 

thiobencarb concentration was 0.162 |o,g/L in 1991 and 

0.200 (j,g/L in 1992, again with no appreciable change 

in concentration during 1991 -92.

The State of California's regulations for the 

release of ricefield water has required progressively 

longer holding times in recent years. The increase in 

holding time for ricefield water from 1990 to 1992 

with carbofuran application was zero to 24 days, with 

molinate was 19 to 28 days, and with thiobencarb has 

remained at 30 days. The increase in holding time 

allows for further degradation and volatilization of 

the pesticides in the ricefields. The shallow ricefields 

provide a large surface area and warm temperatures 

(greater than 30 °C) for volatilization. As a mechanism for loss, volatilization is important only for

713
n LU

tr

LLI 

0.

DC 

0 

0

DC

z

0

LJJ

o

z

0

o

» I I I I: I I 111 I I I 1,1 ],: I J I I I II .vi:: II I: I I I III

MAY JUNE JULY

Figure 2. Concentrations of molinate, carbofuran, and thiobencarb measured at Colusa Basin Drain, 

California, 1990-92. The 1990 data for molinate were obtained from California Regional Water Quality 

Control Board, Central Valley Region, 1991b.

714
molinate because molinate vapor pressure is high, 

more than 100 times greater than the vapor pressure 

for carbofuran or thiobencarb. Photolysis, an 

important degradation mechanism for molinate and 

thiobencarb, is enhanced by the large surface area of 

the ricefields. The photodegradation products for 

molinate, 2-keto and 4-keto molinate, were present in 

the CBD and Sacramento River samples (Domagalski 

and Kuivila, 1991). In contrast, carbofuran degradation occurs primarily by abiotic hydrolysis and 

microbial degradation. These degradation processes 

are enhanced by pH values greater than 7 and warm 

temperatures that prevail in the ricefield water 

(Crosby, 1983). The measured decreases in pesticide 

concentration during 1990-92 can be attributed to the 

increased holding times for ricefield water.

The measured concentrations of molinate and 

carbofuran at each of the sampling sites for 1991 are 

shown in figure 3. The highest concentrations were 

measured at the CBD with progressively decreasing 

concentrations downstream. The maximum concentration for molinate decreased from 14.9 |ig/L at CBD 

to 1.17 |4,g/L at Sacramento to 0.347 |ig/L at Rio Vista. 

The maximum concentration for carbofuran decreased 

from 0.308 |ig/L at CBD to 0.086 |ig/L at Sacramento 

to 0.038 |ig/L at Rio Vista. The maximum concentration of thiobencarb measured in 1991 at CBD was 

0.162 |ig/L (fig. 2). Thiobencarb was not detected at 

the Sacramento River sites at Sacramento or Rio Vista 

in 1991.

The decrease in pesticide concentration downstream was caused primarily by dilution and 

degradation. The greatest dilution was between CBD 

and the Sacramento site after the release of ricefield 

discharge water into the high volume Sacramento 

River. The maximum contribution of ricefield 

discharge water to the total flow of the Sacramento 

River is about 30 percent (California State Water 

Resources Control Board, 1984). The decrease in 

maximum concentrations of molinate and carbofuran 

from CBD to Sacramento in 1991 was by a factor 

of about 12 and 4, respectively. The thiobencarb 

concentration decreased from 0.162 |ig/L to less than 

0.025 |ig/L, a factor of 6 or greater. If the decrease in 

carbofuran concentration is attributed entirely to dilution, the greater decrease in molinate and thiobencarb 

concentrations can be attributed to degradation. 

Photolysis and volatilization of molinate and 

thiobencarb occur more readily in CBD than in the 

Sacramento River because of the higher temperatures

and the higher surface area to volume ratio of the CBD 

compared to the Sacramento River.

The decrease in pesticide concentrations from 

the Sacramento to the Rio Vista sites apparently is 

caused primarily by dilution, because the patterns 

of decrease in molinate and carbofuran from the 

Sacramento to Rio Vista sites are similar. Because 

channel depths increase with a reduction in surface 

area to volume ratios, photolysis and volatilization of 

molinate in the river probably are not as important as 

in the shallow agricultural drains. Water samples for 

pesticides were collected at Chipps Island in 1990 and 

1991. Molinate, carbofuran, and thiobencarb concentrations measured on June 5, 1990, were 0.410 |ig/L, 

0.089 fig/L, and 0.044 |ig/L, respectively. In 1991, the 

molinate concentrations measured were 0.083 M£/L 

(July 25) and 0.073 |ig/L (August 1); carbofuran and 

thiobencarb were not detected (detection limits for 

carbofuran and thiobencarb were 0.015 |ig/L and 

0.025 |ig/L, respectively).

Toxicity of the rice pesticides to aquatic 

life in the Sacramento River has been a concern 

(California Regional Water Quality Control Board, 

Central Valley Region, 199la). The acute toxicity of 

carbofuran is greater than that of molinate or thiobencarb. Annual water-quality performance goals of the 

California State Water Resources Control Board are 

daily maximums and apply to all freshwater environments. The 1992 performance goals reflect the 

comparatively high acute toxicity of carbofuran; the 

performance goal is 0.4 |ig/L for carbofuran contrasted 

to 10 |j,g/L for molinate and 1.5 |ig/L for thiobencarb 

(California Regional Water Quality Control Board, 

Central Valley Region, 1993). All rice-pesticide concentrations measured in 1992 during this study were 

less than these performance goals. The increased 

holding time for ricefield water has apparently been 

successful in decreasing the rice pesticide concentrations in the CBD and Sacramento River. However, it 

is important to note that these performance goals are 

based on the toxicity of a single pesticide. In the 

natural environment, aquatic organisms commonly are 

exposed to several pesticides at the same time, which 

can create additive or even synergistic toxic effects.

SUMMARY

During 1990-92, the maximum molinate 

concentration measured in the CBD has decreased by 

a factor of three each year. Carbofuran concentrations

715
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MOLINATE CONCENTRATION, IN MICROGRAMS PER LI

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i i ii i T^i i ] i i i j } n j.i j i i i

r^(crameir|tS 

^Mo'tistiii

10 20 30 

MAY

9 19 29 

JUNE

9 19 29 

JULY

B

Figure 3. Concentrations of (A) molinate and (B) carbofuran measured at Colusa Basin Drain and 

the Sacramento River at Sacramento and Rio Vista, California, 1991 Continued.

717
decreased similarly from 1990 through 1991, but 

decreased only slightly from 1991 through 1992. 

Thiobencarb had no appreciable change in concentrations measured during 1991-92. The decreases in 

pesticide concentration can be attributed to degradation and loss by volatilization in the ricefield as a 

result of increased holding times of irrigation water. 

The pesticide concentrations decreased downstream 

from the CBD along the Sacramento River. Dilution 

seems to be the primary cause of the decreasing 

pesticide concentrations. However, between CBD 

and Sacramento, the molinate and thiobencarb 

concentrations decreased more than the carbofuran 

concentration, indicating that further loss of molinate 

and thiobencarb probably occurred in the CBD by 

photolysis and volatilization.

REFERENCES

California Department of Pesticide Regulation, 1990, 

Pesticide use data: Computer tapes available from 

California Department of Pesticide Regulation, 

Sacramento, CA 95814.

California Regional Water Quality Control Board, Central 

Valley Region, 199 la, 1989 rice season toxicity monitoring results: Final report, 162 p.

 1991 b, Consideration of approving Department of Food 

and Agriculture's 1991 management practices for rice 

pesticides: Staff report, 37 p.

 1993, Staff report on the rice pesticide control program: 

Staff report, 31 p.

California State Water Resources Control Board, 1984, Rice 

herbicides: molinate and thiobencarb: Special Projects 

Report No. 84-4sp., 176 p.

Chaudry, O.K., and Ali, A.N., 1988, Bacterial metabolism 

of carbofuran: Applied and Environmental Microbiology, v. 54, p. 1414-1419.

Crosby, D.G., 1983, The fate of herbicides in California rice 

culture, in Miyamoto, J., and others, eds., International 

Union of Pure and Applied Chemistry Pesticide Chemistry: Human Welfare and the Environment: New 

York, Pergamon Press, 339-346 p.

Domagalski, J.L., and Kuivila, K.M., 1991, Transport and 

transformation of dissolved rice pesticides in the Sacramento River Delta, California, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 664-666.

Edwards, T.K., and Glysson, G.D., 1988, Field methods for 

measurement of fluvial sediment: U.S. Geological 

Survey Open-File Report 86-531, 118 p.

Nicosia, S., 1989, Summary of carbofuran study, 1988: 

California Department of Food and Agriculture, Staff 

memorandum, 42 p.

Sharom, M.S., Miles, J.R.W., Harris, C.R., and McEwin, 

F.L., 1980, Persistence of 12 insecticides in water: 

Water Research, v. 14, p. 1089-1093.

Smith, A.D., 1990, Analytical support of California rice 

pesticide monitoring program for carbofuran: Princeton, N.J., FMC Corporation Agricultural Chemical 

Group Report P-2512, p. 35.

Soderquist, C.J., Bowers, J.B., and Crosby, D.G., 1977, 

Dissipation of molinate in a rice field: Agricultural 

and Food Chemistry, v. 25, p. 940-945.

718
Abstract

Method for Estimating the Activity of

Carbofuran-Degrading Microorganisms

in Environmental Samples

By William J. Meyers 1 anc/Kathryn M. Kuivila1

INTRODUCTION

A method was developed to determine the 

presence and activity of carbofuran degrading 

organisms in aqueous field samples. Carbofuran is 

a broad spectrum systemic insecticide used on rice 

in the Sacramento Valley of California. Following 

the release of treated rice irrigation water, concentrations of carbofuran can be detected in agricultural drains and in the Sacramento River. 

Microbial populations influence the degradation 

rate of carbofuran in soils, but little is known about 

the importance of these organisms in the degradation of carbofuran in aqueous systems. The 

method described uses laboratory-scale microcosms containing an environmental culture and 

carbon-14 ( I4C) ring-labeled carbofuran to evaluate the ability of microorganisms to degrade 

carbofuran. Degradation is indicated by the evolution of l4C-carbofuran-7-phenol and carbon-14- 

carbon dioxide ( I4C-CO2). Ethyl acetate was used 

to extract the I4C ring-labeled compounds from 

aqueous solution. l4C-carbofuran was resolved 

from l4C-carbofuran-7-phenol by means of thin 

layer chromatography, and the I4C quantified 

using a radioisotope plate reader. Sodium hydroxide base traps were used to capture the I4C-CO2 

evolved from the cultures and quantified using a 

scintillation counter. The method was tested using 

three different aqueous cultures. Overall recoveries of I4C were 66.8 percent plus or minus 6.3 

percent at 95-percent confidence interval (n=24).

'U.S. Geological Survey, Sacramento, Calif.

Pesticides are applied to ricefields of the Central 

Valley of California to prevent the growth of plants, 

algae, and insects that reduce rice yields. Rice growers 

are required to retain treated irrigation water on the 

field following pesticide application to promote dissipation before releasing the water into agricultural 

drains that discharge into the Sacramento River. During the spring and summer months, rice return flows 

can account for as much as 30 percent of total flow in 

the Sacramento River (California State Water 

Resources Control Board, 1984). Therefore, the water 

quality of the Sacramento River and the SacramentoSan Joaquin Delta can be greatly affected by the quality 

of rice return water. Pesticide inputs into the river have 

been identified as a possible cause of the recent decline 

in the striped bass population (Phillips, 1987). Acute 

and chronic toxicity tests using rice return water implicate carbofuran as one of the compounds causing toxicity to the freshwater invertebrate Ceriodaphnia dubia 

(Norberg-King and others, 1991).

Carbofuran (2,2-dimethyl-2,3-dihydrobenzofuranyl-7-N-methyl carbamate) is a broad spectrum 

systemic insecticide used on a wide variety of crops in 

the United States. The Sacramento Valley the northern 

third of the Central Valley of California-is one of the 

principal rice producing areas in the United States with 

as many as 500,000 acres in production annually 

(California Regional Water Control Board, Central 

Valley Region, 1991). In 1990, about 40,000 kg of 

carbofuran were used in the Sacramento Valley to 

control the rice water weevil (California Department of 

Pesticide Regulation, 1990). Despite preventive 

measures to dissipate this insecticide, concentrations of 

carbofuran have been detected in the Sacramento 

River. The results of a Lagrangian sampling study 

(Domagalski and Kuivila, 1991) indicate that carbofuran then can be transported conservatively down the 

Sacramento River.

719
The influence of microbial populations on the 

degradation of carbofuran in soils has been well documented (Harris and others, 1988; Parkin and Shelton, 

1992), but little is known about the importance of these 

organisms in the degradation of carbofuran in aqueous 

systems. This paper describes a microcosm method 

that was developed to determine the presence of carbfuran-degrading organisms in aqueous field samples 

and to estimate the importance of these organisms in 

the degradation of carbofuran. The paper was prepared 

as part of the San Francisco Bay Toxic Contaminants 

Study.

The degradation products of carbofuran have been 

well characterized (Metcalf and others, 1968). The 

carbamate moiety is very susceptible to alkaline 

hydrolysis, producing carbofuran-7-phenol (fig. 1).

M

CO + H2O

OH

Figure 1. Main degradation products of (I) carbofuran. 

Structure (II) is carbofuran-7-phenol. The main products of 

abiotic hydrolysis (H) and microbial degradation (M) are 

indicated.

Microorganisms also can hydrolyze carbofuran to 

produce carbofuran-7-phenol and may be able to use 

the carbamate moiety as a sole source of nitrogen 

(Karns and others, 1986; Chaudry and Ali, 1988). 

Microbial activity in soil can result in an accumulation 

of carbofuran-7-phenol (Harris and others, 1984; 

Rajagopal and others, 1984; Chapman and others, 

1985). Organisms also have been isolated that use 

carbofuran as a sole source of carbon, breaking the ring 

structure and fully mineralizing the compound to 

carbon dioxide (CO2 ) and water (H2O) (Chaudry and 

Ali, 1988; Ramanand and others, 1988). By the adding 

of carbon-14 ( I4C) ring-labeled carbofuran to microcosms containing environmental water samples, degradation can be evaluated by examining the production of 

l4C-carbofuran-7-phenol and the evolution of I4C-CO2 .

MATERIALS AND METHODS

Microcosms were contained in 500-mL jars with 

30-mL beakers suspended just below the inner lid by 

means of a wire. The 30-mL beaker held 1 mL of 0.5 N

(normal) solution of sodium hydroxide (NaOH) to trap 

CO2 . Holes in the lid were covered by a Teflon-lined 

septum to allow for aseptic sampling of the contents of 

the airtight jar. The jars initially contained 50,000 

counts per minute of I4C ring-labeled carbofuran 

(specific activity 43.2 u,Ci/mmol) and analytical-grade 

carbofuran to a concentration of 2.5 ppb. Carbofuran 

and l4C-carbofuran were aseptically delivered dissolved in acetone to the bottom of the dried, sterilized 

jars. The jars remained open for 1 hour to allow the 

acetone to evaporate. Aqueous field samples were 

mixed 50/50 with a nitrogen-free, phosphate-buffered, 

mineral salts medium (NFMS) (MgSO4 7H2O, 0.2 g; 

K2HPO4, 6.68 g; KH2PO4 , 22.10 g; FeSO4 7H2O, 

0.001 g; CaSO4, 0.04 g; distilled water to 1 L) 

(Ramanand and others, 1988), and the pH of each 

culture was adjusted to 6.75. Soil samples to be used in 

the microcosms were first diluted by adding 5 g of soil 

to 50 mL of NFMS and stirred overnight. The pH of the 

soil culture also was adjusted to 6.75 before beginning 

the experiment. An Achromobacter species isolate with 

known carbofuran hydrolyzing activity (Karns and 

others, 1986; Derbyshire and others, 1987; Tomasek 

and Karns, 1989) was used as a positive control. The 

Achromobacter inoculum was added to a NFMS supplemented with glucose and succinate (glucose, 1 g; 

sodium succinate, 0.8 g; added to 200 mL of NFMS) 

added as carbon sources. All microcosm jars were 

incubated at 28 °C on an orbital shaker for the duration 

of the experiment.

At periodic intervals, 1 -mL aliquots of the aqueous 

culture were removed from the jars, extracted, and 

analyzed using thin layer chromatography (TLC) 

(fig. 2). The aliquots were acidified to pH less than 1 

and extracted with 1-mL aliquots of ethyl acetate. 

Extractions consisted of 30 seconds on a vortex mixer 

followed by 15 minutes in a sonicator bath. The organic 

layer then was pipetted off and the extraction repeated 

a second time. The resulting 2 mL of ethyl acetate 

extract then was slowly evaporated to about 50 u,L 

using a stream of purified air. TLC was used to separate 

I4C-carbofuran from its metabolites. Each 50 u,L of 

concentrated sample was spotted 1.5 cm apart and 2 cm 

from the bottom edge of an activated silica gel 60 TLC 

plate (20 x 20 cm x 0.25 mm). After the spots dried, the 

plate was developed using a 1:1 hexane-diethyl ether 

mixture. Resolution for the compounds was high 

with approximate retardation factors of Rf = 0.25 and 

Rf = 0.50 for carbofuran and carbofuran-7-phenol, 

respectively.

720
COHCEHTRATOH

SPOT SAMPLE 

OH TIC PLATE DETELOP TLC PLATE

IHAGE TLC 

PLATE

ETHEK/HEXAHE RADIOISOTOPE 

DETECTOR

Figure 2. Quantitative analytical method to separate I4C ring-labeled carbofuran from its 

degradation product l4C-carbofuran-7-phenol using thin layer chromatography (TLC).

Separated compounds were nondestructively 

quantified directly from the TLC plate using a radioisotope detector. The developed plate was first wrapped 

in a single layer of 1.5-fim Mylar film. The TLC plates 

then were-counted for approximately 1,000 minutes 

using a 1.6 x 3.2 x 0.25-mm-resolution plate. The digitally imaged TLC plates were quantified after subtracting background counts. NaOH periodically was 

removed aseptically from the suspended beaker in the 

microcosm jar, and 14C counts were determined by a 

scintillation counter.

RESULTS

The microcosm experiments were done with three 

different aqueous cultures. The first culture was a soil 

sample collected from an organic farm in Yolo County, 

Calif. (Yolo). The farm soil had no known history of 

exposure to carbofuran. The second culture was from

110

an aqueous sample taken from the Colusa Basin Drain 

(CBD), a major agricultural drain servicing a large 

rice-growing area. These waters had been exposed to 

carbofuran, and it is these waters that will be evaluated 

by the methods developed in this project as part of a 

larger study. The third culture tested was a pure culture 

of the carbofuran hydrolyzing Achromobacter species. 

The Achromobacter species served as a positive control for the detection of carbofuran-7-phenol. Also 

included in the experiment was a sterile control for the 

CBD sample. The control was identical to the CBD 

culture, except that it had been sterilized by autoclaving. The sterile CBD sample represented the amount of 

carbofuran-7-phenol evolved through abiotic hydrolysis.

In the Achromobacter species microcosm experiment, greater than 80 percent of the total counts recovered 

from the TLC plate were in the form of 14C-carbofuran7-phenol after 12 days (fig. 3). The percentage of

Achromobacter 

inoculum added

0

CO

u

H 

<J

tt 

w

OH

TIME, IN DAYS

Figure 3. The evolution of l4C-carbofuran-7-phenol as the percentage of total I4C 

recovered after extraction and thin layer chromatographic analysis. (Each point 

represents the average of two replicates (n=2). Bars represent one standard deviation. 

CBD, Colusa Basin Drain.)

721
radio-labeled carbofuran-7-phenol produced by this 

culture continued to increase to 100 percent by 36 days. 

In contrast, the amount of carbofuran-7-phenol 

detected in the CBD microcosm was not significantly 

different from the sterile controls (95-percent confidence interval (n=2)). This indicates that microbial 

degradation was not responsible for the small production of carbofuran-7-phenol in the CBD culture. This 

also demonstrates the importance of having a sterile 

control for each microcosm treatment. The Yolo microcosm was inoculated after 27 days with cultures taken 

from the Achromobacter microcosm. Nine days later, 

microbial degradation was indicated by the detection of 

74 percent of total I4C counts recovered as l4C-carbofuran-7-phenol in the Yolo microcosm. The rapid 

hydrolysis of carbofuran in the Yolo microcosm after 

adding the Achromobacter organisms indicated that 

microcosm conditions were suitable for microbial 

activity and the carbofuran was bioavailable. Therefore, the absence of microbial degradation products 

probably was due to a lack of carbofuran degrading 

organisms in the Yolo soil. Overall percentage recoveries were obtained by comparing the 14C counts in 1 mL 

of culture counted immediately on a scintillation 

counter to the I4C counts recorded after extraction and 

analysis of the TLC plate. The average recovery/ 

extraction efficiencies for all samples shown in figure 3 

was 66.8 percent ±6.3 percent at 95-percent confidence 

interval (n=24).

CONCLUSIONS

The microcosm method can estimate the microbial 

degradation of carbofuran from various field samples. 

The analytical procedures described allow for the processing of up to 12 samples (20 x 20 cm x 0.25 mm 

TLC plate) simultaneously. During the experiments the 

sodium hydroxide base traps were sampled and I4C 

counts were determined as an estimate of ring-cleavage 

carbofuran. The results indicated a problem with the 

experimental method, perhaps because of an increase 

in the volatility of carbofuran-7-phenol. Further study 

is needed to determine the presence and importance of 

carbofuran-degrading microorganisms in the agricultural drains and waterways of the Sacramento River. 

Additional research also could determine the effects of 

environmental conditions, such as dissolved organic 

carbon and nitrogen concentration, on these degrading 

populations.

REFERENCES

California Department of Pesticide Regulation, 1990, Pesticide use data: Computer tapes available from California 

Department of Pesticide Regulation, Sacramento, CA 

95814.

California Regional Water Quality Control Board, Central 

Valley Region, 1991, 1989 rice season toxicity monitoring results: Final report, 162 p.

California State Water Resources Control Board, 1984, Rice 

herbicides: molinate and thiobencarb: Special Projects 

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319.

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of carbofuran: Applied and Environmental Microbiology, v. 54, no. 6, p. 1414-1419.

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Aronson, D.A., eds., U.S. Geological Survey Toxics 

Substances Hydrology Program Proceedings of the 

technical meeting, Montcrey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 664-666.

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treatment: Journal of Environmental Science and 

Health, v.B 19, no. l,p. 1-11.

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A.B., 1988, Enhanced soil microbial degradation of 

carbofuran and fensulfuron a factor contributing to 

the decline in effectiveness of some soil insect control 

programs in Canada: Journal of Environmental Science 

and Health, v. B23, no. 4, p. 301-316.

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1986, Metabolism of carbofuran by a pure bacterial 

culture: Pesticide Biochemistry and Physiology, v. 25, 

p. 211-217.

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Metabolism of 2,2-dimethyl-2,3-dihydrobenzofuranyl7-N-methyl Carbamate (FuradanCR) ) in plants, insects, 

and mammals: Journal of Agriculture and Food Chemistry, v. 16, no. 2, p. 300-311.

722
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Francisco Bay-Delta and their possible biological 

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723
724
Distributions of Organochlorine Compounds in

Sediments and Livers of Striped Bass (Morone Saxatilis)

in San Francisco Bay-Delta Estuary

By Wilfred E. Pereira1 , Frances D. Hostettler1 , John R. Cashman2

and Richard S. Nishioka3

Abstract

A preliminary assessment was made in 

1992 of chlorinated organic compounds in 

sediments and in livers of striped bass from the 

San Francisco Bay-Delta Estuary. Samples of 

sediment and striped bass livers contain DOT 

(ethane, 1,1,1 -trichloro-2,2-bis (p-chlorophenyl)-) 

and its degradation products, DDD(ethane, 1,1- 

dichloro-2,2-bis(p-chlorophenyl)-) and DDE 

(ethylene, 1,1 -dichloro-2,2-bis (p-chlorophenyl)); 

polychlorinated biphenyls (PCB's); alpha and 

gamma chlordane, and cis and trans nonachlor. 

In addition, livers of striped bass contained small 

concentrations of DCPA (dimethyl 

tetrachloroterephthalate), a pre-emergent 

herbicide. The major source of DDT in the bay 

probably is from runoff from the San Joaquin and 

Sacramento Rivers. A point source of DDT near 

Richmond also may be responsible for the high 

concentrations of DDT in the Central Bay. Ratios 

of pentachloro to hexachloro isomers of PCB's 

suggest that sources of PCB's to the South, 

Central, and North Bays differ. Concentrations of 

alpha chlordane in livers of striped bass were 

greater than those of gamma chlordane suggesting 

a greater environmental stability and persistence 

of alpha chlordane. Trans nonachlor, a minor 

component of technical chlordane, was present in

'U.S. Geological Survey, Menlo Park, CA.

2 IGEN Research Institute, Seattle, WA.

 ^Department of Integrative Biology, Cancer Research 

Laboratory, and Bodega Marine Laboratory, University of 

California, Berkeley, CA.

greater concentrations than alpha and gamma 

chlordane and cis nonachlor. Trans nonachlor is 

more resistant to metabolism than alpha and 

gamma chlordane and cis nonachlor, and serves 

as an environmentally stable marker compound. 

Chlorinated organic compounds are 

bioaccumulated in livers of striped bass. These 

compounds may contribute to the decline of 

striped bass in San Francisco Bay-Delta Estuary.

INTRODUCTION

San Francisco Bay Estuary is the largest estuary 

in the Western United States. It is located at the mouth 

of the Sacramento and San Joaquin Rivers and drains 

about 40 percent of the State of California. The 

estuary is affected by urban, industrial, and 

agricultural activities. Agricultural practices in the 

Central Valley of California have resulted in non-point 

source contamination of the estuary by several firstgeneration organochlorine pesticides. These 

compounds are transported from farmlands by the 

Sacramento and San Joaquin Rivers into San 

Francisco Bay. For example, the Colusa Basin drain 

which discharges to the Sacramento River at Knights 

Landing, conveys runoff and agricultural return-flows 

from about 1 million acres of farmland. Chlorinated 

organic compounds, including DDT, chlordane, 

Dacthal, nonachlor, and PCB's, have been reported in 

fish and sediments from the Colusa Basin Drain 

(Turek, 1990). Many of these lipophilic organic 

compounds are associated with the suspended 

particulate load of the Sacramento and San Joaquin 

Rivers. These compounds are eventually dispersed 

throughout the estuary by tides and prevailing winds

725
to become incorporated into estuarine sediments and 

accumulated in the lipophilic tissue of biota. In 

addition, several chlorinated organic compounds such 

as polychlorinated biphenyls (PCB's) and DOT from 

point sources located on the shore of the bay have also 

contributed to estuarine contamination.

Striped bass (Morone saxatilis) were first 

introduced into the San Francisco Bay Estuary from 

Chesapeake Bay in 1879. These bass supported a 

commercial fishery until the 1930's. After the 

commercial fishery was terminated, the estuary was 

able to support a large striped bass population for 

recreational fishing. However, the striped bass 

population in the estuary has declined significantly 

since the mid-1970's (Arthur and others, 1991). Some 

of the causes for the decline in the striped bass 

population include (1) reduction in egg production, 

(2) loss of larval fish into water diversion projects, (3) 

food limitations in the striped bass food chain, and (4) 

toxic effects of pollutants (Whipple and others, 1983).

Striped bass adults are anadromous and migrate 

in spring from the Pacific Ocean and San Francisco 

Bay to their spawning grounds in the lower San 

Joaquin and Sacramento Rivers. Spawning occurs 

primarily in May. Each year hundreds to thousands of 

dead striped bass are observed in the Sacramento-San 

Joaquin estuary and Carquinez Strait (Kohlhorst, 

1975). It has been suggested that liver dysfunction 

may be an important aspect of the pathology offish in 

this die-off (Brown and others, 1987; Young and 

others, 1993). Urquhart and Knudsen (1987) 

reported that striped bass collected from the Delta 

contained hydrocarbons, chlorinated hydrocarbons 

and heavy metals. Recently it has been shown that the 

livers of moribund striped bass were greatly 

contaminated by organic chemical compounds 

compared to the livers of healthy fish caught in the 

Delta and the Pacific Ocean (Cashman and others, 

1992). Organochlorine insecticides may cause high 

mortality in fish eggs and fry with drastic effects on 

survival (Holden, 1975; D. Ostrach, University of 

California, Davis, CA, oral commun. 1992), but the 

effects of organochlorine compounds on striped bass 

and fry are not well understood.

Because the hepato-biliary tract is the primary 

site of bioaccumulation and metabolism of organic 

contaminants and the potential site of chlorinated 

organic compounds, livers of striped bass were 

selected as the target organs for this study. This report 

describes some results of a preliminary assessment of

the distributions of chlorinated organic compounds in 

bottom sediments and livers of striped bass in the San 

Francisco Bay-Delta Estuary.

MATERIALS AND METHODS

Site description

Sample collection sites for bed sediments and 

striped bass are shown in figure 1 and table 1. In 

January 1992, bed sediment was collected from 17 

different sites in North (San Pablo), Central, and South 

San Francisco Bay. Mature striped bass were captured 

in mid-May 1992, from the Sacramento River at 

Knights Landing and from the San Joaquin River.

122"30' 122*15' 122°00' 121*45' 121*30'

 Sediment Samples 

o Striped Bass Samples

Figure 1. Location map showing sampling sites for bottom 

sediments and striped bass (Morone saxatilis) in San 

Francisco Bay-Delta Estuary.

726
Table 1. Sampling sites for bottom sediments in San 

Francisco Bay

Site Number 

(See Fig. 1) Site Name

25 Mouth of Coyote Creek

26 Mouth of Redwood Creek

27 East end, South side, San Mateo Bridge

28 Oyster Point entrance

29 Army Street, Pier

30 Shoreline, Alameda

31 Oakland Outer Harbor

32 Berkeley Marina

33 Richmond Long Wharf

34 Larkspur Channel, End #2

35 Rat Rock, China Camp

36 East of Pinole Point

37 Petaluma, Rev. #12

38 Sonoma Creek Mouth

39 East of Petaluma, #4

40 Richardson Bay (Shallow)

41 Mid-channel, East of San Bruno Shoal

The fish were procured by personnel from the Department of Fish and Game for purposes of spawning (in 

captivity) so that viable fry could be raised and used to 

repopulate the rivers and lakes of California.

Sample collection and analysis

Bed sediments

Samples of surficial bed sediment were 

collected with a Van Veen grab sampler. The upper 

7 cm of sediment were sampled using a stainless steel 

coring barrel. The sample was stored in a clean glass 

jar and refrigerated. Samples were air dried and 

analyzed by procedures described earlier (Hostettler 

and others, 1989). The aromatic fraction containing 

the PCB's, DDT, and chlordane was analyzed by gas 

chromatography-mass spectrometry (GC-MS).

Striped Bass

In mid-May, five male and two female fish were 

captured by electrofishing. The fish were used for 

spawning at the Central Valley Hatchery of the 

California Department of Fish and Game. In addition, 

five fish which were reared at the Bodega Marine 

Laboratories, served as controls. Control fish were

offspring of adults from the delta and were raised in 

captivity. The fish were dissected, and livers, blood 

and other tissue were rapidly removed and frozen. 

The livers were analyzed by a previously described 

procedure (Cashman and others, 1992). Extracts were 

chromatographed on a florisil column. A hexaneethylacetate fraction containing the compounds of 

interest was analyzed by GC-MS.

Gas chromatography-mass spectrometry

All samples were analyzed on a Finnigan Mat 

Magnum gas chromatography-ion trap mass 

spectrometer system, using a 30m long DBS-MS 

capillary column. The linear temperature program 

was started at 60 °C and increased to 30 °C at a rate of 

6 °C per minute. All compounds were confirmed by 

comparing their mass spectra and retention indices 

with authentic standard compounds.

RESULTS AND DISCUSSION

Chlorinated organic compounds in sediments

DDT and its degradation products.

DDT was first introduced as an insecticide about 

1945. Because of its toxicity, it was banned in 1972. 

It is estimated that approximately 1900 kg (or 2 metric 

tons) of DDT and its degradation products entered San 

Francisco Bay in 1967 from the drainage waters of the 

principal agricultural areas of California (Risebrough 

and others, 1968; Bailey and Hannum, 1967). DDT, 

PCB's, and chlordane residues were reported in 

sediments from 26 streams tributary to San Francisco 

Bay (Law and Goerlitz, 1974). Low concentrations of 

DDT and metabolites (1-6 ng per gram) were reported 

in sediments of San Francisco Bay by the National 

Status and Trends Program of the National Oceanic 

and Atmospheric Administration (NOAA) (1987). A 

former DDT-processing plant also was located near 

Richmond, and it represents a point-source of DDT 

within San Francisco Bay.

Distributions of total DDT (DDT+DDD+DDE) 

in sediments of San Francisco Bay are shown in 

figure 2 and table 2. Concentrations ranged from 1.4 

to 9 ng/g of sediments. The source of DDT in the 

North Bay probably is from run-off from the 

Sacramento and San Joaquin Rivers. The San Joaquin 

River has among the highest bed sediment

727
122*30' 122151

122*00'

Suisun Bay

37°45'

Carquinez 

Strait

Richmond

!*>

Total DDT 

in nanograms per gram

^m^Mmmmm

37°30'

Figure 2. Concentrations of total DOT (DOT + DDE + 

ODD) in bottom sediments in San Francisco Bay, 1992.

concentrations of DOT, DDE, and DDD of major 

rivers in the United States (Gilliom and Clifton, 1990). 

The high concentrations of DDT near Richmond and 

across the bay at Larkspur Channel may be derived 

from the point source of DDT near Richmond. High 

concentrations of DDT in sediments are also found 

near Oakland. A sediment sample collected in the 

South Bay near the mouth of Coyote Creek contained 

the greatest concentration of DDT. Coyote Creek 

receives effluents from the San Jose/Santa Clara Water 

Pollution Control Plant.

Polychlorinated Biphenyls (RGB's).

PCB's were used worldwide as dielectric fluids in 

capacitors and in other industrial applications. It is 

estimated that of 5x105 metric tons of PCB's produced 

in the United States during 1930-70, approximately 

3xl05 metric tons, or 78 percent has been lost to the 

environment (Nisbet and Sarofin, 1972). The National 

Status and Trends Program of NOAA (1987) found that 

concentrations of PCB's in sediments collected from 4 

locations within San Francisco Bay ranged from 9 to 

61 ng/g. Unpublished data of the United States 

Environmental Protection Agency (USEPA) (Phillips

and Spies, 1988) from 10 sites in South Bay sediments 

indicate concentrations of PCB's in the range of from 

below detection limits to 34 ng/g. Schmidt and others 

(1971) report that one of the major sources of PCB's to 

California Coastal waters are urban sewage outfalls. 

They estimated that the City of Richmond Water 

Pollution Control Plant, the East Bay Municipal Utility 

District (EBMUD), and the City of San Francisco 

Southeast Sewage Treatment Plant discharged about 2.6 

kg/d of PCB's, or about 1 metric ton per year into San 

Francisco Bay (Schmidt and others, 1971).

Concentrations of PCB's in sediments collected 

from the 17 sites in San Francisco Bay are shown in 

figure 3 and table 2. Concentrations ranged from 1.3 to 

8.1 ng/g total PCB's (sum of tetra, penta and 

hexachlorobiphenyl isomers). These concentrations are 

of the same order of magnitude as concentrations of 

DDT and its degradation products. The ratio of the 

pentachloro to hexachloro isomers was examined to 

determine source inputs of PCB's. This ratio ranged 

from 0.4 to 1.0 in the South Bay sediments, and from 

1.1 to 4.3 in the Central and North Bay sediments. The 

ratios suggest different sources of PCB's to the South 

Bay and to the Central and North Bay sediments.

122*3<r lEW

Larkspur 

Channel ilflll Richmond

P:Si% 

-,,x y,,,,11113.^

San HilL 

Francisco 4ij |||

Total PCB's 

in nanograms per gram

I o

San Mateo 

Bridge

Coyote Creek {

Figure 3. Concentrations of total RGBs (tetra + penta + 

hexachloroisomers) in bottom sediments in San Francisco 

Bay, 1992.

728
Table 2. Concentrations of chlorinated organic contaminants in bottom sediments from San Francisco Bay in 1992

[All values in nanograms per gram dry weight]

COMPOUND

DDE

ODD

DOT

TOTAL DOT

GAMMA CHLORDANE

ALPHA CHLORDANE

TRANS NONACHLOR

CIS NONACHLOR

TOTAL CHLORDANE

TETRACHLORO PCB

PENTACHLORO PCB

HEXACHLORO PCB

TOTAL PCB

DCPA

#25

3.9

1.9

3.2

9

0.7

<0.1

0.5

0.3

1.5

3

2.3

2.2

7.5

<0.1

#26

1.5

1.3

0.2

3

0.1

<0.1

0.1

<0.1

0.2

1.5

3.2

3.4

8.1

<0.1

#27 #28

0.7 1.1

0.7 3.1

0.1 0.2

1.5 4.4

<0.1 0.1

<0.1 <0.1

<0.1 0.7

<0.1 <0.1

<0.1 0.8

0.5 0.9

0.8 1.3

1.4 1.7

2.7 3.9

<0.1 <0.1

SITE

#29 #30 #31 #32 #33

0.7 0.5 1.3 1.4 1.8

0.1 0.8 2.6 2.7 6.3

<0.1 <0.1 3 1.5 0.7

0.8 1.3 6.9 5.6 8.8

<0.1 <0.1 <0.1 <0.1 <0.1

<0.1 <0.1 <0.1 <0.1 <0.1

<0.1 <0.1 0.1 <0.1 <0.1

<0.1 <0.1 <0.1 <0.1 <0.1

<0.1 <0.1 0.1 <0.1 <0.1

0.8 0.5 0.9 1 1.8

1 0.8 2.9 1.5 2.8

2.4 1.5 2.3 1.3 1.7

4.2 2.8 6.1 3.8 6.3

<0.1 <0.1 <0.1 <0.1 <0.1

#34 #35 #36

1.8 1.5 <0.1

2.8 2.4 <0.1

2.5 0.5 <0.1

7.1 4.4 <0.1

<0.1 0.1 <0.1

<0.1 <0.1 <0.1

<0.1 0.1 <0.1

<0.1 <0.1 <0.1

<0.1 0.2 <0.1

1 0.7 <0.1

1.1 1.2 <0.1

0.6 0.8 <0.1

2.7 2.7 <0.1

<0.1 <0.1 <0.1
Chlordane

Technical chlordane is used as a pesticide in 

residential and agricultural areas. It is estimated that 

more than 6.3x10 metric tons of chlordane have been 

produced since 1946 (Dearth and Kites, 1991). About 

20 to 50 percent of all the technical chlordane ever 

produced still exists unaltered in the environment. In 

1988, the sale and use of technical chlordane were 

suspended. Technical chlordane contains a mixture of 

more than 140 compounds. The major components of 

technical chlordane are alpha and gamma chlordane 

and cis and trans nonachlor.

A high incidence (92 percent of the samples) of 

chlordane residues in sediments of tributaries to San 

Francisco Bay has been reported (Law and Goerlitz, 

1974). Trace levels of cis and trans chlordane and of 

trans nonachlor in sediments from Islais Creek, on the 

eastern side of the San Francisco peninsula, were 

reported (Chapman and others, 1986). Distributions of 

total chlordane (gamma chlordane+alpha 

chlordane+fra/r? nonachlor + cis nonachlor) in this 

study of sediments in San Francisco Bay are shown in 

table 2 and figure 4. Chlordane residues are present

12T30' 1221T 122W

Total Chlordane 

in nanograms per gram

: ...... Oakland San - "" " ^

I Francisco >0.1!ilM

Figure 4. Concentrations of total chlordane (alpha + 

gamma chlordane + cis + trans nonachlor) in bottom 

sediments in San Francisco Bay, 1992.

mainly in the South Bay. These chlordane residues 

probably originate from domestic use of chlordane as a 

termiticide and insecticide. The highest concentration 

of chlordane was found in the sediment from the 

vicinity of the mouth of Coyote Creek which receives 

effluents from the San Jose/Santa Clara Water Pollution 

Control Plant.

Chlorinated organic compounds in livers of 

striped bass

DOT and its degradation products

DOT and its degradation products have been 

found in bluegills and common carp from the San 

Joaquin River and tributaries (Saiki and Schmitt, 1986). 

In carp, the concentrations of total DDT approached the 

highest levels reported by the National Pesticide 

Monitoring Program for fish from other intensively 

farmed watersheds of the United States from 1980- 

1981. Total DDT in striped bass have been reported 

since the early 1980's (Crosby and others, 1986).

Concentrations of DDT, ODD and DDE in 

livers of striped bass and control striped bass are 

shown in table 3. Ultratrace levels of DDT and its 

degradation products were found in some of the 

control samples, possibly due to transfer of DDT from 

female fish to offspring via the egg-yolk lipid (see 

Materials and Methods) (Burdick and others, 1964). 

In addition, samples 9 and 10 (female striped bass) 

contained lower levels of DDT than samples 4, 5, 6, 7 

and 8 (male striped bass). In general, concentrations 

of DDT compounds were in the order 

DDE>DDD>DDT

Chlordane

Chlordane residues have been reported in 

bluegills and common carp from the San Joaquin 

River and tributaries (Saiki and Schmitt, 1986), and in 

striped bass (Whipple and others, 1983). Concentrations of chlordane in livers of striped bass and 

control samples from this study are shown in table 2. 

In general, concentrations of chlordane are smaller in 

livers of female striped bass (sample 9 and 10) than in 

livers of male striped bass (samples 4,5,6 and 7). 

Chlordane was not detected in any of the control 

samples. Concentrations of alpha (cis) chlordane were 

greater than those of gamma (trans) chlordane in 

livers of striped bass (table 2). Alpha chlordane is 

environmentally more stable and hence more 

persistent than gamma chlordane, and thus dominates gamma chlordane in environmental samples.

730
Table 3. Concentrations of chlorinated organic contaminants in livers of striped bass (Morone saxatilis) in 1992

[All values in nanograms per gram wet weight]

SAMPLES

COMPOUND

DDE

ODD

DOT

TOTAL DOT

GAMMA CHLORDANE

ALPHA CHLORDANE

TRANS NONACHLOR

CIS NONACHLOR

TOTAL CHLORDANE

TETRACHLORO PCB

PENTACHLORO PCB

HEXACHLORO PCB

TOTAL PCB

DCPA

#4

103

35

8.8

147

4.4

5.7

10

4.8

25

13

50

77

140

1.9

#5

131

47

11

189

5.6

9.4

18

9.2

42

1.9

113

155

270

<0.1

#6

297

86

13

396

6.1

7.4

11

4.5

29

7.4

18

11

36

8.7

#7

172

40

13

225

5.1

7.1

15

7.4

35

9.3

56

84

149

5.8

#8

137

39

11

187

1.3

2.1

6.2

3.1

13

2.9

19

22

44

0.7

#9

104

32

6.1

142

3.4

5

6.3

3

18

5

13

17

35

3.1

#10 #12 #13 

62 3.4 <0.1 

34 1.5 0.2 

7.3 0.8 0.2 

103 5.7 0.4 

5.3 <0.1 <0.1 

6.2 <0.1 <0.1 

7.1 <0.1 <0.1 

1.4 <0.1 <0.1 

20 <0.1 <0.1 

24 <0.1 <0.1 

35 <0.1 <0.1 

8 <0.1 <0.1 

67 <0.1 <0.1 

2 <0.1 <0.1 
The ratio of alpha chlordane to gamma chlordane is 

about 0.7 to 1.0 in technical chlordane. However, the 

ratio is about 1.2 to 1.7 in striped bass livers again 

indicating the overall environmental stability of the 

alpha isomer. The major components in technical 

chlordane include gamma chlordane (24 percent), 

alpha chlordane (19 percent), and trans nonachlor (7 

percent). In striped bass livers, the concentration of 

trans nonachlor is greater than alpha and gamma chlordane. Therefore, trans nonachlor is probably more 

resistant to metabolism than alpha and gamma chlordane. The resistance of trans nonachlor over alpha 

and gamma chlordane to metabolism in human tissues 

extracts has been reported (Sovocool and Lewis, 

1975).

Polychlorinated Biphenyls (PCB's)

Stevens (1980) reported concentrations of 

PCB's in striped bass from the Delta for the period 

1964-1976. Crosby and others (1986) postulated that 

concentrations of PCB's in striped bass might be 

sufficient to account for reproductive problems, 

including the resorption of eggs.

Concentrations of PCB's in the livers of striped 

bass and control samples are shown in table 3. PCB's 

were not detected in any of the control samples. Total 

concentrations of PCB's in livers of female striped 

bass (sample 9 and 10) were generally lower than 

samples 4,5,7, and 8 (male striped bass).

DCPA (Dacthal, dimethyltetrachloroterephthalate).

DCPA is a pre-emergent herbicide that is widely 

used in the San Joaquin and Sacramento River Basins 

on crops such as onions, cabbage and cotton. DCPA 

has been found in bluegills and common carp from the 

San Joaquin River and its tributaries (Saiki and 

Schmitt, 1986). Concentrations of DCPA in livers of 

striped bass are shown in table 3. DCPA was 

identified in six out of seven striped bass samples 

analyzed. DCPA was not detected in control fish.

CONCLUSIONS

Bottom sediment from San Francisco Bay and 

striped bass (Morone soxatilis) from the Sacramento 

and San Joaquin Rivers are contaminated with 

organochlorine compounds derived from agricultural 

and industrial sources. Although many of these 

compounds are banned from use, they continue to

persist in the estuary and bioaccumulate in the livers 

of striped bass. These chlorinated organic compounds 

may be responsible in part for the hepatotoxic effects 

observed and may be contributing to the continuing 

decline of the striped bass population in San Francisco 

Bay. Striped bass appear to be useful bioindicators of 

contamination by organochlorine compounds in San 

Francisco Bay-Delta Estuary.

ACKNOWLEDGMENTS

We acknowledge the financial Support of the 

California Department of Fish and Game (Agreement 

FG1381) administered by the Striped Bass Stamp 

Committee. This work is a result of research funded 

in part by a grant from the National Sea Grant College 

Program, National Oceanic and Atmospheric 

Administration, U.S. Department of Commerce, under 

grant number NA89AA-D-SG138, project number 

R/F-145, through the California Sea Grant College, 

and in part by the California State Resources Agency. 

The authors wish to thank the crew of the RV Polaris 

for assistance with sampling of bottom sediments.

REFERENCES

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Coll ins, G. 1991, 1990 Striped bass egg and larvae 

management studies San Francisco Bay-Delta Estuary 

executive summary: U.S. Department of the Interior, 

Bureau of Reclamation, Mid-Pacific Region and 

Denver Office, Denver, Colorado, p. 1-18.

Bailey, T.E., Asce, A.M., Hannum, J.R., 1967, Distributions 

of pesticides in California: Journal of the Sanitary 

Engineering Division of the American Society of Civil 

Engineers, v.93, no. 27, p. 27-43.

Brown, C. L., Young, G., Nishioka, R.S., Bern, H.A., 1987, 

Preliminary report of the physiological status of striped 

bass in the Carquinez Strait die-off: Fisheries 

Research, v.6, p.5-16.

Burdick, G.E., Harris, E.J., Dean, H.J., Walker, T.M., Skea, 

J., Colby, D., 1964, The accumulation of DDT in lake 

trout and the effect on reproduction: Transactions of 

the American Fish Society, v. 93, p. 127-136

Cashman, J.R., Maltby, D.A., Nishioka, R.S., Bern, H.A., 

Gee, S.J., Hammock , B.D., 1992, Chemical 

contamination and the annual summer die-of of striped 

bass (Morone saxatilis) in the Sacramento-San Joaquin 

Delta: Chemical Research in Toxicology, v.5, no.l., 

p. 100-105.

732
Chapman, P.M., Dexter, R.N., Cross, S.F., Mitchell, D.G., 

1986, A field trial of the sediment quality triad in San 

Francisco Bay: Rockville, Maryland, National 

Oceanic and Atmospheric Administration, NOAA 

Technical Memorandum NOS OMA 25, 134 p.

Crosby, D.G., Hogan, K., Bowes, G.W., Foster, G.L., 1986, 

The potential impact of chlorinated hydrocarbon 

residues on California striped bass, in Report of the 

Cooperative Striped Bass Study, California: State 

Water Resources Control Board, Sacramento, 

Technical Supplement I, Section I, 152 p.

Dearth, M.A., Hites, R.A., 1991, Complete analysis of 

technical chlordane using negative ion mass 

spectrometry: Journal of Environmental Science and 

Technology, v.25, no. 2, p. 245-254.

Gilliom, R.J., Clifton, D.G., 1990, Organochlorine pesticide 

residues in bed sediments of the San Joaquin River, 

California: Water Resources Bulletin, v.26, no. 1, p. 11 - 

23.

Holden, A.V., 1975, Monitoring persistent organic 

pollutants, in Moriarty, F., ed.,"Organochlorine 

Insecticides Persistent Organic Pollutants", 

Academic Press, London, p. 1-27.

Hostettler, F.D., Rapp, J.B., Kvenvolden, K.A., Luoma, 

S.A., 1989, Organic markers as source discriminants 

and sediment transport indicators in South San 

Francisco Bay California: Geochimica et 

Cosmochimica Acta, v. 54, p. 1563-1576.

Kohlhorst, D.W., 1975 , The striped bass (Morone saxatilis) 

die-off in the Sacramento-San Joaquin estuary in 1973 

and a comparison of its characteristics with those of 

1971 and 1972 die-off: Sacramento, California, 

California Department of Fish and Game, Anadromous 

Fish Branch, Adm. Rep. No. 74-13.

Law, L.M., Goerlitz, D.F., 1974, Selected chlorinated 

hydrocarbons in bottom material from streams 

tributary to San Francisco Bay: Pesticides Monitoring 

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transport of PCB's in the environment: Environmental 

Health Perspective, v. 1, p. 21-38.

NOAA, 1987, National Status and Trends Program for 

Marine Environmental Quality: Progress report and 

preliminary assessment of findings of the benthic 

surveillance project 1984: Rockville, Maryland, 

National Oceanic and Atmospheric Administration, 

81 p.

Phillips, D.G.H., Spies, R.B., 1986, Chlorinated 

hydrocarbons in the San Francisco estuarine 

ecosystem: Marine Pollution Bulletin, v. 19, no 9, 

p. 445-453.

Risebrough, R.W., Huggett, R.J. Griffin, J.J., Goldberg, 

E.D., 1968, Pesticides Transatlantic movements in 

the northeast trades: Science, v. 159, p. 1233-1236.

Saiki, M.K., Schmitt, C.J., 1986, Organochlorine chemical 

residues in bluegills and common carp from the 

irrigated San Joaquin Valley Floor, California: 

Archives of Environmental Contamination and 

Toxicology, v.15, p.357-366.

Schmidt, T.T., Risebrough, R.W, and Gress, F., 1971, Input 

of polychlorinated biphenyls into California costal 

waters from urban sewage outfalls: Bulletin of 

Environmental Contamination and Toxicology, v. 6, 

no. 3, p. 235-243.

Sovocool, G.W., Lewis, R.G. 1975, The identification of 

trace levels of organic pollutants in human tissues: 

Compounds related to chlordane/heptachlor exposure, 

in Trace substances in environmental health-IX; 

Proceedings of the University of Missouri 9th Annual 

Conference on Trace Substances in Environmental 

Health, June 10-12: p. 265-280.

Stevens, D.E. 1980, Factors affecting striped bass fisheries 

of the west coast, in Clepper, H., (ed.), Marine 

Recreational Fisheries: Washington, D.C., Sport 

Fishing Institute, p. 15-28.

Turek, S.M., 1990, Colusa Basin drain water quality 

literature review: California Department of Water 

Resources, Northern District, Memorandum Report, 

72 p.

Urquhart, K.A.F., Knudsen, D.L., 1987, Striped bass health 

monitoring final report: California Department of 

Fish and Game, Final report for intcragency 

agreement, 6-170-300-0, Stockton, California State 

Water Resources and Control Board, 136 p.

Whipple, J.A., Crosby, D. G., Jung, M., 1983, Third 

progress report. Cooperative Striped Bass Study: 

Tiburon, CA, California State Water Resources 

Control Board Special Projects Report No. 83-35 sp: 

S.W Fisheries Center, 208 p.

Young, G., Brown, C.L., Nishioka, R.S., Folmar, L.C., 

Andrews, M., Cashman, J.R., Bern, H.A., 1993, 

Histopathology, blood chemistry, and physiological 

status of normal and moribund striped bass (Morone 

saxatilis) involved in summer mortality ("die-off") in 

the Sacramento-San Joaquin Delta of California: 

Journal of Fish Biology, (in press).

733
734
A Survey of Molecular Marker Compounds in Sediments 

of San Francisco Bay, California

By Frances D. Hostettler1 , John B. Rapp1 , Wilfred E. Pereira1 , 

and Keith A. Kvenvolden 1

Abstract

An areal survey of surficial sediments in 

San Francisco Bay has been conducted to 

evaluate the presence of extractable organic 

compounds. Molecular marker compounds 

studied include hydrocarbons (PAH's), 

chlorinated pesticides, and other organic 

compounds that provide information on sources 

of organic input into the Bay. Fairly uniform 

source profiles are seen throughout the Bay. 

Biomarker profiles contain mature constituents 

indicating anthropogenic influences and extensive 

sediment reworking. The dominant input 

signatures in San Francisco Bay sediment are 

those of anthropogenic PAH's from combustion 

and other sources, and long chain «-alkanes and 

^-aldehydes from terrigenous vascular plants. A 

comparison of the sums of the combustion PAH's 

and the terrigenous «-alkanes shows that an 

anthropogenic influence is dominant in the most 

urban parts of the Bay close to shore, and in midBay channel areas, and a terrigenous signature is 

dominant nearshore in San Pablo Bay and at the 

southernmost station in South Bay.

INTRODUCTION

San Francisco Bay is one of the world's largest 

urbanized estuarine systems. Its water and sediment 

receive organic input from a wide variety of sources. 

These sources include land, street, and river run-off; 

industrial, agricultural, and sewage discharges;

'U.S. Geological Survey, Menlo Park, CA.

atmospheric deposition; shipping traffic; and other 

anthropogenic inputs (Nichols and others, 1986), as 

well as autochthonous marine material such as algae, 

phytoplankton, and zooplankton. Some of the large 

array of organic chemicals introduced from these 

sources are toxic and, therefore, need to be identified 

and their accumulations monitored. Identifying the 

sources of input, both toxic and nontoxic, is a helpful 

step in assessing toxic accumulations as well as 

providing insight into estuarine sediment dynamics 

(Hostettler and others, 1989).

This present study is part of an on-going 

investigation of sources of organic constituents in 

surficial sediments in San Francisco Bay. Surficial 

sediments were collected from 17 sites throughout 

southern, northern (San Pablo), and central San 

Francisco Bay, along with one gravity core from 

Richardson Bay, an embayment within the Central 

Bay (fig. 1). These sediments were analyzed for 

several categories of organic compounds in order to 

evaluate distributions, concentrations, and possible 

source signatures (that is, a compound or group of 

compounds common to one source that identifies that 

source). The use of specific organic compounds to 

determine likely sources of organic matter in 

sediments is well established (Volkman and others, 

1993; Brassell and Eglinton, 1986). The classes of 

compounds investigated in this study include 

hydrocarbons (n -alkanes, branched and cyclic 

alkanes and alkenes, biomarkers (molecular fossils), 

polycyclic aromatic hydrocarbons (PAH's), DOT and 

related pesticides, and n -aldehydes. These 

compounds are collectively referred to as molecular 

marker compounds.

735
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Figure 1. Map of study area, San Francisco Bay.

736
EXPERIMENTAL METHODS RESULTS AND DISCUSSION

Sample Collection

Grab samples of surficial sediments were 

collected in January, 1992 with a Van Veen sampler 

from 17 sites throughout the Bay (fig. 1) at water 

depths of 2 to 13 m. Sediment from the top 7 cm were 

used for analysis. A gravity core, 112 cm deep, 

collected from Richardson Bay in November, 1991, 

was cut into vertical increments 8 cm thick, and seven 

representative horizons were analyzed. All samples 

were refrigerated immediately after collection, then 

transported back to the laboratory where they were 

frozen until analysis.

Extraction and Fractionation

Analysis focused on the unbound lipid 

constituents of the sediments. Sediment subsamples 

(100 g, air-dried, ground to pass a 32-mesh sieve) were 

extracted three times with a total of 400 mL of 

dichloromethane on a wrist-action shaker, and were 

centrifuged and filtered after each extraction. Each total 

extract was evaporated to about 5 mL, the sulfur was 

removed by filtering through HCl-activated copper, and 

the solvent was exchanged for hexane and the extract 

fractionated by liquid chromatography on a silica gel 

and alumina column. Four fractions were collected: a 

hexane fraction containing aliphatic hydrocarbons and 

biomarkers; a 20 percent benzene in hexane fraction 

containing aromatic hydrocarbons, PAH's, and 

chlorinated pesticides; a benzene fraction containing 

mainly w-aldehydes; and a methanol fraction containing 

more polar molecules. The first three fractions were 

examined by gas chromatography (GC) and gas 

chromatography/mass spectrometry (GC/MS). A 

method blank was carried through the entire procedure 

and found not to contain any of the analytes of interest.

Gas Chromatography and Mass Spectrometry

Procedures for GC and GC/MS have been 

described previously (Hostettler and others, 1989). In 

addition, the PAH's and chlorinated pesticides were 

quantitated on an ion-trap mass spectrometer. The gas 

chromatograph was maintained at 50 °C. for 

4 minutes, and programmed at 6 °/min to 300 °C. The 

injector temperature was 280 °C. The capillary 

column (30m, 0.25 fim i.d., containing a 0.25 jim 

bonded phase DB-5) was interfaced directly to the ion 

source of the mass spectrometer.

The amount of extractable organic matter 

(EOM) from each sample is shown in table 1. It 

ranged from 30 fig/gm at Pinole Point (SF-36) to 

391 fig/gm at Coyote Creek mouth (SF-25). The 

EOM from SF-36 was anomalously low; the average 

value for all the other samples was 245 fig/gm. Three 

fractions were examined in order to identify the major 

constituents and evaluate possible source signatures. 

The constituents of this extracted matter are discussed 

below according to the chromatographic fractions into 

which they were separated.

Hexane Fraction

Gas Chromatography. The hexane fractions 

show fairly uniform profiles of aliphatic hydrocarbons 

throughout the Bay. A typical chromatogram is given in 

fig. 2a. Two dominant features are obvious at all sites. 

The first is the w-alkane distribution. The w-alkanes 

range from C|3 to C38 or higher. However, members 

C25-C3|, with a strong odd-over-even carbon number 

predominance (OEP), are the most prominent. These nalkanes are known to originate from the epicuticular 

waxes of higher (vascular) plants (Eglinton and 

Hamilton, 1967) and are, therefore, a signature for 

terrigenous input. All of the sites in this study show an 

w-alkane maximum at C29 Differences in maxima of 

these odd-carbon-number alkanes have been noted in 

other geochemical studies to be related to changes in 

watershed forest-types (Meyers and Eadie, 1993). Thus 

the observed uniformity of C29 predominance may 

indicate either a dominant local terrigenous plant source 

or that the majority of higher plants contributing to this 

signal have the same «-alkane maximum. However, a 

second terrigenous signal, the w-aldehydes, found in the 

benzene extracts and discussed below, shows promise 

of being somewhat more discriminative with respect to 

particular plant sources.

The second dominant feature of the 

chromatograms is the Unresolved Complex Mixture 

(UCM)--the chromatographically undifferentiable 

"hump" that underlies most of the chromato gram. A 

UCM indicates old petrogenic input and (or) evidence 

of biodegradation. Its magnitude, which is rather 

difficult to quantitate, in this setting gives some 

indication of anthropogenic contamination. Samples 

from all sites have a similar large UCM, except for the 

sample from Pinole Point, SF-36, which has 

essentially no UCM.

737
Table 1. Site information and aliphatic hydrocarbon parameters, San Francisco Bay

Site

Water 

Depth

Location (meters)

EOMa m/z191 e 

OEP296

Prc

Ph

Phc

18

XTERRd ^

m

s a

PC31o?R

Surface Sediments:

SF-25

SF-26

SF-27

SF-28

SF-29

SF-30

SF-31

SF-32

SF-33

SF-34

SF-35

SF-36

SF-37

SF-38

SF-39

SF-40

SF-41

Core:

SF-42

SF-43

SF-44

SF-45

SF-46

SF-47

SF-48

Coyote Creek mouth

Redwood Creek mouth

San Mateo Bridge, SE

Oyster Point Entrance

Army Street pier

Alameda shoreline

Oakland Outer Harbor

Berkeley Marina

Richmond Longwharf

Larkspur Channel

Rat Rock, China Camp

Pinole Point, East

Petaluma

Sonoma Creek mouth

Petaluma, East

Richardson Bay, shallow

San Bruno Shoal, mid-channel

Richardson Bay 0-8cm

8-1 6cm

32-40cm

56-64cm

88-96cm

96-1 04cm

104-1 12cm

3

3

2

3

13

3

4

3

4

6

2

4

2

2

4

4

10

391

255

132

292

261

179

232

183

332

232

297

30

233

209

224

202

247

190

227

262

259

298

247

256

4.5

3.8

7.8

5.4

4.1

4.0

4.0

8.8

8.0

9.9

6.2

10.5

8.2

4.3

5.5

5.5

12.1

10.1

6.1

2.4

5.9

4.8

5.7

6.1

0.9*

1.0*

.9*

.7

.7

22

.7

.7

.6

.8

.8

1.1

.6*

.7*

.8

.4*

.7*

0.8

.6

.6

.6

.6

.6

.6

1.3

1.3

1.3

2.5

2.3

1.1

2.5

2.8

1.5

1.9

1.6

0.9

1.3

2.0

1.2

2.6

1.2

2.9

3.7

4.4

4.6

2.9

2.7

2.4

2718 1.1

1704

906

1197

1110

582

1019

887

1338

1517

1812

867

1769

1841

.3

.2

.2

.3

.2

.2

.2

.3

.3

.3

).3

.2

.3

1527 1.3

1104 1.5

1536 1.3

1184 1.4

1377 1.4

1561 1.4

1347 1.4

1722 1.2

1194 1.3

1121 1.3

0.56

.56

.55

.57

.57

.57

.57

.57

.55

.54

.57

.48

.55

.56

.56

.54

.52

0.56

.56

.55

.56

.55

.56

.56

EOM « Extractable organic matter (ng/g dry weight)

- 4C28+4C30

c Pr - Pristane; Ph « Phytane; 18 « n-C^g. * indicates significant interference from br20 

d Sum of n-alkanes C25,27,29,31 (n9/9 drV weight) 

8 Biomarker ratios defined in text
Other features of the gas chromatograms that are 

present include lower range w-alkanes and the 

isoprenoids pristane and phytane. With only two 

exceptions (site 30, Alameda shoreline and site 36) the 

amounts of these hydrocarbons are very low relative to 

the terrigenous w-alkanes. Sources of these compounds 

might be petroleum contamination or, more likely, 

microbial or algal detritus, which is known to contribute 

C1 5 , C1 7 , and pristane. The relatively low amounts 

present are not necessarily an indication of the 

magnitude of that source's contribution, however. 

Planktonic hydrocarbons have been found to be 

remineralized preferentially to other bulk organic matter 

within the water column (Meyers and Eadie, 1993) or 

during or soon after deposition in surface sediments 

(Prahl and Carpenter, 1984).

Ratios of pristane to phytane have been used to 

evaluate the oxic/anoxic character of the depositional 

environment, but there are significant restrictions on the 

interpretations; values close to I are considered difficult 

to interpret (ten Haven and others, 1987; Powell, 1988). 

In this sample set, the values at many sites can only be 

estimated because of the interference of 2,6,10- 

trimethyl-7-(3-methylbutyl)-dodecane (referred to as 

br20 in fig. 2), another indicator of algal input 

(Cranwell and others, 1987). This interference is 

particularly prevalent in the South Bay. The 

pristane/phytane ratio values seem to be close to or 

slightly lower than 1 throughout this sample set. This 

result is consistent with values commonly observed in 

marine sediments (Powell, 1988). The one site where 

the pristane/phytane distribution is unique is site 30; 

this has an over-whelming abundance of pristane. This 

pristane abundance most likely is a result of the biota in 

the water column and (or) the sediment. The sediment 

from this site contains copious amounts of shell debris 

from oysters. A major source of pristane is formation in 

the digestive tract of copepods from phytol, the alcohol 

side chain of chlorophyll a (Meyers and Eadie, 1993); 

oysters may filter-feed on these zooplankton or their 

debris. Thus, the pristane probably derives from this 

feeding cycle. In addition, the extract from this site 

contained uniquely high levels of phytol and sterols, 

also indicative of this type of recent biogenic input.

GC/MS: Biomarkers. Triterpane/hopane and 

sterane constituents of sediments, although relatively 

minor components, can provide much useful 

information. These biomarkers, or "molecular fossils", 

are derived from biogenic precursors. Hopanoids are 

considered to originate from bacteria and blue-green 

algae; the relative proportions of the various hopane and 

hopene isomers and their progressive diagenetic

counterparts are widely used as maturity and source 

indicators. So also are steranes related to sterols, which 

are derived from higher plants and animals as well as 

from planktonic algae (Brassell and Eglinton, 1986). 

The biomarkers are quantified by monitoring m/z 191 

(hopanes) and 217 (steranes) ions using GC/MS 

Selected Ion Monitoring (SIM). Three traditional 

maturity indicators from the biomarker chromatograms 

are considered: a(3C3 1 hopane S/(S+R), which has a full 

maturity value of 0.6 (Mackenzie, 1984); the ratio of 

trisnorhopanes, Tm/Ts, which here might have maturity 

or source implications (Seifert and Moldowan, 1978); 

and aaaC29 sterane S/(S+R), which has a full maturity 

value of 0.5 (Mackenzie and others, 1980).

In this sample set the distributions of both 

hopanes and steranes are very uniform throughout the 

Bay (see fig. 2b and 2c for typical mass chromatograms 

and table 1 for measured biomarker parameters). With 

one exception (site 36, east of Pinole Point), the hopane 

and sterane profiles show very mature constituents. The 

dominant hopanes all have the 17ct(H),21 p(H) (that is, 

ct|3) configuration, as opposed to a |3{3 configuration 

present in the immature biological precursors, and the 

ct{3 C3 1 S/(S+R) ratio is very close to its full maturity 

value. The Tm/Ts ratios are all tightly clustered. The 

sterane profile shows significant maturity in that 

diasteranes dominate steranes, and the ctact-C29 

S/(S+R) ratio is close to maturity. Because this degree 

of maturity is not likely to be a characteristic of surficial 

estuarine sediment, these biomarkers are probably 

derived from old petrogenic sources and reflect highly 

reworked or mixed sediments. The core in this study 

shows high maturity in the hopane profile, slightly 

lower maturity in the sterane profile, and no significant 

change downcore. It appears, therefore, that even to a 

depth of about I meter the sediment in this area is wellmixed and (or) heavily bioturbated, a condition typical 

of much of San Francisco Bay sediment (Nichols and 

Thompson, 1985). The slightly lower maturity of the 

steranes is perhaps attributable to small amounts of 

biogenic admixture. In general, the mature biomarker 

profile indicates an anthropogenic overprint on 

naturally lower levels of immature autochthonous 

biogenic input. It is interesting that at the one 

anomalous site, SF-36, the relative amount of 

extractable organic matter is very low, and the 

hopane/sterane profile is significantly less mature. 

The three maturity parameters are significantly lower 

than at all the other sites. The hopane profile contains 

high proportions of the hopene diploptene and the C2y, 

C29, C^Q, and C^iPP-hopanes, all of which are derived 

from recent biogenic input. This profile is identical

739
LLJ 

O

CQ

Hopanes

13 14

15 16

br20 

\

UK IF

TIME

TERPANES and TRITERPANES (m// 191)

1 C2J tricyclic terpane

2 CM tricyclic terpane

3 Cy tricyclic terpane

4 Triplet: CM tetracyclic terpane 

	Cjg tricyclic terpanes

5 Cjg tricyclic terpanes

6 C29 mcyclic terpanes

7 18a(H)-22,29.30-trisnorneohopane (Ts) <^

8 22,29,30-trisnorhop-17(21)-ene Q

9 17a(H)-22,29.30-trisnorhopane (Tm) 2

10 17p(H)-22,29.30-trisnorhopane D

11 17p(H)-28.30-bisnorhopane CQ

12 17a(H),2lp(H)-30-norhopane <

13 17p(H).21a(H)-30-normoretane

14 18o&18p(U)-olcanane

15 l7a(H).21p(H)-hopane

16 17p(H).21p(H)-30-norhopane

17 17p(H).21a(H)-morctane

18 17a(H).21p(H)-homohopane(22S,22R)

19 17p(H),21p(H)-hopane

20 17p(H).21p(H)-hop-22(29)-ene (diploptene)

21 17a(H),21p(H)-bishomohopane(22S,22R)

22 17p(H),21p(H)-homohopane

23 17a(H).21 P(I I )-tri shomohopane(22S ,22R)

STERANES and DIASTERANES (m/z 217)

A 13 P(H), 17a(I I )-diacholcsune(20S ,20R)

B 5a(H),14a(H),17a(H)-cholcsune(20S,20R)

C 24-ethyl-13P(1I), 17a(II)-diacholesUne(20S,20R)

D 24-methyl-5aOI), 14a(H). 17a(II)-cholcsune(20S,20R)

F. 24-cthyl-5a(H).14a(H),17a(H)-cholestanc(20S,20R)

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i i i i i i i i i i i . i i i 

ME -

Figure 2. Gas chromatography (GC) and GC/mass spectrometry (GC/MS) of the hexane fraction of a typical San Francisco 

Bay site, (a) Numbers represent carbon numbers of n-alkanes, br=branched, Pr=pristane, Ph=phytane, (b) GC/MS Selected 

Ion Monitoring (SIM) of m/z 191; Key shown, (c) GC/MS of m/z 217; Key shown.

740
to that observed at depths significantly lower than the 

depth interval for the above-mentioned core in a study 

currently in progress. Either site 36 has somehow not 

been subjected to the anthropogenic accumulations 

common to most of the rest of the Bay or has recently 

been scoured by dredging.

20 Percent Benzene Fraction

This fraction contains mostly PAH's; several of 

these compounds are considered highly toxic. The PAH 

suite in this sample set (fig. 3 and table 2) is dominated 

by pyrene, fluoranthene, and phenanthrene throughout 

the Bay. Possible sources of these compounds include 

combustion of fossil fuels (that is, of pyrogenic origin, 

deposited primarily through aeolian transport), street 

run-off, and petroleum (from sources such as spills, 

harbor traffic, and run-off from terrigenous 

contamination). Combustion and street run-off give 

very similar distributions of PAH's; this distribution is 

reported in many similar environments throughout the 

world (Laflamme and Hites, 1978). These sources can 

be distinguished from petrogenic input in that the 

former have a higher abundance of 4- and 5-ring PAH's 

relative to 2- and 3-ring PAH's and contain mostly nonalkylated PAH's. In contrast, the petrogenic distribution 

has higher levels of alkylated PAH's (for example, a 

ratio of methylated phenanthrene isomers to the parent 

PAH, MP/P, is in the range of 0-1 for combustion or 

pyrogenic sources and from 2-6 for petrogenic sources 

(Prahl and Carpenter, 1984)). It is very likely that the 

PAH suite in the Bay sediments contain contributions 

from several sources, but the dominant signature is that 

of combustion products/street runoff. With only one 

exception (at the anamolous site 36), the MP/P ratio is 

significantly less than 1. This PAH distribution can be 

considered to indicate anthropogenic input. The 

combustion PAH's are also the dominant signature in 

the total extract at many of the sites.

The relationship Zcombustion PAH/ 

Zterrigenous alkanes (ZCOMB/ZTERR; Prahl and 

Carpenter, 1984) gives an indication of whether an 

anthropogenic or terrigenous signature dominates at a 

given site. Similar ratios have been used previously 

(Hostettler and others, 1989) to show trends in input 

signatures. Figure 4 and table 2 show that the highest 

levels of PAH input are in the central urbanized part of 

the Bay. The sediments of the North Bay, which is less 

heavily urbanized, contain lower levels of 

allochthonous input in general; of that, there is more 

terrigenous input relative to anthropogenic input. 

Lower overall levels of input are also evident in midBay or channel stations, but that may reflect the more 

sandy nature of these sediments. Organic constituents 

associate more with finer-grained particles, that is, 

with mud more than sand.

Many other organic constituents are also 

detected in this fraction, although at generally much 

lower concentrations than the PAH's. Chlorinated 

hydrocarbons such as DOT and its derivatives, other 

chlorinated pesticides, and PCB's are found at low 

concentrations at all the sites and in the core. Analysis 

of these compounds is reported as part of another 

study in this volume (Pereira and others, 1996).

100 Percent Benzene Fraction

This fraction was found to contain primarily a 

suite of long chain ^-aldehydes (see fig. 3). As with the 

A?-alkanes, the high molecular weight w-aldehydes, Q-^T 

C^ 2, which have a strong HOP (even-over-odd carbon 

number preference), are known to be derived from 

terrigenous vascular plants (Rieley and others, 1991). 

However, unlike the /7-alkanes which show a single 

maximum at Cjq, the w-aldehydes seem to have at least 

three separate groupings of homologs. These groupings 

are C"22 24ar>d ^26 28 30' b°m of which have a strong 

EOP, and a low molecular weight group, approximately 

C| 2 -C| X , which has no odd or even predominance but 

progressively higher concentrations to C ]X . This latter 

group is probably not derived from vascular plants, but 

because it is in highest relative concentration at SF-30 

where there are input sources from Bay organisms, it 

could come from a microbial, algal, or animal source. 

As for the first two groups, they vary in dominance 

around the Bay. This distribution could reflect changes 

in specific land-plant input, unlike the w-alkane 

distribution which shows no differentiation throughout 

the Bay. In a preliminary examination, when leaves 

from two types of trees common to the Bay area, 

eucalyptus and oak, were subjected to the same type of 

analysis, the eucalyptus leaves were found to have

^26 28 and me oak to nave ^28 30 as tne dominant 

aldehydes. This preliminary study indicates that an 

expanded look at aldehydes in other Bay area plant 

species could be useful in differentiating specific kinds 

of vascular plant input to Bay sediments. The fact that 

the aldehydes are preserved so well in the sediments 

when they are known to deteriorate rapidly in oxic 

environments implies that they are protected through 

the deposition stage, unlike the low-range alkanes but 

similar to the refractory high-range /7-alkanes, and then 

preserved in the probably anoxic sediments.

741
HI 

O

D 

CD

Eucalyptus leaves 26

24

 .. i.

 TIMP -

JL.

2

li

8

30

ll ., ..1 ,.

Oak leaves 28

24 26

-L Id . .

TIM p . k

,

30

±. 1 .

Figure 3. Gas chromatograms of benzene fractions from a typical site, (a) Fraction 20B, polycyclic aromatic hydrocarbon 

(PAH) peaks labeled: N=naphthalene, P=phenanthrene (M=methyl derivatives); and C=Combustion PAH, from left to right: 

fluoranthene, pyrene, benz(a)anthracene, chrysene, benzofluoranthenes, benz(e) and benz(a)pyrene, indeno(c,d)pyrene and 

benzo(ghi)perylene. (b) 100B fraction, long chain n-aldehydes, carbon numbers indicated. Eucalyptus and oak leave extracts 

also shown.

742
Table 2. Concentrations of polycyclic aromtatic hydrocarbons (PAH's) in surficial sediments from San Francisco Bay, in nanogram

PAH

naphthalene

2-methylnaphthalene

1 -methylnaphthalene

acenaphthylene

acenaphthene

9[H]fluorene

phenanthrene

anthracene

2-methylphenanthrene

1 -methylphenanthrene

fluoranthrene

pyrene

benz[a]anthracene

chrysene

benzo[b]fluoranthene

benzo[k]fluoranthene

benz[a]pyrene

perylene

indeno[1 ,2,3-cd]pyrene

dibenz[a,h]anthracene

benzo[g,h,i]perylene

#25

11

6.9

3.6

2.3

4.2

7.0

68

9.6

15

10

170

247

73

84

189

59

153

69

233

12

164

#26

14

7.9

4.6

4.6

10

12

133

20

24

17

278

333

108

146

254

95

233

90

385

19

236

#27

4.7

4.1

2.3

1.9

4.8

6.4

65

8.5

13

8.9

135

171

52

69

109

51

105

38

113

5.8

116

#28

20

11

6.9

9.3

23

24

152

80

30

52

468

355

165

216

372

96

327

135

747

35

299

#29

38

16

15

38

30

81

623

258

104

226

728

417

287

589

714

624

457

205

1107

81

418

#30

4.6

2.9

2.3

1.6

9.6

7.5

85

26

14

11

133

163

54

68

107

45

103

36

98

7

101

#31

16

13

8.5

7

26

25

236

62

39

30

433

438

148

195

286

93

279

107

410

23

241

Site

#32

12

7.5

4.4

5

11

12

143

44

28

22

284

325

110

133

189

83

204

75

232

16

179

#33

15

9.7

5.5

5.4

12

13

142

43

26

13

261

320

106

123

217

55

212

79

259

17

196

#34

12

6

3.3

4

22

11

165

31

18

14

343

394

107

120

236

78

234

97

375

23

230

#35

6.5

5.8

3.1

2.8

6

8.3

82

21

14

9

192

250

79

96

165

57

162

83

208

14

173

#36

0.5

0.7

0.4

0.3

0.3

1.6

8.6

2.6

4.9

5

41

50

33

30

35

14

30

39

16

2.6

10

Total PAH 1591 2424 1084 3623 7056 1080 3116 2119 2130 2523 1638 325 

ICOMB* 

MP/P**

1372 2149 921 3045 5341 872 2523 1739 1749 2009 1223 229 

0.7 0.5 0.6 0.7 0.5 0.6 0.6 0.7 0.7 0.3 0.6 1.4 

* Sum of "combustion" PAH. See Prahl and Carpenter, 1984; total is the same except this omits benz[e]pyrene

** Methyl phenanthrene ratio. See Prahl and Carpenter, 1984
122°30' 122°15' 122°00'

38°00'

" "vj«*:

Richardson Bay /

37°45'

37°30'

^x:sr-f< -^ 

»Sonoma Creek mouth

Jfctt-X- KT . . . . *?N. V'.T .^^^^^^x ';::^illliilllill|:|?S^ Carquinez Strait *j£i§t5^ 

............................... ^^

Oakland

:!:::;:;:i<|j:;:;:;:x:::5iii^ San

Francisco

San

Francisco 

Bay Area

Input:

0 Terrigenous (n-alkanes) 

O Anthropogenic (PAH's)

::::

San Mateo ^gSii^S 

Bridge y

,^ 

?^

5 MILES

8 KILOMETERS

 ^... ............ Jrr

Coyote Creek

zzzzzzztzszzzz

Figure 4. Proportions of terrigenous and anthropogenic signals at study sites in San Francisco Bay.

744
CONCLUSIONS

This areal study of sediments throughout north, 

central, and south San Francisco Bay provides a 

baseline for future work on contaminants and sources 

of organic input into the Bay. In general, fairly 

uniform source profiles were observed throughout the 

Bay. Unique source profiles were observed at only 

two sites: the first at SF-30 where input from local 

marine organisms gave a high maximum for pristane, 

and the second at SF-36 where immature biogenic 

input was prominent. At all the other sites, however, 

including the core, biomarker profiles are dominated 

by mature constituents that probably were derived 

from old petrogenic input, indicating anthropogenic 

influenge and highly reworked or mixed sediment.

This study showed that two source signatures 

dominate in the sediments: a terrigenous signal from 

vascular plants of high molecular weight OEP nalkanes and EOF //-aldehydes, and an anthropogenic 

signal from PAH's from combustion and (or) street 

run-off sources. The terrigenous signal is dominant at 

the southernmost site in South Bay and in North Bay, 

closer to shore, and the anthropogenic signature is 

dominant in the highly urbanized areas and in the midBay channel areas.

REFERENCES

Brassell, S.C., and Eglinton, G., 1986, Molecular

geochemical indicators in sediments, in Sohn, Mary, 

cd.. Organic Marine Geochemistry, p. 10-32.

Cranwell, P.A., Eglinton, G., and Robinson, N., 1987,

Lipids of aquatic organisms as potential contributors to 

lacustrine sediments II: Organic Geochemistry, 

v. 11, p. 513-527.

Eglinton, G., and Hamilton, J.R. 1967, Leaf epicuticular 

waxes: Science, v. 156, p. 1322-1335.

Hostettler, F.D., Rapp, J.R., Kvenvolden, K.A., and Luoma, 

S.N., 1989, Organic markers as source discriminants 

and sediment transport indicators in south San 

Francisco Bay, California: Gcochimica ct 

Cosmochimica Acta, v. 53, p. 1563-1576.

Laflammc, R.E., and Hitcs, R.A., 1978, The global 

distribution of polycyclic hydrocarbons in recent 

sediments: Geochimica ct Cosmochcmica Acta, v. 42, 

p. 289-303.

Mackenzie, A.S. 1984, Applications of biological markers 

in petroleum geochemistry, in Brooks, J., and Welte, 

D., cds., Advances in Petroleum Geochemistry, 

London Academic Press, v. 1, p. 115-214.

Mackenzie, A.S., Patience, R.L., Maxwell, J.R.,

Vandcnbroucke, M., and Durand, B., 1980, Molecular 

parameters of maturation in the Toarcian shales, Paris 

Basin, France, 1 Changes in the configuration of 

acyclic isoprenoid alkanes, stcranes, and tritcrpancs: 

Geochimica ct Cosmochimica Acta, v. 44, p. 1709- 

1721.

Meyers, P.A., and Eadie, B.J., 1993, Sources, degradation 

and recycling or organic matter associated with sinking 

particles in Lake Michigan: Organic Geochemistry, 

v. 20, p. 47-56.

Nichols, F.H., Clocrn, J.E., Luoma, S.N., and Pctcrson, 

D.H., 1986, The modification of an estuary: Science,

v. 231, p. 567-573.

Nichols, F.H., and Thompson, J.K., 1985, Time scales of 

change in the San Francisco Bay benthos, in Cloern 

J.E., and Nichols, F.H., cds.. Temporal dynamics of an 

estuary, San Francisco Bay, p. 121- 138.

Pcrcira, W.E., Hostettler, F.D., Cashman, J.R., and

Nishioka, R.S., 1996, Distributions of organochlorinc 

compounds in sediments and livers of striped bass 

(Morone saxatilis) in San Francisco Bay-Delta Estuary, 

in Morganwalp, D.W., and Aronson, D.A., cds, U.S. 

Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Powcll, T.G. 1988, Pristanc/phytanc ratio as environmental 

indicator: Nature, v. 333, p. 604.

Prahl, F.G., and Carpenter, R., 1984, Hydrocarbons in

Washington coastal sediments: Estuarinc, Coastal and 

Shelf Science, v. 18, 703-720.

Rieley, G., Collier, R.J., Jones, D.M., and Eglinton, G., 

1991, The biogeochemistry of Ellcsmcre Lake, U.K., 

I Source correlation of leaf wax inputs to the 

sedimentary lipid record: Organic Geochemistry, 

v. 17, p. 901-912.

Seifert, W.K., and Moldowan, J.M., 1978, Applications of 

sterancs, terpancs and monoaromatics to the 

maturation, migration and source of crude oils: 

Gcochimica ct Cosmochcmica Acta, v. 42, p. 77-95.

ten Haven, H.L., de Leew, J.W., Rullkottcr, Sinninghe, J., 

and Damstc, J.S., 1987, Restricted utility of the 

pristane/phytane ratio as a palcoenvironmental 

indicator: Nature, v. 330, p. 641-643.

Volkman, J.K., Barrett, S.M., Dunstan, G.A., and Jeffrey, 

S.W., 1993, Geochemical significance of the 

occurrence of dinostcrol and other 4-methyl stcrols in a 

marine diatom: Organic Geochemistry, v. 20, p. 7-15.

745
746
Distributions and Benthic Flux of Dissolved Sulfides in 

the Toxic Water Column of San Francisco Bay, California

by James S. Kuwabara1 , Yvonne R. Hunter1 , and Cecily C.Y. Chang1

Abstract

Chemical speciation affects the 

bioavailability of metals and organic ligands to 

planktonic and benthic organisms, as well as the 

partitioning of these solutes between phases. Our 

previous work in San Francisco Bay indicated 

that sulfide complexation with metals can be of 

particular importance because of the 

thermodynamic stability of these complexes. 

Although the water column of the Bay is 

consistently well-oxygenated and typically 

unstratified with respect to dissolved oxygen, the 

kinetics of sulfide oxidation could exert at least 

transient controls on metal speciation. Our initial 

data on dissolved sulfides in the main channel of 

both the northern and southern components of the 

Bay consistently indicate submicromolar 

concentrations (from < 1 to 162 nanomolar) as one 

would expect in an oxidizing environment. 

However, chemical-speciation calculations over 

the range of observed sulfide concentrations 

indicate that these trace concentrations in the Bay 

water column can markedly affect chemical 

speciation of ecologically significant trace metals, 

such as cadmium, copper and zinc. In contrast to 

longitudinal concentration gradients for 

dissolved organic carbon, dissolved sulfide 

concentrations were typically elevated at depth 

suggesting a primary benthic source. Subsequent 

flux core and in-situ flux chamber experiments 

support this hypothesis.

'U.S. Geological Survey, Menlo Park, CA.

INTRODUCTION

Trace contaminants enter major estuaries such 

as San Francisco Bay from a variety of point and 

nonpoint sources and can then be repartitioned 

between solid and aqueous phases or altered in 

chemical speciation. Recent toxicological data have 

consistently demonstrated that chemical speciation of 

trace contaminants can significantly affect the 

availability of those contaminants to aquatic 

organisms (Gavis and others, 1981; Anderson and 

others, 1978; Kuwabara and others, 1986). In 

previous studies of the South San Francisco Bay water 

column, two potentially important mechanisms for 

controls on trace-metal speciation were noted 

(Kuwabara and others, 1989). First, a strong 

correlation between two dissolved trace metals 

(copper and zinc) and dissolved organic substances 

was observed. Chemical speciation calculations were 

used to examine the importance of metal-organic 

complexation reactions. Secondly, on the basis of 

measurements of sulfide in the oxic water column of 

the ocean (Cutter and Krahforst, 1988; Luther and 

Tsamakis, 1989), we concluded that sulfide reactions 

in the water column should be considered as a 

controlling mechanism for trace-metal speciation in 

estuarine waters. Previous water-column studies in 

the Bay had indicated dissolved-oxygen (DO) 

concentrations near saturation (Smith and Herndon, 

1980). One might reasonably question why sulfides 

could be an important factor in this oxidizing 

environment. However, the thermodynamic data for 

metal-inorganic ligand complexes (Hogfeldt, 1983; 

Dyrssen and Wedborg, 1989) indicated that 

submicromolar levels of dissolved sulfides could 

impose a pronounced effect on metal speciation. 

Therefore, the purpose of this paper is to: (1) describe 

the distribution of dissolved-sulfide concentrations in 

the northern, central and southern components of San

747
Francisco Bay, and (2) because initial water column 

measurements indicated a benthic source for dissolved 

sulfides, to present subsequent benthic flux 

measurements for dissolved sulfides and dissolved 

organic carbon.

MATERIALS AND METHODS

Two aspects of this study involved the collection 

of water column and benthic flux core samples. 

Water-column samples were collected from the main 

channel of the northern, central and southern 

components of San Francisco Bay (fig. 1). The North 

Bay was sampled during a high discharge event that

38°

37°45'

37'30C

122"30:' 122"15'

occurred during the spring 1991 when 173.0 to 

471.1 m3 /s flowed from the Bay delta during March 

12-15. The water-column samples were collected at 

three U.S. Geological Survey (USGS)- designated 

stations (stations 4, 8, and 18 with depths of 14, 15, 

and 32 m, respectively; see fig. 1). Station numbers 

increase in the North Bay toward the Golden Gate 

Bridge. The overall sampling period occurred during 

a multiyear drought. During the drought, the average 

delta outflow was 185.6 m3 /s between October 1, 1990 

and May 31, 1991, the normally wet season. During 

the predrought conditions of October 1, 1984 - 

May 31, 1985, the average delta outflow was 

440.1 m3/s. We sampled the South Bay during neap

122°

2 METER DEPTH 

10 METER DEPTH 

SAMPLING STATIONS 

 WATER COLUMN 

O BENTHIC FLUX

Figure 1. Study site showing water column and benthic flux core sampling stations, San 

Francisco Bay, California.

748
tides before, during, and after the spring 

phytoplankton bloom of April 1991 at USGSdesignated stations 27, 32, and 36 (fig. 1). Station 

numbers increase in the South Bay with distance from 

the Golden Gate Bridge. Because of tidal variations, 

depths at each station varied with sampling date 

(table 1). At each station, samples were collected 

from the Research Vessel Polaris, by means of 10-L 

fluoroethylene polymer (FEP)-lined sampling bottles 

designed for trace metal sampling at three depths: 1 m 

from the surface, at approximately 50 percent of 

maximum depth, and 2 m from the bottom (hereafter 

designated shallow, mid-depth, and deep samples, 

respectively). The exception to this procedural 

scheme was made at the shallowest sampling station 

36, where no mid-depth sample was collected.

Benthic flux experiments were made in the 

spring of 1992 and involved measurements from flux 

core and in-situ flux chambers at two sites in the South 

Bay: a shoal station 46 and a channel station 29A 

(fig. 1). This phase of the study was done in 

collaboration with macronutrient flux work being done 

by J.M. Caffrey of the U.S. Geological Survey, Menlo 

Park, and D.E. Hammond and W.M. Berelson at the 

University of Southern California. Methods for fluxcore sampling have been previously described (Kemp 

and others, 1990). One modification that should be 

noted is that our samples were collected only before 

and after the flux-core incubation period because of 

sample volume constraints. Therefore, the flux 

determinations assume a linear change in solute 

concentration over the incubation period.

Dissolved sulfide concentrations were measured 

on site by square-wave voltammetry (Kuwabara and 

Luther, 1993), which involved the deposition of 

sulfide species at a hanging mercury electrode 

(-0.40 V) followed by a cathodic scan to -1.0 V to strip 

and detect the sulfide species. By the method of 

addition, only total dissolved sulfide analysis was 

done, as opposed to sulfide-speciation measurements, 

to (1) avoid sample contamination at low anticipated 

sulfide concentrations, and (2) complete the replicate 

sample analyses within 2 hours of sampling. Samples 

also were collected for measurement of pH and 

concentrations of DO (by colorimetry and the 

membrane electrode method), and dissolved organic 

carbon (by the ultraviolet/oxygen/persulfate oxidation 

method). Depth profiles for salinity (in Practical 

Salinity Units (PSU)) and temperature were obtained 

using a data-acquisition system with an arrangement

of multiple probes to monitor various water-quality 

properties. Barometric pressure was determined on 

site for DO-concentration corrections and verified 

against measurements made at San Francisco 

International Airport (National Weather Service, oral 

communication, 1991).

RESULTS AND DISCUSSION

Water-Column Distributions

On the basis of temperature, salinity, and 

barometric pressure corrections, more than 80 percent 

of the DO measurements were within 90 to 

110 percent of saturation, indicating that the water 

column at our sampling sites was an oxic environment 

(table 1) in which the presence of "free" sulfides (I^S, 

HS") was not thermodynamically favored. Measured 

pH values were fairly uniform throughout the 

sampling period, with more than 60 percent of the 

values were 7.9 to 8.0 (table 1).

In the North Bay, the dissolved organic carbon 

(DOC) concentrations increased significantly in the 

direction of freshwater sources and generally 

decreased with depth (table 1). In the South Bay, the 

DOC concentrations also increased with distance from 

the mouth of the estuary. However, the horizontal 

DOC gradient was not as pronounced as it was in the 

North Bay (table 1), and, as with salinity, no vertical 

DOC gradient was observed in the South Bay, which 

received little freshwater input.

In contrast to DOC distributions, no discernable 

longitudinal gradient was noted for dissolved sulfides 

in the North or South Bay (table 1). Relative to 

concentrations of surface samples, the sulfide 

concentrations were elevated (although < 1 (iM) at 

depth. A significant (at the a=0.05 level) positive 

correlation between relative depth and measured 

sulfide concentration indicated a benthic source of 

sulfide (Kuwabara and Luther, 1993).

Benthic Fluxes

Sulfide benthic fluxes were consistently positive 

(out of the sediment) over the course of our sampling 

period in 1992 (table 2). This was consistent with our 

hypothesis of a primary benthic source, based on 

initial concentration distributions in the water column. 

Flux experiments during spring 1993 (a post-drought

749
period) examine the possibility of a water-columnsulfide source resulting from the spring phytoplankton 

bloom.

In contrast to sulfide fluxes, DOC fluxes 

changed direction over the period of study, going from 

negative to positive (table 2). Trace metal flux 

measurements will provide information on the 

association of the trace metals with the benthic 

exchange of sulfide and dissolved-organic ligands.

Kuwabara and Luther (1993) used chemicalspeciation calculations to demonstrate the potential 

importance of dissolved sulfides as ligands in the oxic 

San Francisco Bay water column. Their analysis 

indicated that, as the dissolved sulfide concentration 

declined below the dissolved copper concentration 

(30 nanomolar), copper complexation with humic 

material, with stability constants that are lower than 

stability constants associated with copper sulfide

Table 1. Water-column data for the northern and southern components of San Francisco Bay, California.

[Dissolved oxygen (DO) is given as percent saturation; salinity (SAL) in practical salinity units (PSU); water- column temperature (TRMP) in degrees 

Centigrade; dissolved organic carbon concentration (DOC) in milligrams carbon per liter and dissolved sulfide concentration in nanomolar. Symbol "*" 

indicates that the measurement was not made.]

Date

March 6, 1991

March 12, 1991

March 15, 1991

April 8, 1991

May 6, 1991

Station 

number

27

5

9

32

5.5

8

36

5

4

8

12

8

10

13

18

15

30

27

6

11

32

5

8

36

7

27

4.5

9

32

6

10

36

5

Depth 

(m)

2

102

100

2

106

100

2

109

1

89

88

1

90

86

1

102

99

2

93

97

2

104

99

2

99

2

101

1007.98

2

104

106

2

100

DO

109

7.87

7.84

111

7.87

7.85

103

7.87

89

7.90

7.82

90

7.91

7.93

104

7.93

7.94

102

7.97

7.88

104

7.99

7.98

100

7.94

108

7.96

27.0

110

7.94

7.95

110

7.77

PH

7.89

29.4

29.4

7.87

26.3

26.7

7.97

23.2

7.92

2.9

5.0

7.92

18.4

19.5

7.94

30.8

30.9

7.95

26.7

26.6

8.01

26.7

26.7

7.94

24.9

7.93

27.0

14.8

7.95

26.2

26.2

7.84

23.2

SAL 

(PSU)

29.3

13.1

13.1

25.3

12.7

12.7

23.0

12.7

1.9

12.7

12.7

15.8

12.6

12.5

30.8

12.0

11.9

26.7

14.8

14.8

26.6

15.1

15.1

24.8

14.3

27.0

14.8

2.7

26.1

15.1

15.1

22.9

14.3

TEMP

13.1

2.0

1.8

12.7

2.9

2.6

12.4

2.2

12.6

3.3

3.1

12.7

2.3

*

12.0

1.1

1.2

14.8

1.9

1.9

15.0

1.9

2.2

14.3

2.2

14.8

2.6

21.9

15.0

3.4

3.5

14.3

3.6

DOC

1.6

2.2

2.7

2.4

2.7

35.2

2.2

17.1

3.4

8.7

12.2

2.5

7.1

12.6

1.6

<l

37.3

1.6

4.5

25.5

2.1

2.5

30.8

2.6

30.1

2.7

29.8

3.1

27.6

17.8

4.4

16.2

Dissolved 

Sulfide

2.0

<1

2.8

5.9

1.2

3.3

5.2

11.1

7.1

23.2

12.5

16.0

750
complexes, became increasingly important. Hence, 

the concentration of uncomplexed, and presumably 

more bioavailable, copper increased. This range of 

calculated copper ion activity is ecologically relevant 

because it brackets toxic copper concentrations for 

phytoplankton groups (for example, negative log of 

the copper concentration (pCu) of 8.5 to 10.9 for 

diatom species; Gavis and others, 1981). The 

competition between DOC and dissolved-sulfide 

complexation for control of trace-metal speciation was 

evident when dissolved-sulfide concentrations were 

above 5 percent of the dissolved copper 

concentrations. Similar to copper, an analogous 

sulfide effect on cadmium and zinc speciation over 

observed concentration ranges for these trace metals 

was also reported although the dominant, nonsulfide 

species for cadmium and zinc differed (Kuwabara and 

Luther, 1993). Given the observed dissolved-sulfideconcentration range in these oxic waters, there is 

clearly a potential for ecologically significant effects 

on trace-metal speciation.

Table 2. Benthic fluxes with 95 percent confidence intervals 

using n=3 analytical replicates per sampling interval of 

dissolved organic carbon and sulfides in South San 

Francisco Bay.

[Dissolved organic carbon (DOC) flux in millimoles per square meter per 

hour; and sulfide flux in micromoles per square meter per hour. Paired flux 

measurements indicate determinations made on replicate cores.]

Sampling Date

March 17, 1992

April 13, 1992

May 11, 1992

February 25, 1992

March 17, 1992

April 13, 1992

May 11, 1992

DOC Flux

Shoal Station 46

-0.94±0.43

-0.8010.33

.08±0.11

1.92±0.15

.45±0.08

Channel Station 29A

-.83+0.17

-.11 ±0.55

-.64±0.71

-.16+0.21

-.13±O.I5

.55±O.IO

Sulfide Flux

0.14±0.03

.16±0.03

.37±0.09

.42±0.09

.42±0.03

.25+0.03

.17±O.OI

.09+0.03

.3610.02

.57+0.02

1.0210.07

REFERENCES

Alpine, A.A., Weinke, S.M., Cloern, J.E., and Cole, B.E., 

1988, Plankton studies in San Francisco Bay, IX 

Chlorophyll distributions and hydrographic properties 

of South San Francisco Bay, 1984-1986: U.S. 

Geological Survey Open-File Report 88-319, 86 p.

Anderson, M.A., Morel, F.M.M., and Guillard. R.R.L., 

1978, Growth limitation of a coastal diatom by low 

zinc ion activity: Nature, v. 276, no. 5683, p. 70-71.

Cutter, G.A., and Krahforst, C.F., 1988, Sulfide in the 

surface waters of the western Atlantic Ocean: 

Geophysical Research Letters, v. 15, no. 12, p. 1393- 

1396.

Dyrssen, D., and Wedborg, M, 1989, The state of dissolved 

trace sulphide in seawater: Marine Chemistry, v. 26, 

no. 3, p. 289-293.

Gavis, J., Guillard, R.L.L., and Woodward, B.L., 1981, 

Cupric ion activity and the growth of phytoplankton 

clones from different marine environments: Journal of 

Marine Research, v. 39, no. 1, p. 315-333.

Hogfeldt, E., 1983, Stability constants of Metal-Ion 

Complexes, Part A Inorganic Ligands. Pergamon 

Press, Oxford. 310 p.

Kemp, W.M., Sampou, P., Caffrey, J., Mayer, M.,

Henriksen, K., and Boynton, W.R., 1990, Ammonium 

recyling versus denitrification in Chesapeake Bay 

sediments: Limnology and Oceanography, v. 35, no. 7, 

p. 1545-1563.

Kuwabara, J.S., Davis, J.A., and Chang, C.C.Y., 1986, Algal 

growth response to particle-bound orthophosphate and 

zinc: Limnology and Oceanography, v. 31, no. 3, 

p. 503-511.

Kuwabara, J.S., Chang, C.C.Y., Cloern, J.E., Fries, T.L., 

Davis, J.A., and Luoma S.N., 1989, Trace metal 

associations in the water column of South San 

Francisco Bay, California: Estuarine, Coastal and Shelf 

Science, v. 28, no. 3, p. 307-325.

Kuwabara, J.S., and Luther, G.W., III, 1993, Dissolved 

sulfides in the oxic water column of San Francisco 

Bay, California: Estuaries, v. 16, no. 3, in press.

Luther, G. W., and Tsamakis, E., 1989, Concentration and 

form of dissolved sulfide in the oxic water column of 

the ocean: Marine Chemistry, v. 27, no. 3-4, 

p. 165-177.

Smith, R.E., and Herndon, R.E., 1980, Physical and 

chemical properties of San Francisco Bay waters, 

1977: U.S. Geological Survey Open-File Report 

80-1191, 51 p.

751
752
Past and Present Research on Metal Transport In 

St. Kevin Gulch, Colorado

SyBriant A. Kimball 1

Abstract

To prepare for mitigation of the effects of 

acidic mine drainage on upland watersheds, 

process-oriented research at St. Kevin Gulch has 

focused on chemical reactions that affect metal 

transport and partitioning among phases. We 

operationally defined dissolved and colloidal 

transport phases and the speciation of iron 

because the cycle of iron can affect the cycling of 

other metals. With these phases defined, our 

approach was to study chemical reactions in the 

context of hydrologic transport. By establishing 

the hydrologic setting with tracer-dilution injections, we studied temporal and spatial variability 

in metal concentrations resulting from combined 

hydrologic and chemical processes. Temporal and 

spatial aspects of variation were combined in an 

instream pH- modification experiment to evaluate 

the kinetics of reactions involving metals. These 

data are the focus of a simulation model that 

combines transport and reactive chemistry of 

metals. Ongoing studies of the ecological system 

are in anticipation of higher pH and lower metal 

concentrations that should folow remediation by a 

passive treatement system.

INTRODUCTION

Many streams in upland watersheds are affected 

by acidic mine drainage (Moore and Luoma, 1990). 

Efforts are underway to mitigate the effects of this 

acidic drainage. However, successful mitigation 

requires an understanding of the processes that affect

'U.S. Geological Survey, WRD, Salt Lake City, UT 84104.

metals in these streams. Since 1986, the Toxic 

Substances Hydrology Program of the U.S. Geological 

Survey has supported field-based research on 

processes that affect metals in streams. Objectives of 

this research include: (1) The study of instream 

chemical and biological processes in the context of 

transport processes that affect them, (2) determination 

of the role of sediment and colloids in the processes 

affecting metals, (3) quantification of the temporal and 

spatial variability of processes, and (4) simulation of 

the reactive transport of metals to evaluate field observations. The purpose of this report is to present the 

major findings from past studies of St. Kevin Gulch 

and to describe the focus of present studies.

St. Kevin Gulch is an upland watershed in the 

Rocky Mountains; its elevation is 2,500 to 2,800 m 

(meters) above sea level and it has an area of about 

10 2 km (square kilometers, fig. 1). Most annual precipitation falls as snow; snowmelt runoff that occurs in 

late May or early June is the dominant hydrologic 

event each year. Bedrock is a quartz-biotite-feldspar 

schist. Mining of silver and zinc sulfides in vein 

deposits mostly occurred around 1920. St. Kevin 

Gulch has been a valuable research site because of its 

combination of physical and chemical characteristics: 

(1) Metal concentrations are sufficiently high to be 

quantified by routine methods and yet not so high as to 

be serious health and safety concerns, (2) a 100-m 

reach is a point source of acidic inflow, (3) physical 

and chemical processes that affect metal concentrations occur downstream from the principal acidic 

inflows, (4) a single algal species, Ulothrix sp., 

predominates (McKnight, 198S), and (5) a natural 

wetland is present just upstream from the confluence 

of St. Kevin Gulch and Tennessee Creek (fig. 1).

Downstream changes in the chemical characteristics and discharge in St. Kevin Gulch result from 

acidic inflows (fig. 2). Upstream from the principal

753
.26

EXPLANATION

Sample site with distance 

in meters from lithium 

chloride injection site

1/2 1 KILOMETER

1/2 1MILES

Figure 1. Location of (a) St. Kevin Gulch, near Leadville, 

Colorado, and (b) location of pH-modification experiment in 

1988

acidic inflows (0-363 m, fig. 1) the chemical composition mostly results from natural weathering in the 

basin. Between 363 m and 501 m, discharge of the 

acidic inflows is only about 0.8 L/s (liters per second) 

compared to 6 L/s in the stream, but the inflow chemistry strongly affects instream pH and metal concentrations (figs. 2a and 2b). The confluence with the 

nonacidic Shingle Mill Gulch is a short distance 

downstream from the acidic inflows, at 501 m. At the 

confluence, discharge approximately doubles (fig. 2c) 

and pH increases from the nonacidic inflow. Downstream from 526 m there only are a few acidic inflows 

from small mines; these inflows do not substantially 

change instream metal concentrations upstream from 

the wetland. The mass loading of manganese (Mn) 

(fig. 2d) shows that the acidic inflows between 363 m 

and 501 m are virtually a point source to the stream.

PAST STUDIES

We have taken advantage of the characteristics 

of St. Kevin Gulch to conduct process-oriented 

research, studying well-known chemical reactions in

to 6

§ 5

- A

°- 3

?

- '(a)' *'

- \ .

- 

, 1 , . . *.**! 

v.

 STREAM

 INFLOW

... 1 .... 1 ...

-

'

_

-

, '

500 1.000 1,500

z 

..; £E

t 1 r

0.1

 .t%l ! 

v 

 ' 1

500 1,000 1,500

Q

s§

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CO CC

Q LU

12 

8

4

0

. I . . > . i . . i

r 'c> .- ---

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~

-

. i . . . . i . . . . 

500 1,000 1,500

< to

« UJ

io

1.5

0.5 -

0.0

_.- .' *.

0 500 1,000 1,500 

DISTANCE FROM INJECTION, IN METERS

Figure 2. Downstream profiles of (a) pH, (b) manganese, 

(c) discharge, and (d) manganese mass flow in St. Kevin 

Gulch, near Leadville, Colorado, August 1987.

the context of hydrologic transport. Using tracerdilution injections to establish the hydrologic background, we have carried out experiments with 

intensive spatial and temporal sampling (table 1).

Methods for Assessing Iron-Rich Systems

Particular methods have been used to define 

processes that affect metals in acidic mine drainage. 

We will review some of these methods and then discuss how these methods aid the study of instream 

processes.

Setting the hydrologic framework.~\Jse of 

experimentally injected chemical tracers helps to 

define relevant physical characteristics of a stream 

With the physical characteristics defined, it is possible 

to distinguish between physical and chemical effects 

on instream metal concentrations (Stream Solute 

Workshop, 1990; Bencalaand others, 1990; Broshears 

and others, 1993). However, a seasonal study has a 

greater temporal scale, and it is not always practical to 

inject tracers to set the hydrologic framework. In such 

cases, natural conservative solutes can substitute for

754
Table 1. Summary of experiments in St. Keven Gulch, 1986 through 1993

Date Description Results

August 1986

May 1987 

August 1987

August 1988

August 1989 

August 1990

April-August 

1990

August 1991

July 1993

LiCl injection for 36-hour diel, 

synoptic sampling, and hydrologic 

characterization.

NaCl injection at high flow.

LiCl injection for synoptic sample 

for filtered and particulate concentrations.

LiCl injection for diel sampling; 

nighttime synoptic sampling; pHmodification experiment.

Multiple tracer injections to define 

loss of streamflow.

Hillslope interactions; injection of 

radioactive phosphorous for identification of nutrient pathways.

Seasonal sampling at fixed sites 

using natural conservative tracers

Hillslope interactions; diel 

sampling of streamwater and alluvial water.

Effects of alluvium on instream pH 

modification.

1. Documentation of iron photoreduction reaction (McKnight and others, 

1988).

2. Hydrologic characterization (Broshears and others, 1993).

3. Initial steady-state simulation (Kimball and others, 1991).

I. Hydrologic characterization during snowmelt runoff (unpublished data).

I. Comparison of rates for hydrologic and biogeochemical processes; 

evaluation of particulate concentrations (Kimball and others, 1994).

1. Mechanisms of iron photoreduction reaction (Kimball and others, 

1992b).

2. Synoptic data set without effects of photoreduction (B.A. Kimball, U.S. 

Geological Survey, 1988, unpublished data).

3. Temporal and spatial data on metal response to increasesd pH (Kimball 

and others, 1992a; Kimball and others, 1994).

4. Stream-side sorption experiments (Smith and others, 1991).

I. Quantification of losing reach (Zellwegcr and Maura, 1991).

1. Water exchange between stream and alluvium (Harvey and Bencala,

1993).

2. Phosphate uptake (Tate and others, 1991).

1. Seasonal variation of metal concentrations (B.A. Kimball, U.S. Geological Survey, 1990, unpublished data).

I. Effects of substream on diel patterns (B.A. Kimball, U.S. Geological 

Survey, 1991, unpublished data).

1. (B.A. Kimball, U.S. Geological Survey, 1993, unpublished data).

injected tracers (Bencala and others, 1987). Although 

this procedure does not establish absolute values of 

discharge as does the tracer-dilution method, it 

accounts for the hydrologic effects by a discharge ratio 

so that chemical effects can be studied.

Defining trasport phases. In addition to 

defining the hydrologic framework, it is necessary to 

define transport phases for the metals. The cycle of 

iron (Fe) affects many metals and is strongly affected 

by precipitation (Kimball and others, 1991) and by 

photoreduction (McKnight and others, 1988; Kimball 

and others, 1992b). At pH values greater than about 

2.2, Fe oxyhydroxides commonly precipitate to form 

colloids, which affect the cycling of other metals by 

sorption (Smith and others, 1991). Colloids in St. 

Kevin Gulch can contain as much as 130 ppm (parts 

per million) arsenic (As), 230 ppm copper (Cu),

600 ppm lead (Pb), and 1,200 ppm zinc (Zn). However, Ranville (1992) determined that most of the suspended sediments in St. Kevin Gulch are primarily 

aggregates of colloidal (40-nanometer diameter) Fe 

oxyhydroxides and Fe oxyhydroxysulfates. In opposition to precipitation, photoreduction of Fe III can dissolve colloids and return ferrous iron (Fe II) to the 

stream (McKnight and others, 1988). Other metals that 

may be sorbed to the colloidal Fe also can be released. 

This dynamic cycling of Fe on a daily time scale 

affects the transport and transformation of other metals and can be studied only by adequate sampling for 

Fe phases and species.

Operational definitions of phases can be defined 

by using multiple filtrations. The definitions that are 

most meaningful in terms of the Fe colloids include 

the following: (1) filtered concentration (representing

755
"dissolved" solutes) determined by filtration through a 

0.001-^m (micrometer) pore-size membrane, 

(2) colloidal concentration, determined by filtration 

through a 0.45-^m pore-size membrane and then subtracting the dissolved concentration, and (3) suspended particulate concentration determined by an 

unfiltered sample and then subtracting the colloidal 

and dissolved concentrations. These definitions are 

time consuming to obtain when we could not take the 

time for the sequence of filiations. For intensive temporal and spatial sampling, we generally obtained only 

an unfiltered and a 0.1-jim filtered sample (McKnight 

and others, 1988; Kimball and others, 1992a).

Instream Processes

With these definitions of transport phases, we 

have described instream processes affecting metals in 

the context of transport.

Photoreduction of iron Intensive temporal 

sampling showed the importance of the photoreduction process on the cycling of Fe and other 

metals. Quantification of discharge allowed the calculation of mass flow and flux of Fe II from the 

streambed to the water column in response to light 

intensity (see fig. 3 in Kimball and others, 1992b).

Coupling of rates for hydrologic and chemical 

processes Intensive spatial sampling along a 

1,800-m reach allowed us to compare rates of transport to rates of chemical reaction by using steady-state 

solute transport simulation (fig. 3). If chemical reactions are relatively fast compared to rates of transport, 

the reactions can affect instream concentrations. The 

solid line in figure 3a illustrates conservative transport 

of Fe. Adding first-order rate constants to simulate Fe 

removal indicates the relative importance of chemical 

reaction (dashed line, fig. 3a). The increase of colloidal Fe (fig. 3b) corresponds to the decrease of filtered 

Fe. Loss of colloidal Fe (dashed line simulation, 

fig. 3b) is from sedimentation of colloidal aggregates.

Response to pH modification Combining 

intensive temporal and spatial sampling, we modified 

the chemistry of St. Kevin Gulch to document the 

kinetics of metal reactions as pH was increased in two 

steps (fig. 4a). With an increase of pH to near 6.0, aluminum (Al) was completely partitioned from the filtered to the colloidal phases in the water column 

(fig. 4b). This is likely from the rapid formation of an 

Al oxyhydroxysulfate complex (Kimball and others, 

1994). Sorption onto colloids affected the concentra12

10

8

6

4

2

0

(a) FILTERED IRON _

 CONSERVATIVE

- REACTIVE

 SAMPLED, PRECISION INDICATED 

BY ERROR BARS

500 1,000 1,500

~. 2.0 

O

UJ

O

z 

o 

o

1.5

1.0

0.5

0.0 -

(b) COLLOIDAL IRON

0 500 1,000 1,500 

DISTANCE FROM INJECTION, IN METERS

Figure 3. Variation of sampled and simulated (a) filtered 

iron and (b) colloidal iron with downstream distance in 

St. Kevin Gulch, near Leadville, Colorado.

tions of Cu. However, Mn and Zn were little affected 

at this level of pH (Kimball and others, 1994).

Exchange of streamwater and subsurface 

water By injecting tracers both in the stream and in 

the hillslope, the exchange of water between the 

stream and the alluvium was documented (Harvey and 

Bencala, 1993). This changes the traditional concept 

of the stream as a pipe of water leaving a watershed. 

Instead, the stream continues to interact with the 

watershed, which can affect instream metal concentrations because metals are present in substantial concentrations in the alluvium (J.W. Harvey and B.A. 

Kimball, U.S. Geological Survey, 1991, unpublished 

data, Salt Lake Oty, Utah). The scale of this interaction 

can be on the order of 1 -m stream segments.

Treatment of metals by natural wetlands 

Before entering Tennessee Creek, the metal-rich water 

from St. Kevin Gulch passes through a natural wetland. Initial sampling of inflow and outflow water 

from this wetland indicated that metals were being

756
.-./

09:00 12:00 15:00 18:00

z ^

_ D

2 -

D CC

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^ *. '

0 -

 ' "7 "i ' i 

- (b)

; --" , //-

" FILTERED H 

. COLLOIDAL 1

: 

i , , A-~i

| y r ]

 

- ^-^' -

 V»* '* 

1

09:00 12:00 15:00

TIME, IN HOURS

18:00

Figure 4. Variation of (a) pH and (b) filtered and colloidal 

aluminum with time during pH-modification experiment in 

St. Kevin Gulch, near Leadville, Colorado, August 1988.

treated by interaction with organic matter and minerals 

of the wetland. Metal flux and geochemical processes 

in this wetland have been studied by Walton-Day 

(1992) to determine whether the wetland is removing 

Fe, Mn, cadmium (Cd), Cu, Pb, and Zn from surface 

water flowing through the wetland. Careful measurement of metal fluxes indicates that only Fe was 

removed from surface water. Most metals were 

untreated by the wetland, and essentially passed 

through despite prolonged physical contact with the 

wetland environment.

PRESENT STUDIES

In our current work, we have used mathematical 

models to provide a frame of reference for evaluating 

data from field studies. We developed computer simulations for conservative and reactive solute transport in 

St. Kevin Gulch. Reactive transport simulation takes 

the study of biogeochemical processes beyond batch 

experiments and places it in the context of hydrology.

Sim ulation of conservative transport Ph ys i ca I 

effects on the transport of solute mass must be quantified to distinguish the chemical and biological effects. 

Simulation of transport in mountain streams presents 

certain unique physical aspects, the most prominent 

being transient storage (Bencala and Walters, 1983). 

The transport model developed by Bencala and 

Walters (1983) was improved by using a new algorithm and by execution on new generations of 

computers (Runkel and Broshears, 1992; Runkel and 

Chapra, 1993). These improvements were the basis for 

building the simulation model for reactive solute 

transport.

Simulation of reactive solute transport Data 

from the pH-modification experiment in 1988 (as in 

fig. 4) provide a unique opportunity to evaluate our 

understanding of rates for metal reactions. The 

complexity of reactive chemistry is simulated by 

coupling MINTEQA2 (Allison and others, 1991) with 

the transport code (Runkel and others, 1996; Broshears, 1996).

Our understanding of geochemical reactions in 

St. Kevin Gulch has provided the background to study 

remediation of acidic mine drainage. The State of 

Colorado plans to install a passive treatment system 

near the mine dump in St. Kevin Gulch, creating the 

opportunity to study the changes that occur in the 

recovering stream. Chemical changes in the water 

column, in the bed sediment, and in the alluvium will 

combine to affect the stream ecosystem. The current 

dominance of the blue-green algae, Ulothrix sp. 

(McKnight, 1988) will be affected by chemical 

changes, indicating the ecosystem response. Combining this remediation study with our past and present 

studies will complete a comprehensive processoriented study at the St. Kevin Gulch site.

REFERENCES CITED

Allison, J.D., Brown, D.S., and Novo-Gradac, K.J., 1991, 

M1NTEQ2/PRODEFA2, a geochemical assessment 

model for environmental systems Version 3.0 User's 

Manual: EPA/600/3-91/021, U.S. Environmental Protection Agency, 106 p.

Bencala, K.E., McKnight, D.M., and Zcllwegcr, G.W., 

1987, Evaluation of natural tracers in an acidic and 

metal-rich stream: Water Resources Research, v. 23, 

p. 827-836.

Bencala, K.E., McKnight, D.M., and Zcllwegcr, G.W., 

1990, Characterization of transport in an acidic and 

metal-rich mountain stream based on a lithium tracer 

injection and simulations of transient storage: Water 

Resources Research, v. 26, p. 989-1000.

757
Bcncala, K.E., and Walters, R.A., 1983, Simulation of

solute transport in a mountain pool-and-riffle stream 

A transient storage model: Water Resources Research, 

v. 19, p. 718-724.

Broshcars, R.E., Bcncala, K.E., Kimball, B.A., and

McKnight, D.M., 1993, Tracer-dilution experiments 

and solute-transport simulations fora mountain stream. 

Saint Kevin Gulch, Colorado: U.S. Geological Survey 

Water-Resources Investigations Report 92-4081, 18 p.

Broshcars, R.E., Kimball, B.A., and Runkel, R.L., 1996, 

Simulation of reactive transport during a pH modification experiment in a mountain stream affected by acid 

mine drainag, in Morganwalp, D.W., and Aronson, 

D.A., ed., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Harvey, J.W. and Bcncala, K.E., 1993, The effect of 

strcambcd topography on surface-subsurface water 

exchange in mountain catchments: Water Resources 

Research, v. 29, no. I., p. 89-98.

Kimball, B.A., Broshears, R.E., Bencala, K.E., and

McKnight, D.M., 1991, Comparison of rates of hydrologic and chemical processes in a stream affected by 

acid mine drainage, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxics Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, California, March 11-15, 1991: 

U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 407-412.

Kimball, B.A., Broshcars, R.E., McKnight, D.M., and 

Bcncala, K.E., 1994, Preliminary evaluation of the 

effects of an instream pH modification on transport of 

sulfide-oxidation products, in Alpers, C.L, and Bowles, 

D., cds.. The Environmental Geochemistry of Sulfidc 

Oxidation: Washington, D.C., American Chemical 

Society Books, p. 224-243.

Kimball, B.A., McKnight, D.M., Broshears, R.E., and

Bencala, K.E., I992a, Effect of instream pH modification on aluminum, in Kharaka, Y. K. and Maest, A. S., 

eds., Water-Rock Interaction, Volume I, Low Temperature Environments: Rotterdam, A.A. Balkema, 

p. 393-396.

Kimball, B.A., McKnight, D.M., Wetherbee, G.A., and 

Harnish, R.A., I992b, Mechanisms of iron photo 

reduction in a metal-rich, acidic stream (St. Kevin, 

Gulch, Colorado, U.S.A.): Chemical Geology, v. 96, 

p. 227-239.

McKnight, D.M., 1988, Metal-tolerant algae in St. Kevin 

Gulch, Colorado, /;;, Mallard, G.E., ed., U.S. Geological Survey Toxic Substances Hydrology Program 

surface-water contamination Proceedings of the 

technical meeting, Denver, Colorado, February 2-4, 

1987: U.S. Geological Survey Open-File Report 87- 

764, p. 113-117.

McKnight, D.M., Kimball, B.A., and Bencala, K.E., 1988, 

Iron photoreduction and oxidation in an acidic mountain stream: Science, v. 240, p. 637-640.

Moore, J.N. and Luoma, S.N., 1990, Hazardous wastes 

from large-scale extraction: Environmental Science & 

Technology, v. 24, p. 1278-1285.

Ranville, J.F., 1992, Factors influencing the clectrophorctic 

mobility of suspended sediments in acid mine drainage: Golden, Colorado School of Mines, unpublished 

Ph.D. dissertation T-4213, 270 p.

Runkel, R.L., and Broshcars, R.E., 1992, One-dimensional 

transport with inflow and storage (OTIS) A solute 

transport model for small streams: Boulder, Colo., 

Center for Advanced Decision Support for Water and 

Environmental Systems, 85 p.

Runkel, R.L. and Chapra, S.C., 1993, An efficient numerical solution of the transient storage equations for solute 

transport in small streams: Water Resources Research 

v. 29, p. 211-215.

Runkel, R.L., Bcncala, K.E., and Broshears, R.E., 1996, An 

equilibrium-based simulation model for reactive solute 

transport in small streams, in Morganwalp, D.W., and 

Aronson, D.A., cd., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Smith, K.S., Ranville, J.F., and Macalady. D.L., 1991, Predictive modeling of copper, cadmium, and zinc partitioning between strcamwater and bed sediment from a 

stream receiving acid mine drainage, St. Kevin Gulch, 

Colorado, in Mallard, G.E., and Aronson, D.A., cds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, California, March 15, 1991: U.S. Geological Survey Water-Resources Investigations Report 

91-4034, p. 380-386.

Stream Solute Workshop, 1990, Concepts and methods for 

assessing solute dynamics in stream ecosystems: 

Journal of the North American Bcnthological Society, 

v. 9, p. 95-119.

Tate, C.M., McKnight, D.M., and Spaulding, S.A., 1991, 

Phosphate uptake by algae in a stream contaminated by 

acid mine drainage, St. Kevin Gulch. Leadville, Colorado, in Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: U.S. Geological Survey Water-Resources Investigations Report 

91-4034, p. 387-391.

Walton-Day, K., 1992, Hydrology and geochemistry of a 

natural wetland affected by acid mine drainage 

affected by acid mine drainage St. Kevin Gulch, Lake 

County, Colorado: Golden, Colorado School of Mines, 

unpublished Ph.D. dissertation T-4033, 299 p.

Zellweger, G.W. and Maura, W.S., 1991, Calculation of 

conservative-tracer and flume-discharge measurements 

on a small mountain stream, in Mallard, (J.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 434-437.

758
Iron and Zinc Budgets in Surface Water for a Natural

Wetland Affected by Acidic Mine Drainage,

St. Kevin Gulch, Lake County, Colorado

By Katherine Walton-Day 1

Abstract

Iron and zinc mass input and output in surface 

water were estimated for a wetland receiving 

acidic mine drainage near Leadville, Colorado, to 

evaluate (1) whether the wetland is removing the 

metals from surface water, and (2) the processes 

causing the observed metal fate. The results 

indicate that about 560 and 850 kilograms of iron 

were removed from surface water from late spring 

through fall 1988 and from late spring through 

midsummer 1989, respectively. Iron removal 

occurred primarily through settling of iron oxyhydroxide particles greater than 0.1 micrometer. The 

results also indicate that zinc was not removed 

from surface water by the wetland. All of the zinc 

is in the filtered fraction. Either contact time 

between the water and the wetland substrate is 

insufficient for the reactions that could remove 

dissolved zinc, such as sulfate reduction, to occur, 

or zinc removal reactions do not occur in this 

wetland.

INTRODUCTION

Studies of wetlands used to treat acidic mine drainage (AMD) commonly compare concentrations of 

metals at inflow and outflow locations to assess treatment efficiency. Wieder (1988) suggested instead that 

studies of wetland-treatment systems should consider 

the mass or mass rate of metal transported into and out 

of these systems. Comparison of inflow and outflow 

concentrations fails to include effects of hydrologic 

processes, such as addition of clean water by springs 

and evapotranspiration, which cause dilution and

'U.S. Geological Survey, Denver, Colo.

concentration, respectively, that are unrelated to chemical interactions between water and substrate. Both 

processes are important in studies of natural wetlands. 

In this study, a natural wetland that receives AMD was 

studied to (1) use stream discharge and water-quality 

data to estimate iron and zinc mass input and output, 

and any net retention of either metal in the wetland over 

the time intervals studied; and (2) identify hydrologic 

and geochemical processes causing current (1988-89) 

metal transport and fate of metals in the wetland.

STUDY-SITE DESCRIPTION

The wetland has an area of about 26 ha and is 

located about 7 km northwest of Leadville, Colo. (fig. 1). 

The wetland receives 99 percent of its surface water 

from six sources: Temple Gulch, Gleason Gulch, North 

Spring, St. Kevin Gulch, South Spring, and Porcupine 

Gulch Ditch (which flows intermittently) (fig. 1). The 

weathering of pyrite in the mineral deposits in St. Kevin 

Gulch produces AMD. A principal source of AMD 

(pH = 2.7, iron and zinc concentrations > 100 mg/L) 

emanates from mine-waste piles located about 2 km 

upstream from the discharge of St. Kevin Gulch into 

the wetland. The AMD is diluted by the water in 

St. Kevin Gulch to a pH of 3.5 to 4.5, and to iron and 

zinc concentrations of up to 10 mg/L. Previous work in 

the wetland indicates that most metal loading occurs 

during spring runoff (May through July). During this 

period, flow from St. Kevin Gulch comprises from 50 

to 85 percent of the flow into the wetland. In addition, 

discharge from St. Kevin Gulch into the wetland is 

intermittent and ceases during low-flow periods.

Hydrologic budgets for the two time periods of 

interest were previously constructed (Walton-Day, 

1991), and the results are listed in table 1. The budgets 

balanced within 4 percent in 1988 and about 10 percent 

in 1989; therefore, the budgets account for all major 

components of flow into and out of the wetland.

759
39°18'

39°17'30'

106*22'

T

106e2V30'

|

South 

Spring

Wetland* 

Outflow

EXPLANATION

- N APPROXIMATE 

, OUTLINE

- - ' OF WETLAND

d^ SPRING

- SAMPLING SITE

X

PARSHALL FLUME 

AND SAMPLING SITE

\ 

V..

STREAMGAGE AND 

SAMPLING SITE

STREAM WITH 

INTERMITTENT OR 

POORLY CHANNELIZED 

FLOW

STREAM WITH 

PERENNIAL FLOW

STUDY 

*M*«0 °DENV 

IfAOVlUE

COLORADO

V2 MILE

I 

0.5 KILOMETER

Figure 1. Location of wetland showing flume, streamflow gage, and sampling sites. Location of the 

channel for Porcupine Gulch is approximate.

Table 1. Hydrologic budget for St. Kevin wetland

[Volume of water in thousands of cubic meters. 1988 budget is from 

May through October. 1989 budget is from May through July. Totals 

rounded to two significant figures. Modified from Walton-Day, 1991, 

table 1]

Surface Ground 

water water Precipitation Storage Total

1 988 inputs 

1 989 Inputs

660 

950

7 

6.3

20

24

61 

1. 2

750 

980

Surface Ground Evapotranswater water piration Total

1 988 outflows 

1 989 outflows

620 

990

70 

3.1

92 

87

780 

1,100

Surface-water flow dominates the hydrologic budget. 

Most of this flow (80 percent or greater) occurred 

during May and June of each year. These results justify 

the assumption that surface water is the primary component in the element budgets presented herein. Exceptions to this assumption are described in the Results 

and Discussion section.

METHODS

Iron and zinc mass budgets for the wetland were 

constructed using surface-water discharge and waterquality data from May through October 1988, and May 

through July 1989. Surface-water flow into and out of 

the wetland was monitored by fiberglass Parshall 

flumes located on the five inflow streams (except

760
Porcupine Gulch Ditch) and by a streamflow gage at 

the wetland outflow (fig. 1). Stream stage was recorded 

at each site every 30 minutes from about May 1 8- 

November 10, 1988, and from about May 4-September 30, 1989. A combination of stage-discharge relations and rating curves developed over the 

measurement periods were used to calculate discharge 

at the sites as detailed by Walton-Day ( 1 99 1 ). A continuous record was not obtained for Porcupine Gulch 

Ditch, but was developed using data from the other 

streams (Walton-Day, 1991). Errors in discharge 

measurements were estimated to equal ±10 percent 

(Walton-Day, 1991).

Surface-water samples were collected every 3 to 4 

weeks from June 1 November 10, 1988, and every 2 

weeks from May 4-July 13, 1989. Grab samples were 

collected in the narrow, turbulent throats of the flumes 

where the water was well mixed. At the streamflow 

gage, samples were composited from three to seven 

grab samples collected at various locations across the 

narrowest section of the stream. Two aliquots~one 

unfiltered and acidified (representing total recoverable 

metal concentrations), and the other filtered (0. 1 pm 

membrane) and acidified (representing filterable metal 

concentrations) were collected for analysis of iron and 

zinc concentrations. Iron (1988 and 1989 samples) and 

zinc (1988 samples) concentrations were determined 

using inductively coupled plasma-atomic emission 

spectrometry (Lichte and others, 1987). Zinc concentrations in 1 989 samples were determined using an 

inductively coupled plasma-mass spectroscopy 

system. Combined analytical and sampling errors were 

estimated to be ±5 percent (Walton-Day, 1991).

Iron and zinc budgets in surface water for the two 

time periods were calculated as follows:

RT = Z MFiNFLOWS - MFouTFLow » 0) 

where 

RT is retention through time, the net gain or loss of a

constituent from the wetland in kilograms, and 

MF is the calculated load (mass flow) of the

constituent at the inflows and outflow for the

period of time considered.

When RT is positive, the constituent is accumulating in 

the wetland. When RT is negative, the constituent is 

being flushed from the wetland. When RT is zero, or 

there is no significant difference between the mass of 

the constituent in the inflows and outflows, then the 

wetland has no effect on transport of the constituent.

Constituent loads were calculated using either a 

concentration calculated by use of the continuous discharge record or average constituent concentration; 

when the regression relation between the natural logarithms of measured concentrations and associated 

stream discharge was significant (95 percent confidence level), then the regression was used to calculate 

constituent concentration with the following equation:

CCA LC = eb -Qm, (2) 

where 

CCALC 's tne concentration of the species (in

micrograms per liter), calculated from the 

discharge Q (in liters per second); 

m is the slope; and 

b is the y intercept of the regression line.

The load in a stream through time is

(3)

where

HAM (= MF,NFLOW or MFOUTFLOW) is the load (in

kilograms) of a species of interest over the time

period of interest, 

8.64 x 10"5 is a conversion factor that incorporates the

number of seconds in a day and converts

micrograms to kilograms, and 

Qi is the average daily discharge of the stream over n

days.

When the regression relation between concentration 

and discharge was not significant at the 95 percent confidence level, then average concentrations were used to 

calculate constituent load by the following equation:

n

MFSTREAM = 8.64xlO-5 -CAVGZ Qi, (4) 

where i = i 

MFSTREAM is the load (in kilograms) of a species over

the time period of interest, 

CAVG (*n micrograms per liter) is the mean

concentration of the species based on samples

collected at the stream over the period of

interest, and 

Qi is the average daily discharge of the stream over n

days.

The error in the estimates of load (where load is the 

product of concentration and discharge) is ± 15 percent, 

the sum of the errors of the concentration and discharge 

measurements.

761
Table 2. Average concentrations of iron and zinc in surface water in St. Kevin Gulch and tributaries, 1988-89

[All concentration are in micrograms per liter. Unfilt. indicates concentrations in unfiltered water samples. Filt. indicates concentrations

in filtered water samples]

Element 

and

fraction

Iron

(unfilt.)

Iron

(fill.)

Temple Gulch

1988 1989

850 1,200

270 180

Porcupine Gleason Gulch 

Gulch

Ditch

1989 1988 1989

160 410 350

50 150 60

North Spring South Spring

1988 1989 1988 1989

670 420 230 260

160 180 150 110

St. Kevin Gulch Wetland Outflow

1988 1989 1988 1989

1,350 1,650 320 270

870 850 180 220

Zinc 

(unfilt.)

Zinc 

(fill.)

100

95

290

250

13

14

150

160

250

230

4,200

4,000

4,500

4,600

2,000

2,000

2,000

2,000

6,300

6,100

4,100

4,100

1,800

1,900

2,500

2,600

RESULTS AND DISCUSSION

Comparison of average values in table 2 shows that 

concentrations of unfiltered iron are greater than concentrations of filtered iron at almost every site, indicating that paniculate iron is a major component of the 

total iron load entering the wetland. Unfiltered and filtered zinc concentrations do not significantly differ relative to the analytical error. In a few instances, average 

concentrations of zinc in filtered samples are slightly 

greater than in unfiltered samples. This discrepancy 

reflects analytical and rounding errors and is considered negligible. Consequently, only the loading estimates for unfiltered zinc are shown on table 3. Table 2 

also shows that inputs to the wetland other than St. 

Kevin carry elevated concentrations of metals. These 

inputs represent minor contributions to the total load 

during spring runoff but are the major contributors to 

metal loads when flow into the wetland from St. Kevin 

Gulch ceases. The data in tables 2 and 3 indicate that 

average concentrations were used to calculate all loads 

except loads of unfiltered zinc in South Spring, of unfiltered iron and zinc at the St. Kevin Gulch inflow to the 

wetland, and of unfiltered zinc at the wetland outflow. 

The linear regression relations between the natural logarithms of discharge and constituent concentration are 

significant at the 99 percent confidence level for all 

relations in table 3 except for unfiltered zinc concentrations at the wetland outflow, which are significant at the 

95 percent confidence level. Loads of all metals (table 4) 

were greater in 1989 than in 1988 despite the fact that 

loads were calculated over a shorter time period in 

1989. The difference occurs because the continuous

Table 3. Regression relations used to calculate unfiltered iron 

and zinc concentration at locations where the relations between 

the natural logarithms of discharge and concentration were 

statistically significant

[Unfilt. indicates relations developed from concentrations measured in 

unfiltered water samples. * indicates that linear relationship between the 

natural logarithms of discharge and concentration was not significant, and 

the mean values in table 2 were used to calculate the metal load. C(-AI,O 

calculated concentration of the species in micrograms per liter; e, base of 

the natural logarithm; b, y-intercept of the regression line between the 

natural logarithms of discharge and concentration; Q, average daily 

discharge; and m, slope of the regression line between the natural 

logarithms of discharge and concentration. All relations significant at the 

99 percent confidence level, except zinc at wetland outflow, is significant 

at the 95 percent confidence level]

South spring St. Kevin Gulch Wetland outflow

Iron (unfilt.) * 

Zinc (unfilt.) CCA1 c = e6 94Q

C f*n i^rv 1 2 * CALC ~ e V

recorders were installed earlier in 1989 than in 1988, 

and, therefore, a higher proportion of the spring runoff 

was included in the 1989 calculations than in the 1988 

calculations. This difference illustrates the importance of 

spring runoff to the loads carried into and out of the wetland. By including 2 extra weeks of spring runoff in the 

1989 calculation, loads are 20 to 80 percent greater than 

in 1988 even though the period after July was not 

included in the 1989 calculation. The data in table 4 also 

show that mass inflows differ significantly from outflows 

only for iron in both years. There is no significant difference between input and output loads of zinc for either 

year because the differences between input and output 

loads are not greater than the ±15 percent error associated with the calculations.

762
Table 4. Total loads in surface-water input and output of the 

wetland, 1988 and 1989

[Loads in kilograms. 1988 loads calculated for May through October. 

1989 loads calculated for May through July. Unf ill. indicates loads are 

based on concentrations from unfiltered water samples. Fill, indicates 

loads are based on concentrations from filtered water samples. Bold 

type for a value in the "Net" row indicates that the value is significant. 

Net is significant when smaller value +15 percent is less than larger 

value-15 percent]

1988

Iron load

Input 

Output 

Net

Unfilt.

760 

200 

560

Filt.

360

no

250

Zinc load

Unfilt.

1,600 

1,500 

100

1989

Input 

Output 

Net

uoo

250 

850

470 

2IO 

260

2,000 

2,400 

-400

The difference in net retention of filterable and 

unfiltered iron (table 4) indicates that the settling of 

iron oxyhydroxide particulates in the greater than 

0.1 -iron accumulated in the wetland sediments. Net 

removal of filterable iron (table 4) may be caused by 

(1) precipitation of iron oxyhydroxides, (2) aggregation and settling of colloidal iron oxyhydroxide floe, 

(3) formation of iron sulfides, (4) sorption, and/or (5) 

organic complexation of iron. However, the relatively 

low permeability of the wetland substrate and short 

residence times of water in the wetland restrict the 

amount of water in contact with the wetland substrate 

in which processes 3-5, above, occur. Therefore, most 

iron removal probably occurs through physical settling 

of particulate iron oxyhydroxides and through aerobic 

processes that promote the formation of iron oxyhydroxides in the filterable size fraction. The lack of 

significant, current (1988-89), accumulation of zinc 

conflicts with recent research on wetlands that indicates zinc accumulation in wetland sediments through 

sulfate reduction (Machemer and Wildeman, 1992). 

The most likely explanations are that (1) loading rates 

of zinc onto the wetland are too great, and residence 

times too short to allow sufficient contact time between 

zinc and the wetland substrate where sulfate reduction 

occurs, and/or (2) that sulfate reduction is not occurring 

in this wetland.

The difference between the budget results for iron 

and zinc relates to the dominant forms of each metal in 

surface water and to the mechanisms that can remove 

each metal in a wetland environment. Most of the iron

is present in the particulate or suspended-sediment 

fraction of the inflow loads and can be removed from 

surface water simply by settling where water velocities 

in the wetland decrease. This process requires no 

interaction between the water column and wetland 

substrate. Conversely, all of the zinc is present in dissolved form. The processes that could remove zinc in 

the wetland include sulfate reduction and precipitation 

of sulfide minerals, and sorption. Both of these processes require contact between the water and wetland 

substrate. The lack of detectable zinc removal in this 

wetland, at least within the range of error of the measurements, indicates that the contact time between the 

water and the wetland sediments in not sufficient for 

these processes to occur, and/or that the processes do 

not occur to a significant extent in this wetland. The 

assumption that surface water is the only component of 

the element budgets is valid for zinc but not for iron. 

Additional data in Walton-Day (1991) indicate that 

when ground-water transport was included in the 1988 

budget, the amount of iron that appeared to be accumulating in the wetland, based on the results of the surface-water mass budget, was balanced by an equivalent 

amount of iron being transported downward out of the 

wetland sediments into the underlying aquifer. This 

result does not invalidate the budgets presented here, 

but indicates that iron dynamics in the wetland are 

more complicated than the simple surface-water budget suggests. Additional research is needed to understand these dynamics.

SUMMARY

Calculations of annual loads of iron and zinc in a 

wetland receiving drainage from St. Kevin Gulch show 

that loads of iron in inflow significantly and consistently exceeded loads in outflow, whereas loads of zinc 

in inflow and outflow did not differ significantly. The 

wetland sediments removed 560 and 850 kg of iron 

from surface water in 1988 and 1989, respectively. 

Settling of particulate iron oxyhydroxide material from 

the greater than 0.1 -jim size fraction accounts for 60 to 

70 percent of the iron removed from surface water during transport through the wetland. This process also 

may remove filterable iron. This budget only includes 

the surface-water components of element transport 

and, therefore, excludes the complex dynamics that 

affect iron transport in ground water in this wetland.

763
The lack of detectable zinc removal probably 

results from inadequate residence time of water with 

the wetland substrate during the period of maximum 

loading (spring runoff); that is, the hydraulics of this 

natural wetland system do not allow sufficient contact 

time between the substrate and most of the water for the 

reactions to occur that can remove zinc. It is also possible that processes that can remove zinc, such as sulfate reduction, simply are not occurring at a sufficient 

rate to affect zinc transport in this wetland.

REFERENCES

Lichte, F.E., Golightly, D.W., and Lamothe, P.J., 1987,

Inductively coupled plasma-atomic emission spectrometry, in Baedecker, P.A., ed., Methods for geochemical 

analysis: U.S. Geological Survey Bulletin 1770, 

p. B1-B10.

Machemer, S.D., and Wildcman, T.R., 1992, Adsorption 

compared with sulfide precipitation as metal removal 

processes from acid mine drainage in a constructed 

wetland: Journal of Contaminant Hydrology, v. 9, 

no. 1/2, p. 115-131.

Walton-Day, Katherine, 1991, Hydrology and geochemistry 

of a natural wetland affected by acid mine drainage, 

St. Kevin Gulch, Lake County, Colorado: Golden, 

Colo., Colorado School of Mines, Department of Geology and Geological Engineering, Ph.D. dissertation 

T-4033, 299 p.

Wieder, R.K., 1988, Determining capacity for metal retention in man-made wetlands constructed for treatment of 

coal mine drainage, Mine drainage and surface mine 

reclamation, in U.S. Bureau of Mines Information Circular 9183, Mine water and mine waste, p. 375-381.

764
Tracer Injections in Small Streams 

Why and How We Do Them

By Gary W. Zellweger1

Abstract

By using simple techniques and imagination, 

researchers can use conservative-tracer injections to collect hydrologic data on small streams. 

Stream hydrologic properties such as discharge, 

which cannot be measured in small streams by 

other means, are measurable with a precision 

as good as two percent by use of conservative 

tracers.

This paper briefly describes some of the 

hydrologic information that can be gained from 

continuous injection experiments and describes 

tracer selection criteria, injection equipment, and 

sampling and tracer analysis.

INTRODUCTION

Researchers can use tracers to learn more about 

stream hydrology than just an occasional measurement of discharge or time of travel. Hydrologists have 

successfully used tracers to collect discharge data, 

calculate dissolved loads, calibrate flumes, and follow 

the flow paths of surface water and ground water 

through subsurface zones, by using tracer injections 

that may last from a few minutes to several days, 

depending on the application. Calculations of hydrologic properties using tracer data range from simple 

equations to complex mathematical models. Because 

conservative tracers travel with water, tracers allow 

us to "label" water and observe its movement.

The purpose of this paper is to explain how hydrologic information can be collected from continuous 

injection of conservative tracers, chiefly anionic tracers 

injected into small streams, and to provide details and 

guides on the techniques of tracer injection. The intent 

is convince the reader of the value of tracer injections 

and encourage their use in hydrologic studies.

'U.S. Geological Survey, Menlo Park, Calif.

WHY WE DO THEM

Tracer experiments may have begun long ago 

when the first hydrologist stirred up the muddy bottom 

of a lazy stream and watched the turbidity cloud move 

downstream, spreading as it flowed, traveling faster 

through the middle of the stream and slower at the 

edges, gradually dispersing until it was invisible. This 

quaint image captures the essence of tracers. They provide the ability to observe the movement of water and 

measure hydrologic properties of streams and stream 

channels. Tracer research is not new; the historical use 

of tracers was reviewed by Spencer and Tudhope 

(1958).

A conservative tracer, by definition, does not react 

with streamwater or the stream channel in any manner 

that either increases or decreases the mass of tracer. 

This property allows the researcher to use conservative 

tracers to acquire knowledge about the hydrology of 

small streams.

Knowledge that can be gleaned from a wellplanned and executed tracer experiment includes time 

of travel, instantaneous discharge, temporal and 

longitudinal variation in discharge, dispersion, rates 

of ground-water inflow and stream outflow (by difference in discharge), flow in the subsurface zone, losses 

and gains of dissolved compounds such as heavy metals or nutrients, stream and subchannel transient 

storage, and mixing of surface water and groundwater. Samples of anionic tracers can be analyzed with a 

precision of about I percent, and some hydrologic 

properties can be calculated on the basis of 

concentrations of tracers with a precision as good as 

two percent (Zellweger, 1993).

HOW WE DO THEM

Tracer injections are composed of three parts: 

tracer selection, delivery, and sampling and analysis.

765
Tracer Selection

Among the criteria to consider when selecting 

tracers are measurability, solubility, availability, 

background concentration, cost, and conservativeness. For most applications conservativeness of the 

tracer is not merely important, it is crucial.

The choice of tracers is generally limited to the 

stable anions: chloride, bromide, and sulfate, or 

organic dyes. Of these, chloride as sodium chloride 

and sulfate as magnesium sulfate (Epsom salt), share 

the advantage of being inexpensive. Infrequently, sulfate may not be conservative when dissolved oxygen 

concentration is low, and the dissolved organic carbon 

concentration is great enough that bacteria might 

reduce sulfate.

The characteristics of several fluorescent dyes 

were described by Smart and Laidlaw (1977). 

Rhodamine WT is commonly used for water tracing; 

its use is described by Wilson and others (1986). 

However, rhodamine is not always conservative. 

Rhodamine fluorescence decreases rapidly in acidic, 

metal-rich, sunlit streams (Zellweger and others, 

1987). Bencala and others (1983) found that rhodamine was lost in a pool-and-riffle, gaining stream, 

where a large amount of the surface water passed 

through streambed sediment.

In low order, pool-and-riffle streams with coarse 

gravel beds and low discharge, a significant portion of 

flow may pass over and through the sediment beds. 

Under these conditions dyes tend to be lost whereas 

anionic tracers remain conservative.

In higher order streams that are less pool-and-riffle, 

have greater discharge, and have a greater flow to bedsurface ratio, dye loss to sediment is less important. 

Under these conditions dyes perform well; anionic 

tracers also perform well, but the mass required, milligrams per liter for ions verses micrograms per liter for 

dyes, is a disadvantage.

Anions and dyes can be thought of as complementary tracers anions for small streams and dyes for 

larger streams.

Delivery System

In a continuous injection tracer study, a concentrated solution of tracer is added to a stream at a steady 

rate. The delivery system includes a reservoir of injectate and a pump for delivering the injectate to the 

stream.

Reservoirs can range in size from I gal bottles to 

30 gal plastic garbage cans to a 12-ft-diameter child's 

collapsible wading pool. When tracers are dissolved in 

streamwater, the rate of dissolution becomes slower as 

the solution concentration approaches saturation; 

exceeding 80 percent saturation is a slow process. 

Exact measurement of tracer or water is not necessary 

because the injectate is usually sampled, diluted, and 

analyzed along with the samples. Or, since relative 

concentrations are often all that is needed, the injectate 

can be diluted and used as the standard for measuring 

the stream samples.

The injection pump must be capable of delivering 

injectate at a constant rate for as long as several days, 

depending on the application. The delivery rate 

should have a precision similar to that of the sample 

analyses, 1 percent. Metering pumps are recommended, because they are more reliable than passive 

gravity systems.

Many types of metering pumps which use 120 

V a-c can be found in laboratory equipment catalogues. 

Lacking stable 120 V a-c power, battery-operated 

equipment must be used. Metering pumps that operate 

on a 12-V battery have been used to inject as much as 

50 mL/min with precision better than 1 percent. Pumps 

are available with rotation sensors that data-logging 

devices can use to control and record pump speed.

The injection site should be located so that the tracer 

will be transversely mixed before it reaches the first 

sampling point. Because the stream length required to 

mix the tracer is proportional to the square of the stream 

width (Ward, 1973, Kilpatrick and Cobb, 1985), injection into locations where streams are narrow mix in 

shorter distances than where streams are wide.

Background concentration of tracer, contamination 

of samples, and imprecision in the analysis of low concentrations of tracer can cause loss of resolution in the 

calculation and interpretation of the results of tracer 

studies. These difficulties decrease as the concentration 

of tracer used in the study increases. Within the restrictions of cost, volume of injectate, and environmental 

concerns, better results are achieved with greater 

amounts of tracer.

Sampling and analysis

Streamwater samples for measuring the concentration of conservative tracers do not need to be filtered 

when collected; however, the analytical method or

766
instruments may require that samples be filtered when 

analyzed. New polyethylene bottles are not highly contaminated with tracer anions; rinsing them with sample 

water before filling should remove the highly soluble 

tracer ions. Because it is easier to collect additional 

samples than to interpolate between scattered data or to 

repeat experiments, it is advisable to collect abundant 

samples, including samples of background tracer concentrations.

Chloride, bromide, and sulfate concentrations can 

be readily measured by ion chromatography. Chloride 

and bromide can also be analyzed colorimetrically, 

however the method forms the same color with chloride as with bromide, so that the analysis is difficuttto 

interpret if both ions are present. Both methods have a 

precision of about 1 percent.

CALCULATIONS

Simple to complex hydrologic properties can be 

calculated from the results of a continuous tracer injection. After a continuous injection begins the tracer 

concentrations at locations downstream of the injection 

point increase from background values, which could be 

zero, to the steady state or "plateau" values.

The calculation of discharge is an example of a 

simple calculation (Kilpatrick and Cobb, 1985) that 

requires the injectate concentration, injection rate, the 

background concentration, and a plateau concentration 

downstream to the injection point. More elaborate calculations, such as determining dispersion equation 

parameters, can be made using data collected as the 

concentration increases to the plateau (Bencala, 1984).

SUMMARY

A large amount of hydrologic information can be 

obtained from a continuous injection of conservative 

tracers by use of techniques that have been field tested 

and are easily learned. The methods are simple to use

and are adaptable to field conditions. Supplies are 

inexpensive, and equipment is not difficult to use. 

Calculations of some hydrologic properties can be 

made with an accuracy as good as about 2 percent.

REFERENCES

Bencala, K.E., 1984, Interactions of solutes and streambed 

sediment, 2 A dynamic analysis of coupled hydrologic and chemical processes that determine solute 

transport: Water Resources Research, v. 20, no. 12, p. 

1804-1814.

Bencala, K.E., Rathbun, R.E., Jackman, A.P., Kennedy, 

V.C., Zellweger, G.W., and Avanzino, R.J., 1983, 

Rhodamine WT dye losses in a mountain stream 

environment: Water Resources Bulletin, v. 19, no. 6, 

p. 943-950.

Kilpatrick, F.A., and Cobb, E.D., 1985, Measurement of 

discharge using tracers: U.S. Geological Survey Techniques of Water- Resources Investigations, book 3, 

chap. A16, 52 p.

Smart, P.L., and Laidlaw, I.M.S., 1977, An evaluation of 

some fluorescent dyes for water tracing: Water 

Resources Research, v. 13, no. 1, p. 15-33.

Spencer, E.A., and Tudhope, J.R., 1958, A literature survey 

of the salt-dilution method of flow measurement: Institute of Water Engineers Journal, v. 12, p. 127-138.

Ward, P.R., 1973, Predictions of mixing lengths for river 

flow gaging: Journal of Hydraulics Division, American 

Society of Civil Engineers, 99, no. HY7.

Wilson, J.F., Cobb, E.D., and Kilpatrick, F.A., 1986, Fluorometric procedures for dye tracing: U.S. Geological 

Survey Techniques of Water-Resources Investigations, 

book 3, chap. A12, 34 p.

Zellweger, G.W, Bencala, K.E., McKnight, D.M., Hirsch, 

R.M., and Kimball, B.A., 1987, Practical aspects of 

tracer experiments in acidic, metal enriched streams: 

U.S. Geological Survey Open-File Report 87-764, 

p. 125-130.

Zellweger, G.W., in press, Testing and comparison of four 

ionic tracers to measure stream flow loss by multiple 

tracer injection: Hydrological Processes, v. 7, p. xxx.

767
768
Processes Controlling Dissolved Copper Concentrations 

during an Instream pH-Modification Experiment

By Kathleens. Smith 1

Abstract

Changes in dissolved copper (Cu) concentrations as a function of pH were observed during an 

instream pH-modification experiment conducted 

in St. Kevin Gulch, Colorado, a stream affected by 

acidic mine drainage. Understanding the processes 

that control dissolved Cu concentrations is important because dissolved Cu can pose health risks 

and contaminate receiving waters. Comparison of 

dissolved Cu concentrations during the instream 

experiment with changes in dissolved Cu 

concentrations in inbottle simulation experiments 

performed with streamwater in the absence of 

streambed sediment shows good agreement. These 

data indicate that processes controlling Cu 

attenuation with increasing pH occur primarily in 

the water column. Sorption reactions and (or) 

coprecipitation reactions with iron and (or) aluminum precipitates forming in the water column in 

response to increasing pH may be the processes 

controlling Cu partitioning in these systems. 

Predictive computer modeling of Cu sorption onto 

hydrous ferric oxide slightly underestimates Cu 

attenuation observed in the instream system.

INTRODUCTION

St. Kevin Gulch is a small subalpine stream 

affected by acidic mine drainage. An instream pHmodification experiment was conducted at St. Kevin 

Gulch to examine the response of chemical species in 

this stream to pH changes (Kimball and others, 1994). 

The ambient pH of St. Kevin Gulch is about 3.5.

U.S. Geological Survey, Denver, Colo.

As pH increases, several reactions can take place, 

including precipitation, sorption, and coprecipitation 

reactions. In the pH range of the instream pH-modification experiment (pH about 3.5-6) several major 

elements, especially iron (Fe) and aluminum (Al), will 

likely be controlled by precipitation reactions and 

many trace elements will be controlled by sorption and 

(or) coprecipitation reactions. Reactions that appear 

to control Cu concentrations during the instream pHmodification experiment are examined in this paper. 

Smith (1991) has demonstrated that pH-dependent Cu 

adsorption by iron-rich sediment seems to control Cu 

concentrations in experiments conducted with streamwater and streambed-sediment mixtures collected from 

St. Kevin Gulch.

The objective of this paper is to compare information from the application of different approaches used 

to evaluate processes that control Cu concentrations 

under changing pH conditions. These approaches 

include inbottle simulation experiments, computermodel predictions, and field observations during an 

instream pH-modification experiment.

STUDY METHODS

On August 25, 1988, the instream pH of a reach of 

St. Kevin Gulch was increased from ambient (about pH 

3.5) to about pH 6 by injecting Na2CC>3 dissolved in 

streamwater at a site along the stream (fig. 1). There 

were two steps in the pH modification. The first pH step 

involved constant injection of the Na2CO3 solution for 

a 3-hour period during which time the pH increased to 

just above 4. The second pH step included constant 

injection of the Na2CO3 solution at an increased rate for 

a 3-hour period during which time the pH increased to 

almost 6. Water samples were continuously collected at 

four sites downstream from the injection point during 

the course of the instream pH-modification experiment.

769
Unfiltered and filtered (0.1 um nitrocellulose filters) 

water samples were collected, and pH and temperature 

were measured. Study site 1 (fig. 1) is 24 m from the 

injection point. More detailed information about the 

injection procedure can be found in Kimball and others 

(1994).

On August 24,1988, a series of inbottle simulation 

experiments were conducted to examine pH-dependent 

partitioning of Cu between water and sediment from 

St. Kevin Gulch. Composite samples of streamwater 

and of iron-rich streambed sediment suspended in 

streamwater were collected along the reach of stream 

used for the instream pH-modification experiment.

Dispensed into a series of polyethylene bottles was 

40 mL of either the bed sediment/stream water suspension or of streamwater with no sediment. The pH was 

adjusted in the bottles by addition of differing amounts 

of 0.1 mol/L Na2CO3 to obtain a pH range from ambient to pH >6. No other substances were added. The bottles were allowed to sit for approximately 6 hours at 

stream temperature with occasional shaking. Care was 

taken to shield the bottles from sunlight. After the 

6-hour period, part of the contents in each bottle was 

separated by filtration through 0.1 -um nitrocellulose 

filters. The filtrates were saved for chemical analysis 

and final pH's were measured in the remaining portions

106°23'30"

390 18'30" -

39°17'30" -

- Injection point

498 

 x 

Location of $/

SITED

24 Site 1

Injection point 

70 Site 2

251 Site3

B

498 Site 4

250 METERS 

_J

500 FEET

.498

EXPLANATION

SAMPLING STATION-- 

Number indicates distance 

downstream from site 0, 

in meters

Figure 1. Map showing the reach of St. Kevin Gulch, Colo., used for an instream pH-modification experiment, the 

injection point of Na2CO3 , and downstream sampling locations. (From Kimball and others, 1992, fig. 1.)

770
in the bottles. Determination of the solute concentrations remaining in solution at each pH was used to 

determine pH-dependent partitioning reactions. In the 

inbottle simulation experiments conducted with 

stream water in the absence of the sediment, solute loss 

represents solubility reactions and (or) metal sorption 

(and (or) coprecipitation) onto particulates precipitated 

in the water column as a result of pH adjustment. Water 

samples were analyzed using inductively coupled 

argon plasma-atomic emission spectrometry (ICPAES) and ion chromatography.

Adsorption of Cu onto hydrous ferric oxide (HFO) 

for conditions during the instream pH- modification 

experiment was modeled using the Generalized TwoLayer Sorption Model (GTLM) and data base of equilibrium surface-complexation constants of Dzombak 

and Morel (1990). The same HFO properties were used 

in these modeling exercises as were used by Dzombak 

and Morel to fit the surface-complexation constants. 

The GTLM evokes a two-site model for cation sorption 

that includes a subset of high-affinity cation-binding 

sites. It assumes that M2+ is the only metal species that 

sorbs to HFO and that HFO is the sole sorbent material.

Computer-model simulations were conducted 

using the equilibrium speciation program MINTEQA2, 

version 3.00 (Allison and others, 1991). This program 

incorporates the GTLM and computes adsorption

reactions and solution equilibria simultaneously in a 

self-consistent manner. St. Kevin Gulch water chemistry from site 1 (fig. 1) at one of the sampling times 

was input to MINTEQA2. The pH was fixed at desired 

values (incremental values ranging from 3.5-6). 

Paniculate suspended-iron values were determined by 

subtracting the Fe concentration in filtered water from 

the Fe concentration in unfiltered water. In order to 

calculate the amount of sorbent HFO present, 15 particulate suspended-iron concentrations, which were 

measured during the times when the instream pH was 

>5 at site 1, were averaged (0.65 mg/L iron). It should 

be emphasized that all of the model input was derived 

either from chemical analysis of St. Kevin Gulch water 

or from information provided by Dzombak and Morel 

(1990). None of the model parameters were modified, 

and no fitting parameters were used in these computer 

simulations.

PROCESSES THAT CONTROL COPPER 

CONCENTRATIONS UNDER CHANGING 

pH CONDITIONS

Variation of dissolved Cu with stream pH and 

paniculate suspended-iron concentrations at study site 1 

(fig. 1) is depicted in figure 2. This figure illustrates the 

effect of both pH and paniculate iron on dissolved Cu

O

0.3 0.8

O

6.0

5.0

4.0

3.0

14

HOURS

Figure 2. Variations in concentrations of dissolved copper and paniculate suspended iron, and stream pH at study 

site 1 during an instream pH-modification experiment, St. Kevin Gulch, Colo.

I

Q_

LJ

hCO

771
concentrations. As the stream pH increases, the amount 

of paniculate suspended iron increases and the dissolved Cu concentration decreases. There is a corresponding shift of part of the dissolved Cu phase to a 

particulate phase. This shift indicates either precipitation, coprecipitation, and(or) pH-dependent sorption of 

the Cu. Dissolved Cu concentrations as a function of 

pH are shown in figure 3 for both the instream pHmodification experiment at study site 1 (fig. 1) and the 

inbottle simulation experiments. Dissolved Cu concentrations in the inbottle simulations conducted with 

streamwater in the absence of streambed sediment 

(hollow triangles) provide a very close match to the 

dissolved Cu concentrations observed during the 

instream pH-modification experiment (hollow circles). 

This indicates that most of the Cu attenuation processes 

are occurring in the water column rather than by 

interaction with bed sediment.

Smith (1991) has demonstrated that iron-rich 

streambed sediment can control Cu attenuation by pHdependent adsorption reactions in laboratory simulations using natural materials from St. Kevin Gulch.

^ 0.3

In both the inbottle simulation experiments discussed in 

this paper and the experiments conducted by Smith 

(1991), St. Kevin Gulch streambed sediment was placed 

in direct contact with streamwater and the two materials 

were mixed together; As such, these experiments maximized the contact between water and sediment. 

Although there is abundant iron-rich streambed sediment along the reach of St. Kevin Gulch used for the 

instream pH-modification experiment, there is no evidence of dissolved Cu attenuation by the bed sediment 

(fig. 3). Possible reasons that dissolved Cu concentrations do not appear to be controlled by interaction with 

bed sediment may include (1) hydrologic controls that 

prevent equilibration of dissolved constituents in the 

water column with the bed sediment (especially over 

short distances), (2) the presence of a stagnant liquid film 

at the water/bed-sediment interface that retards diffusion 

of dissolved constituents, (3) the physical proximity of 

suspended sediment to dissolved constituents, and (4) 

differences in sorption site-binding energies between the 

suspended particulates and the bed sediment. In order for 

Cu to adsorb onto HFO surfaces, Cu ions are thought to

O

cr LU

UJ

0.2

LJ 

0_

0.1

LU

>

_1

o (/) 

oo

Q 0.0

O 

A

Instream pH modification experiment

Inbottle simulation in absence of streambed sediment

Inbottle simulation in presence of streambed sediment

O

A O

o

A

A

3.0 4.0 5.0 6.0 7.0

Figure 3. Dissolved copper concentration as a function of pH during an instream pH-modification experiment 

(hollow circles) and in inbottle simulation experiments with streamwater in the presence (filled triangles) and 

absence (hollow triangles) of streambed sediment, St. Kevin Gulch, Colo.

772
o

0.

cr:

O cr

o LU

GL

0.2

LJ cn

Q_ 2

Q_ <

O tr

o o 0.1

odLJ 

0.0

Mode Predictio

O

nstream pH-Modification 

Experiment

3.0 4.0 5.0

Figure 4. Comparison of computer model predictions (solid curve) and observations from an instream 

pH-modification experiment (hollow circles) for dissolved copper as a function of pH, St. Kevin Gulch, Colo.

form an inner-sphere surface complex. To form an innersphere surface complex, waters of hydration between 

surface atoms and the Cu ion must be removed. The 

presence of a stagnant liquid film (possibility 2) at the 

water/ bed-sediment interface might interfere with the 

formation of inner-sphere surface complexes by retarding the diffusion rate of dissolved ions and waters of 

hydration.

The possibility that iron-rich suspended particulates 

control partitioning of Cu between the dissolved and 

particulate phases was tested using a MINTEQA2 computer simulation. The results are illustrated in figure 4. 

The computer simulation predicts no adsorption 

between pH 3.5 and 5, which corresponds to observations from the instream pH modification experiment. 

However, at pH >5.5 the computer simulation slightly 

underestimates Cu partitioning. This could be due to 

several considerations. First, the concentration of particulate iron is very small (averaging 0.65 mg/L, which corresponds to 0.001 g/L HFO, 5.8 x 10'8 mol/L of highenergy binding sites, and 2.3 x 10"6 mol/L of low-energy 

binding sites under conditions where dissolved Cu is 2.1 

x 10"6 mol/L). Surface-complexation models are not as 

reliable under conditions of high adsorbate/adsorbent 

ratios. Dzombak and Morel (1990) have tried to overcome this problem by developing a method to include 

the effects of surface precipitation into the GTLM. This 

is accomplished by adjusting the surface complexation 

constants by a constant factor that incorporates a solub'ility product for ferric hydroxide and a surface precipitation reaction for the hydroxide of the sorbing cation. 

Incorporation of surface precipitation effects into the 

GTLM was not done in the present study. Another explanation for computer model underestimation of Cu attenuation is that HFO is not the sole sorbent material. 

Aluminum hydroxide precipitates were observed forming during the second pH step of the pH-modification 

experiment (Kimball and others, 1992). It is likely that 

some dissolved Cu may be sorbed and (or) coprecipitated by these aluminum hydroxide precipitates.

773
SUMMARY REFERENCES

Subtle changes in dissolved Cu concentrations 

were observed during an instream pH- modification 

experiment conducted along a reach of St. Kevin 

Gulch, Colorado. There is close agreement between 

changes in dissolved Cu concentrations as a function of 

pH during the instream experiment and changes in 

dissolved Cu concentrations in inbottie simulation 

experiments performed with streamwater in the 

absence of streambed sediment. These data indicate 

that processes controlling Cu attenuation with increasing pH occur primarily in the water column. Sorption 

reactions and (or) coprecipitation reactions with iron 

and (or) aluminum precipitates forming in the water 

column in response to increasing pH likely control 

Cu partitioning in these systems. Predictive computer 

modeling of Cu sorption onto hydrous ferric oxide 

slightly underestimates Cu attenuation observed in the 

instream system. However, concentrations of particulate suspended iron are small, which reduces reliability 

of the model predictions.

ACKNOWLEDGMENTS

The work described in this paper is part of a much 

larger effort for the Upper Arkansas River SurfaceWater Toxics Project. I would like to thank Briant 

Kimball and Katie Walton-Day for their helpful 

reviews of this manuscript. Analytical support was 

provided by Paul Briggs, David Fey, and John 

McHugh.

Allison, J.D., Brown, D.S., and Novo-Gradac, K.J., 1991, 

MINTEQA2/PRODEFA2, a geochemical assessment 

model for environmental systems: Version 3.0 user's 

manual: U.S. Environmental Protection Agency Report 

EPA/600/3-91/021, 106 p.

Dzombak, D.A., and Morel, F.M.M., 1990, Surface complexation modeling: Hydrous ferric oxide: New York, 

John Wiley and Sons, 393 p.

Kimball, B.A., Broshears, R.E., McKnight, D.M., and 

Bencala, K.E., 1994, Effects of instream pH modification on transport of sulfide-oxidation products, in 

Alpers, C.N., and Blowes, D.W., cds., Environmental 

geochemistry of sulfide oxidation: Washington, D.C., 

American Chemical Society Symposium Series 550, 

p. 224-243.

Kimball, B.A., McKnight, D.M., Broshears, R.E., and 

Bencala, K.E., 1992, Effect of instream pH modification on aluminum, in Kharaka, Y.K., and Maest, A.S., 

eds., Water-rock interaction: Seventh International 

Symposium on Water-Rock Interaction, Park City, 

Utah, July 13-18, 1992, Proceedings: Rotterdam, 

A.A. Balkema, v. 1; p. 393-396.

Smith, K.S., 1991, Factors influencing metal sorption onto 

iron-rich sediment in acid-mine drainage: Golden, 

Colorado School of Mines, unpublished Ph.D. thesis, 

239 p.

774
An Equilibrium-Based Simulation Model 

for Reactive Solute Transport in Small Streams

By Robert L. Runkel 1 , Kenneth E. Bencala2 and Robert E. Broshears3

Abstract

A reactive solute transport model that 

considers the physical and aphysical processes 

affecting solutes in small streams is presented. 

The model is developed by coupling an existing 

solute transport code with a model of chemical 

equilibria. The resultant model considers a 

variety of physical and chemical processes 

including advection, dispersion, transient storage, 

precipitation/dissolution, sorption/desorption and 

the deposition of particulate metal oxides.

INTRODUCTION

Numerous streams and rivers in the United 

States are adversely impacted by acid mine drainage. 

In addition, many streams are naturally acidic. Given 

acidic conditions, trace metals are more soluble and 

may be present in significant concentrations within the 

water column. This is an important phenomenon, as 

many trace metals are toxic to fish and other aquatic 

organisms. Due to the acid mine drainage problem, 

investigators are currently studying mountain catchments. Some of the ongoing research involves fieldscale experiments to quantify the various processes 

affecting trace metals in streams. These experiments 

often involve the addition of an acid or a base to alter 

the rates of pH-dependent processes.

Two such experiments are those described by 

Broshears and others (1996) and McKnight and 

Bencala (1989). In the first experiment, a sodium

University of Colorado, Center for Advanced Decision 

Support in Water and Environmental Systems (CADSWES). 

2 USGS, Mcnlo Park, CA. 

3 USGS, Lakewood, CO.

carbonate solution was injected into Saint Kevin 

Gulch, a small mountain stream located near Leadville, Colorado. This injection resulted in an increase 

in pH that was accompanied by precipitation of iron 

and aluminum. In addition, copper was removed from 

the water column via sorption reactions. A second pH 

perturbation experiment was conducted on the Snake 

River, a naturally acidic stream located near the continental divide in north central Colorado, and involved 

an injection of sulfuric acid to decrease pH. This 

depression of pH resulted in an immediate increase in 

dissolved iron concentrations, due to the dissolution of 

hydrous iron oxides and the photoreduction of ferric 

iron.

Both of the experiments above involve complex 

processes that are difficult to quantify. Mathematical 

models provide a means of analyzing the processes 

that affect trace metals in natural systems. Solute 

transport models, for example, are used to study the 

physical processes that affect water quality. Models of 

chemical equilibria, meanwhile, are used to quantify 

chemical reactions in batch systems, wherein the 

effects of transport are not considered. Recently, 

complex ground water simulation models have been 

developed that couple hydrologic transport and 

chemical equilibrium. These models are used to 

quantify the nature and extent of various chemical and 

biological processes, in the context of hydrologic 

transport. Although many models are currently available for conventional contaminants, few models 

consider both physical transport and chemical 

equilibria for the case of trace metals in surface 

waters. In the research described herein, an 

equilibrium-based solute transport model for small 

streams is described.

775
MODELING EFFORTS

Many water quality models have been 

developed to assess the fate and transport of conventional contaminants, such as biochemical oxygen 

demand (BOD) and organic nitrogen, and environmental parameters, such as dissolved oxygen and temperature. Closely related to the water quality models 

are the solute transport models described by several 

authors (Zand and others, 1976; Bencala and Walters, 

1983). As opposed to modeling eutrophicationdissolved oxygen dynamics, these models are concerned with the transport of inorganic solutes including 

salts and trace metals. Like the water quality models, 

solute transport models describe the physical processes of advection and dispersion and some specific 

chemical and biological reactions. Simple solute 

transport models representing physical processes alone 

are useful tools for analyzing data obtained from tracer 

injection experiments. More complex models are 

developed by adding kinetic and/or equilibrium-based 

reactions. Examples of such nonconservative models 

include the work of Somlyody (1978), who described 

a model for heavy metals based on first-order reaction 

kinetics; and the research of Bencala (1983), in which 

a kinetic sorption model was developed for strontium.

A major shortcoming of most solute transport 

models is the simplistic nature in which reactive 

chemistry is considered. Existing models rely primarily on the specification of kinetic rate constants and 

on simple partition-coefficient representations of sorption phenomena. For the case of inorganic solutes, the 

database of kinetic rate constants is strikingly sparse, 

and many sorption reactions are thought to obey more 

complex sorption laws. While these transport models 

provide an accurate description of physical transport, 

they fail to include the degree of chemical sophistication needed to model pH perturbation experiments like 

those described above.

Chemical equilibrium models, in contrast, are 

useful in describing chemical equilibria in batch 

systems, but do not consider transport processes. 

When the chemical reactions of interest are relatively 

fast and reversible or the time scale of study is of 

sufficient length, chemical equilibria may be assumed. 

Models of chemical equilibria are founded on the 

mass-action equations that describe the equilibrium 

ratios of chemical species. Given the total mass of the 

chemical components and a set of equilibrium 

constants, the chemical species present at equilibrium 

are determined.

Fortunately, many chemical reactions are 

'sufficiently fast' relative to hydrologic processes, 

such that chemical equilibrium may be assumed. 

Given this local equilibrium assumption (LEA) 

(Rubin, 1983), it is possible to negate the disadvantages of both models by using a transport model to 

describe physics and an equilibrium model to describe 

chemistry. Indeed this is the approach taken by 

numerous authors for solute transport in groundwater 

systems (for example, Cederberg and others, 1985; 

Engesgaard and Kipp, 1992). Surprisingly, few 

models using the transport/equilibrium approach exist 

for surface waters.

The present work is directed towards 

developing a general model that expands the earlier 

work of Chapman and others (1982), who studied a 

short stream reach using a model that couples a semianalytical solution of the advection-dispersion equation with the MINEQL chemical equilibrium code 

(Westall and others, 1976). Several differences are of 

note. First, the model developed below uses a numerical solution of the transport equations, rather than the 

convolution integral approach employed by Chapman. 

In addition, the governing transport equations include 

a transient storage mechanism. Second, the model is 

applied over a much larger spatial scale. Finally, 

special attention is applied to the process of dissolution from stationary surfaces.

MODEL DEVELOPMENT

The general approach taken here is to couple an 

existing solute transport code with a chemical equilibrium model. The resultant model considers the 

physical processes of advection, dispersion, transient 

storage and the deposition of colloidal metal oxides, 

as well as the chemical processes of speciation, 

precipitation/dissolution and sorption. Consideration 

of these processes provides a general modeling system 

for the simulation of inorganic constituents in streams 

and rivers.

The solute transport model used for this work is 

developed by Runkel and Broshears (1991). This 

model is based on a one-dimensional advectiondispersion equation with additional terms to account 

for lateral inflows and transient storage. Transient 

storage is a useful empirical representation of a storage phenomenon that has been noted in small streams. 

Additional details of the transient storage approach are 

provided by Bencala and Walters (1983). In general,

776
the solute transport model may be thought of as a set 

of partial-differential equations describing the physical 

transport of multiple solutes. The chemical equilibrium model is based on MINTEQ (Allison and others, 

1991), a sophisticated geochemical model. Given 

analytical concentrations of the chemical components, 

MINTEQ computes the distribution of chemical 

species that exist within a batch reactor at equilibrium. 

These equilibrium computations include the precipitation and dissolution of solid phases as well as sorption 

processes. The mass-action equations describing 

equilibria form a set of nonlinear algebraic equations.

Given a coupled set of algebraic and partialdifferential equations, several solution techniques are 

available. A common method is known as the sequential iteration approach (Yeh and Tripathi, 1989), 

wherein each model time step is divided into a 

'reaction' step and a 'transport' step. During the 

reaction step, the algebraic equations describing 

chemical equilibria are solved for each segment in the 

stream network. In doing so, the equilibrium model 

determines the solute mass in dissolved, precipitated 

and sorbed forms. Based on this information, a transport step is taken in which the partial-differential 

equations describing transport are solved. In this step, 

the mobile phases of each solute are transported downstream. Because the solution steps neglect the 

coupling of the transport and chemistry, the procedure 

iterates until a specified level of convergence is 

achieved.

Assumptions

Several assumptions are embodied in the 

development of the surface water equations that 

follow. The assumptions inherent to the problem are 

summarized as follows:

'Chemical Equilibria. ~ Perhaps the most 

important assumption for the model developed here is 

the premise that complexation, precipitation and sorption reactions are in a state of chemical equilibrium. 

This assumption is based on the premise of local equilibrium, wherein chemical processes are considered 

sufficiently fast relative to hydrologic transport.

 One-Dimensional Transport. ~ A second assumption is that the stream is of small to moderate width 

and depth, such that one-dimensional transport is 

applicable. Under this assumption, solute mass is

assumed to be uniformly distributed over the stream's 

cross-sectional area. Given one-dimensional transport, the physical processes of advection, dispersion, 

lateral inflow and transient storage are considered.

 Mobile and Immobile Phases. ~ Solute mass for 

each chemical component is distributed among five 

distinct phases. The first three phases represent dissolved, precipitated and sorbed mass present in the 

water column. These three phases are mobile, in that 

they are subject to the effects of transport processes. 

The final two phases represent precipitated and sorbed 

mass that resides on the bed of the stream or is 

attached to stationary debris in the stream channel. As 

such, these final two phases are immobile and are not 

affected by transport mechanisms.

'Precipitation. ~ Dissolved mass in the water column 

may form precipitates if the solution becomes oversaturated with respect to the defined solid phases. Any 

precipitated mass initially resides in the water column 

and is subject to transport, until settling processes 

cause it to become immobile or redissolution returns it 

to the dissolved phase. This assumption states that 

precipitation occurs in the water column exclusively, 

and precipitation directly to the immobile bed is 

excluded. Precipitated mass may accumulate on the 

bed, however, as mass is deposited via settling.

 Dissolution. ~ When the aqueous solution is undersaturated, dissolution occurs preferentially from the 

water column. Under this scenario, all of the precipitate in the water column is allowed to dissolve before 

precipitate on the bed is subject to dissolution. This is 

a reasonable assumption, considering the intimate 

contact between precipitates in the water column and 

the flowing waters.

 Transport of Precipitates and Settling. Any

precipitates formed in the water column are sufficiently small, so as to be transported at a velocity that 

is identical to that for the dissolved species. This 

assumption obviates the need for a more sophisticated 

treatment of solids, such as sediment transport modeling. Transported precipitates are subject to the force 

of gravity and settle at a rate defined by a settling 

velocity. The settling rate of a particle is unaffected by 

other solids in solution (that is, unhindered settling).

777
 Sorption/Desorption. ~ Dissolved species may sorb 

to solid phases present in the water column or to sorption sites present on the bed of the stream. Conversely, sorbed species may desorb from sites in the 

water column or on the bed.

Governing Equations

With these assumptions in hand, the fundamental equations governing reactive solute transport in 

surface waters are presented. These equations are 

formulated for each chemical component defined 

within the equilibrium model. In short, chemical 

components are the fundamental building-blocks from 

which all chemical species are derived. As defined in 

MINTEQ, chemical reactions involve two or more 

components that react to form a chemical species. In 

general, the components must linearly combine to 

form every possible species in the system, and no 

component may be formed as a combination of other 

components (Allison and others, 1991)

To begin, the total component concentration, T, 

is defined by

r= c + pw + ph + sw + sb a)

where C, Pw Pb, Sw and S^ are the total 

dissolved, mobile (water-borne) precipitate, immobile 

(bed) precipitate, mobile sorbed, and immobile sorbed 

concentrations. The total component concentration 

(T) is composed of five separate phases, as given by 

Equation and depicted in Figure 1, where the 

surface water system is represented as two separate 

compartments. The first compartment represents the 

water column wherein the three mobile phases (C, Pw 

and Sw) are found. The second compartment is the 

immobile substrate, such as the streambed or debris, 

onto which water-borne precipitated and sorbed 

phases deposit. This compartment contains the two 

immobile phases, P^ and Sb.

Figure 1 provides a convenient summary of the 

processes affecting each phase. The dissolved phase 

(C), for example, is affected by transport processes, as 

represented by the transport operator, L(). In addition, 

the dissolved concentration is affected by precipitation/dissolution and sorption/desorption reactions 

within the water column (interactions with Pw and 5H,). 

Finally, C is affected by dissolution of precipitate from 

the immobile substrate and by sorption/desorption 

from immobile sorbents (interactions with P^ and 5^). 

Note that the settling process affects both the

precipitated and sorbed phases in the water column 

(Pw and Sw), as given by the settling velocities vpl and 

vs | [LT], respectively.

A differential equation describing conservation 

of mass for the total component concentration (T) is 

derived by developing mass balance equations for the 

five phases and summing them to yield

_//C 4-p\_L c (2)

where t is time, sext is a source/sink term representing 

external gains and losses and L() is the transport 

operator given by

(3)

where

C - concentration of an arbitrary component

phase (for example, T, Sh or Pf,)[M L"3 ], 

CL - solute concentration in lateral inflow

associated with C [M L"3 ], 

Cs- storage zone solute concentration

associated with C [M L'3 ],

>j

A- stream channel cross-sectional area [L ],

Q- volumetric flowrate [L3 T"'],

x- distance [L],

D- dispersion coefficient [L2 T"' ],

quw lateral inflow rate [L3 T"' L" 1 ], and

a - storage zone exchange coefficient [T"'].

Inspection of Equation reveals that T is a function of 

Pfr and Sfr. Equations for these phases are given by

dJjL 

dt

dS, 

dt

v ~d< (4)

(5)

where

fh_ source/sink term for dissolution from the

immobile substrate, 

gjj- source/sink term for sorption/desorption

from the immobile substrate, 

vpl- settling velocity for the precipitated phase

[LT 1 ],

778
Water Column

Immobile Substrate

Figure 1. Conceptual surface water system used to develop the governing differential equations for 

reactive solute transport. (Symbols as defined in text)

vs j- settling velocity for the sorbed phase

[LT1 ],

dj- effective settling depth [L], 

P- total precipitate (Pb + PM.) [ML"3 ], and 

S- total sorbate (Sh + Sw>) [ML~3 ].

Assuming chemical processes also affect solutes 

in the storage zone, a similar set of equations are 

developed for the total storage zone component concentration, Ts

(6)

(7)

(8)

where

source/sink term for dissolution from the 

storage zone immobile substrate,

gsfr- source/sink term for sorption/desorption

from the storage zone immobile substrate, 

sse\t ~ source/sink term representing external

gains and losses to the storage zone, 

vp2- settling velocity for the storage zone

precipitated phase [LT 1 ], 

vs2- settling velocity for the storage zone sorbed

phase [LT1 ],

d2- effective storage zone settling depth [L], 

As- storage zone cross-sectional area [L ], 

Ps- total storage zone precipitate concentration

[ML'3 ], 

Ss- total storage zone sorbate concentration

[ML'3 ], 

Pslj- mobile storage zone precipitate

concentration [ML ], and 

Ssh- mobile storage zone sorbate concentration

[ML'3 ].

Solution of Equations - proceeds as 

follows. First, the chemical equilibrium model is 

called twice for each segment in the stream network. 

On the first call, the total component concentrations, T, 

are supplied to the equilibrium model, and the total 

dissolved, precipitated and sorbed concentrations are

779
calculated for the main channel. For the second call, 

the total storage zone concentrations (Ts) are used as 

input, and the various phases are determined for the 

storage zone in a similar fashion. Given the output 

from the equilibrium model, the source/sink terms (//,, 

gfrfsfr 8sfr sext arjd ssext) are computed as described 

by Runkel (1993). Equations (4), (5), (7) and (8) are 

then solved for Ph, Sh, P^ and Ss^ respectively, and 

Equation (2) is solved for T, using the Crank-Nicolson 

method and the decoupling procedure described by 

Runkel and Chapra (1993). Finally, Equation (6) is 

solved for the total storage zone component concentration, 7^..

CONCLUSION

The model presented here provides a descriptive 

tool for analyzing the fate and transport of trace metals 

in streams and rivers. Work is presently underway to 

incorporate additional capabilities, including an 

expanded analysis of sorption/desorption reactions 

and the consideration of iron redox processes.

REFERENCES

Allison, J.D., Brown, D.S., and Novo-Gradac, K.J., 1991, 

MINTEQA2/PRODEFA2, a geochemical assessment 

model for environmental systems Version 3.0 user's 

manual: Washington, D.C., United States Environmental Protection Agency, 106 p.

Bencala, K.E., 1983, Simulation of solute transport in a 

mountain pool-and-riffle stream with a kinetic mass 

transfer model for sorption: Water Resources 

Research, v. 19, no. 3, p. 732-738.

Bencala, K.E. and Walters, R.A., 1983, Simulation of solute 

transport in a mountain pool-and-riffle stream A 

transient storage model: Water Resources Research, 

v. 19, no. 3, p. 718-724.

Broshears, R.E., Kimball, B.A. and Runkel, R.L., 1996, 

Simulation of reactive transport during a pH modification experiment in a mountain stream affected by acid 

mine drainage, in Morganwalp, D.W., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Ccderberg, G.A., Street, R.L., and Leckic, J.O., 1985, A 

groundwater mass transport and equilibrium chemistry 

model for multicomponent systems: Water Resources 

Research, v. 21, no. 8, p. 1095-1104.

Chapman, B.M., 1982, Numerical simulation of the

transport and speciation of nonconservative chemical 

reactants in rivers: Water Resources Research, v. 18, 

no. l,p. 155-167.

Engesgaard, P. and Kipp, K.L., 1992, A geochemical transport model for redox-controlled movement of mineral 

fronts in ground-water flow systems A case of nitrate 

removal by oxidation of pyrite: Water Resources 

Research, v. 28, no. 10, p. 2829-2843.

McKnight, D.M. and Bencala, K.E., 1989, Reactive iron 

transport in an acidic mountain stream in Summit 

County, Colorado A hydrologic perspective: 

Geochimica et Cosmochimica Acta, v. 53, no. 9, 

p. 2225-2234.

Rubin, J., 1983, Transport of reacting solutes in porous 

media Relation between mathematical nature of 

problem formulation and chemical nature of reactions: 

Water Resources Research, v. 19, no. 5, p. 1231-1252.

Runkel, R.L., and Broshears, R.E., 1991, One-dimensional 

transport with inflow and storage (OTIS) a solute 

transport model for small streams: Boulder, University 

of Colorado, Center for Advanced Decision Support 

for Water and Environmental Systems, Technical 

Report 91-01,85 p.

Runkel, R.L., 1993, Development and application of an 

equilibrium-based simulation model for reactive solute 

transport in small streams: Ph.D. Dissertation, 

University of Colorado, Boulder, Colorado, 202 p.

Runkel, R.L. and Chapra, S.C., 1993, An efficient numerical solution of the transient storage equations for solute 

transport in small streams: Water Resources Research, 

v. 29, no. 1, p. 211-215.

Somlyody, L., 1978, An effort for modelling the transport of 

micropollutants in rivers: IAHS AISH Publ., 125, 

p. 39-49.

Westall, J.C., Zachary, J.L., and Morel, F.M.M., 1976,

MINEQL: A computer program for the calculation of 

chemical equilibrium composition in aqueous systems, 

Cambridge, Water Qual. Lab., Dep. of Civ. Eng., Mass. 

Inst. of Technol., Tech. Note 18.

Yeh, G.T. and Tripathi, V.S., 1989, A critical evaluation of 

recent developments in hydrogeochemical transport 

models of reactive multichcmical components: Water 

Resources Research, v. 25, no. 1, p. 93-108.

Zand, S.M., Kennedy, V.C., Zellweger, G.W. and Avanzino, 

R.J., 1976, Solute transport and modeling of water 

quality in a small stream: Journal of Research. U.S. 

Geological Survey, v. 4, no. 2, p. 233-240.

780
Simulation of Reactive Transport During a pH 

Modification Experiment in a Mountain Stream Affected

by Acid Mine Drainage

By Robert E. Broshears, 1 Briant A. Kimball2 and Robert L. Runkel3

Abstract

A field experiment in a mountain stream in 

Colorado has helped to quantify hydrologic and 

chemical processes that control the transport of metals 

in water affected by acid mine drainage. Injection of 

sodium carbonate raised instream pH from 3.5 to a 

maximum of 5.8. Responses of aluminum and iron 

concentrations during the injection indicated the 

formation of aluminum and iron hydroxysulfate solid 

phases. Simulation of these responses using a reactive 

transport model reproduced the general aspects of 

observed patterns of concentration changes. 

Discrepancies between simulated and observed 

concentrations indicated the importance of chemical 

interactions between the water column and the 

streambed and kinetic restraints on the attainment of 

equilibrium.

INTRODUCTION

Diverse interactive physical, chemical, and 

biological processes affect the mobility and ultimate 

fate of metal constituents in surface water. The 

processes of advection, dispersion, 

precipitation/dissolution, sorption/desorption and 

oxidation/reduction contribute to the patterns of 

concentration observed in streams downgradient from 

contaminant sources such as acid mine drainage. 

While field measurements are necessary to

U.S. Geological Survey, Lakewood, Colo. 

2 U.S. Geological Survey, Salt Lake City, Utah 

3 U.S. Geological Survey, Denver, CO; University of Colorado, Center for Advanced Decision Support in Water and Environmental Systems (CADSWES)

characterize physical transport in streams, 

determination of chemical equilibria and rates of 

chemical and biological reactions is commonly 

relegated to the controlled conditions of the laboratory. 

An understanding of the natural world, however, 

requires that interactive processes be studied in the 

field, where the complexities of stream history, 

variable boundary conditions, and distributions of 

residence times in biogeochemical microenvironments 

combine in ways not duplicated in the laboratory.

This paper describes a field experiment 

conducted in Saint Kevin Gulch, a mountain stream 

receiving acid mine drainage near Leadville, Colorado 

(fig. 1). During the experiment, injection of sodium 

carbonate raised pH in the stream. Aluminum and iron 

concentrations in filtered samples decreased at higher 

pH values. These decreases in aluminum and iron 

were attributed to the precipitation of hydroxysulfate 

solid phases. In the case of iron, formation of these 

complexes apparently was favored by increased rates 

of oxidation at higher pH. The primary interpretive 

tool for the study was a reactive solute transport model 

called OTEQ (Runkel and others, 1996). OTEQ 

couples OTIS (Runkel and Broshears, 1991), an 

advection-dispersion model that includes transient 

storage, with MINTEQA2 (Allison and others, 1991), 

a model that calculates speciation under conditions of 

geochemical equilibrium.

EXPERIMENTAL METHODS

In August 1988, instream pH in Saint Kevin 

Gulch was increased step-wise from 3.5 to 5.8 by 

injecting a concentrated solution of sodium carbonate, 

with sodium chloride included as a conservative tracer. 

As the pulse of increased pH moved downstream, the

781
106-2T30"

39-1830-

iOB-21-30"

A

Study Area

COLORADO

LOCATION MAP

Mine tailings

05 

I

1 MILE

I 

05 l KILOMETER

Location of 

map B

Site\° x Injection point

-70 Site 2

/ ^-- 

24 Site 1

B

250 METERS

500 FEET

498

EXPLANATION

SAMPLING STAT1ONNumber indicates distance 

downstream from site 0, 

in meters

Figure 1. Map showing location of study area and sampling sites.

response of major ions and trace metals was 

documented by analyzing water samples collected at 

sites I through 4, located at 24, 70, 251, and 498 

meters, respectively, from the injection point (fig. I).

In describing the transport and fate of aluminum 

and iron, we distinguish two phases. The "filtered" 

phase remains in water passing through an 0.1 [im 

filter; the "particulate" phase is the concentration 

difference between an unfiltered sample and a filtered 

sample. The particulate phase may consist of solids 

that have aggregated into particles with a diameter

greater than the nominal pore size. The particulate 

phase may also include constituents that have sorbed 

to these particles. The "total" concentration of each 

element is the sum of filtered and particulate 

concentrations. Metal concentrations were determined 

by inductively coupled argon plasma atomic emission 

spectroscopy on samples acidified in the field to less 

than pH 2.0 with ultrapure nitric acid. Concentrations 

of sulfate and other anions were determined by ion 

chromatography.

782
PHYSICAL TRANSPORT

Sodium was geochemically conservative 

throughout the experiment and was used as a tracer in 

defining physical transport. Injection of sodium 

carbonate and sodium chloride caused a distinct 

concentration pulse delineating the injection period 

(fig. 2a) The sequential arrival of the sodium pulse at 

each sampling site permitted definition of subreach 

travel times, while the asymmetry of the pulse allowed 

for the calibration of model parameters describing 

transient storage (Bencala and Walters, 1983; Stream 

Solute Workshop, 1990). Instream flow increased 

from 12.3 L/s at site 1 to 13.2 L/s at sites 3 and 4. 

Transient storage parameters were comparable to 

those developed in other studies of mountain streams 

(Broshears and others, 1993). Aluminum and iron

were subject to the same regime of physical transport 

documented for sodium. By accounting for these 

processes in a rigorous manner, we were able to 

quantify the chemical and biological processes 

influencing aluminum and iron within the context of 

physical transport.

pH PERTURBATION

Prior to the experiment, background pH values 

were 3.4 to 3.5, with an increase to 3.8 at site 4. The 

injection began at 0900 hours and soon resulted in a 

pH of 4.2 at site 1 (fig. 2b). An increased injection 

rate begun at 1200 hours eventually resulted in a pH of 

5.8 at this site. After the injection ended, pH at site 1 

returned to near pre-injection levels. At the three

£ 30

w fi20

O*H

H

10

§ 0

Sitel

Symbols - Observed 

Lines - Simulated

8 12 16 20

TIME, IN HOURS

Figure 2a. Observed and simulated concentrations of sodium during experiment.

783
Sitel

D

4_

* Qu I6

Site 2

Site 3

Site 4

Observed 

Simulated

l_

8 12 16 

TIME, IN HOURS

Figure 2b. Observed and simulated pH during the experiment.

20

784
downstream sampling sites, the pH pulse was subdued 

relative to the upstream profile. These observations 

suggested reactions that liberated protons as the perturbation was transported through the reach. Plausible 

reactions that liberate protons include hydrolysis of 

aluminum and iron. The streambed may also react 

directly with carbonate species to buffer the change in 

pH.

RESPONSE OF ALUMINUM

At site 1, the background concentration of 

filtered aluminum was about 3.1 mg/L. As pH 

increased from 3.5 to 4.2, the aluminum concentration 

at this site did not change (fig. 3). When pH increased 

above 5.0, the concentration of filtered aluminum 

declined substantially, reaching 0.01 mg/L at pH 5.8. 

The decrease in filtered aluminum was accompanied 

by an increase in particulate aluminum to a 

concentration of 3.1 mg/L. Thus, the total 

concentration of aluminum in the water column 

remained more or less unchanged. These observations 

were consistent with the formation of a nonsettling 

aluminum precipitate during the injection. A slight 

decrease in sulfate concentration during the same 

interval indicated that this anion may have been 

incorporated into the aluminum precipitate. After the 

injection stopped, filtered aluminum concentration 

increased and particulate aluminum concentration 

decreased, both eventually returning to pre-injection 

values. However, for approximately 30 minutes after 

the injection stopped, filtered aluminum concentration 

was higher than the background level. This spike in 

aluminum concentration suggested the return of 

aluminum into the water column from a pH-dependent 

source on the streambed.

Relative to site 1, sites 2-4 exhibited a smaller 

decrease in filtered aluminum concentrations during 

the interval of higher pH (fig. 3), and less particulate 

aluminum was formed. There was a slight decrease 

through time in total aluminum relative to preinjection levels. The aluminum mass in the postinjection spike in filtered aluminum increased with 

downstream distance. These observations suggested 

the formation of a slowly settling aluminum 

precipitate, with decreasing rates of formation at the 

lower pH of the downstream sites. The spike in 

aluminum concentration likely was attributable to the 

dissolution of recently settled precipitate as the pH of 

the overlying water decreased following the injection.

RESPONSE OF IRON

Before the injection, background concentration 

of filtered iron declined gradually from 1.1 mg/L at 

site 1 to 0.75 mg/L at site 4 (fig. 4). Background 

particulate iron concentration was virtually unchanged 

along the reach. These observations were consistent 

with approximately equal rates of aggregation of ironbearing chemical precipitates and physical settling of 

particulate iron along the reach. When pH increased 

to 4.2 at site 1, there was a slight decline in filtered 

iron (to 1.0 mg/L) and a concomitant increase in 

particulate iron (to 0.1 mg/L). As pH increased to 5.8, 

filtered iron decreased to 0.4 mg/L and particulate iron 

increased to 0.7 mg/L. Thus, total iron concentration 

in the water column at site 1 was conservative during 

the injection period. These observations were 

consistent with formation of a nonsettling iron 

precipitate at higher pH.

The concentration of filtered iron decreased in 

the downstream direction throughout the experiment 

(fig. 4). The pulse of particulate iron also diminished 

with downstream distance. These observations were 

consistent with a high rate of pH-dependent 

production of particulate iron immediately 

downstream from the injection and subsequent settling 

of particulate iron during transport to sites 3 and 4.

GEOCHEMICAL REACTIONS AND 

REACTIVE TRANSPORT SIMULATIONS

Relations between pH and concentrations of 

filtered metals and sulfate during the experiment were 

evaluated by Kimball and others (1994). Mass balance 

analyses were consistent with the formation of 

aluminum and iron hydroxysulfate solids with the 

following stoichiometry:

and

Calculated instream ion activity products for the 

proposed aluminum and iron solids were 10"2-34 and 

10°'316, respectively.

For the OTEQ simulations, the carbonate 

boundary condition was adjusted to reproduce the 

observed pH at site 1; dynamic exchanges of carbon 

dioxide with the atmosphere were not modeled

785
Q- 

<«

s

O

 s

55 

ONM

H

^

H

O

u

4-

2-

0

8

6U 

4

2

0

8

D CB

Sitel

Site!

Site 3

_ Site 4

Observed 

Simulated

12 16 

TIME, IN HOURS

20

Figure 3. Observed and simulated aluminum concentration during the experiment.

786
W

H

N-l 

N-l

OM

H

W

^

O

O

a:

1.2

0.8

0.4

- Site 1

1.2

0.8

O 0.4

0.0 

1.2

0.8 

0.4

0.0

1.2

0.8 

0.4 

0.0

D

- Site 2

- Site 3

on

D D

n Observed 

 Simulated

D

Site 4

8 12 16 

TIME, IN HOURS

Figure 4. Observed and simulated iron concentration during the experiment.

20

787
directly. The ferric iron boundary condition was varied 

to reflect changes in relative rates of photoreduction 

(McKnight and others, 1988; McKnight and Bencala, 

1989) and pH-sensitive oxidation (Singer and Stumm, 

1970) during the experiment. Only the aluminum and 

iron hydroxysulfate solids described above were 

allowed to precipitate and settle from the water column.

At site 1, simulated and observed values for pH, 

filterable aluminum and filterable iron were in general 

agreement. Differences between simulated and 

observed values increased in the downstream 

direction, as shown in the plots for sites 2-4 (figs. 2-4). 

The increasingly poor correspondence between 

simulated and observed pH values with downstream 

distance may be attributable to the importance of 

geochemical interactions between the water column 

and the streambed. Static simulations with 

MINTEQA2 support the presence of a large reservoir 

of potentially exchangeable hydrogen ions on the 

streambed under ambient conditions; these hydrogen 

ions could have provided a substantial buffering 

capacity during the experiment. Failure to simulate 

pH profiles in part explained the unsatisfactory 

simulation of aluminum concentration at downstream 

sites. Discrepancies between simulated and observed 

iron concentration indicated that iron concentration 

may have been controlled by a solid phase that was 

less soluble than the iron hydroxysulfate solid 

described above. Uncertainty in patterns of iron redox 

chemistry throughout the day also contribute to 

difficulties in modeling the speciation of this element.

A second source of disagreement between 

observed and simulated concentrations involved the 

post-injection spike of aluminum. Because 

equilibrium chemistry was assumed, simulated 

aluminum that settled to the streambed during 

intervals of higher pH dissolved immediately upon 

restoration of lower pH values. Compared with 

simulated profiles, measured spikes were more 

dispersed in time, indicating a kinetic restraint on 

dissolution of the freshly settled aluminum solid.

We remain optimistic that the fate and transport 

of trace metals in surface waters can be characterized 

by a reactive transport model. Simulations to date 

have demonstrated the importance of streambed 

processes and continue to direct attention to areas of 

needed research. Further work is warranted on such 

modeling, including a more rigorous conceptualization 

of chemical interactions between the water column 

and bed sediments and the incorporation of kinetic 

restraints on the attainment of equilibrium.

REFERENCES CITED

Allison, J.D., Brown, D.S., and Novo-Gradac, K.J., 1991, 

MINTEQA2/PRODEFA2, a geochemical assessment 

model for environmental systems Version 3.0 user's 

manual: Washington, D.C., United States 

Environmental Protection Agency, 106 p.

Bencala, K.E. and Walters, R.A., 1983, Simulation of solute 

transport in a mountain pool-and-riffle stream A 

transient storage model: Water Resources Research, 

v. 19, no. 3, p. 718-724.

Broshears, R.E., Bencala, K.E., Kimball, B.A., and

McKnight, D.M., 1993, Tracer-dilution experiments 

and solute-transport simulations for a mountain stream, 

Saint Kevin Gulch, Colorado: U.S. Geological Survey 

Water-Resources Investigations Report 92-4081, 18 p.

Kimball, B.A., Broshears, R.E., McKnight, D.M., and 

Bencala, K.E., 1994, Effects of instream pH 

modification on transport of sulfide-oxidation 

products, in Alpers, C.N., and Blowes, D.W., eds., 

Environmental Geochemistry of Sulfide Oxidation, 

American Chemical Society, ACS Symposium Series 

550, p. 224-243.

McKnight, D.M. and Bencala, K.E., 1989, Reactive iron 

transport in an acidic mountain stream in Summit 

County, Colorado A hydrologic perspective: 

Geochimica et Cosmochimica Acta, v. 53, no. 9, 

p. 2225-2234.

McKnight, D.M., Kimball, B.A., and Bencala, K.E., 1988, 

Iron photoreduction and oxidation in an acidic 

mountain stream: Science, v. 240, p.637-640.

Runkel, R.L., Bencala, K.E., and Broshears, R.E., 1996, An 

equilibrium-based simulation model for reactive solute 

transport in small streams, in Morganwalp, D.W. and 

Aronson, D.A., ed., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Runkel, R.L., and Broshears, R.E., 1991, One-dimensional 

transport with inflow and storage (OTIS) a solute 

transport model for small streams: Boulder, University 

of Colorado, Center for Advanced Decision Support 

for Water and Environmental Systems, Technical 

Report 91-01,85 p.

Singer, PC., and Stumm, W, 1970, Acid mine drainage the 

rate determining step: Science, v. 167, p. 1121-1123.

Stream Solute Workshop, 1990, Concepts and methods for 

assessing solute dynamics in stream ecosystems: 

Journal of the North American Benthological Society, 

v. 9, no. 2, p. 95-119.

788
Difficulties with Intermediate-Scale Experiments

for Studies of Iron Chemistry in Streams

Affected by Acidic Mine Drainage

By Diane M. McKnight 1 , Kenneth E. Bencala2 , Richard A. Harnish1 , and Robert L. Runkel 1

Abstract

Iron chemistry in streams affected by acidic 

mine drainage is driven by many dynamic 

processes, such as precipitation, dissolution, 

photoreduction of ferric iron and microbial 

oxidation of ferrous iron. These processes can 

be studied with a variety of approaches, ranging 

from controlled laboratory experiments to field, 

stream-scale perturbation experiments. Experiments that are intermediate between laboratory 

and field experiments using sediments and water 

from the Snake River have been conducted. These 

experiments have been useful in demonstrating the 

general nature of iron biogeochemistry. However, 

difficulties reproducing results occur both within 

and among experiments. The heterogeneous 

nature of the stream sediments and seasonal 

variations in trace phases of iron oxides confound 

interpretation of the results and limit the predictive 

value of such experiments.

INTRODUCTION

Numerous alpine and subalpine streams in the 

Rocky Mountains of the central United States are 

enriched in trace metals from natural weathering of 

pyrite and (or) from acidic mine drainage. In Colorado, 

an estimated 600 km of stream reaches are contaminated 

by acidic mine drainage, and much of this contamination 

originates from abandoned mines located on remote 

mountain sides (Moran and Wentz, 1974). In these 

streams, the rocks of the streambed are often covered 

with abundant hydrous iron oxides; accordingly, iron 

chemistry is important in understanding the geochemistry and ecology of these acidic mountain streams.

'U.S. Geological Survey, Boulder, Colo. 

2U.S. Geological Survey, Menlo Park, Calif.

Iron chemistry in acidic mountain streams is controlled by diverse reactions and is dynamic over time 

(McKnight and Bencala, 1989). Precipitation/dissolution, oxidation/reduction, complexation, and sorption 

are all important reactions. Reaction rates vary during 

the day because photoreduction of dissolved ferric iron 

and iron oxides drives variation in ferric iron precipitation. The oxidation of ferrous iron is microbially mediated at the low pH of these streams; oxidation can vary 

during the day because of temperature changes, or it can 

be restricted to the water/sediment interface. Also, concentrations of dissolved humic substances and important 

complexing ligands typically increase and iron concentrations become diluted during spring snowmelt.

The dominant phases of iron oxides are amorphous 

rather than crystalline, can have solubility products 

ranging over several orders of magnitude, and are more 

surface reactive than crystalline phases. Uncertainty in 

the thermodynamic constants associated with amorphous iron oxides precludes definitive predictions from 

chemical-equilibrium calculations alone.

For these reasons, geochemical characterization of 

actual streams is required to understand changes in 

their chemistry under varying conditions. In this paper, 

we examine the results of experiments that use 

streambed material to characterize microbial oxidation 

of ferrous iron and dissolution of hydrous iron oxides 

with decreasing pH. These experiments are intermediate in scale between controlled laboratory studies and ' 

stream-scale experiments.

SITE DESCRIPTION AND METHODS

The experiments were conducted using streambed 

material and water from the Snake River, a small 

mountain stream in Summit County, Colo. (fig. 1). The 

Snake River is acidic (pH 3.2-4.5) and metal enriched 

from the weathering of disseminated pyrite in schist 

and gneiss in the eastern side of the watershed 

(Theobald and others, 1963). This stream has iron and

789
aluminum concentrations comparable to streams 

affected by acidic mine drainage in the Colorado 

Rocky Mountains (McKnight and Bencala, 1990). 

The streambed is composed of cobbles coated with 

iron oxides, and some of the rocks contain pyrite 

(McKnight and Bencala, 1989). Hydrologic experiments using inorganic tracers have shown that flow 

through the hyporheic (substream) zone is continuously exchanging with flow in the open channel 

(Bencala and others, 1990).

Most of the experiments were conducted using 

streambed material and water collected from the SN2 

site on the Snake River (fig. 1), which is located just 

above the confluence of the stream with Deer Creek. In 

the 1992 experiment, water and streambed material were 

also collected from the SN1 site about 2 km upstream 

from the confluence. Streambed material and water 

were collected before sunrise in 1985 and at midday in 

1992 in plastic buckets and stored in the dark. Table 1 

summarizes the experiments that were conducted.

105-5?

 3700 TOPOGRAPHIC CONTOUR. 

CONTOUR INTERVAL 375 

METERS. NATIONAL 

GEODETIC VERTICAL 

DATUM OF 1929.

Base from U S Geological Survey 

Montezuma. I 24.000,1958

Figure 1. The Snake River and Deer Creek watersheds, showing the area containing a bog 

iron-ore deposit.

790
Table 1. Summary of intermediate-scale experiments 

conducted in the Snake River.

Process

Microbial oxidation of 

ferrous iron

Dates

July 1 985 

August 1 992

Experimental Design

Observe decrease in 

ferrous iron in the dark.

Dissolution of iron oxide November 1984 

with a pH decrease May 1985

November 1985

Observe increase in iron 

concentration with serial 

exposures to low-pH water.

For the oxidation experiments, rocks of variable 

size were placed on metal racks in 4-L glass beakers 

filled with stream water, which was stirred with a magnetic stirrer rotating below the rack. In the 1985 experiment, two beakers were first exposed to sunlight and 

then placed in the dark, and two were left in the dark. 

In the 1992 experiment, the beakers were filled with 

stream water collected at midday and then placed in the 

dark. The samples collected from the beakers were 

immediately filtered using a 0.4-jam Nuclepore filters 

and an Antlia filtration unit. The total dissolved iron 

and ferrous iron concentrations were determined by the 

2,2 bipyridine colorimetric method (McKnight and 

Bencala, 1988).

The dissolution experiments were conducted on 

November 8, 1984; May 5, 1985; and November 7, 

1985. The rocks and fine sediments were sequentially 

leached with stream water acidified with I^SC^ (pH

2.5 to ambient). The experiments were conducted 

indoors in duplicate; no incident sunlight was allowed 

to fall on the buckets. One liter of acidified stream 

water was placed with rocks and fine sediment in a

26 cm diameter plastic bucket and agitated for 5 minutes. The stream water was then poured from the 

bucket and replaced with another 1 L of stream water; 

aliquots were immediately filtered using GFC glassfiber filters and an Antlia filtration unit. During the 

experiments, pH and Eh were measured with an Orion 

Model 701 pH meter and an Orion 967800 combination redox electrode. The samples were analyzed for 

iron by inductively coupled plasma spectrometry, and 

iron redox states were measured colorimetrically during the 1985 experiments.

RESULTS AND DISCUSSION

The design of the oxidation experiments was based 

on the premise that the bacterial populations that oxidize the ferrous iron in the Snake River reside in the

epilithon~the layer of algae and bacteria that covers 

the streambed substrate. Because of the importance of 

hyporheic flow in these mountain streams, there is sufficient contact of water with the microbial populations 

in the streambed to accomplish substantial decreases in 

ferrous iron concentration over a period of several 

hours of darkness when there is no longer a source of 

ferrous iron from photoreduction.

The results of the two oxidation-rate experiments 

are shown in figures 2 and 3. In the 1985 experiment, a 

progressive decrease in ferrous iron concentration 

occurred when the sunlight-exposed beakers were 

placed in the dark. The initial iron concentrations were 

different in the two beakers that were maintained in the 

dark from the time of sample collection before sunrise. 

In the beaker that had the lower initial iron concentration, no decrease in ferrous iron occurred. However, a 

decrease occurred in the beaker with the higher concentration. The data showing a decrease in ferrous iron 

once the light-exposed beakers were placed in the dark 

was used to calculate an oxidation rate. This rate 

agreed reasonably well with the rate calculated from 

the decrease in ferrous iron in the stream itself at night 

(McKnight and Bencala, 1988).

In the 1992 experiments, no significant decrease of 

ferrous or total iron occurred during the experiments 

with water and rocks from either the main site or the 

upstream site. The results from duplicate beakers were 

very similar. The main difference in the design of the 

1985 and 1992 experiments was that, in 1985, two of 

the beakers were initially exposed to sunlight, whereas 

in the 1992 experiment the water and rocks were taken 

directly from the stream and placed in the dark. A possible reason why this difference should account for the 

failure to observe oxidation of ferrous iron in the 1992 

experiment is that there had been a longer period for 

populations of ferrous oxidizers to adjust to the experimental conditions in the 1985 study.

The most straightforward explanation for the difference in results of the experiments is that, in the 1992 

experiment, the number of active ferrous oxidizing 

bacteria transferred with the streambed rocks was not 

adequate to cause a measurable decrease in ferrous iron 

during the 4-hour experiment. The rate of ferrous iron 

oxidation is only first order with respect to the ferrous 

iron concentration if there is a sufficient microbial population. If the beakers had been incubated for more 

than 3 hours, a microbial population might have developed to oxidize the ferrous iron.

791
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 DARKNESS

Figure 2. Results of an oxidation rate experiment in the light and in the dark 

conducted in 1985 (modified from McKnight and Bencala, 1988).

792
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The results of the three dissolution experiments are 

presented in figure 4. In all three experiments, the iron 

concentrations were greatest at the lowest pH value, as 

would be predicted from chemical-equilibrium calculations. The results of these experiments also were generally consistent in showing that the dissolution of iron 

oxides with a decrease in pH was very rapid, occurring 

within the first 5-minute batch exposure. This rapid 

response is the same as was observed in an instream 

experiment in which the pH of the Snake River was 

lowered from 4.2 to 3.2 for a 3-hour period (McKnight 

and Bencala, 1989). In the stream experiment, photoreduction was an important process that presumably did 

not occur in these experiments. As in the oxidation 

experiments, the duplicate samples of streambed material gave similar results in the iron oxide dissolution.

In the detailed response, however, some significant 

differences between results of the experiments were 

observed, including some peculiar aspects of the iron 

speciation. One of the experiments done at the end of 

summer (November 8,1984) took place when the hydrologic conditions were similar to those existing during the 

instream pH-perturbation experiment (McKnight and 

Bencala, 1989) and the pH was 3.92. In this experiment, 

an initial peak in the iron concentration resulted, which 

closely agreed with the instream experiment results, 

indicating the presence of a trace soluble-oxide phase. 

When the batch experiment was repeated at the same 

time the next year (November 7, 1985), the pH of the 

stream water was lower (pH 3.5) and the initial peak did 

not appear. The reason for this is probably that the soluble-trace phase represented in the initial peak evidenced 

in the instream experiment and the 1984 batch experiment would have already dissolved at this low pH.

The batch experiment was also conducted at the 

beginning of snowmelt (May 5, 1985), when most of 

the upstream reach of the Snake River was covered 

with ice and snow. The initial peak did not occur in this 

experiment either. The presence of the trace solubleoxide phase may be related to the diel cycle of redox 

and precipitation/dissolution processes. Such a trace 

soluble-oxide phase may not be present in detectable 

quantities following extended periods under snow and 

ice without exposure to UV-radiation to drive the photochemical component of the diel cycle observed during 

periods of open water. Another difference between the 

spring batch experiments and the second fall batch 

experiment was that, in the spring, most of the released 

iron was ferrous iron, whereas in the fall, ferric iron

was the predominant form and the ferrous iron concentration did not vary much with changes in pH. This 

result is intriguing, in that a reductant that is independent of sunlight is necessary to explain the ferrous iron 

in the spring experiment. One possible source of ferrous iron in this experimental system may be oxidation 

of pyrite by ferric iron released by dissolution of the 

iron oxide. The abiotic oxidation of pyrite by ferric 

iron is rapid. Trace amounts of pyrite have been found 

in streambed material from this site (McKnight and 

Bencala, 1988).

The results of the batch experiments consistently 

indicate an increase in dissolved iron with a decrease in 

pH. However, the results emphasize that detailed 

responses to a change in stream pH may vary seasonally, possibly because of variations in stream chemistry 

and in iron-containing solid phases in the streambed. 

These variations may be only weakly related to predictable seasonal changes, such as the development of 

an ice cover and decreases in pH at low flow. In order 

to predict reactive transport by means of a solutetransport model, the model would have to include 

important iron reactions and would have to accommodate changes in initial conditions, such as variability in 

concentrations and characteristics of solid phases. In 

addition to spatial variability, seasonal variability may 

constrain the predictive application of solute transport 

models.

CONCLUSIONS

Possible explanations have been presented that 

provide some insight into iron dynamics in the Snake 

River. However, the results of these experiments have 

the common feature of not being definitive. The variation in results of the different experiments was substantial and, although possibly attributable to variations in 

characteristics of streambed material, it would have 

been difficult to make the independent measurements 

of the characteristics that may have controlled the 

response, such as the abundance of a trace soluble 

phase of iron oxide or of viable ferrous oxidizing bacteria. Therefore, the overall conclusion is that the transition from the intermediate-scale to the stream-scale 

is, in some ways, more problematic than the transition 

from carefully controlled laboratory experiments to the 

stream scale.

794
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Figure 4. Results of three dissolution experiments.

795
REFERENCES

Bencala, K.E., McKnight, D.M., and Zellweger, G.W., 

1990, Characterization of transport in an acidic and 

metal-rich mountain stream based on a lithium tracer 

injection and simulations of transient storage: Water 

Resources Research, v. 26, no. 5, p. 989-1000.

McKnight, D.M., and Bencala, K.E., 1988, Diel variations 

in iron chemistry in an acidic stream in the Colorado 

Rocky Mountains, U.S.A.: Arctic and Alpine Research, 

v. 20, no. 4, p. 492-500.

McKnight, D.M., and Bencala, K.E., 1989, Reactive iron 

transport in an acidic mountain stream in Summit 

County, Colorado A hydrologic perspective: Geochimica et Cosmochimica Acta, v. 53, p. 2225-2234.

McKnight, D.M., and Bencala, K.E., 1990, The chemistry 

of iron, aluminum, and dissolved organic material in 

three acidic, metal-enriched, mountain streams, as controlled by watershed and in-stream processes: Water 

Resources Research, v. 26, no. 12, p. 3087-3100.

Moran, R.E., and Wentz, D.A., 1974, Effects of acid-mine 

drainage on water quality in selected areas in Colorado, 

1972-73: Denver, Colorado Water Pollution Control 

Commission, Colorado Water Research Circular 25.

Theobald, P.K., Lakin, H.W., and Hawkins, D.B., 1963, The 

precipitation of aluminum, iron and manganese at the 

junction of Deer Creek with the Snake River in Summit 

County, Colorado: Geochimica et Cosmochimica Acta, 

v. 27, 121-132.

796
Metal Concentrations in Sedges in a Wetland Receiving

Acidic Mine Drainage from St. Kevin Gulch,

Leadville, Colorado

ByB.M. Erickson, 1 P.M. Briggs, 1 andT.R. Peacock1

Abstract

As of the 1989 growing season, there were no 

visible signs of metal stress on the wetland vegetation receiving the acid mine drainage (AMD) from 

of St. Kevin Gulch, Colorado. However, the AMD 

has affected the metal concentrations of the sedges 

growing on the wetland. Sedge concentrations of 

cadmium (Cd), copper (Cu), iron (Fe), lead (Pb), 

manganese (Mn), and zinc (Zn), known to be high 

in the AMD waters, are higher in the area of direct 

inflow than the sedge baseline geometric means 

(GM) reported for the wetland. Concentrations 

higher than the GM in other areas of the wetland 

are evidence for other sources of contamination to 

the wetland. Maximum cattle dietary levels 

reported in the literature for Cd (0.5 milligrams per 

kilogram (mg/kg)) and Zn (500 mg/kg) are 

exceeded in some areas of the wetland and could 

present a nutritional problem for the cattle grazing 

this wetland each year. Vegetation toxicity levels 

for Cu (20-100 mg/kg) and Pb (30-300 mg/kg) are 

not exceeded, but Pb concentrations are higher in 

the inflow area and may indicate accumulation is 

occurring. Fe deficiency, common in vegetation 

and foodstuffs, was not detected in the sedges of 

this wetland. Mn concentrations above the 500 

mg/kg level reported in the literature as resulting 

in signs of visible stress in vegetation are exceeded 

in most areas of the wetland. The Fe/Mn ratio 

necessary for healthy vegetation (1.5-2.5), occurs

U.S. Geological Survey, Denver, Colo.

only in the areas with the higher Fe concentrations. 

Presently the only element being removed from 

the AMD waters by the wetlands is Fe.

If the sedges accumulated higher levels of Fe, 

the ratio of Fe/Mn would not necessarily be 

improved. Fe and Mn are physiological antagonists, 

and at toxic Mn levels, the absorption of Fe process 

could reverse resulting in more stress and eventual 

loss of biomass.

INTRODUCTION

Acid mine drainage from St. Kevin Gulch, Leadville, Colo., drains into a 26-ha subalpine wetland. The 

wetland is a sedge meadow with two ecosystem components: (1) sedges growing on top of hummocks the drier 

component (Carex canescens L.); and (2) sedges growing between or off hummocks the aquatic component 

(Carex utriculata Boott.^. The purpose of this study was 

to evaluate the affects of the AMD on the wetland 

sedges. To accomplish this purpose, two main tasks were 

identified. The first task was to establish the expected 

baseline trace-element concentrations in the sedges to 

delineate areas of possible effects. The second task was 

to examine seasonal variation and establish biomass production estimates to determine the amount of metals 

being removed from the AMD by the wetland. Baseline 

trace-element concentrations in Carex canescens and 

Carex utriculata can be found in Erickson and others 

(1991a;1991b). Seasonal results for all trace-elements 

and estimates of biomass production are currently being 

compiled. This paper presents Carex utriculata baseline 

and seasonal concentrations of Cd, Cu, Fe, Pb, Mn, and 

Zn. Data for Carex canescens are not included because 

the data resulting from both tasks indicated little statistical difference in metal concentrations between the two 

species.

797
STUDY PROCEDURES

The wetland was originally divided into 19 

sampling cells, 122-m on a side, for the baseline task 

(fig. 1). Because the wetland is leased for cattle grazing, and samples of total growth were needed for the 

seasonal and biomass production estimates tasks, field

exclosures were constructed in each of the 19 wetland 

cells in 1989 (fig. 1). Steel post and barbed-wire exclosures measuring approximately 1.8 m on a side were 

constructed. In the center of each exclosure, a 1-m 

diameter circle of flexible PVC tubing was secured 

with landscape pins to define the sampling area. 

Spring, summer, and fall samples were collected by

EXPLANATION 

Baseline sampling cells

Acid mine-drainage inflow area

Elevated element concentation areas

from baseline task 

Poorly channelized flow

Seasonal sampling exclosures

North 

Spring C

LEADVILLE 6 MILES

Figure 1. Location of baseline sampling cells and seasonal sampling exclosures on the 

wetland receiving acid mine drainage waters of St. Kevin Gulch, Colo.

798
clipping the above-ground part of the two species of 

sedges from a different quarter of the circle for each 

season, bagged separately, and weighed in the laboratory.

In the laboratory, the sedge samples were washed in 

distilled water and dried in a forced-air oven at approximately 50 °C (degrees Celsius) for 1 week. A small set 

of sample splits were taken for estimates of variance due 

to preparation technique and (or) analytical precision. 

The samples were then pulverized and a small split of 

the raw material saved for atomic absorption analyses. 

The remaining sample material was ashed in a muffle 

furnace at 450 °C. All samples were weighed before and 

after ashing for percent-ash calculations. All samples 

were processed and analyzed in a random sequence to 

eliminate any systematic biases. Trace-element concentrations presented in this paper were determined on the 

ash by a 40-element inductively coupled plasma-atomic 

emission spectroscopy (ICP-AES) method developed by 

Crock and others (1983). Before statistical processing, 

all qualified values were replaced with a real value equal 

to 0.7 times the lower limit of detection according to 

methods described by Miesch (1976). All results were 

converted to dry-matter equivalents and converted to 

logarithms prior to statistical analysis. Summary statistics (tables 1 and 2) are reported as the geometric mean 

(GM), geometric deviation (GD), and baseline or 

expected range.

RESULTS

Baseline and Seasonal Comparisons

Conclusions reported in Erickson and others 

(1991b) for the baseline study are: (1) the greatest 

amount of variance in the trace-element concentrations 

between the sedges can be explained as the geographic 

differences between the 19 cells, (2) sedge samples, 

with trace-element concentrations above the baseline 

GM values, cluster at the inflow area of the St. Kevin 

Gulch AMD (fig. 1), indicating the vegetation has been 

affected, and (3) clusters of sedge samples with traceelement concentrations above the GM and occurring in 

other areas of the wetland are evidence that St. Kevin 

Gulch AMD is not the only source of contamination to 

the wetland (fig. 1).

Samples for the baseline task were collected in the 

fall of 1988. A comparison of the summary statistics for 

Cd, Cu, Fe, Pb, Mn, and Zn from the baseline task 

(table 1) to the fall seasonal results (table 2) shows that 

the baseline or expected range of concentrations for Cd, 

Mn, and Pb, are similar. The differences in the concentrations of Cu, Fe, and Zn between the two tasks are

Table 1. Baseline summary statistics for metal 

concentrations in Carex utriculata collected within the 19 cells 

in the Leadville, Colo., wetland

(From Erickson and others (1991 a). Concentrations are in milligrams 

per kilogram, dry-matter basis)

Element Detection Geometric Geometric Baseline or 

ratio mean deviation expected range

Cd

Cu

Fe

Mn

Pb

Zn

38:94

94:94

94:94

94:94

58:94

94:94

0.34

6.20

270

320

0.96

1 70

2.00

1.70

1.90

1.70

2.40

2.30

0.09 - 1 .3

2.I - 18

78 - 930

120 -860

0.17 -5.5

32 -860

Table 2. Seasonal summary statistics for metal concentrations in Carex utriculata, collected within the 19 exclosures 

constructed on the Leadville, Colo., wetland 

[Concentrations are in milligrams per kilogram, dry-matter basis]

Element Detection Geometric Geometric Baseline or 

ratio mean deviation expected range

Spring

Cd

Cu

Fe

Mn

Pb

Zn

13:19

19:19

19:19

19:19

11:19

19:19

0.49

17

350

420

0.68

160

2.80

1.70

2.10

1.70

2.40

2.20

0.06 - 3.8

3.9 -35

88 - I400

I44- I200

O.I2 -3.3

33 - 780

Summer

Cd

Cu

Fe

Mn

Pb

Zn

Cd

Cu

Fe

Mn

Pb

Zn

5:19

18:19

18:19

18:19

7:19

18:19

8:19

19:19

19:19

19:19

19:19

19:19

0.24

6.6

1 70

260

0.54

IOO

Fall

0.39

4.9

880

420

1. 90

1 80

1.90

1.50

1.70

1.80

2.00

2.30

2.40

1.70

1.70

2.00

2.10

2.30

O.I 3 -0.46

4.3 - 10

99 -300

1 45 -480

0.27 - I.IO

44-230

O.I 6 -0.94

2.9 - 8.2

520- 1 500

220 - 820

0.90 -4. 1

76 - 400

probably due to the geographical difference between 

sampling sites. The baseline cell samples were collected 

from a 122-m2 area, whereas the seasonal exclosure 

samples were collected within a 1-m-diameter circle.

The flow from St. Kevin spreads out in the areas of 

cells 9,10,8,11,12,13, and 14 (fig. 1, inflow area), and 

Cd, Cu, Fe, Mn, Pb, and Zn concentrations in the baseline samples were higher than the wetland GM in this 

area. Seasonal sample data graphed in figures 2,3, and 4 

also illustrate increased element concentrations in the 

inflow area of the wetland. Higher concentrations on 

the graphs, in areas other than the inflow, tend to occur 

in the same cell areas as the baseline higher concentration clusters of figure 1.

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Seasonal Composition

Loss of biomass usually occurs before visible signs 

of metal stress appear in vegetation (Kabata-Pendias and 

Pendias, 1984). Within the boundaries of the wetland 

during the baseline (1988) and seasonal (1989) tasks, 

neither condition existed. However, baseline results 

indicate that Cd and Zn concentrations in sedge samples 

from the areas of inflow of St. Kevin AMD waters (fig. 

1), South Spring, and outflow exceed the recommended 

cattle dietary levels (0.5 and 500 mg/kg, respectively) 

reported by Dollhopf and others (1990). Results for the 

Cd seasonal data (fig. 2) also indicate that cattle dietary 

levels are exceeded in the spring sedge samples collected in the exclosures located within the baseline wetland cells 10,11, 13, 15, 16, 17, and 18. Summer sedge 

samples exceed the levels in exclosure samples from 10, 

11, 15, 17, and 18. During the fall, Cd levels are 

exceeded in the sedge samples from the exclosures in 

cells 4, 10, 13, 14, and 17. With the exception of cell 4, 

all Cd concentrations exceeding the recommended levels for cattle occur in areas receiving St. Kevin AMD. 

The higher Cd concentrations in the exclosure samples 

from cell 4 are probably affected by the mine tailings 

located just north of the wetland (Walton-Day and Erickson, 1990). Fall concentrations of Zn exceed cattle 

dietary levels in the sedges from the exclosures in cells 

4,8,10,14, and 17. Spring exclosure samples exceed the 

Zn levels in cells 10 and 15. Zinc concentrations in the 

summer sedges are all below the 500 mg/kg level.

Kabata-Pendias and Pendias (1984), reported that 

normal levels of concentrations for Cu in vegetation 

are 5 to 30 mg/kg, deficient levels 2 to 5 mg/kg, and 

toxic levels 20 to 100 mg/kg, depending on species and 

other environmental parameters. All Cu concentrations 

in the seasonal sedge samples from the Leadville wetland were below toxic levels except the spring samples 

from exclosures in cells 10 and 18 (fig. 2). Fall concentrations of Cu appear to be in the deficient range, but 

there is a continuous decrease in concentration 

throughout the growing season, and Cu is known to be 

very mobile and to translocate back to the roots in the 

fall. This decease in Cu concentration over the growing 

season was not observed in Carex aquatilis at Peru 

Creek, Colo., by Cooper and Emerick (1987). Differences in species and (or) substrate are two possible 

explanations.

Pb concentrations increase throughout the growing 

season, except in the samples from exclosures 10 and 

18 (fig. 3). Emerick and others (1987), did not find an 

accumulation of Pb in sedges in the Peru Creek wetland 

study, but the comparatively high concentrations in

cells 4, 10, 14, and 17 indicate that accumulation may 

be occurring in the area of AMD inflow on this wetland. The Pb concentrations in the sedges in this wetland are still below the 30-300 mg/kg toxic level for 

vegetation reported by Kabata-Pendias and Pendias 

(1984).

Vegetation uptake and transport of Fe do not occur 

readily; deficiency is generally the greater problem. 

Nutritional needs of cattle range from 50 to 100 mg/kg, 

and fodder concentrations can range from 18 to 1,000 

mg/kg. Concentrations of Fe in figure 3 indicate deficiency is not a problem in this wetland. Fe toxicity in 

vegetation can occur, and when it does, it is usually in 

vegetation growing on strong acid soils, acid sulfate 

soils, or flooded soils (Kabata-Pendias and Pendias, 

1989). Results from the study of the inflow and outflow 

water chemistry for this wetland indicate that Fe is presently the only metal being retained in the wetland (Walton-Day, 1991). Fe toxicity could occur in the future.

Mn is taken up and translocated rapidly in vegetation and is most mobile on acid and flooded soils. 

Normal concentrations for vegetation in general are 

between 30 and 300 mg/kg, but concentrations as high 

as 1,840 mg/kg in forage have had no adverse affects 

on cattle. Deficiencies can occur below 25 mg/kg and 

toxic affects on vegetation are often visible at concentrations over 500 mg/kg. Mn concentrations in the 

sedges from this wetland exceed the toxic threshold in 

all exclosures except 2, 3, 4, 11, and 17 (fig. 4) during 

some part of the growing season.

The Fe/Mn ratio necessary for healthy vegetation 

should range between 1.5 and 2.5 (Kabata-Pendias and 

Pendias, 1989). The only spring sample with a ratio 

between 1.5 and 2.5 was collected from the exclosure 

in cell 17. Summer samples from exclosures 1 and 3, 

and fall samples from exclosures 1,6,8, 13, 14, 16, and 

17, had healthy Fe/Mn ratios. The greater number of 

samples with a healthy Fe/Mn ratio occurring in the 

fall, is probably related to the higher fall Fe concentrations, as shown in figure 4. Because Fe is the only 

element accumulating in the wetland, the wetland 

Fe/Mn ratio might be expected to improve in the future. 

However, further retention of Fe by the wetland could 

also increase the imbalance because Fe and Mn are 

physiological antagonists (Kabata-Pendias and Pendias, 1989). Deficient and normal Mn concentrations 

can both antagonize Fe absorption, and toxic levels can 

reverse the process. Because the 500 mg/kg Mn toxic 

level capable of causing visible stress in vegetation is 

already exceeded in the wetland sedges, loss of biomass could occur in the future.

803
SUMMARY

St. Kevin Gulch AMD has affected the vegetation 

of the wetland, but there were no visible signs of stress 

or loss of biomass during the 1988 and 1989 growing 

seasons. Baseline and seasonal task concentrations of 

Cd, Cu, Fe, Mn, Pb, and Zn in Carex utriculata were 

elevated above the GM in the area of inflow to the wetland. Results from the seasonal task also support the 

baseline task conclusion that St. Kevin is not the only 

source of contamination to the wetland. Cd and Zn concentrations exceed the recommended dietary levels for 

cattle in the areas of inflow of AMD and the outflow for 

the wetland. Cu and Pb do not present a problem at this 

time, but Pb might be accumulating in the area of 

inflow to the wetland. Fe deficiency in vegetation is not 

a problem in this wetland. The level of Mn that usually 

results in visible signs of stress is exceeded in most 

areas of the wetland, and the Fe/Mn ratio necessary for 

healthy vegetation occurs only in areas with higher Fe 

concentrations. Fe is the only element being removed 

from the AMD by the wetland, but continued Fe accumulation will not necessarily improve the Fe/Mn ratio 

in vegetation. Fe and Mn are physiological antagonists 

and at toxic Mn levels, the absorption of Fe process 

could reverse, resulting in more stress and eventual loss 

of biomass.

REFERENCES CITED

Crock, J.G., Lichte, F.E., and Briggs, P.E., 1983, Determination of elements in National Bureau of Standards Geologic Reference Materials SRM 278 obsidian and SRM 

688 basalt by inductively coupled argon plasma-atomic 

emission spectrometry: Geostandards Newsletter v. 7, 

p. 335-340.

Cooper, D.J., and Emerick, J.C., 1987, The effects of acid 

mine drainage on a Carex aquatilis fen., in Proceedings 

of 8th Annual Meeting of the Society of Wetland Scientists, Seattle, Washington, May 26-29, 1987: p. 96-100.

Dollhopf, D.J., Goering, J.D., Rennick, R.B., Morton, R.B., 

Zavitz, T.L., and Munshower, F.F., 1990, Performance 

of a natural wetland on control of acid mine drainage 

for 50 years, in Proceedings of 5th Billings Symposium

on Disturbed Land Rehabilitation 11, March 25-30, 

1990: Reclamation Research Unit Publication 

#009003, p. 228-238.

Emerick, J.C., Cooper, D.J., Huskie, W.W., and Lewis, W.S, 

1987, Documentation and analysis of the effects of 

diverted mine water on a wetland ecosystem, and 

construction of a computerized data base on acid mine 

drainage in Colorado: Colorado School of Mines, 

Department of Environmental Science and Engineering, Report submitted to Colorado Mined Land Reclamation Division, Colorado Department of Natural 

Resources, Contract #C-188078: 95 p.

Erickson, B.M., Briggs, P.H., Kennedy, K.R., and Peacock, 

T.R., 199la, Analytical results for sedge samples collected on the wetland receiving acid mine drainage 

waters from St. Kevin Gulch, Leadville, Colorado: U.S. 

Geological Survey Open-File Report 91-126, 83 p.

Erickson, B.M., Briggs, P.H., Kennedy, K.R., and Peacock, 

T.R., 1991b, Chemical composition of the wetland 

vegetation receiving the acid mine drainage waters of 

St. Kevin Gulch, Leadville, Colorado, in Mallard, G.E., 

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigation Report 91 -4034, p. 419-421.

Kabata-Pendias, Alina, and Pendias, Henryk, 1984, Trace 

elements in soils and plants: Boca Raton, Fla., CRC 

Press, 315 p.

Miesch, A.T., 1976, Geochemical survey of MissouriMethods of sampling, laboratory analysis, and statistical reduction of data: U.S. Geological Survey Professional Paper 954-A, 39 p.

Walton-Day, Katherine, 1991, Hydrology and geochemistry 

of a natural wetland affected by acid mine drainage 

St. Kevin Gulch, Lake County, Colorado: Colorado 

School of Mines, Department of Geology and Geological Engineering, Ph.D. thesis T-4033. 299 p.

Walton-Day, Katherine, and Erickson, B.M., 1990, Some 

effects of acid-mine drainage on a subalpine wetland, 

Tennessee Park, Colorado, in Gough, L.P., ed, Environmental Forum, 1990: Geoscience investigations that 

emphasize chemical, physical, and biological ecosystem processes: U.S. Geological Survey Open-File 

Report 90-288, p. 12.

804
Uraniferous Waters of Southeastern Colorado: A 

Function of Geology, Climate and Land Use

By Robert A. Zielinski 1 anc/Sigrid Asher-Bolinder1

Abstract

The effect of local geology and land-use 

practices on dissolved uranium in water is 

indicated by reconnaissance sampling of surface 

waters and some springs along the Arkansas 

River valley of southeastern Colorado. The rate 

of increase of uranium concentration with 

distance downriver increases markedly as the 

river flows from predominantly undeveloped 

lands underlain by igneous and metamorphic 

rocks to agriculturally developed lands underlain 

by marine shale and limestone. An additional 

abrupt increase in dissolved uranium is observed 

along the section of river where discharge is 

greatly reduced because of extensive diversions 

for irrigation and where remaining flow is largely 

composed of irrigation return water. Water 

collected along this last section of river is 

anomalously enriched in dissolved uranium 

compared to the average concentration reported in 

water from outside of the irrigated areas. As is 

commonly observed, dissolved uranium shows 

particularly strong positive correlation with 

specific conductance. Uranium also shows strong 

positive correlations with sodium, magnesium, 

sulfate and chloride that contribute heavily to 

total dissolved solids and specific conductance in 

these waters. Lithium, boron, strontium and 

selenium are also positively correlated with 

uranium. The combination of natural weathering 

of uranium-bearing soils and sediments, extensive 

soil leaching by irrigation return waters, and 

evaporative concentration in a semiarid climate 

produce concentrations of dissolved uranium that 

may threaten local water supplies.

'lI.S Geological Survey, Denver, CO

INTRODUCTION

The U.S. Environmental Protection Agency 

(USEPA) is currently (1993) proposing the addition of 

a new primary drinking-water regulation for uranium 

(U) of 20 (ig/L, with implementation scheduled to 

begin in 1996. If accepted, this standard may be 

exceeded in approximately 1,500 public water systems 

(United States Environmental Protection Agency Fact 

Sheet, 1991) and in untold numbers of unregulated 

private water supplies. The addition of a national 

water-quality standard for this previously unregulated 

trace element will produce exceptional demand for 

information regarding the present distribution of 

uraniferous waters as well as the geologic and landuse factors that can affect the concentration and 

chemical form of dissolved U. The largest and most 

current database for U in water is a product of the 

National Uranium Resource Evaluation (NURE) 

Program of the middle 1970's. Within the coverage of 

2° quadrangle units evaluated by NURE, waters in the 

Western U.S., and particularly eastern Colorado, are 

noteworthy for consistently elevated concentrations of 

dissolved U. For example, the mean U concentration 

in ground water in the Lamar quadrangle of 

southeastern Colorado is 19.4 ug/L~ a concentration 

two to four times that of ground water in most other 

NURE quadrangles in the Western U.S. (Shannon, 

1979).

In 1991, the U.S. Geological Survey (USGS) 

Uranium/Radon Investigations Program began a study 

of uranium distribution and mobility in southeastern 

Colorado. The immediate goals of this study were to 

(1) evaluate the local bedrock and soils as sources for 

U, (2) identify the form of dissolved U and the 

geochemical controls on U solubility, (3) evaluate the 

potential for secondary enrichment of U in easily 

disturbed near-surface environments and (4) identify 

land-use practices that affect dissolved U in water.

805
As these goals are met, they will improve the ability to 

predict other areas where similar geology, climate, and 

land use combine to produce high concentrations of 

dissolved U in water that could threaten local water 

supplies. The prospects for addressing the causes of 

high concentrations of dissolved U in water will 

depend upon the extent that land-use practices affect U 

concentrations, and the degree to which these practices 

can be modified.

The area selected for detailed study is confined 

to the Arkansas River valley of southeastern Colorado 

and surrounding lands within approximately 20 km of 

the river. This area includes a reach of some 360 river 

miles (579 km) starting near Leadville, Colorado, and 

ending at the Colorado-Kansas border. The river 

traverses a variety of bedrock units that can be 

conveniently grouped according to geographic 

location that is, predominantly Precambrian age 

igneous and metamorphic bedrock in the mountainous 

terrain from Leadville to Canon City, and 

predominantly Upper Cretaceous age marine shales 

and limestones on the High Plains east of Canon City 

(fig. 1), this study; Tweto, 1979). The type and 

intensity of land use can be generally divided on a 

similar geographic basis. Mountainous areas are 

largely undeveloped but include small areas of mining 

activity or limited agricultural development. In 

contrast, the Arkansas River valley east of Canon City 

has a 120-year history of intensive agricultural 

development sustained by a complex network of 

canals, temporary storage reservoirs, transmountain 

diversions and irrigation wells (Taylor and Luckey, 

1974; Abbot, 1985). Open rangeland exists outside of 

irrigated areas, and an area of intensive urban 

development is centered at Pueblo. The study area, 

therefore, offers ample opportunity to investigate the 

effects of variable geology and land-use in a relatively 

constant semiarid climate.

This paper summarizes preliminary 

observations that were based on detailed examination 

of existing NURE data and reconnaissance sampling 

of surface waters and springs along the Arkansas 

River valley in April 1991. The results begin to define 

the extent and causes of uranium enrichment in local 

waters and to provide guidelines for defining the 

course of future study.

SAMPLE COLLECTION AND ANALYSIS

A total of 77 water samples were collected 

during a 2-week period in mid-April 1991. Sampling 

sites included 19 locations along the Arkansas River 

as well as the mouth of most of the tributary streams. 

Most of the samples (48) were collected by USGS 

Water Resources Division personnel from an 

established network of sites designed to monitor water 

quality and flow in the river and its major tributaries. 

The remaining water samples from smaller tributaries, 

springs, and two saline lakes were collected by the 

authors and associates to provide more complete 

coverage, to include some samples outside of irrigated 

areas, and to begin to assess U enrichment in standing 

bodies of water. Natural flows in April are 

representative of the early spring season and are a 

location-dependent mixture of ground-water seepage 

and early spring runoff. In irrigated areas, these 

natural flows are further affected by releases from 

storage reservoirs, diversions to irrigation canals, and 

return flow from initial stages of irrigation.

The specific conductance, temperature, and pH 

of water samples were measured in situ, and samples 

were filtered through 0.45-(im membranes at the time 

of collection. All water samples were analyzed for 

dissolved uranium by inductively coupled plasmamass spectrometry (ICP-MS) (Date and Gray, 1985), 

for arsenic (As) and selenium (Se) by hydride 

generation-atomic absorption spectrophotometry 

(HGAAS) (Fishman and Friedman, 1985), and for 

alkalinity (as HCO3 "') by titration with standard acid. 

A subset of 39 samples from the eastern part of the 

study area also was analyzed for concentrations of 

dissolved anions by ion chromatography (Fishman and 

Pyen, 1979) and for concentrations of dissolved 

cations by inductively coupled plasma-atomic 

emission spectrometry (ICP-AES) (Lichte and others, 

1987). Estimated precision for most reported elements 

is better than ±10% (relative standard deviation).

RESULTS AND DISCUSSION

Recasting of existing NURE data for the Lamar 

2° quadrangle in southeastern Colorado (fig. 1) 

provides some preliminary information on the 

respective effect of geology and land use on dissolved 

U in water. Almost the entire quadrangle is underlain 

by Upper Cretaceous marine shale and limestone, 

although the southern quarter also includes some

806
106° 105° 104° 103° 102°

39°

38°

37°

EXPLANATION

Tertiary

| | Upper Cretaceous 

[*,.' .] Lower Cretaceous 

gm Precambrian

NM I I

25 50 Kilometers

Figure 1 . Generalized geologic map of the Arkansas River basin of southeastern Colorado. The Lamar 2° quadrangle is 

outlined.

Lower Cretaceous Dakota Sandstone (fig. 1). The 

most obvious difference between the southern quarter 

and the rest of the quadrangle is related to land use. 

The southern quarter includes the Arkansas River 

valley from near Manzanola to the Colorado-Kansas 

border and all the irrigated lands adjacent to the River, 

whereas the rest of the quadrangle consists of largely 

undeveloped rangeland.

Approximately 90 percent of the 825 sampled 

wells in the Lamar 2° quadrangle contain greater than 

l|iig/L U (Shannon, 1979). For this discussion, a value 

of 1 ng/L is considered a reasonable detection limit. If 

this population is subdivided according to latitude, 

water from wells in the southern quarter of the

quadrangle contain a markedly higher average U 

concentration (31.7±42 |ig/L) than water from wells in 

the rest of the quadrangle (18.6±24 ng/L) (fig.2). Only 

276 of the well entries included well depths. If the 

comparison is limited to wells shallower than 100 ft 

(30.5 m), the population is reduced to 175 wells. The 

respective average U concentrations, in micrograms 

per liter, for this small subset are 38.5±39 (59 wells) 

and 18.7±23 (116 wells). The similarity of the 

averages in this subset to those of the entire population 

of sampled wells indicate that most of the wells with 

unreported depths also were shallow. An alternative, 

but less likely, explanation is that there is very little 

difference between the average U concentration in

807
o

LU

13

a

LUtr

180

135

90

45

Wells in Lamar Quadrangle Bounded by 

38°15'-40°N Latitude and 

102°-104°W Longitude

Number=566 

Mean= 18.61 

Std. Deviation=24.13 

Median=10

URANIUM CONCENTRATION, 

MICROGRAMS PER LITER

UJ

O

50

40

30

20

10

Wells in Lamar Quadrangle 

Bounded by 38°-38°15'N 

_ Latitude and 102°-104°W 

Longitude Number=170

Mean=31.67

Std. Deviation=42.15

Median=15

* $ £ -i r, f ? # fr^ *

URANIUM CONCENTRATION, 

MICROGRAMS PER LITER

Figure 2. Frequency distributions of uranium concentration 

in well waters from different parts of the Lamar quadrangle, 

Colorado. Data is limited to wells where ground water 

contains a uranium concentration greater than 1 microgram 

per liter (data from Shannon, 1979).

deep wells (greater than 100 feet) and shallow wells in 

both areas.

If the average U concentration of 18 jug/L in 

ground water in the northern part of the Lamar 

quadrangle is considered representative of 

undeveloped ground water that has interacted with 

Cretaceous bedrock units and derivative soils, then it 

is informative to draw comparisons with U 

concentrations in ground water from Precambrian 

crystalline rocks. NURE data for 10 wells in the 

mountainous half of the Pueblo quadrangle west of

Canon City are too limited, but the concentration of U 

in water from 75 springs exceeds 1 |ug/L and averages 

9.9±22 ng/L (Shannon, 1978).

The above comparisons which are based on 

areally extensive sampling of broadly defined water 

types, indicates that geology and land-use practices 

can affect the concentration of dissolved U in the study 

area. A second more focused investigation of these 

effects is provided by the suite of surface-water and 

spring samples reported herein. These waters, 

collected in April 1991, under moderate-flow, 

prerunoff conditions, provide a relatively 

instantaneous snapshot of downriver changes in 

dissolved U with respect to changes in surrounding 

geology and land use. Sample locations and Uconcentration data are summarized in figure 3.

For the purpose of discussion it is convenient to 

consider three distinct sections of river: In the 

uppermost section from Leadville to Canon City, 

dissolved U in the Arkansas River gradually increased 

from less thanl to 3.3 |tig/L (fig.4). Low U 

concentrations in snowmelt-derived water are not 

greatly modified through contact with predominantly 

igneous and metamorphic bedrock and alluvium over 

120 river miles (193 km). Some slightly elevated U 

concentrations in tributaries (5-20 |tig/L) can be 

attributed to acidic mine drainage near Leadville and 

to drainage from volcanic and volcaniclastic rocks that 

host some uranium mines northwest of Parkdale. 

None of these local tributaries supplied enough 

dissolved U to "spike" dissolved U dramatically in 

nearby sampling sites along the Arkansas River.

In the 100-mile-long (161 km) middle reach 

from Canon City to Manzanola, dissolved U in the 

river again increased regularly from 3.3 to 14 |tig/L, 

but at a faster rate (fig.4). Tributaries and a few 

sampled springs contain moderate U concentrations 

(5-50 ng/L). We propose that generally higher U 

values in tributaries and more rapidly rising U 

concentrations in the Arkansas River reflect the 

transition to (1) Upper Cretaceous marine shale and 

limestone bedrock and (2) the beginning of irrigation 

practices (approximately 30 percent of the irrigated 

acreage borders this section). Our set of river 

sampling sites showed no particular effect of a former 

U extraction mill near Canon City, or of urban 

development near the town of Pueblo.

In the final section of river from Manzanola to 

Lamar, the dissolved U concentration does not 

increase in a linear fashion and concentrations scatter

808
106° 105° 104° 103° 102°

39°

38°

Parkdale

EXPLANATION 

Uranium concentration, micrograms per liter

>5-20 

I

>20-50 >50-100 >100

Figure 3. Location of water samples collected in this study. Sample locations are codified according to the uranium 

concentration measured in the water.

within the 30 to 50 (ig/L range. All the values plot 

well above the linear trend defined by the rate of U 

increase in the middle section of river (fig.4). 

Tributary and spring U concentrations are in the 20 to 

100 u,g/L range with some concentrations exceeding 

100 u,g/L (fig.3). This section of river is distinguished 

from the middle section by markedly decreased 

discharge (10-20 percent of that at Manzanola), and 

increased contribution of recycled irrigation-return 

waters to total discharge (Cain, 1985). The adage that 

"the Arkansas River ends at Manzanola" is based on 

the fact that water diversions create a near zero-flow 

condition during parts of the irrigation season. Below 

Manzanola a much greater percent of discharge is 

either directly delivered drain water, or irrigation 

return indirectly delivered through tributaries and 

ground-water seepage. Ground water recharged to the 

alluvial aquifer is further cycled to the surface by 

pumping from more than 1000 irrigation wells (Taylor 

and Luckey, 1974). We propose that surface waters 

sampled along this section of river and in adjoining 

tributaries contain concentrations of U and other

dissolved constituents that were increased by virtue of 

their extensive use and reuse during irrigation. We 

further propose that the additional enrichment of U in 

river water can be estimated by comparing measured 

concentrations to those predicted by the upper linear 

trend of figure 4. The upper linear trend is based on 

the rate of U increase along the middle section of river 

that traverses similar rock types but supports less 

irrigation and is much less affected by irrigation-water 

return.

The upper linear trend of figure 4 predicts the 

downriver increase in dissolved U concentration under 

moderate-flow conditions and under less intensive 

agricultural development. This trend indicates that, 

prior to extensive irrigation along the Arkansas River 

valley, natural processes plus moderate irrigation 

could have generated 20 (ig/L U water by the time the 

Arkansas River exited Colorado. Uranium 

concentration in water stored in the alluvial aquifer 

also must have increased considering that this water 

was recharged by a mixture of surface water 

containing as much as 20 u,g/L U and ground water

809
150 200 

RIVER MILE

EXPLANATION

a=lgneous + metamorphic bedrock, 

undeveloped lands

b=Sedimentary bedrock,

agriculturally developed lands

Figure 4. Uranium in Arkansas River water, April, 1991, 

plotted as a function of river mile from Leadville, Colorado. 

Visually fitted linear trends labelled a and b define different 

rates of increase of dissolved uranium concentration in two 

sections of river with contrasting geology and land use. 

Linear trend a is for a section of river underlain by igneous 

and metamorphic rocks and containing largely undeveloped 

lands. Linear trend b is for a section of river underlain by 

marine shale and limestone and containing agriculturally 

developed lands. (Abbreviation: ng/L U/mile , microgram 

per liter of uranium per river mile.)

containing an average of 18 |ig/L U. The 

predevelopment concentration of U in Arkansas River 

water cannot be estimated from our data, but 

concentrations could have approached the 18 jig/L 

ground-water average during periods of low flow 

when ground-water seepage provided a large 

component of total discharge.

Dissolved U in all 49 surface water and spring 

samples collected east of Canon City shows a strong 

positive correlation (r= 82) with specific conductance 

(fig. 5) indicating that dissolved U is a soluble and 

mobile species that is largely conserved in solution. 

This relation, which is commonly observed in natural 

waters, suggests that the more easily measured 

property of specific conductance can be used to 

estimate the potential for high U concentration. 

Previous studies of temporal and spatial variation of 

specific conductance in the Arkansas River valley 

(Konikow and Person, 1985; Cain, 1987) may 

indirectly address probable historic variations of trace 

elements such as U. Increases in specific conductance 

are attributed to mineral dissolution and evaporative

concentration, both of which are accentuated as 

surface waters are highly managed and reused during 

irrigation in semiarid climates.

Areally extensive sampling and a probable wide 

variety of applied water compositions make it difficult 

to separate the effects of evaporative concentration 

from mineral dissolution. Cheraw Lake, a highly 

saline closed-basin sink for irrigation return water, 

provides a probable extreme example of the local 

extent of U concentration by evaporation. The lake is 

located 12 km north of La Junta and at the time of 

sampling contained 230 jig/L U in surface water 

(fig. 3). This sample does not plot on figure 5 because 

of its unusually high specific conductance of 

31,000 fiS/cm. Unusually high specific conductance 

associated with dissolved U may indicate addition of 

salts washed in or blown in from marginal salt flats. 

Alternatively, some precipitation of U can occur under 

reducing conditions at the lake bottom. More detailed 

study, including sampling of local input seepage 

waters, is required to address these questions.

Summary statistics for 16 measured variables in 

39 surface-water samples collected east of Canon City 

are presented in table 1. Samples were chosen to span 

the range of specific conductance (Cheraw Lake outlier 

excluded). The large positive correlation between U 

and specific conductance is also accompanied by a large 

positive correlation of U with the major dissolved

250

r^U 200

150

O< 

OCC 100

^ 50

00)

DC

2,000 4,000 6,000 8,000

SPECIFIC CONDUCTANCE, 

MICROSIEMENS PER CENTIMETER

Figure 5. Uranium concentration as a function of specific 

conductance for all surface waters and springs collected 

east of Canon City, Colorado, April 1991 . Surface waters 

collected west of Canon City plot near the origin and are 

omitted for clarity.

810
species sodium, calcium, magnesium, sulfate, and 

chloride that strongly contribute to specific 

conductance. The minor elements lithium, potassium 

and strontium that have geochemical affinity with the 

major dissolved species also correlate with U. The 

minor elements boron and selenium, that are commonly 

concentrated in saline waters of the Western U.S., also 

correlate positively with dissolved U. A previous study 

(Mueller and others, 1991) documented elevated 

concentrations of Se in irrigation-drainage waters along 

the Arkansas River valley of Colorado. The apparent 

behavior of U as a "conserved" element in solution 

indicates that, in this setting, mobilization and transport 

of U dominate over precipitation or sorptive uptake. 

High solubility of U is expected in oxygenated alkalinepH surface waters that contain sufficient concentrations 

of dissolved carbonate to stabilize dissolved U as uranyl 

carbonate complexes (Langmuir, 1978).

A less well-defined correlation is exhibited 

between U and pH, which rarely shows strong 

covariation with specific conductance, and between U 

and dissolved carbonate, silica and manganese, which 

are probably limited by precipitation of calcium 

carbonate, silica, and manganese oxides. Additional 

statistical calculations for smaller subsets of samples 

containing measurable Fe and Ba also show poor 

correlation of these elements with U, again indicating 

solubility control, probably through precipitation of Fe 

oxides and barite. Equilibrium-based, mineralsolubility calculations using the computer code 

PHREEQE (Parkhurst and others, 1980) indicate 

saturation of these waters with calcite, dolomite, 

quartz, ferric hydroxide, and barite, and 

undersaturation-to-saturation with gypsum.

Table 1. Statistical summary of chemical data for 39 surface-water samples from the Arkansas River valley east of Canon 

City, Colorado, collected April, 1991.

[ n, number of values above detection; *, calculated after assigning a value of 0.7of the detection limit to undetected samples; Sp. Cond.(nS/cin), specific 

conductance (microsiemens per centimeter at 25 degrees Celsius); SiC>2, silicon dioxide; U, uranium; dashes indicate value not calculated]

Variable

Sp.Cond.(nS/cm) 

pH

n

39 

39

Minimum

229

7.55

Maximum

8,000 

8.70

Mean

2,681 

8.09

Standard 

deviation

1,983

.25

Correlation 

coefficient 

withU

0.84 

-.37

SiO2

Sodium

Potassium

Calcium

Magnesium

Bicarbonate

Sulfate

Chloride

Boron

Lithium

Manganese

Selenium

Strontium

Uranium

Total Iron 

Barium

Major elements (milligrams/Liter)

39

39

39

39

39

39

39

39

34

39

28

35

39

39

13

22

1.3

9.0

1.0

26

7.0

95

28

1.7

34

1,000

36

580

490

625

5,080

320

15

259

6.9

240

119

283

1,344

70.6

Minor elements (micrograms/Liter) 1

<100

28 

<20

<] 

220 

3.3

2,000

540

620

68

11,000

220

394* 

129 

75.3* 

14.9*

4,335 

45

6.9

265

6.8

147

111

102

,250

85.3

404*

124

127*

13.0* 

3,195 

47

Less frequently detected elements (micrograms/Liter)

<50 

<40

200

870

.31

.86 

.50 

.70 

.77 

.30 

.82 

.87

.96* 

.87 

.06* 

.79* 

.79 

1.00

-.22

-.19

'Other elements sought but consistently below detection limit (micrograms/Liter) include; silver(2), arsenic(2), beryllium(l), bismuth(IO), cadmium(l), cobalt(3), chromium(l), copper(IO), gallium(5), molybdenum( 10), nickel(5), lead(IO), tin(IO), vanadium(6). zinc(3), and zirconium(l).

811
CONCLUSIONS

Preliminary reconnaissance sampling of surface 

waters along the Arkansas River valley of 

southeastern Colorado indicates the existence of both 

geologic and land-use controls on concentrations of 

dissolved U. A plot of U concentrations in the 

Arkansas River as a function of distance downriver 

displays abrupt changes in slope and degree of scatter 

of values that are attributed to the contrasting and 

sequential effects of (I) Precambrian crystalline 

bedrock and derivative soils/sediments, (2) Cretaceous 

sedimentary bedrock and derivative soils/sediments 

and the extensive irrigation thereof. U concentrations 

are consistently highest and show the most scatter in 

the section of river where flow is largely sustained by 

returns of multicycled irrigation water. As is 

commonly observed in natural waters, U correlates 

positively with specific conductance and with other 

major and minor chemical species that are not limited 

by major sorptive uptake or the solubility of mineral 

precipitates. In this setting, specific conductance 

serves as an easily measured, albeit approximate, 

predictor of dissolved U concentration because both 

properties increase as a function of solid/water 

interaction and evaporative concentration. In semiarid 

southeastern Colorado, natural weathering, soil 

leaching, and evaporative concentration can produce 

regional ground water with U concentrations near the 

proposed USEPA drinking water regulation of 

20 jig/L. Locally extensive irrigation in the Arkansas 

River valley further elevates dissolved U 

concentrations and this combination of natural and 

manmade effects could compromise water quality for 

communities that derive drinking water from the 

alluvial aquifer.

REFERENCES

Abbot, P.O., 1985, Description of water systems operations 

in the Arkansas River basin, Colorado: U.S. 

Geological Survey Water-Resources Investigations 

Report 85-4092, 67 p.

Cain, Douglas, 1985, Quality of the Arkansas River and 

irrigation-return flows in the lower Arkansas River 

valley, Colorado: U.S. Geological Survey WaterResources Investigations Report 84-4273, 85 p.

Cain, Douglas, 1987, Relations of specific conductance to 

streamflow and selected water-quality characteristics 

of the Arkansas River basin, Colorado: U.S. 

Geological Survey Water-Resources Investigations 

Report 87-4041,93 p.

Date A.K. and Gray, A.L., 1985, Determination of trace 

elements in geological samples by inductively coupled 

plasma source mass spectrometry: Spectrochimica 

Acta40B,p. 115-122.

Fishman, M.J. and Friedman, L.C.,1985, Methods for 

determination of inorganic substances in water and 

fluvial sediments: U.S. Geological Survey Techniques 

of Water-Resources Investigations, book 5, chap. Al, 

709 p.

Fishman, M.J. and Pyen, G., 1979, Determination of

selected anions in water by ion chromatography: U.S. 

Geological Survey Water-Resources Investigations 

Report 79-101,30 p.

Konikow, L.F. and Person, Mark, 1985, Assessment of 

long-term salinity changes in an irrigated streamaquifer system: Water Resources Research, v. 21, 

no. 11, p. 1611-1624.

Langmuir, Donald, 1978, Uranium solution-mineral 

equilibria at low temperatures with application to 

sedimentary ore deposits: Geochimica et 

Cosmochimica Acta, v. 42, p. 547-569.

Lichte, F.E., Golightly, D.W. and Lamothe, P.J., 1987, 

Inductively coupled plasma-atomic emission 

spectrometry, in Baedecker, PA., ed., Methods for 

Geochemical Analysis: U.S. Geological Survey 

Bulletin 1770, chap. Bl, 10 p.

Mueller, O.K., DeWeese, L.R., Garner, A.J.and Spruill, T.B., 

1991, Reconnaissance investigation of water quality, 

bottom sediment, and biota associated with irrigation 

drainage in the middle Arkansas River basin, Colorado 

and Kansas, 1988: U.S. Geological Survey WaterResources Investigations Report 91-4060, 168 p.

Parkhurst, D.L., Thorstenson, D.C. and Plummer, L.N., 

1980, PHREEQE- A computer program for 

geochemical calculations: U.S. Geological Survey 

Water-Resources Investigations Report 80-96, 210 p.

Shannon, S.S. Jr., 1978, Uranium hydrogeochcmical and 

stream sediment reconnaissance of the Pueblo NTMS 

quadrangle, Colorado: Los Alamos National 

Laboratory Informal Report LA-7341-MS, 76 p.

Shannon, S.S. Jr., 1979, Uranium hydrogeochemical and 

stream sediment reconnaissance of the Lamar 

quadrangle, Colorado, including concentrations of 

forty-three additional elements: Los Alamos National 

Laboratory Informal Report LA-7342-MS, 149 p.

Taylor, O.J.and Luckey, R.R., 1974, Water-management 

studies of a stream-aquifer system, Arkansas River 

valley, Colorado: Ground Water, v. 12, no. 1, p. 22-38.

Tweto, Ogden, 1979, Geologic map of Colorado: U.S. 

Geological Survey Map, scale 1:500,000.

U.S. Environmental Protection Agency, 1991, National 

primary drinking water regulations for radionuclides, 

proposed rule, June, 1991: Washington, D.C., Office of 

Ground Water and Drinking Water, EPA, Fact Sheet, 

Radionuclides in Drinking Water, 570/9-91 -700, 12 p.

812
Dissolved Organic Compounds in the Mississippi River

from Minneapolis-St. Paul, Minnesota

to New Orleans, Louisiana

By Larry B. Barber, II 1 , Jerry A. Leenheer2 , Wilfred E. Pereira3 , and Mary A. Noriega2

Abstract

Dissolved organic contaminants were measured 

in water samples collected from the Mississippi 

River during four sampling surveys to assess spatial 

and temporal variability of these substances in the 

river. A hierarchical analytical approach was taken 

in which both broad spectrum and specific 

compound analyses were used. Specific indicators 

of sewage effluents included anionic surfactants and 

caffeine. These compounds were detected throughout the river, with maximum concentrations occurring near major cities. Other compounds such as 

volatile chlorinated hydrocarbons and ethylenediamine-tetraacetic acid, which are more indicative of 

industrial effluents, also were detected throughout 

the river. Specific organic compound concentrations in water were less than 50 micrograms per 

liter, and were controlled by a variety of environmental processes. Depending on the molecular 

structure of individual chemicals, they can be 

persistent in the Mississippi River or they can be 

removed by biodegradation, volatilization, or sorption.

INTRODUCTION

The Mississippi River (fig. 1), a source of public 

water supply for more than 100 cities, receives wastewaters containing a variety of organic compounds, 

some of which are detrimental to aquatic organisms 

and human health. Runoff and wastewater from urban 

areas, farms, factories, and individual households 

contribute organic compounds to the Mississippi River.

U.S. Geological Survey, Boulder, Colo. 

2U.S. Geological Survey, Arvada, Colo. 

3U.S. Geological Survey, Menlo Park, Calif.

The objective of this study was to (1) determine the 

distributions and concentrations of organic compounds 

along the entire length of the river, (2) evaluate the seasonal variability in contamination patterns, (3) evaluate 

the environmental processes that control the occurrence and fate of the contaminants, and (4) compare the 

present-day water quality of the river with historical 

water quality. This paper summarizes some of the major 

findings of the larger study, in particular the longitudinal distribution of select dissolved constituents.

APPROACH

Organic compounds in the Mississippi River were 

evaluated by collecting water and sediment samples 

between New Orleans, La., and Minneapolis-St. Paul 

(Twin Cities), Minn., during four sampling surveys 

done in June-July 1991 (summer), October-November

1991 (fall), February 1992 (winter), and March-April

1992 (spring). Midchannel grab samples of water were 

collected at about 10-mi intervals as the Research 

Vessel Acadiana traveled upstream from New Orleans 

to the Twin Cities. During the return to New Orleans, 

depth-and-discharge averaged water samples and composite bed sediment samples were collected. The water 

and sediment samples were analyzed for a variety of 

organic compounds, a few of which are presented here.

RESULTS

Dissolved organic carbon (DOC) has important 

geochemical implications because it has the potential 

to (1) increase the water solubility and facilitate the 

transport of organic contaminants, (2) alter rates of biodegradation, (3) form complexes with trace metals, and 

(4) react during water treatment to produce potentially 

toxic byproducts. The DOC longitudinal profiles 

(fig. 2a) show high concentrations (10-12 mg/L) in the 

upper Mississippi River (above the confluence with the 

Ohio River) that decrease downstream, a sharp

813
. MINNESOTA

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45'V

200 MILES 

230 KILOMETERS

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Figure 1. The Mississippi River and major tributaries.

decrease in concentrations below the Ohio River, and 

relatively constant concentrations (3-4 mg/L) in the 

lower Mississippi River (between the confluence with 

the Ohio River and the Gulf of Mexico). The distribution of DOC was influenced by regional cycling of 

natural organic matter, which is affected by climate, 

vegetation, soil type, and biogeochemical factors.

The longitudinal distributions of specific contaminants such as methylene-blue-active-substances 

(MBAS), total organic halide (TOX), volatile organic 

compounds (VOC), ethylenediaminetetraacetic acid 

(EDTA), and caffeine along the Mississippi River were 

more variable, reflecting the effect of multiple inputs 

from municipal, industrial, and agricultural sources.

814
Methylene-blue-active-substances are a composite 

measure of total dissolved anionic surfactants including linear alkylbenzenesulfonates (LAS~the surface 

active ingredient of detergents), LAS biological metabolites and impurities, other synthetic-anionic surfactants, and natural organic compounds such as humic 

and fulvic acids. Because MBAS is an indicator of 

sewage contamination, there is a historical record of 

MBAS measurements dating back to the early 1950's 

that allows long-term trends to be assessed. MBAS was 

present throughout the Mississippi River and its 

tributaries at concentrations ranging from 0.025 to 

0.090 mg/L (fig. 3). Peak MBAS concentrations were 

detected in the vicinity of major cities as the result of

sewage-effluent discharges, and concentrations rapidly 

decreased downstream as a result of dilution, biodegradation, and sorption to sediments. The results of LASspecific analyses for water from the Mississippi River, 

and the biogeochemical processes that control their fate 

are reported by Tabor (1993) and Tabor and Barber 

(1996).

Total organic halogen is a measure of dissolved 

chlorinated organic compounds (volatile and nonvolatile) that originate from disinfection of sewage 

effluents, biofouling control in powerplant-cooling 

waters, and pulp-wood bleaching. Although the 

composition of TOX is variable, it is an index of 

impairment of water quality as indicated by the toxicity

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DISTANCE FROM HEAD OF PASSES, IN THOUSAND MILES

Figure 2. Longitudinal distribution of dissolved organic carbon for midchannel water samples collected 

during three upriver surveys of the Mississippi River during 1991-92.

815
of many compounds determined by this measurement. 

The TOX concentrations in the Mississippi River 

ranged from 10 to 120 u£/L and showed significant 

seasonal differences (fig. 4). TOX is thought to be 

relatively resistant to removal by sorption and biodegradation.

Caffeine, a chemical in beverages such as coffee 

and soft drinks, was used as a specific tracer of domestic wastewater. Caffeine concentrations indicated the 

input of sewage from the Twin Cities into the relatively low volume upper Mississippi River and 

progressive dilution downstream (fig. 5). The Illinois 

River contained elevated concentrations of caffeine 

probably derived from the large population in

Chicago, 111. An increase in caffeine concentrations 

was detected in the Mississippi River below its confluence with the Illinois River.

Ethylenediaminetetraacetic acid is used to form 

water soluble complexes with insoluble metals, and, as a 

consequence, has found use in a wide variety of domestic and industrial applications. The EDTA longitudinal 

profile indicates the upper Mississippi River had greater 

EDTA concentrations than did the lower Mississippi 

River (fig. 6). EDTA and caffeine are present in similar 

concentrations in sewage effluent; however, industrial 

sources and the lack of instream degradation probably 

produce EDTA concentrations in the Mississippi River 

that are higher than those of caffeine.

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Figure 3. Longitudinal distribution of dissolved methylene-blue-active-substances for midchannel water 

samples collected during three upriver surveys of the Mississippi River during 1991-92.

816
Volatile organic compounds such as trihalomethanes, chlorinated ethenes and ethanes, and aliphatic 

and aromatic hydrocarbons come from industrial and 

municipal sources, as well as petroleum fuel used in 

transportation. Many VOC's are U.S. Environmental 

Protection Agency priority pollutants and have maximum contaminant levels (MCL's) that range from 2 to 

200 ng/L depending on the particular compound (U.S. 

Environmental Protection Agency, 1987). Thirty-two 

VOC's were detected and 19 were present in more than 

1 percent (up to 15 percent) of the samples analyzed;

about 23 percent of the samples had detectable VOC's. 

The most frequently detected compounds were dichloroethane, methylene chloride, toluene, trichloroethane and 

tetrachloroethene. Concentrations for these compounds 

ranged from 0.2 to 1.1 ng/L (typically an order of magnitude below MCL's for the individual compounds); the 

only reaches of the river with significant levels of VOC 

were below the Twin Cities, below the confluence with 

the Ohio River, and between Baton Rouge and New 

Orleans, La. (fig. 7). Loss of VOC's from the water as the 

result of volatilization appears to be a rapid process.

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DISTANCE FROM HEAD OF PASSES, IN THOUSAND MILES

Figure 4. Longitudinal distribution of dissolved total organic halogen for midchannel water samples 

collected during three upriver surveys of the Mississippi River during 1991-92.

817
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DISTANCE UPRIVER FROM HEAD OF PASSES IN MILES

Figure 5. Longitudinal distribution of dissolved caffeine for depth and discharge integrated water samples 

collected during three downriver surveys of the Mississippi River during 1991-92.

A Tributary 

Mainstem

1,500 1,000 500

DISTANCE UPRIVER FROM HEAD OF PASSES IN MILES

Figure 6. Longitudinal distribution of dissolved ethylenediamine tetraacetic acid for depth and discharge 

integrated water samples collected during three downriver surveys of the Mississippi River during 1991-92.

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DISTANCE FROM HEAD OF PASSES, IN THOUSAND MILES

Figure 7. Longitudinal distribution of volatile organic compounds for midchannel 

water samples collected during three upriver surveys of the Mississippi River 

during 1991-92. 819
DISCUSSION

The organic compounds in the Mississippi River 

comprise a complex mixture of a wide range of natural 

and synthetic chemicals introduced by a variety of 

sources. Although concentrations of most contaminants measured in this study were below U.S. Environmental Protection Agency regulatory limits, their 

presence was detected along the entire Mississippi 

River. Significant sources of organic compounds in the 

Mississippi River include municipal-wastewater discharge, urban runoff, powerplant cooling-water discharge, pulp-mill effluent, feedlot runoff, commercial 

and recreational river traffic and spills during refueling, 

and discharge from industrial facilities.

Although dissolved contaminant concentrations 

measured in this study indicate the potential for 

adverse effects, the Mississippi River is dynamic, and 

concentrations are continually changing; locations in 

the immediate vicinity of point sources such as municipal discharges are most likely to have the highest concentrations and largest number of organic compounds. 

It is difficult to relate concentrations in the river water 

to drinking water standards because water treatment 

can remove or introduce contaminants.

Distinct spatial patterns were observed in the 

organic-chemical data. The highest concentrations of 

organic compounds were in the vicinity of major 

metropolitan areas, and concentrations were typically 

greatest in the upper Mississippi River where stream 

dilution was lowest. The organic chemistry of major 

tributaries such as the Minnesota, Illinois, Missouri, 

and Ohio Rivers significantly affected the chemistry of 

the Mississippi River. Seasonal differences in organic 

compound concentrations are related to hydrologic,

climatic, biological, and geochemical factors. Concentrations were highest during periods of low flow (fall) 

and lowest during periods of high flow (spring). Likewise, concentrations of biologically labile and volatile 

organic compounds were affected by temperature with 

highest concentrations in the winter, when temperatures were lowest. Depending on the molecular structure of the individual chemicals, some compounds 

were quite persistent in the Mississippi River, whereas 

others were rapidly removed by biodegradation, volatilization, and sorption. Although the results from this 

study represent a significant step forward in understanding the occurrence and nature of organic compounds in the Mississippi River, there still is much to 

be learned about their sources, fate, and effects on 

human health and aquatic ecology.

REFERENCES

Tabor, C.F., Jr., 1993, Occurrence and fate of linear alkylbenzene sulfonate surfactants and related compounds in 

the Mississippi River: Boulder, University of Colorado, 

Unpublished master's thesis, 65 p.

Tabor, C.F., Jr., and Barber, L.B., II, 1996, Linear alkylbenzene sulfonate in the Mississippi River, in Morganwalp, 

D.W., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program-Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

U.S. Environmental Protection Agency, 1987, National 

primary drinking water regulations Synthetic organic 

chemicals: Monitoring for unregulated contaminants: 

Final Rule: Federal Register, v. 52, p. 25690-25717.

820
Linear Alkylbenzene Sulfonate in the 

Mississippi River

By Charles F. Tabor, Jr. 1 ,2 , and Larry B. Barber, II 1 ,

Abstract

Linear alkylbenzene sulfonate (LAS), the 

active component in most detergents, is one of the 

most common synthetic organic compounds 

found in natural waters and sediments. The 

estimated LAS loading in the Mississippi River 

basin is 3,500 kilograms per day. A detailed 

sampling program was undertaken by the U.S. 

Geological Survey to determine the occurrence 

and fate of LAS in a 2,800 kilometer section of 

the Mississippi River from Minneapolis, 

Minn., to New Orleans, La. LAS was 

identified and quantified in 22 percent of the 

water samples at concentrations ranging from 

0.06 to 28.2 micrograms per liter. Most 

dissolved LAS was associated with the sewage 

treatment plant outfalls of large cities along the 

river. LAS was identified in all of the composite 

bottom sediment samples at concentrations of 

0.01 to 20 milligrams per kilogram. Sorption to 

sediment removes 5 to 30 percent of dissolved 

LAS. On the basis of homolog and isomer data, 

biodegradation is the most important removal 

process affecting dissolved LAS.

INTRODUCTION

Linear alkylbenzene sulfonate (LAS) is the world's 

most widely used synthetic surfactant. LAS 

consumption totaled 390 x 106 kg in the United States in 

1991,88 percent of which was used in detergents and 

household cleaners (Modler and others, 1993). Rapaport

'U.S. Geological Survey, Boulder, Colo. 

Department of Geological Sciences University of Colorado 

at Boulder

and Eckhoff(1990) estimate that each person in the 

United States uses about 2.6 g of LAS daily.

Most LAS is disposed of into municipal 

wastewaters and subsequently treated at sewagetreatment plants (STP). Most STP use secondary 

sewage treatment during which an average of 98 

percent of LAS input is removed. Removal occurs by 

either (1) sorption to sludges or (2) biological 

degradation to intermediate compounds or ultimately 

to water, carbon dioxide, and sulfate ion (Rapaport and 

Eckhoff, 1990; Berna and others, 1989). Despite the 

high removal rate, significant amounts of LAS can be 

introduced into surface waters through discharge of 

municipal sewage effluent. Estimated LAS use in the 

Mississippi River basin is 175,000 kg/d (Tabor, 1993). 

Therefore 3,500 kg of LAS could be introduced daily 

to the Mississippi River and its tributaries.

This study was designed to evaluate the 

occurrence and fate of LAS in a 2,800 km reach of the 

Mississippi River between Minneapolis, Minn and New 

Orleans, La. The work reported here was part of the 

U.S. Geological Survey's 5-year study of contaminants 

in the Mississippi River (Meade, 1993) and is a 

summary of the master's thesis of Tabor (1993).

Study Area

The Mississippi River is the largest river in 

North America (fig. 1) and is the sixth largest river in 

the world based on annual discharge (Milliman and 

Meade, 1983). Its watershed covers 3.3 x 106 km2 and 

drains 41 percent of the area of the 48 conterminous 

states (Turner and Rabalais, 1991). The upper river 

consists of a series of 27 locks and dams beginning 

about 10 km upriver from St. Louis and ending in the 

Minneapolis/St. Paul area. The lower section of the 

river is free-flowing to the Gulf of Mexico. Major 

tributaries include the Minnesota, Illinois, Missouri, 

Ohio, and Arkansas Rivers.

821
:oo MILES

30*

cut of

Figure 1. The Mississippi River and its major tributaries.

Linear Alkylbenzene Sulfonate

Commercial LAS is a mixture of homologs 

having alkyl chain lengths ranging from 10 to 14 carbon 

units with isomers of variable phenyl position. The 

phenyl positions range from 2 to 7 for C| 0 to C14 LAS, 

resulting in a total of 26 isomers. The variable length of 

the unsubstituted alkyl chain causes each isomer to have

slightly different characteristics relating to sorption and 

biodegradation. LAS has an aqueous solubility of 0.2 to 

160 mg/L and a distribution coefficient (K^) of < 1000 to 

6000 L/kg (Marcomini and others, 1988). This wide 

range of physical properties makes LAS an ideal 

indicator of environmental processes occurring in water 

and sediment.

822
EXPERIMENTAL

Field Sampling

Water and sediment samples were collected 

from the Research Vessel ACADIAN A on three cruises 

in the summer (June 23-August 7) and fall (September 

24-November 13) of 1991 and the spring (March 25- 

May 10) of 1992. Samples were collected from the 

Mississippi River between Minneapolis, Minn. and 

New Orleans, La. Tributaries having at least 5 percent 

of the discharge of the Mississippi also were 

sampled. Surface grab samples were collected on each 

upriver cruise, and composite water and sediment 

samples were collected on each subsequent downriver 

cruise. Details of the sampling procedures can be 

found in Meade( 1993).

Analytical

LAS was identified using a solid-phaseextraction/derivatization/gas-chromatography/massspectrometry method adapted from Trehy and others 

(1990). LAS was isolated using a Cj 8 solid-phase 

extraction cartridge. LAS was derivatized to its 

trifluoroethyl ester and subsequently analyzed using a

Hewlett-Packard gas chromatograph with electron 

impact mass selective detector. Quantitation was 

accomplished using a 9 LAS surrogate standard 

added to each sample before isolation. LAS 

identification was based on the retention times and 

mass spectra of U.S. Environmental Protection 

Agency LAS standard #1288.

RESULTS AND DISCUSSION

Dissolved Linear Alkylbenzene Sulfonate

Dissolved LAS was found in 22 percent of the 

437 water samples at concentrations of 0.06 to 

28.2 u,g/L. Spatial distribution of dissolved LAS in 

grab samples is shown in figure 2. Maximum 

concentrations of dissolved LAS in the Mississippi 

River were found below the STP outfalls of 

municipalities such as Minneapolis and St. Louis. 

Concentrations decreased with increasing distance 

from the outfall. Seventy-two percent of dissolved 

LAS was found within 65 km of the sewage outfalls of 

towns and cities along the river. Several tributaries 

contained measurable concentrations of LAS. Most of 

the water samples containing LAS were collected

v_ 12

0)

CD 10 

Q_

CO

E oj s

O) 

O

O

O 4

CD 2 

O

C 

O

0

2,000

June-July 1991

 

Sept-Oct 1991 

March-April 1992

\i\

1,500 1,000 500

Distance Upriver From Head of Passes (miles)

Figure 2. Spatial distribution of linear alkylbenzene sulfonate in grab samples collected from the Mississippi River.

823
from the Mississippi River above its confluence with 

the Ohio River, indicating that dilution is a secondary 

factor that determines LAS concentration in the river. 

Calculations based on the homolog and isomer 

distributions of dissolved LAS measured in the river 

indicated that biodegradation is the most important 

control of dissolved LAS in the Mississippi.

Sorbed Linear A Iky I benzene Sulfonate

LAS was identified in all 29 Mississippi River 

composite bottom-sediment samples in concentrations 

.of 0.01 to 0.95 mg/kg (dry weight). Of the four 

tributaries sampled the highest concentration of sorbed 

LAS found was 20 mg/kg in Pig's Eye Slough, a canal 

that transports the Minneapolis/St. Paul sewage 

effluent to the Mississippi River. The highest sorbed 

LAS concentrations in the Mississippi River were 

found within 60 km of the Minneapolis area. The 

ubiquitous sorption of LAS to bottom sediments 

indicates that sorption is an important removal process 

of dissolved LAS. Calculations indicate that 5 to 30 

percent of dissolved LAS may be removed from the 

river by sorption.

SUMMARY

Most LAS in the Mississippi River originates 

from the sewage outfalls of cities and towns located 

near the river and its tributaries. Dilution is an 

important control of LAS concentrations in the river. 

Sorption and biodegradation are the main controls on 

the fate of LAS in the river. LAS is a ubiquitous 

component of the bottom sediments, indicating that 

sorption plays a significant role in removal of LAS 

from the dissolved phase. Dissolved LAS homolog 

and isomer distributions indicate that biodegradation 

is the most important removal process of dissolved 

LAS. Biodegradation is the ultimate fate of LAS in 

the Mississippi River.

REFERENCES

Berna, J.L., Ferrer, J., Moreno, A., Prats, D., and Ruiz

Bevia, R, 1989, The fate of LAS in the environment:

Tenside Surfactants Detergents, v. 26, p. 101-107. 

Hand, V.C. and Williams, O.K., 1987, Structure-activity

relationships for sorption of linear alkylbenzene

sulfonates: Environmental Science and Technology,

v. 21, p. 370-373. 

Kimerle, R.A., 1989, Aquatic and terrestrial ecotoxicology

of linear alkylbenzene sulfonate: Tenside Surfactants

Detergents, v. 26, p. 169-176. 

Kimerle, R.A. and Swisher, R.D., 1977, Reduction of

aquatic toxicity of linear alkylbenzene sulfonate

(LAS) by biodegradation: Water Research, v. 11,

p. 31-37. 

Marcomini, A. and Giger, W., 1988, Behaviour of LAS in

sewage treatment: Tenside Surfactants Detergents,

v. 25, p. 226-229. 

Meade, R.H., ed., 1993, Contaminants in the Mississippi

River, 1987-1992: U.S. Geological Survey Water

Supply Paper 2440. 

Milliman, J.D. and Meade, R.H., 1983, World-wide delivery

of river sediment to the oceans: Journal of Geology,

v. 91, p. 1-21. 

Modler, R.F., Willhalm, R., and Yoshida, Y., 1993, Linear

alkylate sulfonates: Chemical Economics Handbook

Marketing Research Report, surface-active agents,

593.9000-9001. 

Rapaport, R.A. and Eckhoff, W.S., 1990, Monitoring linear

alkyl benzene sulfonate in the environment 1973-

1986: Environmental Toxicology and Chemistry, v. 9,

p. 1245-1257. 

Swisher, R.D., 1963, The chemistry of surfactant

biodegradation: Journal of the American Oil Chemists'

Society, v. 40, p. 648-656. 

Swisher, R.D., 1987, Surfactant Biodegradation, Second

Edition: Marcel Dekker, Inc., NY. 

Tabor, Jr., C.F., 1993, The occurrence and fate of linear

alkylbenzene sulfonate in the Mississippi River: a

molecular indicator of sewage contamination: Master's

Thesis, University of Colorado at Boulder. 

Trehy, M.L., Gledhill, W.E., and Orth, R.G., 1990,

Determination of linear alkylbenzenesulfonates in

water and sediment by gas chromatography/mass

spectrometry: Analytical Chemistry, v. 62, p. 2581-

2586. 

Turner, R.E. and Rabalais, N.N., 1991, Changes in

Mississippi River water quality this century:

Bioscience, v. 41, p. 140-147.

824
Organic Carbon Associated With Suspended Silt and

Colloid From the Mississippi River and Some of Its

Tributaries, July 1991-May 1992

By Colleen E. Rostad 1 , Stephanie G. Monsterleet1 , 

LaDonna M. Bishop1 , and Geoffrey S. Ellis 1

Abstract

Suspended silt and colloid samples were 

collected from 16 sites along the Mississippi River and 

from some of its tributaries during July-August 1991, 

October-November 1991, and April-May 1992 to 

determine the organic carbon associated with the 

suspended matter. Sample collection involved 

pumping discharge-weighted volumes of river water 

from a cross section of the river into a continuous-flow 

centrifuge to isolate the suspended silt. Colloid 

remaining in the aqueous phase was isolated by 

ultrafiltration. This is the first time that the presence of 

organic carbon associated with the suspended silt and 

colloid has been investigated along the entire reach of 

the Mississippi River from near Minneapolis, 

Minnesota, to south of New Orleans, Louisiana. The 

organic carbon content associated with the colloids (4 

to 31 percent) was much higher than the organic 

carbon associated with the silt (1 to 17 percent).

INTRODUCTION

Depending on their water solubility, organic 

contaminants in a river system can be located in any of 

several compartments, such as water, silt, colloid, bed 

sediment, or benthic plants and organisms. The more 

water-soluble compounds are shown to be transported 

in the water phase (Pereira and Rostad, 1990; Pereira 

and others, 1990), whereas the less water-soluble 

compounds would be associated with organic carbon 

and would be transported with the organic carbon

associated with the suspended matter-either the silt or 

the colloid.

Samples of silt and colloid were collected along 

the Mississippi River from near Minneapolis, 

Minnesota, to south of New Orleans, Louisiana. The 

sampling sites included four tributaries the 

Minnesota, Illinois, Missouri, and Ohio Rivers. 

Sampling cruises were made during July-August 1991, 

October-November 1991, and April-May 1992. The 

summer and fall cruises were made during periods of 

low flow, in contrast to the spring cruise, which was 

made during high flow on the Mississippi River.

The purpose of this paper is to describe the 

distribution of organic carbon associated with the 

suspended silt and colloid at selected sites on the 

Mississippi River and some of its tributaries for these 

three cruises.

SAMPLE COLLECTION

Representative, discharge-weighted, suspendedmatter samples were collected using techniques 

described by Moody and Meade (1992). Dischargeweighted samples are proportional to the discharge at 

each evenly spaced vertical of the cross section of the 

river at each site. The sampling sites for each cruise 

are shown on a map of the Mississippi River area in 

figure 1.

At each sampling site, the water discharge was 

estimated at each of from 5 to 30 (depending on the 

width of the river) equally spaced verticals along a 

cross section of the river by using the depth profile of

USGS, Arvada, Colo.

825
200 MILES

100 200 KILOMETERS

H«od of Passes 

Gulf Of Mexico

Figure 1. Map showing sampling sites on the Mississippi River and some tributaries. (* denotes a tributary)

826
the river and five depth-integrated velocity measurements. The 17-m research vessel ACADIANA, from 

which the samples were collected, was positioned in 

the river at each of the verticals of the river cross section by using microwave trisponders located on each 

bank. The maximum possible sampling depth for the 

equipment was 5 m. The discharge-weighted pumped 

sample was collected at each vertical from one-half 

the depth or 5 m, whichever was less.

The water was pumped through FEP Teflon 

tubing, rigidly positioned parallel to the flow, by an 

air-driven, double-diaphragm Teflon pump, into a 63- 

(j.m nickel-mesh sieve. The water then flowed into a 

large (40 L) glass funnel, which provided constant 

head pressure into a continuous-flow, high-speed 

centrifuge. The centrifuge was operated at 16,000 rpm 

(relative centrifugal force (RCF) of 15,000 gravities) 

and is described in detail by Leenheer and others 

(1989) and Rees and others (1991). To ensure high 

recovery efficiency of suspended silt, 2-L/min flow 

rate through the centrifuge was used; this rate was 

one-half that used by Ongley and Blachford (1982), 

who reported preferential losses of the organic-rich 

portion of the suspended sediment over the mineralrich portion at a flow rate of 4 L/min. Particles as 

small as 370 nm were recovered in the centrifuge 

(Rees and Ranville, 1990).

All exposed centrifuge surfaces were coated 

with or directly machined from TFA or FEP Teflon to 

minimize sample contamination. Porous TFE Teflon 

was avoided throughout the sample isolation because 

it absorbs trace organic compounds and can retain 

contamination between sites. The suspended silt from 

a total of 500 to 1,000 L of water was deposited on the 

cylindrical wall of the centrifuge bowl, which was 

lined with a removable sheet of 0.35-mm-thick FEP 

Teflon. The liner was easily transferred from the bowl, 

using Teflon tweezers and FEP Teflon gloves, into an 

FEP Teflon sample bag, where the silt sample was 

resuspended by gentle massage of the liner from 

outside the bag. The suspended silt from the 500 to 

1,000 L of river water was then contained in 1 to 2 L 

of water. This solution was transferred to 1-L glass 

jars, preserved with five drops of chloroform, 

transported on ice to the laboratory, allowed to settle 

for at least 1 week, aspirated to remove the 

supernatant, and air dried in the original container. 

The air-dried silt resulting from the sampling process 

was ground in a glass mortar, weighed, and mixed 

well. A portion was taken for organic-carbon analysis.

Colloid remaining in the aqueous phase was 

isolated by ultrafiltration through 30,000-molecularweight cellulose filters. The colloid on the filter plates 

was placed in Teflon bags and resuspended using 

gentle external massage. The approximately 2 L of 

colloid in aqueous suspension was transferred to a 2-L 

Teflon bottle and transported on ice to the laboratory 

after two drops of chloroform was added. The sample 

was transferred to freeze-dry flasks, frozen, and freeze 

dried. A small portion of the fluffy colloid was used 

for organic carbon analysis.

RESULTS

The colloid accounted for about 10 percent by 

weight of the total suspended matter (silt and colloid) 

for most of the river in the summer. In the fall, the 

proportion of colloid increased to 20 percent at most 

sites and 35 percent at two upper river sites. During 

spring, the percentage of colloid by weight of the total 

suspended matter remained high in the upper river, but 

decreased to 5 percent in the lower river.

The organic carbon content, in percent, 

associated with the silt and colloid at each site is 

shown in figure 2a for the July-August 1991 cruise, in 

figure 2b for the October-November 1991 cruise, and 

in figure 2c for the April-May 1992 cruise. The 

samples were analyzed by Huffman Laboratories, 

Golden, Colo. The organic carbon content of 

duplicate samples varied less than 2 percent from the 

mean value.

At almost every site, the organic carbon content associated with the colloid (4 to 31 percent) was significantly higher than that associated with the silt (1 to 17 

percent). The organic carbon associated with the silts 

averaged 4.1 percent in summer, 5.5 percent in fall, 

and 4.9 percent in spring; the organic carbon associated with the colloids averaged 16.5 percent in summer, 16.1 percent in fall and 13.0 percent in spring. 

Tributaries with high suspended-sediment concentration, the Minnesota River (site 2) and the Missouri 

River (site 11), usually had lower organic carbon content, associated with both the silt and the colloid, than 

did other nearby sites. During April-May 1992, the 

organic carbon content was consistently higher at the 

upstream sites (sites 1-7 on figure 2) than at the downstream sites, especially with respect to the colloid. 

The Ohio River, which supplies as much as half of the 

water to the lower Mississippi River, carried colloid 

with higher organic carbon content than that of nearby 

sites.

827
A. Percent Organic Carbon associated with suspended silt and 

colloid, July-August 1991

35

30

25

20

15 H

10

5 

n .1

r

n

1 i

 1 i

~

i

 s

nc

in -

 I . i .

ILT

OLLOID

.V

1 2* 3 4 5 6 7 8 9 10* 11* 12 13* 14 15 16

SAMPLE SITE

B. Percent Organic Carbon associated with suspended silt and 

colloid, October-November 1991

35

30

25

20

15

, 

5^ 1 1

n

 SILT

D COLLOID

I

i I I

Lh 1 1 1

1 2* 3 4 5 6 7 8 9 10* 11* 12 13* 14 15 16

SAMPLE SITE

C. Percent Organic Carbon associated with suspended silt and 

colloid, April-May 1992

35 

30 - 

25 

20 

15 - 

10 - 

5 - 

n -,r|

| i r

i

 '

1 1

r~ 

i

~i r n

1 j - -

 SILT 

D COLLOID

 i

n. jijui

1 2* 3 4 5 6 7 8 9 10* 11* 12 13* 14 15 16

SAMPLE SITE

Figure 2. Percent organic carbon associated with the silt and colloid (A) July-August 1991, (B) October-November 1991 

(C) April-May 1992.

828
The organic carbon content associated with the 

colloids averaged 15.2 percent. This percentage is 

substantially higher than that assumed in previous 

colloidal studies. Estimates of colloidal organic carbon 

content by Pankow and McKenzie (1991) were 10 

percent for river headwaters and lake water, and 

5 percent for rivers downstream of agriculture. The 

river headwater in this study (site 1) had colloidal 

organic carbon content averaging greater than 

25 percent, when not under ice in April-May. Colloid 

from sites along the lock-and-dam lake system, such as 

sites 4-7, had consistently higher organic carbon content 

than colloid from the free-flowing river sites 14-16.

Hydrophobic or non-water-soluble compounds 

which associate with organic carbon will have more 

organic carbon available associated with the colloid 

than with the silt in the river. It is likely that the 

colloid will contain higher amounts of these 

compounds than the silt.

SUMMARY

Organic carbon associated with silt and 

colloid are reported from 16 sites along the 

Mississippi River and some of its tributaries; the 

samples were collected during July-August 1991, 

October-November 1991, and April-May 1992. 

This is the first time that the organic carbon 

associated with the suspended silt and colloid has 

been reported for the entire reach of the Mississippi 

River from near Minneapolis, Minnesota, to south 

of New Orleans, Louisiana. Colloidal organic 

carbon content has not been reported previously. 

Because hydrophobic compounds partition into 

organic carbon, the colloidal transport of 

hydrophobic toxic organic compounds may be at 

least as important as their silt transport in the 

Mississippi River.

REFERENCES

Leenheer, J.A., Meade, R.H., Taylor, H.E., and Pereira, 

W.E., 1989, Sampling, fractionation, and dewatering of 

suspended sediment from the Mississippi River for 

geochemical and trace-contaminant analysis, in 

Mallard, G.E., and Ragone, S.E., eds., U.S. Geological

Survey Toxic Substances Hydrology Program 

Proceedings of the technical meeting, Phoenix, 

Arizona, September 26-30, 1988: U.S. Geological 

Survey Water-Resources Investigations Report 88- 

4220, p. 501-511.

Moody, J.A., and Meade, R.H., 1992, Hydrologic and 

sedimentologic data collected during three cruises at 

low water on the Mississippi River and some of its 

tributaries, July 1987-June 1988: U.S. Geological 

Survey Open-File Report 91-485, 143 p.

Moody, J.A., and Meade, R.H., 1995, Hydrologic and 

sedimentologic data collected during three cruises on 

the Mississippi River and some of its tributaries, from 

Minneapolis, Minnesota, to New Orleans, Louisiana, 

July 1991 to May 1992: U.S. Geological Survey, 

Open-File Report 94-474.

Ongley, E.D., and Blachford, D.P, 1982, Application of 

continuous flow centrifugation to contaminant analysis 

of suspended sediment in fluvial systems: 

Environmental Technology Letters, v. 3, no. 5, p. 219- 

228.

Pankow, J.F., and McKenzie, S.W., 1991, Parameterizing 

the equilibrium distribution of chemicals between the 

dissolved, solid particulate matter, and colloidal matter 

compartments in aqueous systems, Environmental 

Science and Technology, v. 25, No. 12, p. 2046-2053.

Pereira, W.E., and Rostad, C.E., 1990, Occurrence, 

distributions and transport of herbicides and their 

degradation products in the lower Mississippi River 

and its tributaries: Environmental Science and 

Technology, v. 24, no. 9, p. 1400-1406.

Pereira, W.E., Rostad, C.E., and Leiker, T.J., 1990,

Distribution of agrochemicals in the lower Mississippi 

River and its tributaries: Science of the Total 

Environment, v. 97/98, p. 41-53

Rees, T.F., Leenheer, J.A., and Ranville, J.F., 1991, Use of a 

single-bowl continuous-flow centrifuge for 

dewatering suspended sediments effect on sediment 

physical and chemical characteristics: Hydrological 

Processes, v. 5, p. 201-214.

Rees, T.F., and Ranville, J.F., 1990, Collection and analysis 

of colloidal particles transported in the Mississippi 

River, U.S.A.: Journal of Contaminant Hydrology, 

v. 6, p. 241-250.

829
830
Accumulation of Organochlorine Compounds in 

Hydrilla Verticillata Relative to Sediments in the Tidal

Potomac River

By Jessica A. Hopple1 and Gregory D. Foster2

Abstract

Submersed aquatic plants are a sink for 

hydrophobic organic compounds. An average 

accumulation of 71 micrograms per kilogram of 

polychlorinated biphenyls was measured in 

Hydrilla verticillata collected from the freshwater 

reach of the tidal Potomac River, and an average 

of 53 micrograms per kilogram was measured in 

the bed sediment at the same location. Hydrilla 

verticillata accumulated an average of 1.60 

micrograms per kilogram and the bed sediment 

accumulated an average of 0.77 micrograms per 

kilogram for each of the five chlorinated 

pesticides measured. Accumulation of 

chlorinated pesticides and polychlorinated 

biphenyls in Hydrilla verticillata was found to be 

a significant factor in the redistribution of 

contaminants in the tidal Potomac River.

INTRODUCTION

Organochlorine pesticides and polychlorinated 

biphenyls (PCB's) are nonpolar, hydrophobic organic 

compounds with large octanol/water partition coefficients (log KoW>4) and have been shown to partition 

from water onto sediments in aquatic systems (Elzerman and others, 1987). Stamer and others (1984) 

reported mean PCB concentrations of 152 jag/kg in the 

bed sediments of the Schuylkill River Pennsylvania, 

where the concentrations in water were usually below 

detection limits (1.0 jag/kg).

'U.S. Geological Survey, Reston, VA 

2George Mason University, Fairfax, VA

The Potomac River, Washington, D.C., is the 

second largest source of water to the Chesapeake Bay, 

contributing 18 percent of the total freshwater flow 

(Hickman, 1984). The tidal river has a deep channel 

with wide shoals on either side; the sides contain 

abundant submersed aquatic vegetation (Carter and 

others, 1987). An estimated 26 to 30 kg of PCB's are 

transported from the upper Potomac River to the tidal 

Potomac River each year (Foster, 1992).

The purpose of this study was to explore the 

accumulation of chlorinated pesticides and polychlorinated biphenyls (PCB's) in Hydrilla verticillata relative to bed sediments in the tidal Potomac River. 

Samples were collected from two sites in October 

1991 and July 1992 and analyzed for PCB's and 

selected Organochlorine pesticides using gas chromatography. The results of this study will help to determine whether the accumulation of the contaminants in 

plants will have an effect on the transport of hydrophobic organic compounds from the upper Potomac River 

to the Chesapeake Bay. Autumn senescence of 

Hydrilla verticillata and transport downstream of the 

released particulate material may be an important 

means of bulk transport from the area.

MATERIALS AND METHODS

Samples were collected from two sites in the 

tidal Potomac River south of Washington, D.C., 

Piscataway Bay and Greenway Flats (fig. 1). Piscataway Bay is located about 13 km south and Greenway 

Flats is located about 22 km south of the Blue Plains 

Sewage Treatment Plant on the east side of the Potomac River. Piscataway Bay is located near the mouth 

of Piscataway Creek.

831
77*15'

Blue Plains Sewage 

Treatment Plant

Piscataway Bay site 

P92, P91. PS91

Greenway Flats site

G92, G91, GS91 

dian Head

Figure 1. Map showing location of sampling sites in the 

Potomac River (modified from Carter, 1987, p. 4).

Hydrilla verticillata and sediment samples were 

collected in October 1991 from Piscataway Bay and 

Greenway Flats. Hydrilla verticillata was collected 

from the same sites in July 1992. The plant samples 

were collected with a rake and wrapped in aluminum 

foil. The top 5 cm. of bed sediment within the Hydrilla verticillata beds was collected using a sediment 

corer and the sediment placed in baked amber glass 

bottles with Teflon lined caps. All samples were 

stored in an ice filled cooler until placed in a freezer at 

-20 °C.

Samples were lyophilized in a freeze dryer with 

a 2-propanol/dry ice slush trap to prevent back flushing of rough pump oil vapors. Lyophilized sediment 

samples were stored in a freezer at -20 °C after removing snails and other foreign material from samples. 

The dried plant material was homogenized in a stainless steel blender, placed in baked amber glass bottles 

with Teflon lined caps, and stored in a freezer at -20 °C 

until analyzed.

Plant and sediment samples (n>3) were analyzed according to the procedure developed for tissue 

analysis for the U.S. Geological Survey National 

Water Quality Assessment program (Shan, 1992). 

Samples were analyzed for oxychlordane, a-chlordane, a-chlordane, dieldrin, 4,4'-DDT, and the sum of 

the PCB congeners identified.

RESULTS AND DISCUSSION

An average of 90 |ig/kg of total PCB's were 

found in the Piscataway Bay Hydrilla verticillata 

collected in 1992 and 70 ng/kg in samples collected in 

1991 (fig. 2). Greenway Flats Hydrilla verticillata 

accumulated less than Piscataway Bay with 66 fig/kg 

and 57 fig/kg total PCB's in 1992 and 1991, respectively. Piscataway Bay sediment accumulated 

83 fig/kg total PCB's ~ 13 fig/kg more than Hydrilla 

verticillata from the same site in 1991, but less than 

Hydrilla verticillata from 1992. The concentrations 

were higher in the Piscataway Bay sediments than in 

the Greenway Flats sediments, which contained 

23 ^ig/kg total PCB's.

Hydrilla verticillata from Piscataway Bay in

1991 and 1992 accumulated as much as 10 (ig/kg of 

the chlorinated pesticides measured if DOT is 

excluded from consideration (fig. 3). Greenway Flats 

plant concentrations were 7.0 (ig/kg and 5.3 Mg/kg for

1992 and 1991, respectively. Piscataway Bay sediment contained 6.2 ng/kg, less than the 1991 and 1992 

Hydrilla verticillata measurements from the same site. 

Sediment samples from Greenway Flats contained 

1.5 M£/kg, the lowest concentration of chlorinated 

pesticides in this study.

DDT was difficult to quantify, perhaps 

because of thermal decomposition and (or) affinity 

with active sites in the gas chromatographic system 

where DDT can degrade (Bill Foreman, National 

Water Quality Laboratory, U.S. Geological Survey, 

oral commun., 1993). Concentration of DDT should 

be viewed with this in mind.

100

80

60

40

0 20

 5

PISCATAWAY BAY GREENWAY FLATS 

EXPLANATION 

111992 HYDRILLA d 1991 HYDRILLA M 1991 SEDIMENT

Figure 2. Total PCB concentations for Hydrilla verticillata 

and sediment samples collected at Piscataway Bay and 

Greenway Flats, Potomac River, Washington, D.C., 

October 1991 and July 1992.

832
g

14.3

I I

iii 

= 1 =

P92 P91

[~~1 Oxychlordane 

H Dieldrin

PS91 G92 

SAMPLE IDENTIFIER

EXPLANATION 

[U 6-Chlordane

G91

ot-Chlordane

GS91

Figure 3. Pesticide concentrations for Hydrilla verticillata 

and sediment samples collected at Piscataway Bay and 

Greenway Flats, Potomac River, Washington, D.C., October 

1991 and July 1992; P92, Piscataway Bay 1992 Hydrilla 

verticillata; P91, Piscataway Bay 1991 Hydrilla verticillata; 

PS91, Piscataway Bay 1991 sediment; G92, Greenway 

Flats 1992 Hydrilla verticillata; G91, Greenway Flats 1991 

Hydrilla verticillata; GS91, Greenway Flats 1991 sediment.

In 1990, areal coverage of submersed aquatic 

plants in the freshwater tidal Potomac River was estimated at 1,652 hectares (Nancy Rybicki, U.S. Geological Survey, oral commun., 1993). A calculation of the 

mass of PCB's accumulated in submersed aquatic 

plants in 1990 can be made by using the mean biomass 

of plants (180 g/m2) for the calendar year 1984 (Carter 

and Rybicki, 1986). If plant material accumulates as 

much as 71 (ig/kg of total PCB's, the average of the 

four Hydrilla verticillata sample sets, then about 

0.21 kg of total PCB's could be transported from the 

tidal Potomac River each year during autumn senescence.

CONCLUSIONS

The results of this study demonstrate that 

Hydrilla verticillata in the tidal Potomac River is 

accumulating PCBs and chlorinated pesticides.

Accumulation of pesticides and PCBs in 

Hydrilla verticillata is a significant factor in the 

redistribution of organochlorine contaminatnts in the 

tidal Potomac River.

Autumn senescence and transport downstream 

of the released leaf material is an important means of 

bulk transport from the tidal Potomac River towards 

the Chesapeake Bay.

REFERENCES

Carter, V., and Rybicki, N.B., 1986, Resurgence of

submersed aquatic macrophytes in the tidal Potomac 

River, MD, VA, and the District of Columbia: 

Estuaries, v. 9, no. 4b, p. 368-375.

Carter, V., Rybicki, N.B., Jones, R.C., Barko, J.W., Dresler, 

P.V., Hickman, R.E., Anderson, R.T., 1987, Data on 

physical, chemical, and biological characteristics of 

Hydrilla beds, mixed vegetation beds, and unvegetated 

sites in the tidal Potomac River, Maryland and 

Virginia, 1987: U.S. Geological Survey Open-File 

Report 88-709.

Elzerman, A.W., Coates, J.T., 1987, Hydrophobic Organic 

Compounds on Sediment: Equilibria and Kinetics of 

Sorption: Sources and Fates of Aquatic Pollutants, 

p. 263.

Foster, G.D., 1992, Chesapeake Bay Fall Line Loadings 

Survey: Final Report submitted to the State of 

Maryland Department of the Environment, 36 p.

Hickman, R.E., 1984, Loads of Suspended Sediment and 

Nutrients from local nonpoint sources to the tidal 

Potomac River and Estuary, Maryland and Virginia, 

1979-91 water years: U.S. Geological Survey OpenFile Report 84-860, 82 p.

Shan, T.A., 1992, Multiresidue Analysis of Organohalogen 

Contaminants in Tissues of Aquatic Organisms: 

unpublished Master's Thesis, George Mason 

University, Fairfax, Va., 56 p.

Stamer, J.K., Yorke, T.H., and Paterson, G.L., 1984, 

Distribution and transport of trace substances in the 

Schuylkill River basin from Berne to Philadelphia, 

Pennsylvania: U.S. Geological Survey Water-Supply 

Paper 2256.

833
834
Pathways of Methanogenic Biodegradation of CreosoteDerived Aromatic Compounds

By E. Michael Godsy 1 , Donald F. Goerlitz 1 , and Dunja Grbic'-Galic'2

Abstract

The fate of organic compounds in ground water 

is controlled by various transport and biotransformation processes. Possibly the most important, but 

currently the least understood, process affecting 

ground-water quality is biotransformation of 

organic compounds by indigenous microorganisms. 

In this study, the degradation pathways of benzothiophene, quinoline, and naphthalene are determined on the basis of intermediate compounds that 

appear before and just after the onset of methanogenesis. This study revealed that the biodegradation 

process consists of both a major and a minor pathway. The first transformation step of the majorpathway heterocyclic compounds is oxidation and 

cleavage of the heterocyclic ring. After cleavage of 

this ring, the substituent side chains and the remaining homocyclic ring are subjected to various reactions, including oxidation, decarboxylation, desulfurylation or annnonification, and O-methylation. 

These reactions are followed by the reduction of the 

homocyclic ring, cleavage of this ring, p-oxidation, 

and mineralization. The major pathway intersects 

both the benzoic acid and phenol methanogenicdegradation pathways. A minor pathway for heterocyclic compounds starting with the oxidation of the 

homocyclic ring with subsequent ring reduction, 

ring cleavage, degradation of the remaining heterocyclic ring, and mineralization also was observed.

INTRODUCTION

The fate of homocyclic and heterocyclic aromatic 

compounds present in creosote and commonly found 

in subsurface environments is controlled by various

'U.S. Geological Survey, Melo Park, Calif. 

2Stanford University, Stanford, Calif.

transport and biotransformation processes. The most 

important, but one of the least understood processes 

affecting ground-water quality, is biotransformation of 

these contaminants by indigenous microorganisms. 

Even less is known about these transformations under 

methanogenic conditions. Several homocyclic and heterocyclic aromatic hydrocarbons, which contain nitrogen and sulfur in their ring structure, have been shown 

to be susceptible to transformations under methanogenic 

conditions (Pereira and others, 1987) and to be biodegradable to methane (CH4) and carbon dioxide (CO2) 

(Berry and others, 1987a; Godsy and others, 1989; 

Godsy and Grbic'-Galic', 1989); the current knowledge 

has been summarized in four recent review papers 

(Berry and others, 1987a; Grbic'-Galic', 1989; Grbic'- 

Galic', 1990; Schink and others, 1992). In this paper, the 

methanogenic pathways of biodegradation for benzothiophene, indole, quinoline, indene, and naphthalene 

are investigated. The biodegradation pathways for benzothiophene, quinoline, and naphthalene are inferred 

from the analysis of intermediate compounds that appear 

in the growth liquor at the onset and during methanogenesis in laboratory microcosms and from compounds 

appearing in ground-water samples. No intermediates 

were detected during methanogenesis of indole and 

indene except for the end products CH4 and CO2 .

MATERIALS AND METHODS 

Sample Site

The study site is located adjacent to an abandoned 

wood-preservation plant within the city limits of Pensacola, Florida (Godsy and others, 1992). The wood-preservation process consisted of steam-pressure treatment 

of pine poles with creosote or pentachlorophenol (PCP). 

For more than 80 years, large but unknown quantities of 

wastewaters consisting of extracted moisture from the 

poles, cellular debris, creosote, PCP, and diesel fuel from 

the treatment processes were discharged to nearby surface impoundments. These impoundments were unlined 

and in direct hydraulic contact with the sand-and-gravel

835
aquifer. Contamination of the ground water resulted 

from the accretion of wastes from these impoundments 

(Mattraw and Franks, 1986).

Laboratory Microcosms

Large methanogenic microcosms were constructed 

in an anaerobic glove box containing a 10 percent 

hydrogen (H2)-90 percent argon (Ar) atmosphere. The 

microcosms were prepared in 4-L glass bottles containing approximately 4 kg of contaminated aquifer material collected at the site near the centroid of the 

contamination. Microcosms were filled with 2.5 L of 

anaerobic mineral salts solution composed of the following (per liter): KH2PO4,0.75 g; K2HPO4 ,0.89 g; 

MgCl2-6 H20,0.36 g; NH4C1,0.9 g; trace-metal solution (Zeikus, 1977), 9.0 mL ; vitamin solution (Wolin 

and others, 1963), 5.0 mL. The pH was adjusted to 6.0, 

and the medium was then boiled, cooled and dispensed 

under a stream of oxygen (O2)-free Ar gas. The medium 

was then sterilized at 121 °C for 15 minutes. Control 

microcosms for each of the tested compounds were 

autoclaved for 45 minutes and used as killed-cell controls. A positive control was prepared using only the 

contaminated aquifer material and 2.5 L of anaerobically prepared mineral-salts solution to determine if 

organic material potentially sorbed onto the aquifer 

material was biodegradable under the aquifer conditions. One mL/L of sterile amorphous ferrous sulfide 

(FeS) was aseptically added to all microcosms as a 

reducing agent (Brock and O'Dea, 1977).

Microcosms were sampled every 1 to 2 days and 

analyzed by gas chromatography/mass spectrometry 

(GC/MS) and by high-performance liquid chromatography for intermediates as described by Godsy and 

Grbic'-Galic' (1989). Only those compounds that 

appeared in viable microcosms, but not in either control, 

are reported as intermediates. Ground-water samples 

were analyzed for possible intermediate compounds by 

GC/MS as described by Godsy and others (1992). The 

presence of CH4 and CO2 in the microcosms was confirmed by the gas-chromatographic head-space method 

described by Godsy and others (1992).

RESULTS

Degradation Pathway for Sulfur-Heterocycles

Compounds detected in the growth liquor just 

before the onset and during methanogenesis of benzothiophene are shown in figure 1. Compounds that

were found in the autoclaved and organic free controls 

are not shown and presumably do not arise from the 

degradation of the previously mentioned compounds. 

More than 33 intermediate compounds were detected in 

the growth liquor during degradation of this compound. 

On the basis of the study by Pereira and others (1988), 

who demonstrated that the first step in the oxidation of 

quinoline consisted of oxidation of the heterocyclic 

ring, the sulfur (S)-heterocyclic ring may be oxidized to 

2-benzothiophenone or the tautomer benzothiophene2-ol. However, this compound was not detected in the 

microcosms or in ground-water samples. After hydrolytic cleavage of this ring, the substituent side chains 

and remaining homocyclic rings were possibly subjected to various reactions including oxidation, decarboxylation, desulfurylation, and O-methylation with 

ultimate conversion to benzoic acid and phenol. These 

reactions were most likely followed by the previously 

described pathways (Berry and others, 1987a) consisting of the reduction of the homocyclic ring, corroborated by the appearance of cyclohexyl alcohol and 

cyclohexanecarboxylic acid, cleavage of this ring, (3-oxidation, and mineralization to CH4 , CO2, and hydrogen 

sulfide (H2S). A minor pathway, based on relative abundance of intermediates of benzothiophene, was 

observed that started with the oxidation of the homocyclic ring. The first implied intermediate compound was 

not observed in the microcosms but has been found in 

ground-water samples. This reaction is presumably 

followed by the hydrolytic cleavage of the homocyclic 

ring to thiophene-2-ol. Aliphatic sulfur-containing 

intermediates of the thiophene-2-ol transformation 

were also detected in the growth medium, indicating 

further degradation of the remaining S-heterocyclic 

ring, and mineralization.

Degradation Pathway for NitrogenHeterocyclic Compounds

Compounds detected in the growth liquor just 

before the onset and during methanogenesis of quinoline are shown in figure 2. Pereira and others (1988) 

demonstrated that the first step in the oxidation of quinoline consisted of oxidation of the heterocyclic ring; the 

major pathway involves the hydrolytic cleavage of the 

oxidized ring. The transient appearance of aniline in the 

growth liquor could indicate that this compound is 

involved in the major pathway. The minor homocyclicring-oxidation pathway observed during the methanogenesis of benzothiophene also was possibly occurring 

during the methanogenesis of quinoline. The oxidized

836
OXIDATION OF THE

HOMOCYCLIC

RING

OXIDATION OF THE

HETEROCYCLIC

RING

H

H 'H <35

4-Benzothiophenone Benzothiophene-4-ol Benzothiophene-2-ol 2-Benzothiophenone

HYDROLYTIC CLEAVAGE OF 

THE OXIDIZED RING

Thiophene-2-ol

RING CLEAVAGE

Diethyl Disulfide 

Ethane-dithioic Acid 

Thiopropionic Acid

4-Hydroxybenzene Sulfonic Acid

a -Hydroxyphenylacetic Acid

a -Oxophenylacetic Acid

Benzyl Alcohol

Toluene 

Benzoic Acid

Phenol

Cyclohexanecarboxylic Acid 

Cyclohexyl Alcohol

i i 

v

3-Hexene-l-ol, 2-Hexenoic Acid

Hexanoic Acid, Pentanoic Acid

Pentanol, 2-Methyl-l,2-propanediol

4-Butanol, Propionic Acid

ACETOGENESIS 

Acetic Acid, Formic Acid

4

METHANOGENESIS

CH4 C02 H2S

1r

® Proposed intermediate - not detected in microcosms but present in ground-water samples.

* Proposed intermediate - not detected in microcosms or in ground-water samples.

ft Hydrogen sulfide analyses were not performed.

Figure 1. Proposed pathways of benzothiophene transformation under methanogenic conditions in 

aquifer-derived microcosms. All intermediates shown were detected in the culture fluid except for those 

marked by asterisks.

837
OXIDATION OF THE

HOMOCYCLIC

RING

OXIDATION OF THE

HETEROCYCLIC

RING

8-Quinolinone

OH 

8-Quinolinol 2-Quinolinol

H 

2(lH)-Quinolinone

HgC

2,3-Dimethylpyridine 

\

\?

\ \ \

HYDROLYTIC CLEAVAGE OF 

THE OXIDIZED RING

Aniline

Benzoic Acid 

Phenol

RING CLEAVAGE 

ACETOGENESIS

Acetic Acid, Formic Acid

1

METHANOGENESIS 

CH4 C02

® Proposed intermediate - not detected in microcosms but present in ground-water samples, 

ft Ammonia analyses were not performed.

Figure 2. Proposed pathways of quinoline transformation under methanogenic conditions in aquifer-derived microcosms. All intermediates shown were detected in the culture fluid except for those 

labeled by asterisks.

intermediate of the minor pathway, 8-quinolinone <-> 

quinoline-8-ol, was not detected in the microcosms but 

was found in many of the ground-water samples from 

the study site. After hydrolytic cleavage of the oxidized 

homocyclic ring, the remaining compounds were probably subjected to similar reactions that were observed 

during methanogenesis of benzothiophene. This was 

corroborated by the appearance of 2,3-dimethylpyridine. The minor pathway was proposed on the basis 

of the breakdown products that were identified for 

benzothiophene; however, other than 2,3-dimethylpyridine, no other nitrogen-containing aliphatic 

intermediates were identified for quinoline in the

microcosms. Notably, 2,3-dimethylpyridine did not 

persist in the microcosms and does not persist during 

down-gradient travel in the aquifer at the study site and 

is believed to be mineralized to CH4 and CO2 .

During the methanogenesis of indole in the 

microcosms, oxindole was the only intermediate that 

was observed before complete conversion to CH4 and 

CO . Isatin and anthranilic acid, however, were 

detected in ground-water samples from the research 

site. These compounds have been shown to be intermediates in the degradation of indole under denitrifying and, presumably, methanogenic conditions 

(Madsen and Bollag, 1989).

838
Naphthalene

OXIDATION AND

HYDROLYTIC CLEAVAGE OF

ONE RING

I

'OH 

2-Ethylphenol

Ground-water samples from the site contained both 

1-naphthalenol and 2-naphthalenol, possible candidates 

for the first oxidation step for naphthalene. It is not 

certain at this time whether these particular compounds 

arose from the degradation of naphthalene or came from 

creosote itself.

Benzofuran was also observed in the growth 

medium of naphthalene, but the role of this compound 

is inconclusive from these studies. This compound, 

however, probably resulted from a reaction involving 

the hydroxyl-group of 2-hydroxyphenylacetaldehyde 

(an oxidation intermediate of 2-ethylphenol) and the 

carbonyl carbon of the aldehyde group, followed by 

dehydration to yield a furan as proposed by Grbic'- 

Galic'(1989).

RING CLEAVAGE

ACETOGENESIS

Acetate

METHANOGENESIS 

CH4 C02

Figure 3. Proposed pathway of 

naphthalene transformation under 

methanogenic conditions in groundwater aquifer-derived microcosms.

Degradation Pathway for Homocyclic 

Compounds

Intermediate-compound analyses for indene and 

naphthalene transformation were done in the same manner as for the heterocyclic compounds. The analyses on 

microcosms containing indene did not reveal any intermediates other than the end products CH4 and CO2 . The 

microbes in the aquifer sediments used as the inoculum 

were presumably very active toward this compound. 

The analyses for intermediate compounds of naphthalene (fig. 3) revealed 2-ethylphenol, a compound that 

would be in the major degradation pathway resulting 

from the hydrolytic cleavage of one of the naphthalene 

rings. This intermediate would subsequently be subjected to various reactions similar to the above, resulting 

in CH4 and CO2 as the end products.

DISCUSSION AND CONCLUSIONS

Environmental contamination by homocyclic and 

heterocyclic aromatic hydrocarbon compounds is 

widespread; however, little information is available on 

the biotransformation pathways of this class of compounds under methanogenic conditions and even less 

on the environmental fate of these compounds in contaminated ground water. In general, recent studies on 

the methanogenic transformation of aromatic compounds have demonstrated that three major-ringcleavage transformation pathways are used (Young, 

1984; Berry and others, 1987a; Evans and Fuchs, 1988; 

Schink and others, 1992): (1) the benzoic acid, (2) the 

resorcinol, and (3) the phloroglucinol pathways. In 

older reviews, pathways for phenol degradation were 

proposed that proceeded through complete ring saturation to cyclohexanol, conversion to cyclohexanone, 

and subsequent ring cleavage (Evans, 1977; Young, 

1984). These pathways were based on simple analogy 

to the benzoate pathway; however, no experimental 

evidence of their existence has ever been found. An 

alternative pathway for phenol was postulated through 

which the quinoid phenol tautomer was reduced to 

form cyclohexanone. This mechanism takes advantage 

of the polarizing effect of the hydroxyl group on the 

Ti-electron distribution in the aromatic ring (Bakker, 

1977; Schink and Tschech, 1988). Recently, evidence 

has been presented that links the phenol with the benzoate pathway (Knoll and Winter, 1989; Sarak-Genther 

and others, 1989; Haggblom and others, 1990). The 

carboxylation of phenol to 4-hydroxybenzoate was 

proposed as the first step in the degradation. This step 

was then followed by reduction of the hydroxyl group

839
and removal to form benzoate. Because the groundwater consortia used for this part of the study were 

mixed ground-water cultures, it is not surprising that 

the major transformation pathways for benzothiophene 

and quinoline intersected segments of both the benzoate or the phenol pathway. It is not known at this 

point whether phenol or benzoate is more important as 

the key intermediate in the methanogenic transformation of quinoline and benzothiophene in these microcosms; however, Grbic'-Galic' (1989) found that 

phenol was the most persistent simple aromatic intermediate of the methanogenic degradation of naphthalene in a fed-batch reactor. It is also not known at this 

time whether phenol was then degraded by the pathway 

suggested by Schink and Tschech (1988) or by carboxylation to 4-hydroxybenzoate and then degradation via 

the benzoate pathway.

The major pathways for anaerobic transformation 

of benzothiophene and quinoline seem to be initiated 

by the oxidation and then cleavage of the nitrogen 

N-or S-heterocyclic ring. Delocalization of the rc-electrons in the heterocyclic ring results in the heteroatom 

being relatively electron depleted and the carbons in 

the ring being relatively electron enriched. The carbon 

atom proximal to the heteroatom, as a result of the electron delocalization, is then available to nucleophilic 

attack by water. This observation is in line with the 

findings of Berry and others (1987b), who demonstrated that the first step in the methanogenic transformation of indole, the nitrogen analog of 

benzothiophene, was the oxidation of the N-heterocyclic ring to form oxindole. Similarly, Pereira and others 

(1988) were able to demonstrate the oxidation of the 

N-heterocyclic ring of quinoline using H2 I8O to form 

l8O-labeled 2(lH)-quinolinone under methanogenic 

conditions; however, they were not able to demonstrate 

methanogenic fermentation of this compound, probably due to their use of reasonably fresh sewage sludge 

as the inoculum for the microcosms. In this work, it 

was found that the substituent side chains attached to 

the homocyclic ring after the cleavage of the heterocyclic ring were subjected to various transformation 

reactions, as exemplified by the myriad of single-ring 

compounds found. The large number of compounds is, 

undoubtedly, a result of the activity of a complex 

ground-water consortium.

The other, as yet unreported minor pathway for 

heterocycles involves the initial oxidation of the 

homocyclic ring, hydrolytic cleavage of this ring, and 

then attack on the heterocyclic ring. Several years ago,

Vogel and Grbic'-Galic' (1986) showed that the oxidation of an unsubstituted homocyclic aromatic ring was 

possible when they demonstrated the oxidation of benzene to CH4 and CCK by a mixed methanogenic culture. 

The magnitude, importance, and organisms involved in 

the minor transformation pathway for benzothiophene, 

quinoline, and other creosote-derived compounds are 

unknown at this time; however, the mechanism and 

organisms involved in the minor pathway might be similar to those initiating the attack on indene, naphthalene, 

benzene, and toluene. The oxidation of the homocyclic 

ring at positions 5 or 8 for quinoline, or at positions 4 or 

7 for benzothiophene, are consistent with observations 

of organic synthesis reactions, where the carbons of the 

homocyclic ring proximal to the heterocyclic ring are 

most susceptible to attack (Vollhardt, 1987).

The inability to identify intermediate compounds 

during the methanogenesis for some compounds (for 

example, benzofuran and indene) may be because of 

low initial concentrations of the compounds in the 

microcosms. Compounds that could be intermediates 

of these were observed in ground-water samples. 

3(2H)-benzofuranone and both 2,3-dihydro-indene-lone and l,3-dihydro-indene-2-one were observed in 

ground-water samples and could be intermediates in 

the degradation of benzofuran and indene, respectively.

REFERENCES CITED

Bakker, G., 1977, Anaerobic degradation of aromatic compounds in the presence of nitrate: FEMS Letters, v. 1, 

p. 103-108.

Berry, D.F., Francis, A.J., and Bollag, J.M., 1987a, Microbial 

metabolism of homocyclic and heterocyclic aromatic 

compounds under anaerobic conditions: Microbiological 

Reviews, v. 51, p. 43-59.

Berry, D.F., Madsen, E.L., and Bollag, J.M., 1987b, Conversion of indole to oxindole under methanogenic conditions: Applied and Environmental Microbiology, 

v. 53, p. 180-182.

Brock, T.D., and O'Dea, K., 1977, Amorphous ferrous 

sulfide as a reducing agent for culture of anaerobes: 

Applied and Environmental Microbiology, v. 33, 

p. 254-256.

Evans, W.C., 1977, Biochemistry of the bacterial catabolism of aromatic compounds in anaerobic environments: Nature, v. 270, p. 17-22.

Evans, W.C., and Fuchs, G., 1988, Anaerobic degradation 

of aromatic compounds: Annual Reviews of Microbiology, v. 42, p. 289-317.

840
Godsy, E. M., Goerlitz, D.F., and Grbic'-Galic', Dunja, 

1989, Transport and degradation of water-soluble creosote-derived compounds, in Alien, D.T., ed., Intermedia 

Pollutant Transport Modeling and Field Measurements: New York, Plenum Publishing, p. 112-143.

Godsy, E.M., and Grbic'-Galic', Dunja, 1989, Biodegradation pathways for benzothiophene in methanogenic 

microcosms, in Mallard, G.E., and Ragone, S.E., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, Phoenix, Arizona, September 26-30, 1988: U.S. Geological 

Survey Water Resources Investigations Report 

88-4220, p. 559-564.

Godsy, E.M., Goerlitz, D.F., and Grbic'-Galic', Dunja, 

1992, Methanogenic biodegradation of creosote contaminants in natural and simulated ground-water ecosystems: Ground Water, v. 30, p. 232-242.

Grbic'-Galic', Dunja, 1989, Microbial degradation of 

homocyclic and heterocyclic aromatic hydrocarbons 

under anaerobic conditions: Developments in Industrial 

Microbiology, v. 30, p. 237-253.

Grbic'-Galic', Dunja, 1990, Anaerobic microbial transformation of nonoxygenated aromatic and alicyclic compounds in soil, subsurface and freshwater sediments, in 

Bollag, J.M., and Stotzky, G., eds., Soil biochemistry, 

v. 6, New York, Marcel Dekker, p. 117-189.

Haggblom, M.M., Rivera, M.D., Bossert, I.D., Rogers, I.E., 

and Young, L.Y., 1990, Anaerobic biodegradation of 

para-creso\ under three reducing conditions: Microbial 

Ecology, v. 20, p. 141-150.

Knoll, G., and Winter, J., 1989, Degradation of phenol via 

carboxylation to benzoate by a defined, obligate 

syntrophic consortium of anaerobic bacteria: Applied 

Microbiology and Biotechnology, v. 30, p. 318-324.

Madsen, E.L., and Bollag, J.M., 1989, Pathway of indole 

metabolism by a denitrifying microbial community: 

Archives of Microbiology, v. 151, p. 71-76.

Mattraw, H.C., Jr., and Franks, B.J., 1986, Description of hazardous-waste research at a creosote works, Pensacola, 

Florida, in Mattraw, H.C., Jr., and Franks, B.J., eds., 

Movement and fate of creosote waste in ground water, 

Pensacola, Florida: U.S. Geological Survey Toxic

Waste-Ground-Water Contamination Program: U.S. 

Geological Survey Water-Supply Paper 2285, p. 1-8.

Pereira, WE., Rostad, C.E., Updegraff, D.M., and Bennett, 

J.L., 1987, Fate and movement of azaarenes in an aquifer contaminated by wood-treatment chemicals: Environmental Toxicology and Chemistry, v. 6, p. 163-176.

Pereira, WE., Rostad, C.E., Updegraff, D.M., and Bennett, 

J.L., 1988, Microbial transformations of azaarenes in 

creosote-contaminated soil and ground water: Laboratory and field studies: International Conference on 

Water and Wastewater Microbiology, Newport Beach, 

California, p. 5-1 to 5-7.

Sarak-Genther, B.R., Townsend, G.T., and Chapman, P.J., 

1989, Anaerobic transformations of phenol to benzoate 

viapara-carboxylation Use of fluorinated analogs to 

elucidate the mechanism of transformation: Biochemical and Biophysical Research Communication, v. 162, 

p. 945-951.

Schink, B., Brune, A., and Schnell, S., 1992, Anaerobic 

degradation of aromatic compounds, in Winkelmann, 

G., ed., Microbial degradation of natural products: 

Weinheim, FRG., Verlag Chemie, p. 219-242.

Schink, B., and Tschech, A., 1988, Fermentative degradation 

of aromatic compounds, in Hagedorn, S.R., Hanson, 

R.S., and Kunz, D.A., eds., Microbial metabolism and 

the carbon cycle: New York, Harwood Academic Publishers, p. 213-226.

Vogel, T.M., and Grbic'-Galic', Dunja, 1986, Incorporation 

of oxygen from water into toluene and benzene during 

anaerobic fermentative transformation: Applied and 

Environmental Microbiology, v. 52, p. 200-202.

Vollhardt, K.P.C., 1987, Organic chemistry: New York, 

W.H. Freeman and Co., p. 1179-1229.

Wolin, E.A., Wolin, M.J., and Wolfe, R.S., 1963, Formation 

of methane by bacterial extracts: Journal of Biological 

Chemistry, v. 238, p. 2882-2886.

Young, L.Y. 1984, Anaerobic degradation of aromatic 

compounds, in Gibson, D.T., ed., Microbial degradation of organic compounds: New York, Marcel Dekker, 

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Bacteriological Reviews, v. 41, p. 514-541.

841
842
Modeling Steady-State Methanogenic Degradation of 

Phenols in Ground Water at Pensacola, Florida

By Barbara A. Bekins 1 , E. Michael Godsy1 , and Donald F. Goerlitz 1

Abstract

Field and microcosm observations of methanogenic phenolic-compound degradation indicate that Monod kinetics theory governs the 

substrate disappearance but overestimates the 

observed biomass. In this paper, we present modeling results from an ongoing multidisciplinary 

study of methanogenic biodegradation of phenolic 

compounds in a sand and gravel aquifer at Pensacola, Florida. The aquifer is contaminated with 

chemicals and wastes used in wood treatment. 

Field-disappearance rates of four phenols match 

those determined in batch microcosm studies performed by other workers in a related study. The 

degradation process appears to be at steady state 

because the transported contaminants still are disappearing within 150 meters downgradient of the 

source even after a sustained influx over several 

decades. The existence of a steady-state degradation profile of each substrate together with a low 

bacteria density in the aquifer indicate that the bacterial population is exhibiting no net growth. This 

is possibly because of the oligotrophic nature of 

the biomass population in which growth and utilization are approximately independent of the phenolic-compound concentration for most of the 

concentration range. Thus, a constant bacteria 

growth rate should exist over much of the contaminated area, which may, in turn, be balanced by an 

unusually high decay or maintenance rate caused 

by hostile conditions or predation.

'U.S. Geological Survey, Menlo Park, Calif.

INTRODUCTION

The fate of toxics in anaerobic environments has 

potential applications to many ground-water 

contamination problems. Use of unlined wastewaterdisposal ponds at a former wood-treatment facility in 

Pensacola, Fla., led to contamination of a shallow sand 

and gravel aquifer by creosote, PCP, and diesel fuel 

(Goerlitz and others, 1985). A native methanogenicbacteria population has adapted to a sustained influx of 

the water-soluble fraction of these contaminants 

(Godsy and others, 1992b) such that the saturated zone 

has reached an apparent steady-state condition. The 

dissolved contaminants enter at a constant rate and are 

degraded within a fixed distance downgradient in the 

ground-water flow field.

Previous studies of continuous sources have 

obtained different results. Kindred and Celia (1989) 

did a theoretical study of one-dimensional solute 

transport in which the solute reaction rate and 

associated bacterial growth were described by Monod 

kinetics. Their results predict an unbounded increase 

in the mass of bacteria (biomass) near a steady supply 

of substrate unless the model is altered to include 

diffusion limitations caused by build-up of a thick 

biofilm. In addition, unbounded growth in response to 

continuous substrate supply has been observed as 

clogging around well bores in wastewater-injection 

experiments (Ehrlich and others, 1979). In contrast, at 

the Pensacola site, the volume of bacteria, as 

determined by Acridine Orange Direct Counts 

(AODC), is still less than 0.003 percent of the pore 

space (Godsy and others, 1992b), even though there 

has been a continuous influx of substrate for decades. 

Rittman and McCarty (1980) have shown that a steadystate substrate profile can be maintained with a biofilm 

in an aerobic-column reactor. However, in their study, 

mass transport internal to the biofilm was an important

843
rate control, whereas the density of bacteria is much too 

low for biofilm limitations at the Pensacola site.

The objectives of this paper are to (1) verify that 

the disappearance rate of phenols at the field site can be 

modeled using Monod kinetics parameters determined 

from microcosm studies by Godsy and others (1992a), 

and (2) formulate a model for steady-state 

disappearance of a constant influx of contaminants. We

concentrate on the fate of four dissolved phenolic 

compounds: phenol, and 2-, 3-, and 4-methylphenol.

SITE DESCRIPTION

The study site is an abandoned wood-treatment 

facility in the extreme western end of the Florida 

Panhandle within the city of Pensacola (fig. 1). For

87° 14' 30" 87°14'15"

30° 24' 15'

30° 24'

EXPLANATION

WATER-LEVEL CONTOUR. 

0.5 Datum is sea level. 

Interval 0.5 meters.

A A Line of section

(Section shown in fig. 2) 

! Site

0 100 200 METERS I rj r1

0 300 600 FEET

Figure 1. Location of the abandoned creosote works, test sites, and selected drainage features.

844
about 80 years, contaminants from unlined wastedisposal ponds entered the ground water in the sand 

and gravel aquifer.

The aquifer consists of approximately 90 m of 

poorly sorted fine to coarse-grained deltaic sand 

deposits interrupted by discontinuous silts and clays 

(Franks, 1988). The water table is about 1.5 m below 

the land surface. A generalized geologic section 

through the study area is shown in figure 2. The sand 

deposits have a fairly uniform average porosity of 0.38. 

A grain-size analysis of a sediment sample from a 

depth of 6 m at site 39 gave an average grain size of 

0.335 mm with a d10 = 0.1 mm and a d60 = 0.5. 

Sorption of the phenolic compounds is relatively 

unimportant, because the organic fraction of the aquifer 

material is only 0.07 percent.

30

0 100 METERS

I f

0 300 FEET

Vertical Exageration lOx

EXPLANATION 

I | Sand |||| Sandy Clay Clay

Figure 2. Generalized geologic section along line A-A'. 

(Trace shown in fig. 1)

Only the upper 30 m of the aquifer is affected by 

the wood-preserving wastes. Franks (1988) made a 

detailed study of the local ground-water flow. The 

altitude and configuration of the water table in the 

study area is shown in figure 1 (Franks, 1988). Ground 

water generally flows southward toward Pensacola 

Bay, and flow velocities range from 0.3 to 1.2 m/d; the 

higher velocities prevail at shallow depths. No vertical 

component to the flow was found except at very 

shallow depths near the stream.

Concentrations of the four phenolic compounds 

(phenol, and 2-, 3-, and 4-methylphenol) have been 

monitored at the study site for more than 12 years. The 

data indicate that the contaminants were undergoing 

steady-state degradation during this period. Data from 

6 m below the surface was chosen for the analysis 

presented in this paper because this depth is well within 

the methanogenic part of the contaminated zone; it is 

below the depth at which recharge and any floating 

hydrocarbon at the water table affect concentrations; 

and it is above the clay lenses in a zone where the 

assumption of one-dimensional flow applies (fig. 2).

MODEL DESCRIPTION

Monod kinetics describes the utilization of a 

single growth-limiting substrate by the bacteria 

population:

dS = ^m BS 

dt YKS + S

(1)

where S is the limiting substrate concentration, in milligrams per liter; fo,m is the maximum growth rate, in

day ; Y is the biomass growth yield, in milligrams 

biomass per milligram substrate; Kx is the half-saturation constant, in milligrams per liter; and B is the biomass concentration, in milligrams per liter (Monod, 

1950). We assume that the growth is carbon limited at 

this site based on inorganic analyses of the pore waters 

performed by Godsy and others (1992b). Thus, we 

use the single substrate Monod reaction term given in 

eq. 1, and in each simulation, a single phenolic compound serves as the sole carbon source.

The resulting growth of biomass due to utilization of the substrate is

dB 

dt + S

-b\ B (2)

where b is the biomass specific decay or death rate, in 

day' 1 .

Godsy and others (1992a) established laboratory 

microcosms simulating the methanogenic conditions in 

the subsurface environment and determined utilization 

rates for the four phenolic compounds. They fit the 

change in concentration with time of each phenol and 

the associated microbial growth to equations 1 and 2

845
using nonlinear regression analysis. Table 1 shows the 

Monod kinetics constants obtained from the fits for 

each compound.

We assume that the substrate profile at a depth of 

6 m satisfies the one-dimensional, transport equation 

with a Monod reaction term:

dt

_ V _ "~

dx Y®Ks +

(3)

where t is time, in days; x is distance downgradient 

from site 3; D is the dispersion coefficient; v is average 

linear velocity, in meters per day; B is milligrams of 

bacteria per liter of sediment; and 0 is the porosity. A 

one-dimensional model is appropriate for this site 

because the flow direction is primarily horizontal and 

perpendicular to a wide contaminant source. In this 

initial model, the spatial distribution of biomass, B, is 

uniform. This is consistent with results from Godsy 

and others (1992b) who found, at a depth of 6 m, a 

constant density of bacteria within the contaminated 

zone. Bekins and others (1993) showed that the 

observed substrate profile cannot exist at steady-state 

if the bacteria concentration is changing with time. 

Therefore, the model biomass value also does not 

change with time.

Equation 3 was solved numerically using a 

computer code developed by Kindred and Celia (1989). 

The boundary conditions used were

= 0, (4)

where S, is the contaminant concentration 6 m Y

below the ground surface at site 3, the closest site to the 

source. The initial condi tion used was

(5)

RESULTS AND DISCUSSION

Results of simulations indicate that the 

concentrations of phenolic compounds at the field site 

will reach steady-state profiles within 1,000 days. Two 

predicted substrate profiles, after 1,000 days along with 

measured phenolic-compound concentrations of 

Goerlitz (1992) at 6 m below land surface at each site 

are shown in figure 3A. The measured data values are 

from sites 3, 39,40,4, and 37, from left to right, 

respectively. The computed profiles are steady-state 

solutions to a one-dimensional advective-dispersive 

equation with a biological reaction term (eq. 3). 

Biomass is held constant during the simulation. The 

upper curve predicts the field profile that would result 

from the phenol degradation rate that was measured in 

the lab, whereas the lower one corresponds to the rate 

measured for 2-methylphenol. These two rates were 

used because they have the smallest associated errors 

and bracket the rates for the other two compounds 

(table 1). The parameters used in the solution for 

phenol and 3-methylphenol were: S, = 26.0, 

Hm=0.111d-', 7=0.013,^=1.33 mg/L,5=1.5xl(r2 

mg/L, v= 1.0 m/d, D- 1 .Om2/d. For 2- and 4- 

methylphenol, the values used were: S, =13.5 mg/L, 

^im=0.044d-l, 7=0.022, Ks=0.25 mg/L, £=3.0xlO'2 

mg/L, v=1.0 m/d, £>=1.0 m2/d.

The model profiles indicate that the microcosm 

simulations accurately represent the degradation 

process taking place in the field. The validity of the 

Monod kinetics expression for the degradation rate is 

apparent from the field data because the rate of 

decrease in the phenol concentration changes 

dramatically around site 40 (located about 90 m 

downgradient from site 3). When the substrate 

concentration is high (upgradient of site 40), the 

degradation kinetics can be approximated by a zeroTable 1. Kinetic constants for microcosm studies for each of the compounds tested with ± 95 percent confidence intervals. 

[[|im (d"')> maximum specific growth rate; Ks (mg/L), half-saturation constant; b (cT 1 ), specific decay rate; Y (mg/mg), yield coefficient]

Compound Ks (mg/L)

*Y values obtained from protein determinations before and after substrate utilization.

Y (mg/mg)1

Phenol

2-Methylphenol

3-Methylphenol

4-Methylphenol

.lll±0.005

.04410.001

.10310.078

.09910.110

1.3310.07

.25±0.82

.5516.67

3.34+11.1

.00110.012

.00210.008

.00010.019

.00010.032

.013

.022

.026

.025

846
&.

O 0£

BP

CONCENTRA IN MILLIGRAMS I

25 < 

20

15 

I

10 

5

0 

.<;

(A) O phenol 

1. n 2-methylphenol 

o 3-methylphenol 

A 4-methylphenol j

 - phenol & 3-methylphenol

------ 2- & 4-methylphenol 

a,

\ o X '-o 

^

~"BD _ _.

40 80 120 160 200

BACTERIA GROWTH RATE, IN 

MILLIGRAMS MILLIGRAM PE DAY

0.10 

0.08

0.06 

0.04 

0.02 

n

(B) ........... phenol & 3-methylphenol

"" ... ------ 2- & 4-methylphenol

X 't

t i N ** - t , t

40 80 120 160 200 

DISTANCE, IN METERS FROM SITE 3

Figure 3. (A) Computed concentration profiles (lines) and 

measured concentrations (points) of the four phenol solutes 

plotted against distance down-gradient from site 3. (B) Theoretical growth rate computed from the substrate concentration, the Monod growth expression, and the growth 

parameters measured in the microcosm simulations.

order reaction term consistent with the low values of Ks 

observed in the lab studies. When the substrate 

concentration is close to the value of Ks, the 

degradation rate drops as predicted by Monod kinetics. 

Thus, modeling confirms that the degradation rate is 

accurately described by Monod kinetics.

The biomass concentration used for each model 

profile was obtained by performing a least-squares fit 

of the computed concentration to the measured field 

data. The two values of B obtained using this method 

are: 1.5 x 10~2 mg/L for the steeper profile and 3.0 x 

10"2 mg/L for the lower profile (or, in terms of numbers 

of bacteria per 100 g sediment, 4.8 x 106 and 9.6 x 106 

respectively). An important point is that these numbers 

represent estimates of only the active bacteria mass 

controlling the rate-limiting step in the entire 

degradation process.

The fitted biomass estimates can be compared 

the value measured in sediments from a 6-m depth at 

site 39 (located 65 m downgradient from site 3). This 

was determined from analyses of total proteins in 10. g 

subsamples of the aquifer material by modification of 

the method of Bradford (1976) as described by Galli 

(1987). The total biomass concentration was 1.3 ± 

0.07 x 102 mg/L. Recall that the fitted value represents 

only the mass of active bacteria in the limiting step of 

the degradation process. Godsy and others (1992b) 

suggest that methanogenesis is the limiting step based 

on an observed build-up of acetate both in the 

microcosms and in the field. They determined that the 

methanogenic bacteria account for only 0.1 percent of 

the total bacteria population. Moreover, their Most 

Probable Number estimates of the mass of active 

bacteria were about 10 percent of the total biomass 

determined by AODC. Combining these results 

indicates that the active methanogenic biomass should 

be about four orders of magnitude less than the value 

obtained from total protein analyses, or 1.3 ± 0.07 x 

10"2 mg/L. This agrees well with our fitted estimates 

of 1.5-3.0 x 10'2 mg/L.

The most interesting result arising from the 

modeling is the evidence that the growth of bacteria 

within the contaminated region is independent of the 

substrate concentration and that the net growth is zero. 

The theoretical single compound growth rates 

corresponding to the model substrate profiles are 

shown in figure 3B. These growth rates were computed 

from eq. 2 with the parameters in table 1 and a decay 

value, b, of zero. In the first 65 m, the growth rates are 

nearly independent of the substrate concentration. This 

is a consequence of the very low half-saturation values, 

Ks, typical of the oligotrophic populations found in 

aquifers. This nearly constant growth rate may be 

balanced by a standard zero-order decay term (b in eq. 

2) to achieve zero net growth. However, beyond 65 m, 

the predicted growth rates on phenol and 3- 

methylphenol drop off, and beyond 90 m, the growth 

rates on 2- and 4-methylphenol also decrease.

Possibly, as the growth rate on the phenolic 

compounds begins to decrease, other compounds then 

serve as substrates. Candidates for this are oxindole 

and quinolinone. These compounds are oxidized 

intermediates of indole and quinoline. They appear 

after 20 to 30 days in microcosm studies and are found 

at site 3 in the field (Godsy, 1993). They are 

energetically favorable to growth with Monod growth

847
OS

z 

.E^

F

L1J < ^

S3W O£

8 i

z

JV

25 i

20

15

10

5

0

: (A) O phenol

S D 3-methylphenol

*> x ........... phenol & 3-methylphenol

\V N ------ with Haldane inhibition

\\

'«_s '

' ".,. p 

\ -

x'<- .

0^--- -0-Q -

40 80 120 160 200

0.12

0.10

g 0.08

o I 0.06

§s

1ft 0.04

0.02

0

no inhibition

with Haldane inhibition"

40 80 120 160 200 

DISTANCE, IN METERS FROM SITE 3

Figure 4. (A) Computed concentration profile for phenol that 

includes inhibition effects from the total phenol concentration 

shown together with measured concentration and the Monod 

growth expression with and without the modification for 

Haldane inhibition.

constants nearly equal to those measured for phenol 

(Godsy, 1993); however, they persist longer than the 

phenolic compounds before being degraded. Thus, it 

appears that in the field they might support the same 

maximum bacterial growth rate even as the phenolic 

compound concentrations decrease below their halfsaturation constants. Another possibility is that growth 

downgradient is supported by products of cells lysing 

upgradient. Thus, it seems likely that a constant 

bacterial growth rate in the field can be modeled by 

Monod kinetics on a sequence of compounds.

The decay or maintenance rate required to 

balance the predicted growth rate is unusually high. 

Suidan and others (1989) give a value of 0.0192 day" 1 

for a continuous-reactor study whereas Godsy and 

others (1992a) obtained a value of approximately 0.001

day" on the basis of microcosm studies. These values 

are too low to balance the 0.04-0.1 day" 1 growth rates 

shown in fig. 3B.

Inhibition of growth by various compounds was 

considered to be a possible explanation of the low 

bacteria population. Total phenolic-compound 

concentrations exceed 100 mg/L upgradient from site 3 

and are, thus, high enough to slow degradation and 

growth by about 20 percent (Suidan and others, 1989). 

Substrate inhibition may be modeled by modifying the 

biological reaction term in equation 3 according to the 

Haldane inhibition model (Haldane, 1930):

GtJ _ r\*-' " ' ill Lf (6)

ST ~ I/A". T J i

where £7- is the total phenol concentration and Kf is the 

Haldane inhibition constant (in mg/L).

A computed steady-state solution to eq. 6 for 

phenol using a value for Kt obtained by Suidan and 

others (1989) of 363 mg/L is shown in fig. 4A. The 

value of B that best fit the data is 2.4x 10~2 mg/L. The 

remaining model parameters are the same as used for 

fig. 3. The expected biomass growth rate on phenol 

under the Haldane model is shown in figure 4B. For 

comparison, the growth rate for phenol with no 

inhibition is also shown. The inhibited growth rate is 

significantly lower than the growth rate predicted by 

the model with no inhibition; however, it is still higher 

than the decay rate in Suidan and others (1989).

CONCLUSIONS

In summary, we have shown that Monod 

degradation constants measured in simulated 

methanogenic ecosystems accurately describe 

phenolic-compound degradation rates observed in the 

sand and gravel aquifer at the study site. In order to 

describe the natural system however, the model 

requires that there be no net growth in the aquifer 

biomass. This constraint is consistent with the field 

data because the biomass concentration in the aquifer is 

relatively low in spite of the continuous influx of 

substrate over many decades. The explanation for the 

the constant biomass concentration may involve a 

combination of two factors: First, the oligotrophic 

population exhibits growth that is independent of the

848
substrate concentration near the contaminant source. 

Second, where the growth rate begins to decline as the 

phenolic-compound concentrations approach the halfsaturation constants, the biomass may begin to 

consume other substrates. These may be organic 

compounds from lysed cells upgradient, or other 

contaminants that appear to persist for a longer time 

than the phenolic compounds but allow similar 

bacterial growth rates. Finally, a high decay rate is 

required to balance the proposed constant bacterial 

growth rate. This may be due to the hostility of the 

environment, transport of cells downgradient, or 

predation.by other organisms.

REFERENCES

Bradford, M.M., 1976, A rapid and sensitive method for the 

quantitation of microgram quantities of protein 

utilizing the principle of protein dye-binding: 

Analytical Biochemistry, v. 72, p. 248-254.

Bekins, B.A., Godsy, E.M., and Goerlitz, D.M., 1993, 

Modeling steady-state methanogenic degradation of 

phenols in groundwater: Journal of Contaminant 

Hydrology, v. 14, no. 3-4, p. 279-294.

Ehrlich, G.G., Ku, H.F.H., Vecchioli, J., and Ehlke, T.A., 

1979, Microbiological effects of recharging the 

Magothy Aquifer, Bay Park, New York, with tertiary 

treated sewage: U.S. Geological Survey Professional 

Paper, 751-E, 15 p.

Franks, B.J., 1988, Hydrogeology and flow of water in a 

sand and gravel aquifer contaminated by woodpreserving compounds, Pensacola, Florida: U.S. 

Geological Survey Water-Resources Investigations 87- 

4260, 72 p.

Ga'lli, Rene, 1987, Biodegradation of dichloromethane in 

waste water using a fluidized bed bioreactor: Applied 

Microbiology and Biotechnology, v. 27, p. 206-213.

Godsy, E.M., Goerlitz, D.F., and Grbic'-Galic', Dunja, 

1992a, Methanogenic degradation kinetics of phenolic 

compounds in aquifer-derived microcosms: 

Biodegradation, v. 2, p. 211-221.

Godsy, E.M., Goerlitz, D.F., and Grbic'-Galic', Dunja, 

1992b, Methanogenic biodegradation of creosote 

contaminants in natural and simulated ground-water 

ecosystems: Groundwater, v. 30, p. 232-242.

Godsy, E.M., 1993, Methanogenic biodegradation of

creosote-derived contaminants in natural and simulated 

groundwater ecosystems: unpublished Ph.D. 

Dissertation, Department of Civil Engineering, 

Stanford University, 174 p.

Goerlitz, D.F., Troutman, D.E., Godsy, E.M., and Franks, 

B.J., 1985, Migration of wood-preserving chemicals in 

contaminated groundwater in a sand aquifer at 

Pensacola, Florida: Environmental Science and 

Technology, v. 19, p. 955-961.

Goerlitz, D.F, 1992, A review of studies of contaminated 

groundwater conducted by the U.S. Geological Survey 

organics project, Menlo Park, CA, 1961-1990, in 

Lesage, S. and Jackson, R., eds., Groundwater 

Contamination and Analysis at Hazardous Waste Sites: 

New York, Marcel Dekker, Inc., p. 295-355.

Haldane, J.B.S., 1930, Enzymes: New York, Longmans, 

Green, London, 235 p.

Kindred, J.S., and Celia, M.A., 1989, Contaminant 

transport and biodegradation, 2 Conceptual model 

and test simulations: Water Resources Research, v. 25, 

p. 1149-1159.

Monod, Jacques, 1950, La technique de culture continue, 

theorie et applications: Annales de 1'Institut Pasteur 

(Paris), v. 79, p. 390-401.

Rittman, B.E., and McCarty, P.L., 1980, Design of fixedfilm processes with steady-state biofilm model: 

Progress in Water Technology, v. 12, p. 271 -281.

Suidan, M.T., Najm, I.N., Pfeffer, J.T., and Wang, Y.T., 

1989, Anaerobic biodegradation of phenol inhibition 

kinetics and system stability: Journal of Environmental 

Engineering, v. 114, p. 1359-1376.

849
850
Modeling Breakthrough of Nitrogen Heterocyclic

Compounds in Laboratory Columns Containing

Creosote-Contaminated Aquifer Material

By Ean Warren1 andE. Michael Godsy 1

Abstract

Laboratory-column breakthrough curves for 

the nitrogen-containing heterocyclic compounds 

found in the water soluble fraction of creosote 

were modeled and compared assuming local 

equilibrium and nonequilibrium conditions. The 

aquifer material consisted of a low organiccarbon-content (0.024 percent organic carbon), 

low ion-exchange-capacity (2 meq/100 g) clayey 

sand. Retardation factors and partition coefficients 

for quinoline, 2( 1 H)-quinolinone, carbazole, and 

acridine, derived from breakthrough curves, were 

contrasted with those obtained from batch experiments. Retardation factors for quinoline and 

2( 1 H)-quinolinone were comparable to each other, 

although those for carbazole and acridine were 

quite different despite similar chemical structures 

and octanol-water partition coefficients. Transport 

of quinoline, 2(lH)-quinolinone, and acridine is 

governed by organic and ionic interactions; 

however, transport of carbazole is controlled by 

organic interactions alone. In addition, although 

the local equilibrium assumption can be used to 

describe quinoline and 2( 1 H)-quinolinone breakthrough curves, carbazole and acridine breakthrough curves require nonequilibrium kinetics.

INTRODUCTION

Modeling is necessary to predict the fate and transport of organic contaminants in aquifers. Model parameters must be estimated or determined either from field

'U.S. Geological Survey, Menlo Park, Calif.

studies or laboratory experiments. Commonly, model 

input data are generated from properties of water that 

are not comparable to the properties of ground water at 

the site of contamination. Ground-water characteristics, such as pH, must be compared to ionization constants (pKa) to determine the effects of the sediment's 

ion exchange capacity. Simulated transport will not be 

accurate if the parameter at a pH differs from that of the 

ground water. In addition, laboratory studies commonly yield values for parameters using methods that 

do not reflect field conditions. For example, batch or 

column studies are used to determine partition coefficients, even though batch studies, often performed 

without considering the rate at which equilibrium is 

approached, do not consider nonequilibrium partitioning between aqueous and solid phases and column 

studies do not necessarily use representative packing or 

ground-water velocities.

In this paper, we compare the values of partition 

coefficients and retardation factors for quinoline, 

2(lH)-quinolinone, carbazole, and acridine by two 

different laboratory methods on low organic-carboncontent, low ion-exchange capacity aquifer material. 

The partition coefficients and retardation factors vary 

with the measurement method, most notably for acridine. In addition, we compare the results obtained for 

different equilibrium assumptions used in modeling 

column effluent. Column effluents were modeled successfully using the local equilibrium assumption for 

quinoline and 2(lH)-quinolinone and the nonequilibrium assumption for carbazole and acridine. Column 

studies are preferred over equilibrium batch studies 

because column studies yield more data under conditions similar to those in the field. However, to determine sorption rate constants independently, batch 

experiments measuring the rate at which equilibrium is 

attained must be conducted in conjunction with column 

experiments (Ball and Roberts, 1991b).

851
Site description

The study-site aquifer consists of quartz sand and 

gravel with interspersed silt and clay lenses (Franks, 

1988). During 1902-81, wood-treatment wastes were 

discarded into impoundments that were in direct 

contact with the ground water. Creosote and pentachlorophenol wastes migrated vertically into the subsurface. The ground water, at a pH of approximately 

5.9 and moving at a speed of about 1 m/d, continually 

dissolved the more soluble compounds in creosote 

creating an extended water soluble fraction (WSF) 

contamination plume. A major fraction of the WSF 

consists of the nitrogen heterocycles (for example, 

quinoline, carbazole, and acridine, as well as 2(1H)- 

quinolinone, a microbial degradation product) (Godsy 

and others, 1992).

Models

The transport of the WSF in laboratory columns 

can be described by the convection-dispersion equation 

using the local equilibrium assumption (LEA) as proposed by Hashimoto and others (1964) 

dx dx dt

(1)

where

D is the dispersion coefficient [m2/d],

C is aqueous concentration [g/m3],

v is linear velocity [m/d],

x is distance [m],

t is time [d], and

R relates the movement of a contaminant to a conservative 

	tracer [-].

Equation 1 assumes a linear Freundlich isotherm 

(Freundlich, 1926 X

(2)

where

C is mass concentration of contaminant adsorbed onto the

adsorbent [g/g] and 

K(j relates aqueous concentration to sorbed concentration

[m3/g].

Retardation factors, R, and partition coefficients, Kd, 

are related by

(3)

where

pfe is the bulk density of the aquifer material [g/m3] and 

0 is the porosity [-].

Transport commonly must be modeled using the 

nonequilibrium assumption (NEA) as proposed by van 

Genuchten and Wierenga (1976), which accounts for a 

readily mobile fraction and a stagnant or immobile 

fraction of water in the aquifer matrix (subscripts m and 

im, respectively)--

(4)

dx' dx

(5) 

where

/ is the fraction of sorption sites in the mobile region

[-] and

a | defines the transfer rate of the solute between mobile 

and immobile water [d' 1 ].

The model is based on the assumption that convection and dispersion govern transport in the mobile 

water and diffusion controls the transfer of contaminant between mobile and immobile water. The model 

includes two fitted dimensionless parameters: co, the 

ratio of hydrodynamic residence time to characteristic 

time of sorption (Brusseau and Reid, 1991); and (3, the 

relative contribution of the sites accessible to mobile 

water to the overall retardation. The original coefficients, oi| and/, relate to co and (3 as follows:

(6)

l ~ L* (7) 

where

q is the flux rate [m3/m2»d] and 

L is characteristic length [m].

MATERIALS AND METHODS

Aquifer materials were collected from a site 

upgradient from the contaminant plume. The samples 

were washed with 10 mM calcium chloride (CaCl2), 

pH 5.9 before being dried in an oven at 104 °C to prevent dilution by residual water. This CaCl2 concentration and pH closely parallel the ionic strength and pH 

of the natural ground water at the study site. Aquifermaterial characteristics are given in table 1.

852
Table 1. Aquifer material and column characteristics 

[m, meter; meq/100 g, milliequivalents per 100 grams; g/m3 , 

gram per cubic meter; m3/d, cubic meter per day]

Median particle diameter [m]

Percent organic carbon [-]

Cation exchange capacity [meq/IOO g]

Length [m]

Diameter [m]

Porosity [-]

Bulk Density [g/m3 X I0~6]

Flow Rate [nrVd x I06]

0.000375

.024

1.6

0.354

.025

.449

1.361

I40

Organic-carbon content of the aquifer material was 

measured by the Walkley-Black Method (Walkley and 

Black, 1934) as described in Methods of Soil Analysis 

(Black and others, 1965, p. 1372). Cation-exchange 

capacity of the aquifer material was determined by the 

method of Black and others (1965, p. 899) using ion 

chromatography.

Aqueous samples were analyzed for organic 

compounds by reverse phase high-performance liquid 

chromatography (HPLC) using gradient elution of 5 to 

100 percent acetonitrile in 26 mM ammonium acetate, 

pH 7.6 on a C-18 column with a diode-array detector.

The batch studies involved five different concentrations (table 2), in duplicate, of quinoline, 2(1H)- 

quinolinone, and acridine in 15 mM CaCl2 (pH 5.9), 

sterilized through a filter with a 0.45-nm pore size, and 

added to autoclaved 10 mL serum bottles with and 

without aquifer material. The serum bottles were 

sealed with Teflon-lined butyl rubber septa and aluminum crimp seals. The samples were first shaken rapidly 

for 2 days at room temperature (23 ± 2°C) and then 

allowed to equilibrate in the dark at room temperature 

with intermittent shaking before sampling. Samples 

were checked for equilibrium conditions by periodically sacrificing bottles of the same concentration. 

The concentration of the compounds in the bottles with 

aquifer material was compared to the concentration in 

the bottles without aquifer material to determine the 

mass of compound adsorbed onto the aquifer material. 

Samples were checked for sterility at the end of the 

experiment. No bacterial contamination was found.

Columns were prepared using dried aquifer material. 

The aquifer material was first passed through a 1.0 mm 

sieve, removing less than 0.3 percent by weight. The 

sediment was dampened slightly to aggregate the 

material and packed into a glass column, 1 to 3 mm at

a time, as described by Goerlitz (1984). The column 

characteristics are given in table 1. A constant-flow 

HPLC syringe pump was connected to the influent end 

of the column and a variable wavelength UV detector 

and a refractometer were connected to the effluent end. 

All the fittings and tubing were either glass, stainless 

steel, or Teflon. All CaCl2 solutions were degassed and 

filtered before use. Ten mM CaCl2 at pH 5.9 was 

pumped through the column until the column reached 

equilibrium. The compounds were dissolved into 15 

mM CaCl2 at pH 5.9 and run through the column at 

approximately 120 x 10'6 m3/d until the CaCl2 concentration in the effluent equaled that in the influent 

(table 2). This flow rate is approximately equal to a 

linear velocity of 1 m/d in the aquifer.

Parameters-/? for the LEA and /?, p, and co for 

NEA-were fit to the breakthrough curves (BTC's) using 

nonlinear regression analysis. BTC's were analyzed by 

use of the computer program HASHPE (Oravitz, 1984) 

to determine LEA parameters and the computer program CFITIM (van Genuchten, 1981) to determine 

NEA parameters. The dispersion parameter for all 

model simulations was determined from CaCl2 

breakthrough.

Table 2. Concentrations of quinoline, 2(1H)-quinolinone, 

carbazole, and acridine for batch and column studies 

[Concentrations in gram per cubic meter]

Quinoline

2(IH)-Quinolinone

Carbazole

Acridine

1 40

53

35

72

26

19

Batch

30

9.7

9.4

9.3

3.2

2.6

1. 7

.60

.47

Column

2.5

2.5

8.9

IS

RESULTS 

Batch Studies

Linear Freundlich partition coefficients for batch 

studies of quinoline, 2(lH)-quinolinone, and acridine 

were determined by linear regression (with the y-intercept forced to zero) on the aqueous and sorbed concentrations (table 3). The r2 values for acridine and 2(1H)- 

quinolinone were 0.95. The r2 value for quinoline was 

0.85. This value is low because quinoline exhibited 

nonlinear sorption, confirming that reported by Zachara 

and others (1986). Despite this, the local equilibrium 

model, which is based on the assumption of linearity, 

closely reproduced quinoline in the column effluent, 

possibly because of a low selectivity coefficient.

853
Table 3. pKa, log Kow, partition coefficients, and retardation factors for quinoline, 

2(1H)-quinolinone, carbazole, and acridine 

[Partition coefficients in cubic meter per gram x 106]

Partition Retardation factor

Quinoline

2( 1 H)-Quinolinone

Carbazole

Acridine

H^a

4.90

5.29

-5.70

5.60

iuy i\ow

2.03

1.26

3.29

3.40

coefficient 3/

0.133

.231

4.56

Batch 4/

1.40

1.70

14.8

Column 5/

1.32

1.54

4.32

39.6

-' (Stecher and others, 1968, p. 210; American Petroleum Institute, 1979; 1981). 

2/ (Hansch and Leo, 1979).

^ Determined using linear regression on the aqueous and sorbed batch concentrations. 

^ Calculated using equation 3.

^ Determined using local equilibrium assumption for quinoline and 2(lH)-quinolinone 

and nonequilibrium assumption for carbazole and acridine.

Column Studies

Retardation factors for quinoline and 2(lH)-quinolinone were similar (table 3). 2(lH)-Quinolinone, with 

a pKa of 5.29 (table 3), is approximately 20 percent 

ionized and quinoline, with a pKa of 4.9, is approximately 9.1 percent ionized. Zachara and others (1986) 

have shown that sorption of quinoline is dominated by 

ion exchange up to 2 pH units above its pKa. 2(1H)- 

Quinolinone, like quinoline, should be retained by ion 

exchange and organic sorption. The retardation of 

acridine is significantly higher than that of carbazole 

despite the fact that both have two benzene rings fused 

to a nitrogen heterocyclic ring (fig. 1) and have similar

octanol-water partition coefficients (table 3). However, 

the pKa of acridine is 5.6 and that of carbazole is -5.7 

(table 3). Thus, at pH 5.9, the ionized fraction of acridine is 0.33, but carbazole is completely un-ionized. 

The degree of affinity of acridine to charged functional 

groups on the aquifer material (the selectivity coefficient) and the extent of ionization as well as the cationexchange capacity contributes to the exchange capacity. With an acridine concentration of 18 g/m3 (0.10 

meq/L), the column capacity due to ion exchange is 

160. The column capacity is based on the assumption 

of total sorption of the ionized fraction of acridine to 

the aquifer material and complete displacement of 

calcium ions.

Quinoline 2( lH)-Quinolinone

Carbazole Acridine

Figure 1. Chemical structures of quinoline, 2(1H)-quinolinone, 

carbazole, and acridine.

854
1

§P 0.75

1 0.5

8

g

5 0.25

,-J

n

/ /

/ *

f '

I '

r 4

 ^*Mr A A ^^^^^ A i

0 0.5

 - LEA

1 1.5 

PORE VOLUMES 

O CaCU

2.5

A 2(lH)-Quinolinone

Figure 2. Calcium chloride (CaCI2) and 2(1 H)-quinolinone breakthrough curves from column data and 

fitted local equilibrium assumption model (LEA). Because local equilibrium assumption and nonequilibrium assumption models for 2(1 H)-quinolinone can be superimposed on each other, the nonequilibrium

C C assumption model is not shown. Relative concentrations [-] were calculated using -= -£-, where C is

cr co

the effluent concentration, C0 is the original concentration at the start of the experiment, and C\ is the 

influent concentration. Pore volumes [-] were determined by dividing time by the time it takes for onehalf of the relative concentration of CaCI2 to move through the column.

LEA and NEA Models

Breakthrough curves for quinoline, 2(lH)-quinolinone, carbazole, and acridine were modeled using both 

LEA and NEA. The BTC's for 2(lH)-quinolinone and 

the conservative tracer, CaCl 2 , for both LEA and NEA 

fitted models are shown in figure 2. LEA and NEA 

models are indistinguishable.

The ETC for acridine along with LEA and NEA 

fitted models are shown in figure 3. The LEA model 

does not follow the BTC at any time. Carbazole gave 

similar results (data not shown). The results indicate 

that, under flow conditions similar to that at the study 

site, transport of acridine and carbazole is controlled by 

nonequilibrium conditions and transport of quinoline 

and 2(lH)-quinolinone is controlled by equilibrium 

conditions.

Fitted parameters and original coefficients for 

quinoline, 2(lH)-quinolinone, carbazole, and acridine 

using the NEA model are given in table 4. The values 

of p for carbazole and acridine agree with each other.

The value of co for carbazole, however, is about twice 

that for acridine, indicating that the characteristic time 

of sorption contributes more to acridine transport than 

to carbazole transport.

DISCUSSION AND CONCLUSION

Transport of nitrogen heterocycles is governed 

by the retardation and the rate at which equilibrium is 

approached, both of which must be determined to model 

transport and sorption processes properly. The values of 

both parameters used in models depend on ground-water 

pH, water velocities, and contaminant concentrations, 

among other factors. To reduce error in the laboratory 

studies, these variables must be similar to that of the 

ground water at the contaminated site. The effects of ion 

exchange can have a profound effect on transport of 

weakly acidic or basic compounds. Transport under nonequilibrium conditions will also affect initial breakthrough and tailing. Without accurate laboratorygenerated data, models may produce erroneous results.

855
0

 - LEA

40 

PORE VOLUMES

 NEA O Acridine

Figure 3. Acridine breakthrough curves from column data and local equilibrium (LEA) and nonequilibrium (NEA) assumption fitted models. Relative concentrations [-] were calculated using

c-c

, where

C is the effluent concentration, C0 is the original concentration at the start of the experiment, and Cj is 

the influent concentration. Pore volumes [-] were determined by dividing time by the time it takes for 

one-half of the relative concentration of CaCI2 to move through the column.

Table 4. Nonequilibrium assumption parameter values 

for quinoline, 2(1H)-quinolinone, carbazole, and acridine

Quinoline

2(lH)-Quinolinone

Carbazole

Acridine

p

0.0

.0

.60

.6!

fV

0.0

.0

.60

.6!

(0

1 60

1,300

2.6

1.2

-I*

110

900

1.8

.81

- Calculated assuming 6p-6m equals zero (van Genuchten 

and Wierenga, 1976) in equation 6.

^ Calculated using equation 7.

In this study, parameter values determined from 

batch and column studies for quinoline and 2(1H)- 

quinolinone, compounds with low octanol-water partition coefficients, were similar. However, the two methods gave quite different results for acridine. This is to 

be expected because equilibrium batch studies give 

values for LEA parameters and column studies under 

conditions similar to those in the field give values for 

NEA parameters. Batch or column studies are often

used to generate model parameters. Both types of studies have advantages and disadvantages. Batch experiments have been criticized as (1) using a large liquid to 

sediment ratio and improperly represented sediment 

characteristics (Barrow and Shaw, 1975; Sparks, 1985; 

Ball and Roberts, 199la; Kookana and others, 1992), 

and (2) not modeling actual flow-field conditions 

(Freeze and Cherry, 1979; Ball and Roberts, 1991b). 

Column experiments commonly are subject to unrealistic flows and packing (Johnson and Farmer, 1993) 

and cannot determine the specific reason for nonequilibrium transport.

Although retardation factors for quinoline and 

2(lH)-quinolinone, both of which are partially ionized, 

were comparable, those for carbazole and acridine 

were quite different. This difference indicates that ion 

exchange plays a large role in the transport of contaminants in the aquifer material. This study shows that 

transport of quinoline, 2( lH)-quinolinone, and acridine 

is governed by ionic and organic interactions, but that 

carbazole is controlled by organic interactions alone.

856
The LEA model sufficiently describes the column 

effluent for quinoline and 2(lH)-quinolinone. However, 

the NEA model must be used for carbazole and acridine. Brusseau and Rao (1989) suggest that, for values 

of co less than 10, the NEA should be used instead of 

the LEA. The values of co for carbazole and acridine are 

well below the value suggested by Brusseau and Rao, 

justifying the use of the NEA model. Godsy (1993) 

observed that NEA provides a better fit to data when 

the log Kow is greater than 2.5. The specific reason for 

nonequilibrium-diffusion between mobile and immobile water and (or) the kinetics of sorption to organic or 

ionic sites-cannot be determined from BTC analysis 

and must be resolved by rate batch studies. Column 

models show the importance of nonequilibrium in 

describing transport of contaminants. Equilibrium 

batch studies, on the other hand, do not indicate 

whether a contaminant, under field flow conditions, is 

actually at equilibrium. This distinction is important 

when modeling transport in the ground water.

REFERENCES

American Petroleum Institute, 1979, Quinoline: Washington, D.C., American Petroleum Institute, API Monograph Series, Publication 711, 43 p.

American Petroleum Institute, 1981, Carbazole, 9-Methylcarbazole, and Acridine: Washington, D.C., American 

Petroleum Institute, API Monograph Series, Publication 716, 51 p.

Ball, W.P. and Roberts, P.V., 199la, Long-term sorption of 

halogenated organic chemicals by aquifer material, 1 

Equilibrium: Environmental Science & Technology, 

v. 25, no. 7, p. 1223-1237.

Ball, W.P. and Roberts, P.V., 1991b, Long-term sorption of 

halogenated organic chemicals by aquifer material, 

2 Intraparticle diffusion: Environmental Science & 

Technology, v. 25, no. 7, p. 1237-1249.

Barrow, N.J. and Shaw, T.C., 1975, The slow reactions 

between soil and anions, 5 Effects of period of prior 

contact on the desorption of phosphate from soils: Soil 

Science, v. 119, no. 4, p. 311-320.

Black, C.A., Evans, D.D., White, J.L., Ensminger, L.E., and 

Clark, F.E., eds., 1965, Methods of Soil Analysis, Part 

2 Agronomy: Madison, Wisconsin, American Society 

of Agronomy, 15 72 p.

Brusseau, M.L. and Rao, P.S.C., 1989, The influence of sorbate-organic matter interactions on sorption nonequilibrium: Chemosphere, v. 18, no. 9-10, p. 1691-1706.

Brusseau, M.L. and Reid, M.E., 1991, Nonequilibrium 

sorption of organic chemicals by low organic-carbon 

aquifer materials: Chemosphere, v. 22, no. 3-4, 

p. 341-350.

Franks, B.J., 1988, Hydrogeology and flow of water in a 

sand and gravel aquifer contaminated by wood-preserving compounds, Pensacola, Florida: U.S. Geological 

Survey Water-Resources Investigations Report 

87-4260, 72 p.

Freeze, R.A. and Cherry, J.A., 1979, Groundwater: 

Englewood Cliffs, N.J., Prentice-Hall, 604 p.

Freundlich, Herbert, 1926, Colloid and Capillary Chemistry: London, Methuen.

Godsy, E.M., 1993, Methanogenic biodegradation of creosote-derived contaminants in natural and simulated 

ground water ecosystems: Stanford, Calif., Stanford 

University, Ph.D. dissertation, 173 p.

Godsy, E.M., Goerlitz, D.F., and Grbic-Galic, Dunja, 1992, 

Methanogenic biodegradation of creosote contaminants 

in natural and simulated ground-water ecosystems: 

Ground Water, v. 30, no. 2, p. 232-242.

Goerlitz, D.F., 1984, A column technique for determining 

sorption of organic solutes on the lithological structure 

of aquifers: Bulletin of Environmental Contamination 

and Toxicology, v. 32, p. 37-44.

Hansch, Corwin and Leo, A.J., 1979, Substituent Constants 

for Correlation Analysis in Chemistry and Biology: 

New York, John Wiley and Sons.

Hashimoto, I., Deshpande, K.B., and Thomas, H.C., 1964, 

Peclet numbers and retardation factors for ion exchange 

columns: Industrial Engineering Chemical Fundamentals, v. 3, p. 213-218.

Johnson, J.A. and Farmer, W.J., 1993, Batch versus column 

method for determining distribution of organics 

between soil and water phases: Soil Science, v. 155, 

no. 2, p. 92-99.

Kookana, R.S., Gerritse, R.G., and Aylmore, L.A.G., 1992, 

A method for studying nonequilibrium sorption during 

transport of pesticides in soil: Soil Science, v. 154, 

no. 5, p. 344-349.

Oravitz, J.L., 1984, Transport of trace organics with onedimensional saturated flow: Mathematical modeling 

and parameter sensitivity analysis: Michigan Technological University, M.S.C.E. Thesis, Department of 

Civil Engineering.

Sparks, D.L., 1985, Kinetics of ionic reactions in clay 

minerals and soils: Advances in Agronomy, v. 38, 

p. 231-266.

Stecher, P.G., Windholz, Martha, Leahy, D.S., Bolton, 

D.M., and Eaton, L.G., eds., 1968, The Merck Index 

(8th ed.): Rahway, N.J., Merck & Co., Inc., 1713 p.

857
van Genuchten, M.T., 1981, Non-Equilibrium Transport 

Parameters from Miscible Displacement Experiments: 

U.S. Department of Agriculture, U.S. Salinity Laboratory Research Report 119, 88 p.

van Genuchten, M.T. and Wierenga, P.J., 1976, Mass transfer studies in sorbing porous media, I Analytical solution: Soil Science Society of America Proceedings, 

v. 40, p. 473-480.

Walkley, A. and Black, I. A., 1934, An examination of the 

Degtjareff method for determining soil organic matter 

and a proposed modification of the chromic acid titration method: Soil Science, v. 37, p. 29-38.

Zachara, J.M., Ainsworth, C.C., Felice, L.J., and Resch, 

C.T., 1986, Quinoline sorption to subsurface materials: 

role of pH and retention of the organic cation: Environmental Science and Technology, v. 20, p. 620-627.

858
Bioremediation of Petroleum and Metal Contamination 

with Dissimilatory Metal Reduction

By Derek R. Lovley1

Abstract

The possibility that the metabolism of dissimilatory metal-reducing microorganisms might be 

used for removing organic contaminants and 

metals from contaminated environments was 

investigated. The Fe(III) chelator, nitrilotriacetic 

acid (NTA) enhanced the degradation of toluene in 

petroleum-contaminated aquifer material in which 

Fe(III) reduction was the terminal electron-accepting process. Addition of Fe(III) oxide containing 

sediments stimulated toluene degradation in sediments that were originally methanogenic. These 

results suggest that increasing the availability of 

Fe(III) may be a useful bioremediation strategy in 

aquifers that are heavily contaminated with 

aromatic compounds. Studies on the mechanism 

for U(VI) reduction by Desulfovibrio vulgaris 

indicated that cytochrome c3 is the U(VI) and 

Cr(VI) reductase. Cytochrome c3 reduced U(VI) 

in uranium-contaminated mine drainage and 

ground water. This finding indicates that it may 

be possible to genetically engineer microorganisms with enhanced metal-reducing capacity. 

These studies demonstrate that dissimilatory metal 

reduction may be a useful mechanism for 

bioremediating organic and (or) metal contamination in some environments.

INTRODUCTION

The potential for bioremediation of environments 

contaminated with organic compounds is well known.

U.S. Geological Survey, Reston, Va.

The goal of bioremediation is to convert toxic organic 

compounds to nontoxic compound(s), typically carbon 

dioxide. Most studies on the in situ bioremediation of 

organic contaminants in ground water have focused on 

oxidation of organic contaminants coupled to the 

reduction of oxygen. However, Fe(III) is typically the 

most abundant electron acceptor available for organic 

matter oxidation when organic contaminants are introduced into previously pristine shallow aquifers 

(Lovley, 1991). Previous studies have indicated that 

Fe(III) reduction is an important intrinsic process that 

naturally removes organic contaminants from polluted 

ground water (Lovley and others, 1989; Lyngkilde and 

Christensen, 1992; Baedecker and others, 1993). Thus, 

if it were possible to stimulate the activity of Fe(III)- 

reducing microorganisms in contaminated aquifers, the 

rate at which organic contaminants were degraded 

might be enhanced.

In contrast to organic contaminants, metal contaminants are not "biodegradable." However, a number of 

important contaminant metals are less soluble in the 

reduced state than they are in the oxidized state 

(Lovley, 1993b). For example, microbial uranium 

reduction converts the oxidized soluble form of uranium, U(VI) to U(IV) which precipitates from most 

natural waters (Lovley and others, 1991). This reductive precipitation can effectively remove uranium from 

a variety of uranium-contaminated waters (Lovley and 

Phillips, 1992b) and can be used in conjunction with a 

soil-extraction technique for concentrating uranium 

from contaminated soils (Lovley, 1993b). In a similar 

manner, numerous microorganisms can reduce highly 

soluble and toxic Cr(VI) to less soluble, less toxic, 

Cr(III) (Lovley, 1993b).

The purpose of the studies reported here was to 

further investigations on the potential for using dissimilatory metal reduction as a tool for bioremediating 

organic and inorganic contaminants.

859
BIOREMEDIATION OF ORGANIC 

CONTAMINATION

Studies with the dissimilatory Fe(III)-reducing 

microorganism Geohactcr melallireducens (strain 

GS-15) demonstrated that the reduction of Fe(III) 

could be greatly stimulated with the metal chelating 

compound nitrilotriacetic acid (NTA) (data not shown). 

Thus, the potential for using NTA to stimulate the 

oxidation of organic contaminants coupled to Fe(III) 

reduction in ground water was evaluated with aquifer 

material collected from a petroleum-contaminated 

aquifer located in Hanahan, S.C. (Vroblesky and 

Chapelle, 1996). Zones in which Fe(III) reduction or 

methane production were the terminal electronaccepting processes (Tt^APs) were located using 

measurements of dissolved H-> and the TEAPs in the

sediments were confirmed with laboratory incubations 

and the use of radiolabeled tracers. Sediments that had 

been collected by hand augering were incubated under 

strict anaerobic conditions in serum bottles. Heat-killed 

controls were generated by autoclaving sediments. Toluene was selected as a typical aromatic ground water 

pollutant. Toluene and NTA dissolved in oxygen-free 

water were added to the sediments in order to provide 

initial concentrations of 10 (.iM and 4 mM, respectively. 

Metabolism of toluene over time was monitored by 

sampling the headspace with a syringe and needle and 

analyzing the headspace gas for toluene with a gas 

chromatograph equipped with a flame ionization 

detector.

There was a steady loss in toluene concentrations 

over time in sediments from the Fe(III)-reducing zone 

(fig. 1). If the microorganisms in the sediment were

g g

LU Z

ID D

LU hO <

00

9 »-

^ Z

LU -

120000

1ooooo

80000

60000

40000

20000

- NTA-Live

--«-- NTA-Killed

--*-- Killed

 * Live

10 A A A 15

DAYS

Figure 1. Effect of nitrilotriacetic acid (NTA) on the loss of toluene over time in Fe(lll)-reducing sediments. 

The designation killed indicates that the microorganisms in the sediment were killed with heat prior to the 

addition of toluene. Toluene was added at an initial concentration of 10 nM and an equivalent amount of 

toluene was again added to the sediments containing live microorganisms that had been amended with 

NTA at the times designated by the arrows along the x-axis. Integration units specify the numerical area 

calculated by an electronic integrator for the size of the toluene peak detected by the gas chromatograph.

860
40

UJ

o z 

< 30

UJ 

Z Q.

UJ /)

O z

20

10

Toluene-Methanogenic Sediments

Toluene-Fe(lll) Amended Sediments

Methane-Methanogenic Sediments

Methane-Fe(lll) Amended Sediments

800

UJ

< z

600 |§

< i iu

X DC

400 z£ 

UJ CO

200 JZ Q. 

UJ _

10 15

DAYS

Figure 2. Effect of addition of Fe(lll)-oxide containing sediments to sediments in which methane 

production was the TEAR Toluene was added at an initial concentration of 10 nM.

killed with heat, there was little loss of toluene over 

time. After a brief initial lag, the addition of NTA 

greatly accelerated the loss of toluene over time in sediments containing live microorganisms. As the toluene 

was metabolized, more toluene was added and it was 

rapidly degraded. NTA did not stimulate loss of toluene 

in heat-killed sediments.

In contrast to the sediments in which Fe(III) reduction was the TEAP, toluene was not metabolized over a 

15 day incubation period in sediments in which the 

TEAP was methane production (data not shown). Furthermore, additions of NTA did not stimulate toluene 

degradation in the methanogenic sediments (data not 

shown). In order to determine whether the difference in 

rates of toluene degradation in the methanogenic and 

Fe(III)-reducing sediments could be attributed to 

differences in the availability of Fe(III) as an electron 

acceptor, methanogenic sediments were mixed with an

equal volume of methanogenic sediments which had 

been oxidized under air in order to convert Fe(II) compounds to Fe(III) oxides. In the presence of Fe(III) 

oxides, methane production was inhibited (fig. 2), presumably because the TEAP had been switched from 

methane production to Fe(III) reduction. The addition 

of Fe(III) also greatly stimulated the loss of toluene.

These results suggest two mechanisms in which 

the degradation of aromatic hydrocarbons may be stimulated in petroleum contaminated aquifers. When 

Fe(III) oxides are naturally available, it may be possible to enhance contaminant degradation by adding chelators, such as NTA, which make Fe(III) more available 

for reduction. In methanogenic zones of aquifers in 

which Fe(III) is not available for reduction, it may be 

possible to stimulate in situ metabolism of contaminants by emplacing a barrier of sand heavily coated 

with Fe(III) oxide. This "iron curtain" could have the

861
I

Uj °J

3§

H O 

Z O

s^

DC >,

UJ D 

O.

120

100

Ground Water, Enzymes, Hydrogen 

Ground Water, Enzymes, No Hydrogen 

Mine Water, Enzymes, Hydrogen 

Mine Water, Enzymes, No Hydrogen

MINUTES

Figure 3. U(VI) concentrations over time in mine drainage water and uranium-contaminated ground water 

in the presence of hydrogenases and cytochrome 03, with or without added H2 . The initial concentrations of 

U(VI) in the mine water and ground water were 36 and 0.5 nM, respectively.

added benefit of filtering trace metals out of the ground 

water as it moved through it because it is well known 

that many trace metals bind tenaciously to Fe(III) oxide 

coatings (Singh and Subramanian, 1984).

BIOREMEDIATION OF URANIUM AND 

CHROMIUM CONTAMINATION

Microorganisms in the genus Desulfovibrio have 

many desirable characteristics that may make them the 

organisms of choice for metal-treatment systems 

employing reductive precipitation (Lovley and Phillips, 

1992b, Lovley and others, 1993). Furthermore, Desulfovibrio species may be important in reducing metals 

in a variety of sedimentary environments (Coleman 

and others, 1993, Lovley and Phillips, 1992a, Lovley 

and others, 1993).

Recent studies in our laboratory have demonstrated 

that cytochrome c^ is the U(VI) and Cr(VI) reductase 

in Desulfovibrio vulgaris (Lovley, 1993b). Whole cells 

of Desulfovibrio desulfuricans were previously shown 

to effectively treat uranium in mine drainage water 

from a U. S. Department of Interior site and contaminated ground water from a U. S. Department of Energy 

site (Lovley and Phillips, 1992b). In order to determine 

whether cytochrome c^ could also reduce U(VI) in 

these waters, cytochrome c ? was combined with H2 

and hydrogenase as an electron donating system. Cytochrome c^ readily reduced U(VI) in both waters (fig. 3). 

There was no U(VI) reduction when the electron donor, 

H2 , was omitted.

These results demonstrate that cytochrome c^ can 

potentially be used to remove uranium from contaminated environments. For example, cytochrome c ? is

862
readily mass produced and could be employed in fixedenzyme reactors. Furthermore, the gene for cytochrome cj from D. vulgaris has been cloned and 

expressed in the closely related Desulfovibrio desulfuricans (Voordouw and others, 1990) as well as the 

genetically distant Rhodobacter sphaeroides (Cannae 

and others, 1991). Thus, the ability for microorganisms 

to reduce U(VI) and Cr(VI) could potentially be genetically combined with other useful metabolic properties such as denitrification and the ability to degrade 

organic contaminants. This might confer in a single 

organism the ability to remove not only the contaminant metals but also other important contaminants, 

such as organic solvents and nitrate that are typically 

found in wastes bearing these metals. Future studies 

will evaluate whether genetically engineering an 

organism to over-express cytochrome c^ will enhance 

the capacity for metal reduction.

CONCLUSIONS

It is well known that dissimilatory metal reduction 

has an important influence on the organic and inorganic 

geochemistry of a variety of sedimentary environments 

(Lovley, 1991, Lovley, 1993a, Lovleyl993b). The 

studies reported here support the recent suggestion 

(Lovley, 1993a, Lovley, 1993b) that dissimilatory 

metal reduction may also have practical application for 

the remediation of some environments contaminated 

with organic or metal contaminants.

REFERENCES

Baedecker, M.J., Cozzarelli, I.M., Eganhouse, R.P., Siegel, 

D.I., Bennett, P.C., 1995, Crude oil in a shallow sand 

and gravel aquifer, 3 Biogeochemical reactions and 

mass balance modeling in anoxic groundwater: Applied 

Geochemistry, v. 8, p. 569-586.

Cannae, V., Caffrey, M.S., Voordouw, G., and Cusanovich, 

M.A., 1991, Expression of the gene encoding cytochrome c? from the sul fate-reducing bacterium 

Desulfovibrio vulgaris into the purple photosynthetic 

bacterium Rhodobacter sphaeroides: Journal of Bacteriology, v. 286, p. 629-632.

Coleman, M.L., Hedrick, D.B., Lovley, D.R., White, D.C., 

and Pye, K., 1993, Reduction of Fe(III) in sediments by 

sulphate-reducing bacteria: Nature, v. 361, p. 436-438.

Lovley, D.R., 1991, Dissimilatory Fe(III) and Mn(IV) 

reduction: Microbiological Reviews, v. 55, p. 259-287.

Lovley, D.R., 1993a, Anaerobes into Heavy Metal: Dissimilatory metal reduction in anoxic environments: Trends 

in Ecology and Evolution, v. 8, p. 213-217.

Lovley, D.R., 1993b, Dissimilatory metal reduction: Annual 

Reviews of Microbiology, v. 47, p. 263-290.

Lovley, D.R., Baedecker, M.J., Lonergan, D.J., Cozzarelli, 

I.M., Phillips, E.J.P., and Siegel, D.I., 1989, Oxidation 

of aromatic contaminants coupled to microbial iron 

reduction: Nature, v. 339, p. 297-299.

Lovley, D.R., and Phillips, E.J.P., 1992a, Reduction of 

uranium by Desulfovibrio desulfuricans: Applied and 

Environmental Microbiology, v. 58, p. 850-856.

Lovley, D.R., and Phillips, E.J.P., 1992b, Bioremediation of 

uranium contamination with enzymatic uranium reduction: Environmental Science and Technology, v. 26,

p. 2228-2234.

Lovley, D.R., Phillips, E.J.P., Gorby, Y.A., and Landa, E.R., 

1991, Microbial reduction of uranium: Nature, v. 350, 

p. 413-416.

Lovley, D.R., Roden, E.E., Phillips, E.J.P., and Woodward, 

J.C., 1993, Enzymatic iron and uranium reduction by 

sulfate-reducing bacteria: Marine Geology, v. 113, 

p. 41-53.

Lyngkilde, J., and Christensen, T.H., 1992, Fate of organic 

contaminants in the redox zones of a landfill leachate 

pollution plume (Vejen, Denmark): Journal of Contaminant Hydrology, v. 10, p. 291-307.

Singh, S.K., and Subramanian, V, 1984, Hydrous Fe and 

Mn oxides-scavengers of heavy metals in the aquatic 

environment: Critical Reviews in Environmental 

Control, v. 14, p. 33-90.

Voordouw, G., Pollock, W.B.R., Bruschi, M., Guerlesquin, 

F, Rapp-Giles, B.J., and Wall, J.D., 1990, Functional 

expression of Desulfovibrio vulgaris Hildenborough 

cytochrome c3 in Desulfovibrio desulfuricans G200 

after conjugational gene transfer from Escherichia coli: 

Journal of Bacteriology, v. 172, p. 6122-6126.

Vroblesky, D.A., and Chapelle, F.H., 1996, Shifting of terminal electron-accepting processes in a petroleumhydrocarbon-contaminated aquifer, in Morganwalp, 

D.W., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

863
864
Effects of Pb and Terminal-Electron-Accepting

Processes on Organic Acid Concentrations in

Contaminated Aquifer Sediments

By Paul M. Bradley1 , Francis H. Chapelle1 , and Don A. Vroblesky 1

Abstract

The effects of terminal-electron-accepting 

processes (TEAPs) and dissolved Pb 

concentrations on carbon mineralization and 

organic acid accumulation in petroleumcontaminated sediments were examined under 

aerobic, nitrate reducing, sulfate reducing, and 

methanogenic conditions. Pb concentrations in 

the 14 to 60 jag/L range observed in the field 

significantly inhibited CO2 production for all 

TEAPs as well as CH4 production under 

methanogenic conditions. However, 

accumulation of organic acids was more closely 

related to the dominant TEAR Acetic, propionic, 

i-butyric, and n-butyric acids accumulated in 

methanogenic treatments, whereas only acetic 

acid was detected in sulfate reducing treatments. 

As expected, no organic acids were detected 

under nitrate reducing or aerobic conditions. This 

pattern parallels field observations in which high 

concentrations of organic acids (acetic, propionic, 

i-butyric, and n-butyric acids) were detected in 

methanogenic zones, but only low concentrations 

of acetic acid were detected in sulfate reducing 

zones.

INTRODUCTION

Recent investigations have shown that shortchain organic acids are important in the chemical and 

physical diagenesis of contaminated and pristine

'U.S. Geological Survey, Columbia, South Carolina

aquifer systems. The accumulation of organic acids in 

petroleum hydrocarbon-contaminated aquifers has 

been widely observed (Cozzarelli and others, 1990; 

Godsy and others, 1992). In pristine aquifer systems, 

transport of formic and acetic acids from organic-rich 

confining bed sediments provides the electron donors 

for iron reduction, sulfate reduction and 

methanogenesis in adjacent aquifers (McMahon and 

Chapelle, 1991; Chapelle and McMahon, 1991; 

McMahon and others, 1992). Furthermore, the 

production of organic acids in both pristine and 

contaminated aquifers appears to be linked to 

sediment diagenesis and the development of 

secondary porosity (McMahon and others, 1992; 

Hiebert and Bennett, 1992).

Given the importance of these compounds as 

electron donors for microbial respiration in aquifers 

and their potential role in production of secondary 

porosity in aquifers, the factors affecting the 

accumulation of short-chain organic acids in ground 

water merit further investigation. The shallow aquifer 

underlying a petroleum tank farm near Charleston, 

South Carolina is characterized by extensive 

petroleum contamination and localized high 

concentrations of short-chain organic acids. At this 

site, concentrations of acetic, propionic, i-butyric, and 

n-butyric acids as high as 7,910; 110; 3,140 and 

796 ^iM, respectively, have been reported in groundwater contaminated by petroleum hydrocarbon 

compounds (P.B. McMahon, U.S. Geological Survey, 

written commun. 1991).

Water chemistry data collected at the site 

suggests that the high concentrations of organic acids 

may be associated with high concentrations of Pb. For 

example, in January 1992 approximately 40 ^ig/L of 

dissolved Pb was detected in water from a well 

characterized by high organic-acid concentrations.

865
Because heavy-metal inhibition of organic acid 

consumption under methanogenic conditions has been 

reported (Jarrell and others, 1987), this data suggests 

that Pb at the Hanahan site may contribute to organicacid accumulation by inhibiting respiration.

The presence or absence of Pb is not the only 

factor that can affect concentrations of organic acids. 

Methanogenesis is the dominant electron-accepting 

process within the zone of organic-acid accumulation 

at Hanahan, and several studies have reported organicacid accumulation under methanogenic conditions 

(Cozarelli and others, 1990, Godsy and others, 1992, 

Grbic-Galic and Vogel, 1987). Thus the potential 

exists that accumulation of organic acids is also 

related to the dominant terminal electron accepting 

process (TEAP).

Available field data are insufficient to determine 

the relative importance of Pb concentrations and 

predominant TEAPs on the observed distribution of 

organic acids at the Hanahan site. In the studies 

reported here, an integrated laboratory approach is 

used to investigate the importance of Pb 

concentrations and predominant TEAPs on observed 

concentrations of organic acids at this site. The 

results indicate that the distribution of TEAPs, rather 

than concentrations of Pb, is the predominant control 

on the accumulation of organic acids in ground water 

at this site.

MATERIALS AND METHODS

Sediment samples. Sediment was collected 

from a petroleum hydrocarbon-contaminated, surficial 

aquifer underlying a fuel tank farm near Charleston, 

South Carolina. A zone of heavily contaminated 

sediment containing millimolar concentrations of 

organic acids is located immediately north of tank 3. 

The shallow aquifer at the site is characterized by 

medium-grained sands underlain by predominantly 

clayey material at a depth of 6-11 m. At the time of 

sample collection, the depth to water table was 5.0 m. 

Recharge to the water table at the site occurs from 

infiltration of rainwater and from lateral movement of 

ground water from areas southeast of the facility. 

Ground water flows northwestward from tank 3 at 

rates of 60 to 90 m/yr. During 1990-92, the ground 

water north of tank 3 contained 0.2 mg/L benzene and 

up to 2.5 mg/L toluene.

Sediment was collected at the water table from a 

site where concentrations of organic acids were low

(EW-12). Sediment was collected with a flamesterilized hand auger and stored in sterilized jars. Jars 

were filled with ground water from an adjacent 

monitoring well and sealed without headspace in order 

to preserve anaerobic conditions. Additional water 

was collected without headspace in sterile BOD 

bottles for use in microcosm slurries. Sediment and 

water samples were stored at 4 °C in the dark until 

microcosm experiments were initiated.

Water samples. Water samples for analysis of 

short-chain organic acids and inorganic constituents 

were obtained from 5 cm i.d. diameter PVC 

monitoring wells by means of a teflon bailer. Samples 

were placed in precleaned glass bottles, chilled to 

4 °C, and transported to the laboratory for subsequent 

analysis. In the field, ground-water pH and dissolvedoxygen concentrations were determined according to 

Wood (1976) and hydrogen gas concentrations were 

measured as described by Chapelle and McMahon 

(1991).

Microcosm experiments. Production of CO2 

and organic acids was investigated for a combination 

of four TEAPs and six concentrations of Pb. 

Approximately 17.5 g of saturated sediment (12.5 g 

dry sediment) was transferred asceptically to 40 mL 

sterile serum bottles. Stock solutions of Pb were 

prepared by diluting PbCl2 in ground water collected 

from a monitoring well immediately adjacent to the 

site of sediment sampling. The Pb stocks were 

adjusted to pH 6.5 and flushed with He to eliminate 

oxygen. At the time of sediment collection, the ground 

water from EW-12 contained about 14 u,g/L of 

dissolved Pb. A 5 mL aliquot of the appropriate Pb 

stock was added to each microcosm to yield added Pb 

concentrations ranging from 0 to 1,000 jo.g/L (pH = 

6.5). Aerobic treatments were prepared under room 

atmosphere. Methanogenic treatments were created 

by flushing continuously with He during sediment 

transfer and for 5 min prior to sealing with thick butyl 

rubber stoppers. Results of a preliminary study 

indicated that EW-12 sediments that were treated in 

this manner were dominated by methanogenesis. 

Nitrate and sulfate reducing conditions were created 

by adding anoxic NaNO3 or Na2SO4 stock solution to 

yield final concentrations of 2 mM and then flushing 

with He gas as described above. Triplicate live and a 

single biologically inactive (amended with HgCl2 to 

concentration of 2 mM and autoclaved for 20 min at 

121 °C) treatments were created for each combination 

of Pb and TEAP.

866
Analytical procedures. Production of CO2 in 

treatment bottles was quantified as the change in 

concentration (relative to initial values) after a 24-hour 

incubation period. CH4 production was quantified in 

the same manner for methanogenic treatments. 

Headspace CO2 and CH4 concentrations were 

measured on a gas chromatograph with thermal 

conductivity detection. Dissolved-gas concentrations 

were estimated using Henry's Law coefficients 

(Stumm and Morgan, 1981: p 109). Statistically 

significant differences between treatment means were 

determined by one-way ANOVA and Tukey's 

Studentized Range Test (SAS Institute, 1985). The 

dominant terminal electron accepting process 

occurring in the treatment bottles was identified by 

analysis of headspace H2 gas on a gas chromatograph 

equipped with a reduction gas detector according to 

methods described previously (Chapelle and 

McMahon, 1991, Lovely and Goodwin, 1988).

The extent to which organic acids accumulated 

was evaluated for aerobic, nitrate reducing, sulfate 

reducing, and methanogenic treatments containing 0 

and 1,000 fig/L of added Pb. To determine which 

conditions favored the accumulation of organic acids 

in the presence of a readily degradable substrate, 

treatment vials were amended (following completion 

of the CO2 and CH4 production experiment) with 

glucose to a final concentration of 1.0 mM. To 

determine if the pattern of organic acid accumulation 

differed significantly in treatments where petroleum 

hydrocarbons represented the predominant carbon 

substrate, additional sulfate reducing and 

methanogenic treatments were created without 

glucose amendment. Glucose amended and 

unamended treatments were incubated for 

approximately 5 days. Subsequently, organic acids 

were separated by ion-exclusion chromatography and 

quantified by conductivity detection. Treatment 

bottles were mixed thoroughly, and porewater was 

collected and filtered prior to analysis. Acetic, 

propionic, i-butyric, and n-butyric acids were resolved 

on a Dionex ICE-AS 1 column with 1 uM HC1 at a 

flow rate of 0.8 mL/min. Minimum detection limits 

under these conditions were 0.5 uM for all analytes.

RESULTS

Organic acids were not detected in the ground 

water at EW-12, but ground water from MWGS-34B 

contained concentrations of acetic, propionic,

i-butyric, and n-butyric acids of 9,080; 58; 1,462 and 

1,071 uM, respectively (table 1). Both sites were 

characterized by high concentrations of benzene and 

toluene in the ground water. The maximum Pb 

concentration observed at EW-12 was approximately 

20 ug/L lower than at MWGS-34B.

Table 1. Select characteristics of ground-water from the site 

of sediment collection (EW-12) and the site at which 

maximal concentrations of short-chain organic acids were 

observed (MWGS-34B)3

PARAMETER

NO3 ((.iM)

S04 (^iM)

CH4 (nM)

Pb (^ig/L)b

pH

H2 (nM)

Benzene (ppb)

Toluene (ppb)

Acetic Acid (fiM)

Propionic Acid (^M)

I-Butyric Acid (^M)

N-Butyric Acid (uM)

EW-12

<0.5

31

1 800

I4-I9

5.6

I -5°

72

3400

<5

<5

<5

<5

MWGS-34B

<0.5

6

2300

39

5.2

12-13

99

1 900

9IOO

58

1 500

IIOO

"Ftxcept where noted, data are for samples collected in June of 1992. 

Given as the range of concentrations measured at each site from 

January to July of 1992.

cRange of H2 concentrations detected in ground-water from monitoring wells adjacent to HW-l 2 at time of sample collection.

For all TEAPs examined in this study, the total 

CC>2 produced in 24 hours decreased with increasing 

Pb concentration (fig. I). In the aerobic and nitrate 

reducing treatments, Pb concentrations in the range of 

35-60 ng/L significantly (pF = 0.0001) inhibited CO2 

production. Under methanogenic conditions, CO2 

(fig. I) and CH4 (fig. 2) production was significantly 

inhibited in Pb treatments greater than 50 |ig/L. CO2 

production in the sulfate reducing treatments was not 

significantly affected by Pb additions in the range of 0 

to 250 ug/L. In the low Pb treatments, production of 

CO2 varied significantly (pF < 0.0001) with the TEAP 

in the order aerobic > nitrate reducing > sulfate 

reducing > methanogenic. However, for treatments 

containing at least 500 fig/L Pb, CO2 production was 

significantly higher in the aerobic bottles (pF <0.0005) 

but did not differ significantly between nitrate 

reducing, sulfate reducing, and methanogenic 

treatments. The concentrations of H2 gas measured in 

the anaerobic TEAP treatments were in the range of 

concentrations reported for natural sediments under 

nitrate reducing, sulfate reducing, and methanogenic 

conditions, respectively (table 2).

867
o

E 

^.

o

LJ 

O

ID

0

o

Q£ 

Q_

CM 

O 

O

0 200 400 600 BOO 1000

Pb CONCENTRATION (/xg/L)

Figure 1. CO2 produced in 24 hours as a function of the 

concentration of Pb added to aerobic, nitrate reducing, 

sulfate reducing, and methanogenic treatments. For a 

given TEAP different superscripts denote significantly 

different subsets according to Tukey's studentized range 

test (pF <0.05). The background Pb concentration in the 

study sediment was approximately 14 ^g/L.

Table 2. Concentrations of H2 gas (nM) detected in 

anaerobic microcosms incubated with different Terminal 

Electron Accepting Processes (TEAPs) and Pb treatments

TEAP

CH4

SO4

NO3

[Pb]

M9/L

0

1000

0

1000

0

1000

Min.

4.0

10.1

3.0

3.0

0.6

1.1

[H2]

Max.

11.5

13.5

4.0

5.4

0.8

1.7

Mean3

6.8±4.1

1 1 .8±2.4

3.7±0.7

3.9±1.3

0.7±0. 1

1.3±0.3

o

=> 

o 

o

X

0 200 400 600 BOO 1000 

Pb CONCENTRATIONS (/xg/L)

Figure 2. CH4 produced in 24 hours as a function of the 

concentration of Pb added to methanogenic treatments. 

Different superscripts denote significantly different subsets 

according to Tukey's studentized range test (pF < 0.05). 

The background Pb concentration in the study sediment 

was approximately 14

With glucose amendment, organic acids 

accumulated to detectable levels (0.5 |iM) in the 

sulfate reducing and methanogenic treatments but not 

under aerobic or nitrate reducing conditions (table 3). 

Sulfate treatments contained trace amounts of acetic 

acid (<2 nM) and, in the presence of 1,014 ng/L Pb, 

low concentrations of propionic acid. Under 

methanogenic conditions, high concentrations of all 

four acids were observed, and Pb amendment reduced 

the accumulation of all acids except propionic. 

Without glucose amendment, the same pattern was 

observed. Only acetic acid was detected in the sulfate 

reducing treatments, but all four acids were detected in 

the methanogenic treatments. The concentrations of 

organic acids in the methanogenic treatments were 

approximately an order of magnitude higher in the 

glucose-amended treatments than in the unamended 

microcosms.

Table 3. Concentrations of short-chain organic acids in 

microcosms incubated with different Terminal Electron 

Accepting Processes (TEAP), Pb, and glucose treatments3

TEAP Glucose"

CH4 +

-

+

SO4 +

-

+

NO3 +

+

02 +

+

[Pb]

M9/L

0

0

1000

0

0

1000

0

1000

0

1000

[Acet]

MM

691 ±522

97±50

68±34

1±1

2±1

2±1

<0.5

<0.5

<0.5

<0.5

[Prop] 

MM

46±3

8±6

42±21

<0.5

<0.5

11±9

<0.5

<0.5

<0.5

<0.5

[l-But]

48±34

__ c

23±16

<0.5

<0.5

<0.5

<0.5

<0.5

<0.5

<0.5

[N-But] 

MM

'304±231

38±30

10±2

<0.5

<0.5

<0.5

<0.5

<0.5

<0.5

<0.5

aMeasured values are means ±SD of triplicates.

aMeasured values are meansSD for triplicates. 

bGlucose amended (+) or unamended (-) treatments. 

cNot analyzed.

868
DISCUSSION

Pb concentrations above 14 |j,g/L (background 

concentration in collected sediments) inhibited carbon 

mineralization in these petroleum-contaminated 

sediments. The effect of Pb was apparently TEAP 

dependent with the greatest inhibition occurring under 

oxygen and nitrate reducing conditions. The 

percentage decreases in CO2 production in the 

1,000 U£/L added Pb treatments (relative to 

background) were 61, 59, 51 and 45 % for aerobic, 

nitrate reducing, sulfate reducing, and methanogenic 

conditions, respectively. This observation is 

consistent with the findings of previous investigators 

that sulfate reduction is more sensitive to Pb toxicity 

than methanogenesis (Capone and others, 1983). The 

lack of inhibition of CO2 production by low Pb 

concentrations observed in the sulfate reducing 

treatments was probably due to a reduction in the 

concentration of dissolved Pb by precipitation with 

sulfide. Thus, the Pb treatments reported for sulfate 

reducing conditions may be overestimates of the 

actual dissolved concentrations. Even though Pb 

concentrations in the range observed at the tank farm 

inhibited CO2 production in the methanogenic 

treatments, CH4 production remained unaffected. It is 

not clear whether this response is due to a Pb-induced 

1) reduction in fermentative CO2 production, 2) 

elimination of low levels of competitive CO2 - 

producing processes such as 804 reduction (Capone 

and others, 1983), or 3) a shift from organic acidconsuming mechanisms such as acetotrophic 

methanogenesis (simultaneous CO2 and CH4 

production) toward hydrogenotrophic 

methanogenesis (simultaneous CO2 consumption and 

CH4 production).

The 20 |J.g/L difference in maximum Pb 

concentrations observed between EW-12 and MWGS34B in January 1992 initially suggested a role for Pb 

in the build-up of organic acids at MWGS-34B. 

However, several lines of evidence suggest that Pb 

concentrations are not a major contributor to observed 

organic acid accumulation. As discussed above, CO2 

production was reduced by Pb concentrations in the 

range observed at the site, but CH4 production 

remained unaffected (fig. 2). Moreover, addition of Pb 

under methanogenic conditions appears to inhibit 

fermentative organic acid-production (table 2) as well 

as carbon mineralization (fig. 1). Finally, in the more 

recent water-quality data, Pb remained approximately 

14 |iig/L at EW-12 and organic acids were not detected

while at MWGS-34B organic acid concentrations 

remained high but Pb was not detected. The lack of 

detectable Pb in the recent ground-water samples from 

MWGS-34B diminishes the importance of Pb at this 

site. Although Pb may be a significant factor affecting 

carbon mineralization in the petroleum-contaminated 

sediments at the tank farm, it is unlikely that Pb is a 

primary contributor to organic acid-accumulation.

An alternative explanation for the variation in 

organic acid accumulation at the site involves the 

range of carbon substrates degraded by different 

TEAPs. Aerobic microorganisms are able to oxidize a 

wide number of organic substrates, including the short 

chain organic acids, to CO2 (Gottschalk, 1986 p. 149). 

Thus, even considering the presence of anaerobic 

microsites which pervade saturated sediments, it is 

unlikely that short-chain organic acids would 

accumulate under predominantly aerobic conditions. 

Similarly, denitrification allows complete oxidation of 

the organic substrate to CO2 (Gottschalk, 1986 

p. 123). Moreover, given the greater solubility of NC>3 

relative to O2 , it is even less likely that microsites 

would form in which fermentation products could 

accumulate. In contrast, the substrates for sulfate 

reducers and methanogens are primarily the products 

of fermentative bacteria. Among the sulfate reducers, 

two groups are generally distinguished: 1) those 

species which incompletely oxidize higher 

fermentation products to acetate and 2) those species 

which are able to oxidize these organic substrates, 

including acetate, to CO2 (Gottschalk, 1986 p. 260; 

Widdel, 1986; Widdel, 1988; Chapelle, 1993). Thus, 

accumulation of acetate may occur in an anaerobic 

environment dominated by group 1 sulfate reducers, 

but in general accumulation of the short-chain acids to 

high concentrations would not be expected in an 

environment dominated by a mixed community of 

sulfate reducers. Unlike other anaerobic respirers, 

methanogens are able to metabolize only a limited 

number of organic substrates. The acknowledged 

substrates for methanogens include H2 and CO2 , 

formate, acetate, methanol, and methylated amines 

(Gottschalk, 1986 p. 252; Oremland, 1988; Chapelle, 

1993). Degradation of other fermentation products by 

methanogenic systems requires a syntrophic 

association between methanogens and the bacteria 

which cleave these substrates to acetate, CO2 , and H2 

(Archer and Harris, 1986; Gottschalk, 1986 p. 265; 

Widdel, 1986). Thus, accumulation of short-chain 

organic acids may be expected under methanogenic

869
conditions in the absence of acetogenic, 

hydrogenogenic syntrophs. Moreover, since the 

substrate affinity for acetate (Archer and Harris, 1986; 

Widdel, 1986) is much lower for methanogens than 

sulfate reducers, a higher acetate concentration under 

methanogenic conditions may occur as well.

Several studies have observed an accumulation 

of short-chain organic acids in petroleumcontaminated sediments under anaerobic conditions. 

A laboratory investigation of benzene and toluene 

biodegradation under methanogenic conditions 

revealed the production of several aliphatic acids, 

including acetic and butyric acids (Grbic-Galic and 

Vogel, 1987). In a shallow aquifer system, mg/L 

concentrations of acetic acid and unidentified CyC^ 

volatile organic acids were detected in methanogenic 

sediments immediately down gradient from the source 

of creosote- and petroleum-containing waste water 

(Godsy and others, 1992). High concentrations of 

organic acids, primarily acetic acid, were detected in 

the methanogenic zone adjacent to a crude-oil spill 

(Cozarelli and others, 1990).

The results of the current investigation indicate 

that the TEAP is the primary factor contributing to the 

accumulation of short-chain organic acids in 

petroleum-contaminated sediments. As expected, no 

short-chain organic acids were detected in either the 

aerobic or the nitrate reducing microcosms. Organic 

acids accumulated in sulfate reducing and 

methanogenic microcosms under glucose amended 

and unamended conditions. Glucose amendment 

increased the concentration but not the species of 

accumulated organic acids. A similar response to 

glucose amendment was reported for methanogenic 

acid-bog sediments (Goodwin and Zeikus, 1987). All 

four acids accumulated under methanogenic 

conditions, but only trace amounts of acetic acid were 

detected in the sulfate-reducing treatments.

The relation between organic acids 

accumulation and TEAP observed in the lab study is 

consistent with what is observed in the field. In this 

study, H2 gas concentrations were used as a primary 

indicator of the TEAP which dominated the sediment 

microbial processes, because H2 is a short lived 

(rapidly turned-over) and ubiquitous intermediate of 

microbial metabolism. Thus, the H2 concentration 

measured at a site is characteristic of the dominant 

metabolism at that location (Lovely and Goodwin, 

1988) while, in contrast, respiratory products such as 

sulfide and CH4 can be transported over great

distances by ground water flow. The 12-13 nM H2 gas 

concentrations observed at MWGS-34B, the site of 

organic acid accumulation, are in the range 

characteristic of methanogenic sediments (Chapelle 

and McMahon, 1991, Lovely and Goodwin, 1988). 

Moreover, the low 804 (6 f*M) concentration 

measured at MWGS-34B is consistent with 

methanogenesis. The H2 concentrations at EW-12 

(1-5 nM) were indicative of predominantly sulfate 

reducing conditions and organic acids were not 

detected at this site. The higher SO4 concentration at 

EW-12, also, was consistent with sulfate reducing 

conditions. These observations, which are consistent 

with previous studies (Grbic-Galic and Vogel, 1987; 

Cozzarelli and others, 1990; Godsy and others, 1992), 

suggest that the accumulation of high concentrations 

of organic acids, especially the butyrates, primarily 

result from the biodegradation of petroleum 

hydrocarbons in predominantly methanogenic 

environments.

REFERENCES

Archer, D.B., and. Harris, J.E., 1986, Methanogenic

bacteria and methane production in various habitats, 

Anaerobic Bacteria In Habitats Other Than Man: 

Boston, Blackwell Scientific Publications, p. 185-224.

Capone, D.G., Reese, D.D., and Kiene, R.P.. 1983, Effects 

of metals on methanogenesis, sulfate reduction, carbon 

dioxide evolution, and microbial biomass in anoxic salt 

marsh sediments: Appl. Environ. Microbiol. v. 45, 

p. 1586-1591.

Chapelle, F.H., 1993, Bacterial metabolism: Ground-Water 

Microbiology and Geochemistry, John Wiley and 

Sons, New York, p. 71 -104.

Chapelle, F.H., and McMahon, P.B., 1991, Geochemistry of 

dissolved inorganic carbon in a Coastal-Plain aquifer 

I. Sulfate from confining beds as an oxidant in 

microbial CC>2 production: J. Hydrology, v. 127, p. 85- 

108.

Cozarelli, I.M., Eganhouse, R.P., and Baedecker, M.J., 

1990, Transformation of monoaromatic hydrocarbons 

to organic acids in anoxic ground-water environments: 

Environ. Geol. Water Sci., v. 16, p 135-141.

Godsy, E.M., Goerlitz, D.F., Grbic-Galic, D., 1992, 

Methanogenic biodegradation of creosote 

contaminants in natural and simulated ground-water 

ecosystems: Gound Water, v. 30, p. 232-242.

Goodwin, S., and Zeikus, J.G., 1987, Ecophysiological 

adaptations of anaerobic bacteria to low pH Analysis 

of anaerobic digestion in acidic bog sediments: Appl. 

Environ. Microbiol. v. 53, p. 57-64.

870
Gottschalk, G., 1986, Bacterial Metabolism, SpringerVerlag, New York. p. 359.

Grbic-Galic, D., and Vogel, T.M., 1987, Transformation of 

toluene and benzene by mixed methanogenic cultures: 

Appl. Environ. Microbiol, v. 53, p. 254-260.

Hiebert, F.K. and Bennett, P.C., 1992, Microbial control of 

silicate weathering in organic rich ground water: 

Science, v. 258, p. 278-281.

Jarrell, K.F., Saulnier, M., and Ley, A., 1987, Inhibition of 

methanogenesis in pure cultures by ammonia, fatty 

acids, and heavy metals and protection against heavy 

metal toxicity in sewage sludge: Can. J. Microbiol, 

v. 33, p. 551-554.

Lovely, D.R., Goodwin, S., 1988, Hydrogen concentrations 

as an indicator of the predominant electron-accepting 

reactions in aquatic sediments: Geochim. Cosmochim. 

Acta., v. 52, p. 2993-3003.

McMahon, P.B. and Chapelle, F.H., 1991, Geochemistry of 

dissolved inorganic carbon in a coastal plain aquifer, 

2. modeling carbon sources, sinks and 813C evolution: 

J. Hydrology, v. 127, p. 109-135.

McMahon, P.B., Chapelle, F.H., Falls, W.F., and Bradley, 

P.M., 1992, Role of microbial processes in linking 

sandstone diagenesis with organic-rich clays: J. 

Sediment, Petrology, v. 62, p. 1-10.

Oremland, R.S., 1988, Biogeochemistry of methanogenic 

bacteria: Biology Of Anaerobic Microorganisms, John 

Wiley and Sons, New York, p. 641-707.

SAS Institute, 1985, SAS User's Guide: Statistics, Cary, 

NC.

Stumm, W., and Morgan, J.J.. 1981, Aquatic Chemistry: 

John Wiley and Sons, New York, NY, 780 p.

Widdel, F., 1986, Sulphate-reducing bacteria and their 

ecological niches: Anaerobic Bacteria In Habitats 

Other Than Man., Boston, Blackwell Scientific 

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Widdel, F., 1988, Microbiology and ecology of sulfate- and 

sulfur-reducing bacteria: Biology Of Anaerobic 

Microorganisms, New York, John Wiley and Sons, 

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24 p.

871
872
Shifting of Terminal-Electron-Accepting Processes in a 

Petroleum-Hydrocarbon-Contaminated Aquifer

By Don A. Vroblesky 1 and Francis H. Chapelle1

Abstract

Measurements of dissolved hydrogen and 

other biologically active solutes in ground water 

from a shallow petroleum hydrocarboncontaminated aquifer indicate that the distribution 

of microbial terminal electron accepting 

processes (TEAP's), such as methanogenesis, 

sulfate reduction, and ferric-iron reduction, is 

highly dynamic in both time and space. Delivery 

of sulfate to previously methanogenic zones by 

infiltrating recharge or lateral transport can result 

in a TEAP shift from methanogenesis to sulfate 

reduction. Conversely, lack of recharge and 

consumption of available sulfate can result in a 

shift from sulfate reduction to methanogenesis. 

Temporal shifts between sulfate and ferric-iron 

reduction were also observed.

INTRODUCTION

Identifying the dominant terminal-electronaccepting processes (TEAP's) in contaminated ground 

water is an important step in determining the fate of 

contaminants. Although petroleum hydrocarbons are 

most readily degraded under aerobic conditions, it is 

apparent that degradation also occurs under 

methanogenic (Grbic-Galic and Vogel, 1987), sulfate 

(SO4)-reducing (Beller and others, 1992; Edwards and 

others, 1992; Edwards and Grbic-Galic, 1992), ferric 

iron (Fev)-reducing (Lovley and others, 1989), and 

nitrate (NO3 )-reducing (Hutchins and others, 1991) 

conditions. Furthermore, there is evidence that 

biodegradation rates depend on TEAP's. For example,

'U.S. Geological Survey, Columbia, South Carolina

Smolensk! and Suflita (1987) showed that the rate of 

biodegradation of several cresol isomers was faster 

under sulfate-reducing conditions than under 

methanogenic conditions. This observation raises the 

possibility that differences in anaerobic TEAP 

conditions may directly affect microbial degradation 

rates of hydrocarbon contamination in ground water.

Baedecker and others (1993) observed areal 

differences in TEAP conditions in oil-contaminated 

ground water, and they noted that changes in 

concentrations of microbially active constituents in the 

oil-contaminated ground water implied that the TEAP 

was changing from iron reduction to methanogenesis 

over a 5-year period. Little work has been done, 

however, in examining short-term variations of 

TEAP's in contaminated ground water or in 

determining the implications of these changes on 

biodegradation of the contamination. The purpose of 

this paper is to examine spatial and temporal 

variations in TEAP's in a shallow, petroleumhydrocarbon-contaminated aquifer.

Background

Historically, methods used to identify TEAP's in 

ground-water systems have focused on documenting 

the consumption of particular terminal-electron 

acceptors, such as nitrate (Baedecker and Back, 1979) 

or sulfate (Thorstenson and others, 1979), that support 

microbial metabolism. Alternatively, metabolic end 

products, such as methane (CH4 ) (Thorstenson and 

others, 1979; Baedecker and others, 1988), sulfide 

(H2S) (Rightmire and Hanshaw, 1973; Jackson and 

Patterson, 1982), or ferrous iron (Fe2+) (Baedecker and 

others, 1988), have been used as a criteria to identify 

the dominant TEAP's. Because some electron 

acceptors, such as SO4, can be replenished from 

mineral sources, and because most final products

873
(methane, Fe2+ , and sulfide) are transported by flowing 

ground water, these methods are commonly equivocal.

The use of hydrogen (H2 ) -- a transient 

intermediate product of anaerobic microbial 

metabolism ~ as an indicator of predominant TEAP's 

was introduced by Lovley and Goodwin (1988). This 

approach has been used to document the zonation of 

TEAP's in various systems by Chapelle and Lovley 

(1990), Chapelle and McMahon (1991), and Chapelle 

and Lovley (1992). These studies indicate that H2 

concentration ranges of 5 to 25 nmol/L are 

characteristic of methanogenesis, 1 to 4 nmol/L are 

characteristic of SO4 reduction, and 0.1 to 0.8 nmol/L 

are characteristic of Feu reduction. Because H2 is an 

extremely transitory intermediate, with a half life of 

less than a minute (Conrad and others, 1987), it is an 

ideal constituent for documenting temporal as well as 

spatial variations in TEAP's. In this study, H2 

concentrations were used in conjunction with 

concentrations of other microbially active solutes to

79°59'10"

document the spatial and temporal variations in 

TEAP's.

Description of Study Area

The study area consisted of a petroleum-tank 

farm, referred to hereafter as the facility, in Hanahan, 

South Carolina, and the area surrounding the facility 

(fig. 1). The site is underlain by unconsolidated 

Pleistocene sediments composed of quartz sand and 

lenses of clay. The base of the Pleistocene sediments 

is at a depth of about 1 1 m. Below the Pleistocene 

sediment is a dense clay, approximately 65 m thick, 

that vertically isolates the shallow ground- water-flow 

system. A thin layer (about 5 cm) of carbonate 

material covers the bermed basins containing the fuel 

tanks.

The Pleistocene sediment functions as an 

unconfined water-bearing zone, but clay lenses locally 

produce confined-flow conditions. This surficial

79° 59'

 .3. 

EXPLANATION

LINE OF EQUAL CONCENTRATION OF 

BTEX-Oashed wtier« approximate. 

Contour interval in micrograms per liter, 

variable es shown.

DIRECTION DF GROUND-WATER 

MOVEMENT

DATA POINT AND BTEX 

CONCENTRATION IN MICROGRAMS 

PER LITER

 T » ^ 540 \ 40

10 7501 3860 \ 1230 s ^-~ ^^ tfr^V&sx-f- vx*&oai&.,>**y**c<>yxx*x***>^^ ^ m W

32°54'50" -

^^^.i,.^. > -,>,,^ .,>^ .j:.i.. . X,r...f*+MwX*M*'

£ A "-Vr ^Vr^

/^~~x

1 0 . . "^^w B ^C w-"- ^ ^ 

0 1 2030 1850 143°>>

We// / 

AM-4O4

^^\ /~X

Tank Tank 3

600 800 FEET

0 200 METERS

Figure 1. Location of study area in South Carolina, location of wells MW-5A and MW-40, and distribution of benzene, 

toluene, ethylbenzene, and total xylene (collectively termed BTEX) Site location and distribution of BTEX in ground water 

during late June to early July 1991 (one sampling).

874
aquifer is recharged by rainwater infiltration. Periods 

of intense rainfall can increase ground-water levels by 

0.6 m over several days. Ground-water levels 

declined about 1.2 m during a drought between late 

September 1991 to early January 1992.

Ground-water discharge at the site is by 

evapotranspiration and by flow to nearby drains and 

streams. The predominant direction of ground-water 

flow beneath the facility is northwestward (fig. 1). 

Ground water at the site is contaminated with benzene, 

toluene, ethyl benzene, and xylene (collectively 

termed BTEX) (fig. 1).

Although Feu-rich grain coatings are common 

in uncontaminated parts of the aquifer, the sediment in 

petroleum-contaminated parts of the aquifer is white 

to light gray with no Feu-rich grain coatings. The 

substantial precipitation of Fe-u in surface water where 

contaminated ground water discharges attests to 

widespread removal of Fe3+ from the aquifer 

sediments by reduction in the contaminant plume.

Methods

Temporal variations in TEAP's were monitored 

in the surficial aquifer at wells for varying lengths of 

time during June 1991 - January 1993. The two wells 

discussed here are screened at the water table (well 

MW-40A) and from about 4.9 to 7.9 m below the 

water table (well MW-5A). Data used to discuss 

temporal variation in spatial zonation of TEAP's were 

collected from 19 wells in late June to early July 1991 

and from 27 wells in November 1992.

Water samples for analysis of BTEX were collected by 

slowly filling 40-mL glass vials from a bottomdischarge bailer. The vials were allowed to overflow 

several seconds and were then preserved with 

hydrochloric acid and capped with Teflon-lined septa. 

If aeration of a sample was suspected, or if bubbles 

were observed in a vial, the sample was replaced. On 

the day of collection, the water samples for BTEX were 

delivered to a private laboratory for analysis.

Immediately following collection of BTEX 

samples from each well, samples for inorganic ions 

were collected by peristaltic pump. Well water was 

pumped through a 0.45-um filter into sample-rinsed 

polyethylene bottles. Ferrous iron was analyzed using 

a field colorimetric method. NO? and SO4 were 

separated in the laboratory by an ion-exchange 

chromatography and quantified by conductivity 

detection. CH4 samples were collected using a 

syringe with an attached 0.45-jim filter to inject 5 mL 

of water into septated vials. CH4 samples were 

refrigerated to minimize microbial activity. CH4 was 

quantified by flame ionization detection gas 

chromatography. Dissolved CH4 concentrations were 

calculated using Henry's Law coefficients (Stumm 

and Morgan, 1981).

Ground-water H2 samples were collected using 

the bubble-strip method of Chapelle and McMahon 

(1991). A stream of water was pumped from the well 

through a gas-sampling bulb at a rate of about 600 mL 

per minute. An injected bubble of nitrogen in the bulb 

asymptotically collected H2 and other soluble gases 

until equilibrium was achieved. Once equilibrium was 

achieved (less than 5 percent change in 5 minutes, 

which occurred within 15 minutes), gas was extracted 

from the bulb by syringe. H2 was measured on a gas 

chromatograph equipped with a reduction gas detector. 

The detection limit of this method was approximately 

0.01 uL/L. H2 samples were not collected if the 

ground water at the well was found to be aerobic.

RESULTS

Ground water collected from wells MW-5A and 

MW-40A contained differing concentrations of 

several constituents (table 1). Temporal variations in 

the concentrations of H2 were found in the ground 

water at each of the wells (figs. 2 and 3).

In addition, temporal variations in the spatial 

zonation of H2 concentrations were found. Temporal

Table 1. Range of concentrations of relevant microbially affected constituents in ground water at wells MW-5A and MW-40A 

for June 1991-January 1993

[mg/L, milligrams per liter; nmol/L. nanomoles per liter; |amol/L. micromoles per liter; ng/L, micrograms per liter; BTFX, the sum of benzene, toluene, 

ethylbenzene. and total xylene. Parentheses indicate number of samples]_______________________________________________

Well Hydrogen 

(nmol/L) Methane (nmol/L) Sulfate (mg/L) Ferric iron (mg/L) Nitrate (mg/L) BTEX

MW-5A 

MW-40A

1.2-12.5(8) 

0.6-1.4(4)

1,120-1,770(5) 

80-680 (5)

Concentration

0.01-2.6 (8) 

12.1-23.17(5)

6.0-30.0(7) 

2.8-8.4(6)

<0.0 1-0. 12(8) 

0.14-0.99 (5)

9-40 (7) 

40-770 (7)

875
JFMAMJ J ASOND

Figure 2. H2 (A) and SO4 (B) concentrations in ground 

water at well MW-5A for January 16,1992 - January 7, 

1993.

changes in ground-water H2 concentrations differed 

over short distances horizontally (<5 m) and vertically 

(<3.5 m) at this site. Moreover, distinct changes were 

found in the areal distribution of H2 concentrations 

from June-July 1991 to November 1992 (fig. 4).

DISCUSSION

The ground-water-chemistry data from wells 

screened in different areas of ground-water 

contamination show that the predominant TEAP's 

change with time. The changes differed from site to 

site, depending on differences in the ground-water 

chemistry in various parts of the contamination. Well 

MW-5A is an example of a site where concentrations 

of BTEX are low (9-40 fJ.g/L), concentrations of 

dissolved Fe2+ are high (up to 40 mg/L) and 

concentrations of SO4 are low (up to 2.6 mg/L). In 

January 1992, the H2 concentration (1.7 nmol/L) (fig. 

2A) indicated that SO4 reduction was the predominant 

TEAR Supporting evidence for SO4 reduction was the 

presence of dissolved SO4 (0.88 mg/L) and the 

decreasing concentrations of SO4 (fig. 5B) in ground

DATE

Figure 3. H2 (A) and Fe2+ (B) concentrations in ground 

water at well MW-40A for July 13,1992 to January 5,1993.

water at well MW-5A over the next few months. In 

July and November 1992, the H2 concentration 

increased (9.2 and 15.4 nmol/L, respectively), 

indicating that the predominant TEAP in the ground 

water at well MW-5A had changed to methanogenesis 

(fig. 2A). The SO4 concentrations (0.14 and 0.2 

nmol/L, respectively) were apparently too low to 

support SO4 reduction as the predominant TEAP. By 

January 1993, the SO4 concentration had increased to 

2.6 mg/L (fig. 2B). The corresponding decrease in H2 

concentration (to 1.2 nmol/L) indicated that the 

presence of SO4 had resulted in a shift in the 

predominant TEAP from methanogenesis back to SO4 

reduction (fig. 2A).

Well MW-40A is an example of an area where 

ground water became oxygenated after a period of 

heavy rainfall. When the well was first sampled in 

July 1992, the H2 concentration (1.4 Nm/L) was in the 

range characteristic of SO4 reduction (fig. 3A). In 

August 1992, however, a period of rainfall provided 

substantial recharge to the aquifer. By the time well 

MW-40A was sampled in early September 1992, the 

ground water at the site had become oxygenated. The 

corresponding sharp decrease in dissolved Fe2+

876
79°59'10" 79°59'

32°55' -

32°54'50

June-July 1991

EXPLANATION

, WELL AND HYDROGEN 

°- 1 CONCENTRATION, IN 

NANOMOLES PER LITER

BOUNDARY BETWEEN 

TERMINAL-ELECTRONACCEPTING PROCESSES

79°59'10" 79°59'

32°55'

32°541 50"

WELL AND HYDROGEN 

0 1 CONCENTRATION, IN

NANOMOLES PER LITER

BOUNDARY BETWEEN 

TERMINAL-ELECTRONACCEPTING PROCESSES

>-.-» 3 - 9 SULFATE-REDUC

1.9 2.6 2.9

200 METERS

Figure 4. Areal distribution of terminal-electron-accepting processes, based on interpretation of H2 measurements in ground 

water for (A) June-July 1991 and (B) November 1992.

877
concentrations (8.4 mg/L in July 1992 to 2.8 mg/L in 

September 1992) (fig. 3B) indicates that Fe2+ was oxidized to Fe3+ and precipitated out of solution as grain 

coatings. By late September 1992, the dissolved oxygen had been depleted. At least part of the electron 

flow from oxidation of organic matter was sequestered 

by Fe3+ reduction, as implied by the decrease in H2 

concentration to 0.7 nmol/L, in the range of H2 concentration characteristic of Fe3+ reduction (fig. 3A). A 

corresponding increase in dissolved Fe2+ may not have 

occurred because of precipitation of iron sulfide or 

other minerals.

SUMMARY AND CONCLUSIONS

The predominant TEAP in anaerobic ground 

water contaminated with petroleum hydrocarbons at a 

tank farm in Hanahan, South Carolina, varies with 

time as electron acceptors are depleted or added. 

During times when little or no SO4 or oxygen are 

introduced to the ground water by recharge from 

rainfall, the available oxygen, Fe3+ , and SO4 can be 

depleted by respiring bacteria, leaving methanogenesis 

as the predominant TEAP. Introduction of oxygen, 

from rainfall infiltration, into the anaerobic parts of the 

aquifer can cause precipitation of Fe3+ as grain 

coatings. Once the oxygen is depleted, Fe^-reducing 

bacteria can take advantage of the precipitated Fe3+ to 

sequester the electron flow from degradation of 

organic compounds. If the water table is lowered 

below the zone of Fe3+ availability, electron flow can 

be sequestered by less efficient TEAP's, such as SO4 

reduction. Introduction of SO4 into a methanogenic 

part of the aquifer can cause a shift in the predominant 

TEAP to SO4 reduction.

REFERENCES

Acton, D.W., and Barker, J.F., 1992, In situ biodegradation 

potential of aromatic hydrocarbons in anaerobic 

groundwaters: Journal of Contaminant Hydrology, 

v. 9, p. 325-352.

Baedecker, M.J., and Back, W., 1979, Modern marine 

sediments as a natural analog to the chemically 

stressed environment of a landfill: Journal of 

Hydrology, v. 43, p. 393-414.

Baedecker, M.J., Seigel, D.I., Bennett, P., Cozzarelli, I.M., 

1988, The fate and effects of crude oil in a shallow 

aquifer, in, Mallard, G.E., and Ragone, S.E., eds., 

Proceedings of the Technical Meeting of the U.S. 

Geological Survey Toxic Substances Hydrology 

Program, U.S. Geological Survey, Phoenix, Ariz., 

p. 13-20.

Baedecker, M.J., Cozzarelli, I.M., Siegel, D.I., Bennett, 

P.C., and Eagenhouse, R.P., 1993, Crude oil in a 

shallow sand and gravel aquiferm, 3 Biogeochemical 

reactions and mass balance modeling in anoxic ground 

water: Applied Geochemistry, v. 8, no. 6, p. 569-586.

Seller, H.R., Reinhard, M., Grbic-Galic, D., 1992, 

Metabolic by-products of anaerobic toluene 

degradation by sulfate-reducing enrichment cultures: 

Applied and Environmental Microbiology, v. 58, 

p. 3192-3195.

Chapelle, F.H., and Lovley, D.R., 1990, Hydrogen 

concentrations in ground water as an indicator of 

bacterial processes in deep aquifer systems, in, 

Fliermans, C.B., and Hazen, T.C., eds.. Proceedings of 

the First International Symposium on the 

Microbiology of the Deep Subsurface: U.S. 

Environmental Protection Agency, Orlando, Fla., 

p. 2-123-2-124.

 , 1992, Competitive exclusion of sulfate reduction 

by Fe(IIl)-reducing bacteria A mechanism for producing discrete zones of high-iron ground water: 

Ground Water, v. 30, p. 29-36.

Chapelle, F.H., and McMahon, P.B., 1991, Geochemistry of 

dissolved inorganic carbon in a coastal plain aquifer, 

1 Sulfate from confining beds as an oxidant in 

microbial CO2 production: Journal of Hydrology, 

v. 127, p. 85-108.

Conrad, R., Lupton, F.S., and Zcikus, J.G., 1987, Hydrogen 

metabolism in a mildly acidic lake sediment (Lake 

Mendota): FEMS Microbiology and Ecology, v. 45, 

p. 107-115.

Edwards, E.A., and Grbic-Galic, D., 1992, Complete

mineralization of benzene by aquifer microorganisms 

under strictly anaerobic conditions: Applied and 

Environmental Microbiology, v. 58, p. 2663-2666.

Edwards, E.A., Wells, L.E., Reinhard, M., Grbic-Galic, D., 

1992, Anaerobic degradation of toluene and xylene by 

aquifer microorganisms under sulfate-reducing 

conditions: Applied and Environmental Microbiology, 

v. 58, p. 794-800.

Grbic-Galic, D. and Vogel, T.M., 1992, Transformation of 

toluene and benzene by mixed methanogenic cultures: 

Applied and Environmental Microbiology, v. 53, 

p. 254-260.

878
Hutchins, S.R., Downs, W.C., Wilson, J.T., Smithe, G.B., 

Kovacs, D.A., Fine, D.D., Douglass, R.H., and 

Hendrix, D.J., 1991, Effect of nitrate addition on 

biorestoration of fuel-contaminated aquifer field 

demonstration: Ground Water, v. 29, no. 4, p. 571-580.

Jackson, R.E., and Patterson, R.J., 1982, Interpretation of 

pH and Eh trends in a fluvial-sand aquifer system: 

Water Resources Research, v. 18, p. 1255-1268.

Lovley, D.R., Baedecker, M.J., Lonergan, D.J., Cozarelli, 

I.M., Phillips, E.J., and Siegel, D.I., 1989, Oxidation 

of aromatic contaminants coupled to microbial iron 

reduction, Nature, v. 339, p. 297-299.

Lovley, D.R., and Goodwin, S., 1988, Hydrogen

concentrations as an indicator of the predominant 

terminal electron-accepting reactions in aquatic 

sediments: Geochimica et Cosmochimica Acta, v. 52, 

p. 2993-3003.

Rightmire, C.T., and Hanshaw, B.B.,1973, Relationship 

between the carbon isotope composition of soil CO2 

and dissolved carbonate species in ground water: 

Water Resources Research, v. 9, p. 958-967.

Smolenski, W.J., and Sulflita, J.M., 1987, Biodegradation of 

cresol isomers in anoxic aquifers: Applied and 

Environmental Microbiology, v. 53, p. 710-716.

Stumm, Werner, and Morgan, J.J, 1981, Aquatic chemistry 

(2d ed.), N.Y., John Wiley & Sons, 780 p.

Thorstenson, B.C., Fisher, D.W., and Croft, M.G., 1979, The 

geochemistry of the Fox Hills-Basal Hell Creek 

aquifer in southwestern North Dakota and 

northwestern South Dakota: Water Resources 

Research, v. 15, p. 1479-1498.

879
088
Preliminary Assessment of Jet Fuel Contamination

of an Atoll Aquifer

By Gordon W. Tribble1 , Roger W. Lee2 , Duane S. Wydoski3 , 

Jill D. Torikai 1 , and Charles D. Hunt, Jr. 1

Abstract

Jet fuel spilled in 1991 on the atoll island of 

Diego Garcia (Chagos Archipelago, Indian Ocean) 

has contaminated one of the fresh ground-water 

lenses on the island. Freshwater lenses on the 

island are unconfined and strongly influenced by 

tidal oscillations in sea level. The U.S. Geological 

Survey is using a multidisciplinary approach to (1) 

evaluate contaminant migration and, (2) determine 

the geochemical processes that affect contaminant 

degradation. A network of water-level recorders is 

being used to determine the effect of a hydraulic 

control scheme on the water-table configuration 

and contaminant transport. The water table near 

the spill is 1 to 2 feet above mean sea level and 

undergoes diurnal oscillations of as much as 4 feet 

in response to ocean tides.

Geochemical characterization of the aquifer 

will help to resolve ambient and fuel-affected 

levels and types of microbial oxidation. This 

information is useful to determine the need for, 

and type of, remediation of residual 

contamination after efforts to remove the fuel by 

pumping are completed. Initial data from the first 

of three planned trips indicate that natural levels 

of microbial oxidation are sufficient to deplete C>2 

in all aquifer waters. Concentrations of CO2, 

CH4, and H2S indicate that the natural levels of 

microbial oxidation are enhanced by the presence 

of hydrocarbon contaminants. An evaluation of

'U.S. Geological Survey, Honolulu, Hawaii 

"U.S. Geological Survey, Austin, Texas 

5 U.S. Geological Survey, Arvada, Colorado

the rates, mechanisms, and limits of microbial 

oxidation in the aquifer will require additional 

sampling and physical measurements to assess 

hydraulic processes such as aquifer properties and 

the response of ground-water flow to different 

recharge and withdrawal conditions.

INTRODUCTION

Diego Garcia is a low, 20 ft maximum altitude, 

coral atoll of the Chagos Archipelago in the central 

Indian Ocean at latitude 7°S and longitude 72°E 

(fig. 1). It is a part of the British Indian Ocean 

Territory and the site of a U.S. Navy Support Facility 

(Navy) jointly administered by the United Kingdom 

and the United States. The facility is used to support 

military operations in the Indian Ocean and nearby 

regions. The atoll has about 3,000 inhabitants. In 

1991, fuel leaking from a buried pipeline introduced 

contaminants into a water-table aquifer underlying the 

Air Operations area (fig. 1, U.S. Navy, oral commun., 

1991). At the Navy's request, the U.S. Geological 

Survey (USGS) is evaluating the hydrologic and 

geochemical aspects of the fuel spill to provide a 

conceptual basis for aquifer remediation. The USGS 

acknowledges and thanks the U.S. Navy for funding 

and logistic support. This paper describes the 

contaminated site, lists tasks currently underway, and 

presents some preliminary findings with an emphasis 

on the geochemical aspects of the project.

About 98 percent of all drinking water on the 

atoll is supplied by thin freshwater lenses in the 

Cantonment and Air Operations areas of Diego 

Garcia. The water table is typically 5 ft below land 

surface and the aquifer material is unconsolidated 

calcareous sediments. The two freshwater lenses are 

less than about 60 ft thick and less than 1.5 mi2 in area.

881
40° 80° 120° E

7°15'N

40°

0 2,000 KILOMETERS

I_________________I_____

7°20'

LOCATION MAP

7°25'

______72°25' E

I

P * 

DIEGO GARCIA 

ATOLL

72°30' E

r~

Air Operations

*D.

Study area ' "^iS3\ Dieeo Garcia

STUDY AREA LOCATION MAP

EXPLANATION

 -- HORIZONTAL WATER-SUPPLY WELL

 CHEMICAL MONITORING WELL

 WATER-LEVEL OBSERVATION WELL 

A EXPLORATION WELL

Water supply wells

Figure 1 . Location of study area and types of wells on Diego Garcia.

882
Ground-water withdrawal from the lenses averages 

about 1.2 Mgal/d. About 30 percent of pumpage is 

from the Air Operations area, near the site of the spill 

(fig. 1). Although average annual rainfall is 106 in., 

the quantity of freshwater stored in ground water is 

small and the aquifer is highly susceptible to saltwater 

intrusion owing to seasonal and interannual variations 

in rainfall (Hunt, 1991). To avoid saltwater upconing, 

more than 100 shallow horizontal production wells 

have been constructed to skim freshwater from the two 

lenses and distribute the effect of pumpage over a 

broad area. Salinity acceptable for drinking has been 

maintained over the past decade, with chloride concentrations ranging between 30 and 130 mg/L. Before 

May 1991, pumpage from the Air Operations area was 

0.4 Mgal/d (C.D. Hunt, USGS, written commun., 

1993).

Aquifers on atolls respond dynamically to both 

physical and chemical processes. Variable quantities 

of recharge, combined with relatively low storage, 

cause the freshwater part of the aquifer to shrink and 

expand in response to seasonal and interannual 

variations in rainfall (Hunt, 1991). Sea-level 

fluctuations owing to tides are transmitted through the 

aquifer and cause diurnal oscillations of several feet in 

the water table. The oscillations also create a broad 

transition zone between infiltrated rainwater and 

underlying saltwater (Buddemeier and Holladay, 

1977; Hunt and Peterson, 1980; Wheatcraft and 

Buddemeier, 1981; Ayers and Vacher, 1986; Anthony 

and others, 1989). The oxidation of organic matter is 

pervasive in many atoll aquifers, resulting in naturally 

low concentrations of dissolved oxygen, elevated 

concentrations of dissolved inorganic carbon, and 

production of hydrogen sulfide (Buddemeier and 

Oberdorfer, 1986). Changes in water chemistry 

associated with the oxidation of organic matter change 

the saturation state of the water with respect to the 

carbonate minerals that constitute the aquifer. Organic 

matter oxidation and carbonate mineral diagenesis 

cause water chemistry to differ from that expected 

from simple conservative mixing of rainwater and 

seawater.

Several characteristics make this site attractive 

for a multidisciplinary study of contaminant 

hydrogeochemistry: (1) The seawater surrounding the 

atoll provides sharp physical and geochemical 

delineation to the aquifer; (2) constantly warm 

subsurface temperatures provide a stable environment 

favorable to microbial metabolism; (3) aquifer waters

are derived solely from seawater and rainwater, so the 

stoichiometry and extent of specific reactions can be 

deduced from the nonconservative characteristics of 

selected constituents; (4) the calcium carbonate 

mineralogy of the aquifer provides a geochemically 

simple, albeit reactive, matrix; and (5) contamination 

at the site is thought to have originated from a single 

leak, so complications associated with multiple 

releases over time and space are not likely. Logistics, 

however, are difficult because of the distance between 

Diego Garcia and the United States.

PROBLEM

In the mid-1980s, an aircraft-refueling support 

base was constructed at the Air Operations area (fig. 1). 

In May 1991, a leak of JP-5 jet fuel from a pipeline 

beneath the apron was detected. Because of the history 

of use, the leak probably occurred sometime after 

January 1991. Much of the spill is overlain by several 

feet of compacted carbonate sand and gravel fill 

covered with 18 in. of reinforced concrete.

The volume of spilled fuel is unknown. 

Following discovery of the spill, the Navy started a 

program of exploratory drilling and fuel recovery. As 

of January 1993, more than 120,000 gal of fuel has 

been recovered by pumping free-floating fuel from 

shallow pits. Several nearby horizontal water-supply 

wells in the Air Operations area have been shut down 

so that pumping will not entrain fuel towards the 

wells. The wells remain shut down pending the results 

of ongoing studies. The lost production capacity is 

about 0.5 Mgal/d, about 40 percent of the present 

island-wide pumpage of ground water. To offset the 

loss, withdrawal of ground water has been increased 

elsewhere, although this could raise the salinity to 

levels unacceptable for drinking if the increased 

pumpage coincides with a drought. Consequently, the 

Navy intends to restore lost pumping capacity through 

a program of aquifer remediation.

This project has focused on the physical 

processes that determine the transport of contaminants 

and the geochemical processes that affect contaminant 

degradation. A combination of ground-water 

modeling, hydrologic monitoring, and chemical 

sampling is being used to determine these processes 

and to evaluate the effectiveness of remedial actions. 

The USGS is providing the Navy information on 

methods of aquifer protection and the hydrologic 

aspects of free-product recovery. Hydrologic

883
monitoring during free-product recovery is planned for 

use in evaluating the effectiveness of anticipated 

water-table adjustments and for providing large-scale 

information on hydraulic properties of the atoll 

aquifer. After the end of free-product recovery, a 

substantial mass of free fuel is expected to remain in 

the aquifer at residual saturation. In anticipation of the 

need for remediation following free-product recovery, 

the USGS is also evaluating geochemical 

considerations relevant to hydrocarbon degradation 

within the aquifer.

HYDRAULIC DIVERSION 

AND MONITORING

Sparse water-level data indicate that the water 

table is relatively flat but has a small gradient from the 

spill area towards the west (PRC Toups, 1983; C.D. 

Hunt, USGS, written commun., 1991). Withdrawal of 

freshwater from the water-supply wells could lower 

the water table enough to substantially change the 

direction of ground-water flow. In an attempt to 

inhibit movement of fuel from the spill towards the 

water-supply wells, freshwater has been injected since 

April 1992 into several water-supply wells close to the 

spill. This injection is being done to create a mound in 

the water table and modify the hydraulic gradient in 

the aquifer, thus changing the direction of groundwater flow near the spill to divert contaminants to the 

north and away from the water-supply wells. A line of 

six free-product recovery wells is planned for the 

northern edge of the refueling apron; each recovery 

well will have a pump to lower the water table and a 

fuel skimmer (U.S. Navy, oral commun., 1992).

In October 1992, 10 water-level observation 

wells were installed (fig. 1). Each well has been 

equipped with a data logger that records water levels 

at 10 minute intervals. The water-level data are 

downloaded weekly by Navy personnel. The high 

frequency of data collection has been selected to 

measure the tidal dynamics of the aquifer. An 

example of the observed water-level fluctuations is 

shown in figure 2. Well WL-2 is 20 ft deep and about 

300 yards from the lagoon; the depth to water is about 

5 ft below land surface. The water-level data show an 

average head of 1.75 ft above mean sea level. The 

ratio of the amplitudes between water-level 

fluctuations in well WL-2 and ocean tides was 0.72, 

and the time lag between peak high or low tide in the 

ocean and well WL-2 was 20 minutes.

Measurement of water level at 10 minute 

intervals continues at each of the 10 observation wells. 

The recorded data are reduced to average daily water 

levels to construct synoptic, tidally-averaged, watertable maps from which the direction of movement of 

free and dissolved fuel can be inferred. Harmonic 

analysis of the data is being used to derive amplitude 

ratios and time lags between aquifer water levels and 

ocean tides. This information can be used to infer 

aquifer hydraulic properties and to calibrate groundwater-flow models (Ferris and others, 1962; 

Buddemeier and Holladay, 1977; Underwood and 

others, 1992). Of particular interest are changes in the 

water table caused by remedial measures, such as the 

hydraulic diversion now underway and the recovery of 

free fuel, which is anticipated to use pumps to lower 

the water table and fuel skimmers.

GEOCHEMICAL CHARACTERIZATION

Nests of monitoring wells were installed at five 

sites in October 1992 to characterize geochemical 

processes within the aquifer, evaluate hydrocarbon 

degradation, and monitor site restoration (fig. 1). At 

each site, 4 or 5 wells have depths that range from 15 

to 55 ft below ground surface. The wells are 

constructed of 2-in. inside-diameter (i.d.) PVC and 

have a screened interval of 5 ft. At each site, a vapor 

probe is used to sample soil gas from 3 to 3.5 ft below 

land surface. At unpaved sites where rainfall can 

infiltrate, soil-water probes have been installed at 

depths of 2 and 3 ft below land surface to sample 

fluids from the unsaturated zone. Following discovery 

of the spill, several exploratory wells were constructed 

of screened 2 in. (i.d.) PVC to a depth of about 5 ft 

below the water table. These exploratory wells are 

used to detect free and dissolved fuel at the top. of the 

water table.

In February 1993, the monitoring wells were 

sampled using a submersible pump with a casing 

packer. pH, O2 concentration, and temperature of 

water samples were measured in a closed cell at the 

site. H2S concentration was determined by titration 

following field preservation. Samples for 

determination of major ions and nutrients were filtered 

in a field laboratory and analyzed at the University of 

Hawaii. Aqueous and vapor-phase measurements of 

CO2 , O2 , N2 , CH4 and volatile aromatic carbon 

compounds were made using portable gas 

chromatographs equipped with thermal-conductivity

884
LU

LU

_l

5

03

1

LU

O 

CD

LJJ 

LJJ 

LL

LJJ

LJJ

_l

CC 

LJJ

I

4.0

3.5 -

3.0 -

2.5 -

2.0 -

1.5 -

1.0 -

0.5

0.0

12 13 14 15

JANUARY 1993

16 17 18

Figure 2. Water-level oscillations in well WL-2 from January 12-19,1993. Peak high and low water levels follow high and 

low ocean tides by about 20 minutes.

and photoionization detectors. The exploratory wells 

were sampled using disposable bailers. Much of the 

data from the initial trip are not yet available and all 

results are considered as preliminary pending 

additional sampling and analysis. However, several 

trends are evident from the preliminary data.

The approximate areal extent of free-phase and 

dissolved-phase fuel in proximity to the spill is shown 

in figure 3. As of February 1993, fuel was detected 

several hundred feet from the spill and found as a 

dissolved phase as deep as 40 ft below the water table. 

Of particular concern is that free-phase and dissolvedphase fuel have migrated about half the distance from 

the spill to several water-supply wells. Additional 

data are needed to evaluate the effectiveness of the 

hydraulic control system in retarding or diverting the

migration of contaminants towards the water-supply 

wells.

There is evidence that microbial oxidation is 

greater where the aquifer is contaminated by the fuel. 

Microbial respiration will initially consume O2 and 

switch to other metabolic pathways such as sulfate 

reduction (producing H2S) and methanogenesis 

(producing CH4) after O2 is depleted (Chapelle, 1993). 

Cross sections of benzene, CH4, CO2 , and H2S 

concentrations in the saturated zone at monitoring 

sites S-l, S-3, and S-5 are shown in figure 4. Sites S-l 

and S-3 are contaminated with jet fuel, whereas site 

S-5 is uncontaminated. Concentrations of O2 are 

typically less that 0.3 ppm in the ground-water 

samples from the study area. Ground-water samples 

with the highest O2 concentrations (0.6-0.3 ppm) were

885
(A)

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EXPLANATION

---- HORIZONTAL WATER-SUPPLY WELL 

r , ,, ,° 0 WELL WITH NO FREE-PHASE FUEL PRESENT

 WELL WITH FREE-PHASE FUEL PRESENT

llllllllllllllllllll^ * RECOVERY PIT WITH FREE-PHASE FUEL PRESENT

11$$^^

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800 FEET

200 METERS

EXPLANATION

---- HORIZONTAL WATER-SUPPLY WELL 

 JL. 0 NO DISSOLVED FUEL PRESENT

Illllllllit DISSOLVED FUEL PRESENT AT SURFACE

fillip (42) , DISSOLVED FUEL PRESENT AT DEPTH-Number 

:;||||;|;||||||||||||!|||1 Indicates maximum depth of detection, In feet

!!!lf!llf!!l^

:y^:-^Q^imyf:jr " ^"VKSS . :y^* ^:SKS:j('*l'^^:':':S;HS:::S:fe4:S

IL JlHlft0 Jill JC Jl

II " o o ^ m r 

" .... ,o. ..., ,... .... ,... .... ,... I .... ,...

800 FEET

200 METERS

Figure 3. Location of sites with and without (A) free-phase fuel and (B) dissolved fuel, in February 1993.

886
FEET

10 -i

Mean sea level -

-10 -

-20 -

-30-

-40 -

-50 -i

A sli -P OT Land surface UJ ° * CO

" ?'2 0.8 -

- 7.5 

1.9-

- 6.3 2.0 -

57 1.6-

<0.5 -

<0.5-

<0.5 -

<0.5 -

Benzene <o.s -

FEET

10 -i

Mean sea level -

-10 -

-20 -

-30 -

-40 -

.50 -

uJ Land surface UJ c

" 25'6 8.0 -

8.2-

- 3.1 7.9 -

2.5 -I

35 

pCH4 (x1,000) 

£ c

FEET

10 -,

Mean sea level -

-10 -

-20 -

-30 -

-40 -

^ 

03 Land surface CO c

" 2'3 7.2 -

- 3.4 

8.7-

-5.2 10.0-

3-8 7.8 -

\\ s

1.0-

1.1 -

2.4 -

1.3 -

H2S 1 .3 -

400 FEET Vertical scale greatly exaggerated

FEET 

10 -i

Mean sea level -

-10 -

-20 -

-30 -

-40 -

-50 -

W Land surface CO m » 

" 6 '8 4.3 -

- 2.1 

3.3-

- 0.9 3.1 -

"" 1.2-J

pC02 (x100) 

100 METERS

Figure 4. Chemical sections from well sites S-1 to S-5. Benzene concentrations in parts per billion, H2S concentations in parts per m

g atmospheres of partial pressure.
from the contaminated site S-l. Additional sampling 

is needed to determine if this is a result of sampling 

error or incomplete oxic respiration in that part of the 

aquifer. Concentrations of CH4, CO2, and H2S are 

low in both surface seawater and rainwater; the high 

concentrations of CH4, CO2 , and H2S in water from 

all well waters demonstrates that substantial microbial 

respiration is occurring within the aquifer. The contaminated sites S-1 and S-3 tended to have much 

greater concentrations of CH4, CO2 , and H2S than do 

the uncontaminated site S-5. It thus seems that natural 

levels of microbial oxidation are sufficient to deplete 

O2 in aquifer waters, and the oxidation of fuel is proceeding through other metabolic pathways such as 

methanogenesis and sulfate reduction.

Future work needs to focus on determining the 

extent and metabolic pathways of microbial processes 

in both the saturated and unsaturated parts of the 

aquifer. Of particular interest are the microbial 

oxidation of fuel, the factors limiting microbial 

respiration, and the relation between microbial 

oxidation and carbonate-mineral diagenesis. 

Complete analysis of data from the February 1993 

trip, as well as future data as it becomes available 

should permit a thorough characterization of both 

natural and fuel-enhanced geochemical processes in 

the aquifer.

SUMMARY

The USGS has begun a project to investigate a 

subsurface spill of JP-5 jet fuel that has contaminated 

one of the freshwater lenses at Diego Garcia. Part of 

this project is to provide hydrologic data to evaluate 

the migration of both free-phase and dissolved-phase 

fuel constituents towards water-supply wells and to 

facilitate the recovery of free fuel. The hydrologic 

data are obtained from a network of 10 water-level 

observation wells. Preliminary analysis of data reveals 

oscillations in the water table of as much as 4 ft 

associated with the rise and fall of ocean tides. The 

tidal signal is propagated rapidly through the aquifer 

with a relatively high efficiency, indicating that the 

aquifer is moderately to highly permeable.

Another part of this project entails the 

characterization of geochemical processes within the 

aquifer to evaluate methods of remediating residual 

contamination after the free fuel has been removed. Of 

particular interest is the determination of microbial 

processes within the aquifer and the effect of the fuel on 

such processes. Preliminary data indicate that natural 

microbial oxidation of organic matter is occurring

throughout the aquifer and results in the depletion of 

O2, and the production of CO2, H2S, and CH4. Areas 

contaminated by fuel have higher concentrations of the 

products of microbial respiration (CO2 , H2S, CH4) than 

uncontaminated areas, indicating that the fuel is being 

oxidized by microbial activity.

REFERENCES

Anthony, S.S., Peterson, F.L., Mackenzie, F.T., and Hamlin, 

S.N., 1989, Geohydrology of the Laura fresh-water 

lens, Majuro Atoll A hydrochemical approach: 

Geological Society of America Bulletin, v. 101, 

p. 1066-1075.

Ayers, J.F., and Vacher, H.L., 1986, Hydrogeology of an 

atoll island A conceptual model from detailed study 

of a Micronesian example: Ground Water, v. 24, no. 2, 

p. 185-198.

Buddemeier, R.W., and Holladay, G., 1977, Atoll

hydrology Island groundwater characteristics and 

their relationship to diagenesis, in International Coral 

Reef Symposium, Third, Miami, Fla., 1977, Proceedings: Miami, Fla., University of Miami, p. 167-173.

Buddemeier, R.W., and Oberdorfer, J.A., 1986, Internal 

hydrology and geochemistry of coral reefs and atoll 

islands: Key to diagenetic variations, in Schroeder, 

J.H. and Purser, B.H., eds., Reef Diagenesis: Berlin, 

Springer-Varlag, p. 91 -111.

Chapelle, F.H., 1993, Ground-water microbiology and 

geochemistry: New York, John Wiley & Sons, 424 p.

Ferris, J.G., Knowles, D.B., Brown, R.H., and Stallman, 

R.W., 1962, Theory of aquifer tests: U.S. Geological 

Survey Water-Supply Paper 1536-E, 174 p.

Hunt, C.D. Jr, 1991, Climate-driven saltwater intrusion in 

atolls, in Peters, H.J., ed., Ground Water in Pacific Rim 

Countries, Symposium of the Irrigation and Drainage 

Division: American Society of Civil Engineers, 

Honolulu, Hawaii, July 23-25, 1991, p. 43-49.

Hunt, C.D. and Peterson, F.L., 1980, Groundwater resources 

of Kwajalein Island, Marshall Islands: University of 

Hawaii Water Resources Research Center Technical 

Report 126,91 p.

PRC Toups, 1983, Final submittal engineering study to 

evaluate potable water supply alternatives and 

groundwater yield at Diego Garcia, British Indian 

Ocean Territory, Volume II (prepared for U. S. Navy, 

Naval Facilities Engineering Command, Pacific 

Division, Pearl Harbor, Hawaii): Orange, Calif., PRC 

Toups, 7 app.

Underwood, M.R., Peterson, F.L., and Voss, C.I., 1992, 

Groundwater lens dynamics of atoll islands: Water 

Resources Research, v. 28, no. 11, p. 2889-2902.

Wheatcraft, S.W., and Buddemeier, R.W., 1981, Atoll island 

hydrology: Ground Water, v.19, no. 3, 311-320.

888
Microbial Degradation of Nitrotoluenes in Surface Soils 

and Aquifer Sediments, Weldon Spring, Missouri

By Francis H. Chapelle1 and Paul M. Bradley 1

Abstract

Surface soils and underlying aquifer 

material from a nitrotoluene-contaminated site at 

Weldon Spring, Missouri, were evaluated for the 

potential of indigenous microorganisms to 

transform 2,4,6-trinitrotoluene (2,4,6-TNT), 2,4- 

dinitrotoluene (2,4-DNT) and 2,6-dinitrotoluene 

(2,6-DNT). The results indicate that the microbial 

communities are capable of rapidly metabolizing 

2,4-DNT, 2,6-DNT and 2,4,6-TNT. In most 

cases, complete dissappearance of the source 

compound was achieved within 20 to 70 days. 

These results indicate that in-situ bioremediation 

may be a viable alternative for treating 

contaminated soils and ground water. 

Furthermore, the results suggest that the presently 

observed nitroaromatics contamination of ground 

water is unlikely to be a relic of past (50 years 

before present) waste-disposal practices, but 

rather reflects continual leaching of 

nitroaromatics from contaminated soils.

INTRODUCTION

The manufacture, loading, assembling, and 

packing of munitions have resulted in nitrotoluene 

contamination of soils and ground water at many 

sites throughout the United States. The process of 

separating the desired 2,4,6-trinitrotolune (2,4,6-TNT) 

product from the numerous impurities created during 

the manufacturing process generates large quantities 

of "red-water" wastes that contained 2,4-

'(J.S. Geological Survey, Columbia, South Carolina

dinitrotoluene (2,4-DNT)-and 2,6-dinitrotoluene (2,6- 

DNT) as well as other nitrotoluene isomers. During 

World War II, red water was commonly disposed of 

by direct discharge to streams at the Weldon Spring 

munitions manufacturing facility in Missouri. These 

past waste-disposal practices resulted in widepread 

nitroaromatics contamination of soils and ground 

water at the Weldon Spring site.

The presently observed nitroaromatics 

contamination of ground water at the Weldon Spring 

site (Schumacher and others, 1992) may be a relic of 

past waste-disposal practices. However, this 

hypothesis suggests that nitroaromatics are stable in 

the aquifer system and not subject to significant 

microbial degradation. On the other hand, if the 

aquifer system contains microorganisms capable of 

degrading nitroaromatics, a continuous source of 

nitroaromatics (such as leaching from coantaminated 

surface soils) is indicated. In either case, one key to 

resolving the source of the presently observed 

nitroaromatic contamination of ground water is to 

evaluate the potental of indigenous microorganisms to 

degrade nitroaromatics.

Numerous recent investigations have focused on 

microbial degradation of nitrotoluenes (Carpenter and 

others, 1978; Kaplan and Kaplan, 1982; Hallas and 

Alexander, 1983; Liu and others, 1984; Fernando and 

others, 1990; Spanggord and others, 1991; Higson, 

1992; Spiker, 1992; Valli and others, 1992; Funk and 

others, in press). However, these investigations 

utilized artifical sources of microbial communites such 

as sewage and activated sludge systems (Carpenter 

and others, 1978, Hallas and Alexander 1983, Liu and 

others, 1984), land farming and composting (Kaplan 

and Kaplan 1982, Spiker and others, 1992, Funk and 

others, in press), as well as isolated cultures of bacteria 

(McCormick and others, 1976, Spanggord and others,

889
1991) or fungi (Fernando and others, 1990, Valli and 

others, 1992). Surprisingly, however, little is known 

about the capability of native microorganisms to 

transform nitrotoluenes in-situ. In this report, we 

examine the potential for in-situ biodegradation of 

nitrotoluenes by surface soils and aquifer sediments at 

a contaminated site in Weldon Spring, Missouri. We 

also examined the effect of nitrotoluene additions on 

sediment CO2 production to evaluate the impact of 

nitrotoluene contamination on overall microbial 

community metabolism at the site.

MATERIALS AND METHODS

The ability to transform 2,4-dinitrotoluene, 2,6- 

dinitrotoluene and 2,4,6-trinitrotoluene was examined 

using a contaminated surface soil (red tank), an 

uncontaminated surface soil (top soil), fractured 

carbonate bedrock material (carbonate), and material 

from a weathered, semi-consolidated water-bearing 

zone (residuum) that occurs at the top of the carbonate 

aquifer. The highly contaminated red-tank soil was 

collected within an inactive red-water holding-lagoon 

in which TNT concentations range from 1,000 to 

10,000 mg/kg soil. The top soil was collected at an 

adjacent site with no detectable nitrotoluene 

contamination and no history of nitrotoluene 

exposure. The residuum and carbonate aquifer 

materials were collected from areas contaminated by 

red water in the past, and where leaching of 

nitroaromatics from contaminated soils may be 

presently occurring. Concentrations of 2,6-DNT and 

2,4-DNT in the aquifer are presently in the parts per 

billion range.

Sediment samples were incubated statically in 

35 mL serum vials containing 20 g wet soil or 

sediment and 20 mL of nitrotoluene solution. The 

final dissolved concentrations of the nitrotoluenes 

were 100-150 jiM. The treatments were created with 

an air headspace and sealed with Teflon coated butyl 

rubber stoppers. Triplicate experimental samples, a 

single live control (not sterilized and without 

nitrotoluene addition) and a single dead control (with 

nitrotoluene addition and sterilized by amendment 

with 5 mM HgCl2 and autoclaving for 60 min) were 

prepared for each combination of sediment and 

nitrotoluene treatment. The vials were incubated in 

the dark at room temperature and periodically sampled 

for dissolved nitrotoluenes and headspace CO2 for 

approximately 70 days.

Nitrotoluene analyses were made as follows: 

Approximately 12 hours before sample collection and 

analysis, the treatment vials were shaken by hand to 

thoroughly mix the interstitial and standing water. To 

sample, about 0.5 mL of standing water was removed, 

filtered (0.2 urn pore size) and analyzed by high 

pressure liquid chromatography. The sample volume 

was replaced with 99.9 percent O2 gas. Nitrotoluene 

analyses was performed on a Beckman Ultrasphere 

ODS Ci 8 column at 40 °C with a acetonitrileidistilled 

water:methanol mobile phase. The 2,4-DNT and 2,6- 

DNT analyses were made isocratically with a 15:45:40 

acetonitrile:distilled water:methanol mobile phase. 

TNT was determined by gradient elution with a 

15:80:5 acetonitrile: distilled watenmethanol mobile 

phase ramping up to 47:48:5 over 25 minutes and 

holding for 5 minutes (method by Chris Lindley, U.S. 

Geological Survey). All compounds were detected by 

UV absorption at 250 nm. Peak areas were quantified 

by calibration against known standards (Aldrich 

Chemical Co.).

Headspace CO2 concentrations were determined 

as an indicator of microbial community metabolism. A 

0.5 mL headspace sample was removed and analyzed by 

thermal conductivity detection gas chromatography. The 

sample volume was replaced with 99.9 percent O2 . Peak 

areas were quantified by calibration against known 

standards (Scotty II, Scott Gas Co.).

RESULTS

The pattern of microbial-community 

metabolism (CO2 production) varied between soil and 

sediment samples, but, with the exception of the top 

soil (2,4-DNT) treatment, community CO2 production 

did not differ between the TNT, 2,4-DNT, or 2,6-DNT 

treatments for a given soil or aquifer sediment. In the 

residuum, for example (fig. 1), CO2 production was 

similar for each compound. Significantly, however, 

the carbonate and residuum samples produced CO2 

only in the nitrotoluene-amended treatments. Without 

nitrotoluene addition, the CO2 production did not 

differ from that in the sterile controls (fig. 1). In the 

red-tank soils, CO2 production was greatest in the live 

controls (without nitrotoluene amendment). Addition 

of 2,4-DNT, 2,6-DNT or 2,4,6-TNT inhibited CO2 

production in this soil. In contrast, addition of 2,4- 

DNT or 2,4,6-TNT to the top soil did not inhibit CO2 

production relative to that of the live control. 

However, addition of 2,4-DNT did appear to reduce 

the top soil microbial metabolism.

890
RESIDUUM

N

4000 

3500 

3000 

2500 

u 2000 

<< 1500 

Q 1000 

W 500

 24DNT 

T 26DNT

 246TNT

 LIVE 

o DEAD

10 20 30 40 50 60 70

DAYS

Figure 1. Production of headspace CO2 by semi-consolidated aquifer material (Residuum) from Weldon Springs, Missouri. 

Data are means (±SD) of triplicates for microcosms amended with 2,4-DNT, 2,6-DNT or 2,4,6-TNT. Controls are nitrotoluene 

amended , sterilized microcosms (DEAD) and unsterilized microcosms without nitrotoluenes (LIVE).

During the 70-day incubation period, all three 

nitrotoluenes were degraded by the surface soils and 

aquifer sediments (figs. 2-4). The concentrations of 

nitrotoluenes remaining in each treatment are 

expressed as a percent of the initial dissolved 

concentration (100-150 fiM). The carbonate microbial 

community degraded 2,4,6-TNT more rapidly than the 

dintrotoluenes, but the concentrations of all 

nitrotoluenes were reduced by at least 80 percent 

during incubation. There was no observable lag in the 

degradation of nitrotoluenes by the carbonate 

sediments. Although the concentrations of 2,4-DNT 

and 2,4,6-TNT in the residuum microcosms were 

reduced below the detection limit (0.1 u>M), about 

45 percent of the added 2,6-DNT remained at day 70 

(fig. 2). The lag preceding a detectable decrease in 

nitrotoluene concentrations in the residuum treatments 

varied from 8 days (2,4-DNT) to about 22 days (2,4,6- 

TNT). About 95 percent of the 2,4-DNT and 2,4,6- 

TNT in the red tank microcosms was removed without 

a lag by day 30 (fig. 3). After a 20 day lag, the 

concentration of 2,6-DNT decreased to the detection 

limit by the end of the experiment. In contrast, 

approximately 95 percent of the 2,6-DNT and 2,4,6- 

TNT in the top soil microcosms was removed (without 

lag) within 22 days, while complete degradation of 

2,4-DNT was delayed by about 30 days (fig. 4).

DISCUSSION

The results indicate that microbial communities 

associated with surface soils and underlying aquifer 

material at Weldon Springs, Missouri are capable of 

rapidly metabolizing 2,4-DNT, 2,6-DNT and 2,4,6- 

TNT. In most cases, complete dissappearance of the 

source compound was achieved within 20 to 70 days. 

In this study, degradation of the source nitrotoluene 

was followed by accumulation of the associated 

aminonitrotoluenes (data not shown). However, 

subsequent decreases in the dissolved concentrations 

of these intermediates indicates that the aminocompounds are biodegradable as well. These results 

indicate that in-situ bioremediation may be a viable 

alternative for cleanup of nitrotoluene-contaminated 

aquifers and soils.

From an environmental impact point of view, 

degradation of a contaminant is desirable only if the 

products are innocuous or at least less problematic 

than the source compound. Ideally, the contaminant is 

completely degraded to CC>2 (or CH4 and CO2 under 

anaerobic conditions) during bioremediation. At 

present, only the white-rot fungus, Phanaerochaete 

chiysosporium, has been reported to degrade 

nitrotoluenes to CO2 . Approximately 35 percent of

891
125

2.4 DNT RESIDUUM 2,4 DNT

o

I- 

<

o

z 

o 

o

o

Ld

O

in in

0 10 20 30 40 50 60 70

0 10 20 30 40 50 60 70

2.4,6 TNT

10 20

DAYS

Figure 2. Decrease in dissolved concentrations of 

nitrotoluenes in microcosms containing semi-consolidated 

aquifer material (Residuum) from Weldon Springs. 

Concentrations are expressed as a percentage of the initial 

dissolved concentration. Data are means (±SD) of triplicate 

experimental treatments ( ) and sterilized controls (O).

the 14C from labeled 2,4,6-TNT and 2,4-DNT was 

recovered as CC>2 during incubation with P. 

chrysosporium (Fernando and others, 1990, Valli and 

others, 1992). However, this organism has been 

shown to be sensitive to high concentrations of 2,4,6- 

TNT (Spiker and others, 1992). The majority of 

studies have reported accumulation of 

aminonitrotoluenes (McCormick and others, 1976, 

Hallas and others, 1983, Liu and others, 1984, Funk 

and others, in press) or highly toxic and biologically 

inert azoxy polymers (Kaplan and Kaplan 1982, Funk 

and others, in press). Two lines of indirect evidence

125

100

RED TANK

Z 

O

01»- 

z

Ul

o

z 

o 

o

o

LJ

O 2,4,6 TNT

</) 125

(/>

O

Figure 3. Decrease in the dissolved concentrations of 

nitrotoluenes in microcosms containing contaminated Red 

Tank soil from Weldon Springs. Concentrations are 

expressed as a percentage of the initial dissolved 

concentration. Data are means (±SD) of triplicate 

experimental treatments ( ) and sterilized controls (O).

suggest that the Weldon Springs microbial 

communities may be capable of more complete 

degradation. The fact that aminonitrotoluenes initially 

accumulated and subsequently diminished with time 

(data not shown) indicates that these compounds are 

transitory intermediates of nitrotoluene degradation by 

the Weldon Springs microbial communities. 

Moreover, the fact that CO2 production by carbonate 

and residuum sediments was significant only in the 

presence of nitrotoluenes (fig. 1) is highly suggestive 

of complete degradation to CO2 . However, because 

these sediments contain carbonate material, the

892
125

2,4 DNT TOP SOIL

o

o 

o

p 

w

 J 

o

OT 

KJ

P

10 20 30 40 50 60 70

125

100

75

50

25

0 10 20 30 40 50 60 70

DAYS

Figure 4. Decrease in the dissolved concentrations of 

nitrotoluenes in microcosms containing uncontaminated 

Top Soil from Weldon Springs. Concentrations are 

expressed as a percentage of the initial dissolved 

concentration. Data are means (±SD) of triplicate 

experimental treatments ( ) and sterilized controls (O).

potential exists that the observed CO2 production is 

due to carbonate dissolution by acidic intermediates. 

Radiotracer experiments to evaluate complete 

degradation of nitroaromatics to CO2 are currently 

under way.

The primary motivation for undertaking these 

studies was to determine (1) if the residuum and 

carbonate aquifer lack microorganisms capable of 

degrading nitroaromatics, in which case the presently 

observed nitroaromatic contamination of ground water 

could be a relict of past contamination, or (2) if the 

residuum and carbonate aquifers contain microbial

populations capable of nitroaromatics degradation, in 

which case the observed ground-water contamination 

may indicate a continuous nitroaromatics source from 

the contaminated soils. Available evidence supports 

the second hypothesis and suggests that nitroaromatics 

are being continually leached into the aquifer from 

contaminated soils. These results also suggest that 

naturally occuring biodegradation will tend to limit the 

transport of nitroaromatic contaminants by ground at 

this site.

REFERENCES

Carpenter, D.F., McCormick, N.G., Cornell, J.H., and 

Kaplan, A.M., 1978, Microbial transformation of

C-labeled 2,4,6-trinitrotoluene in an activated-sludge 

system: Applied and Environmental Microbiology, 

v. 35, p. 949-954.

Fernando, T., Bumpus, J.A., and Aust, S.D., 1990., 

Biodegradation of TNT (2,4,6-trinitrotoluene) by 

Phanaerochaete chrysosporium: Applied and 

Environmental Microbiology, v. 56, p. 1666-1671. 

Funk, S.B., Roberts, D.J., Crawford, D.L., and Crawford, 

R.L., (1993), Initial phase optimization for the 

bioremediation of munitions-contaminated soils: 

Applied and Environmental Microbiology, v. 59, 

p. 2171-2177.

Hallas, L.E. and Alexander M., 1983, Microbial

transformation of nitroaromatic compounds in sewage 

effluent. Applied and Environmental Microbiology, 

v. 45, p. 1234-1241.

Higson, F.K., 1992, Microbial degradation of nitroaromatic 

compounds, in Neidleman, S.L., and Laskin, A.I., eds., 

Advances in Applied Microbiology: New York, 

Acedemic Press, Inc., p.1-19

Kaplan, D.L., and Kaplan, A.M., 1982, Thermophilic 

biotransformations of 2,4,6-trinitrotoluene under 

simulated composting conditions: Applied and 

Environmental Microbiology, v. 44, p. 757-760.

Liu, D., Thomson, K., and Anderson, A.C., 1984, 

Indentification of nitroso compounds from 

biotransformation of 2,4-dinitrotoluene: Applied and 

Environmental Microbiology, v. 47, p. 1295-1298.

McCormick, N.G., Feeherry, F.E., and Levinson, H.S.,

1976, Microbial transformation of 2,4,6-trinitrotoluene 

and other nitroaromatic compounds: Applied and 

Environmental Microbiology, v. 31, p. 949-958.

Schumacher, J.G., Lindley, C.E., and Anderson, F.S., 

1992, Migration of nitroaromatic compounds in 

unsaturated soil at the abandoned Weldon Spring 

Ordanance Works, St. Charles County, Missouri 

Proceedings from the 16th anual Army Research and 

Development Symposium, CETHA-TS-CR-92062, 

p. 173-192.

893
Spanggord, R.J., Spain, J.C., Nishino, S.F., and Valli, K., Brock, B.J., Joshi, O.K., and Gold, M.H., 1992,

Mortelmans, K.E., 1991, Biodegradation of 2,4- Degradation of 2,4-dinitrotoluene by the lignindinitrotoluene by Psuedomonas sp.: Applied and degrading fungus Phanaerochaete chiysosporium:

Environmental Microbiology, v. 57, p. 3200-3205. Applied and Environmental Microbiology, v. 58,

Spiker, J.K., Crawford, D.L., and Crawford, R.L. 1992. P' 221 '228' 

Degradation of 2,4,6-trinitrotoluene (TNT) in 

explosives-contaminated soils by white-rot fungus 

Phanaerochaete chrysosporium: influence of TNT 

concentration: Applied and Environmental 

Microbiology, v. 58, p. 3139-3144.

894
Depletion of Nitrogen-Bearing Explosives Wastes in a 

Shallow Ground-water Plume near Hawthorne, Nevada

8/A.S. Van Denburgh1 , D.F. Goerlitz2, a/ic/E.M. Godsy2

Abstract

Liquid explosives wastes from the emptying 

and washout of military projectile and bomb 

casings were discharged into unlined disposal 

pits north of Hawthorne, Nevada. The explosives 

involved at this arid study area were ammonium 

picrate, during 1952-58, and trinitrotoluene (TNT) 

plus ammonium nitrate, during 1964-68. Percolation from the pits entered unconsolidated, heterogeneous sedimentary deposits in which clay- to 

gravel-size particles predominate. The water table 

at and downgradient from the pits was from 1-1/2 

to 6-1/2 meters below land surface during the 

study. The shallow ground water moves northwestward at an average linear velocity of about 

130 meters per year.

A shallow plume of ground water containing 

3 to 130 milligrams per liter of dissolved-nitrogen 

species, as nitrogen, extended more than 1,700 

meters northwest of the northernmost pits. Either 

ammonium or nitrate dominated at sample sites 

within 300 meters of the pits; farther downgradient 

in the plume, nitrate was the most abundant. Overall, the plume contained about 3 megagrams of 

dissolved nitrogen downgradient from the pits. 

The proportions of sodium and bicarbonate ions 

increased downgradient, relative to chloride and 

sulfate ions. Saturated sedimentary deposits 

within the plume generally had greater concentrations of organic nitrogen, ammonium, and organic 

carbon than did the adjacent sediments.

The maximum linear velocities for the 

dissolved explosives-waste components (a measure 

of their mobility) ranged from more than 77 meters

'U.S. Geological Survey, Carson City, Nev. 

2U.S. Geological Survey, Menlo Park, Calif.

per year for nitrate to less than 20 meters per year 

for picrate, with velocities of TNT and ammonium 

in between. The downgradient movement of nitrate 

may not have been impeded by microbiological or 

geochemical processes. For the other dissolved 

components, in contrast, the following depletion 

mechanisms are proposed: TNT, microbial transformation to inorganic nitrogen and carbon species, 

and probable sorption of organic transformation 

products; ammonium, sorption, and probable 

oxidation by nitrifying bacteria; and picrate, microbial alteration to picramic acid, then mineralization 

to inorganic nitrogen and carbon species. The reactions that depleted TNT, ammonium, and picrate in 

the ground water may also have enriched the 

proportions of sodium and bicarbonate ions in the 

plume. Overall, the Hawthorne study shows that 

natural microbiological and geochemical processes 

can degrade or retard the transport of large quantities of toxic, chemically complex, nitrogen-rich 

explosives wastes over distances of hundreds of 

meters, leaving nitrate as the only residual nitrogenbearing component of the ground-water plume.

INTRODUCTION

The military ammunition plant near Hawthorne, 

Nev., was commissioned in 1930 to fill, store, and 

empty explosives containers such as projectile and 

bomb casings. A variety of liquid wastes from the 

emptying and washout (demilitarization) processes 

were discharged into unlined disposal pits at several 

sites within the 700-km2 plant during the last six 

decades. Ground water beneath the disposal sites 

moves generally northwestward toward Walker Lake, 

a 140-km2 closed-basin body of saline water that is 

used for recreation (fig. 1). Because much of the lake 

is outside the plant boundary, the U.S. Department of 

the Army asked the U.S. Geological Survey (USGS)

895
38°45'

118°45' 118°30'

38°30'

10 KILOMETERS NEVADA

0 5 10 MILES

EXPLANATION 

- - - Boundary of ammunition plant

Figure 1. Study area and vicinity.

in 1978 to evaluate the extent and magnitude of 

contamination in shallow ground water between the 

then-northernmost demilitarization site and the lake 

(the nearest demilitarization facility upgradient from 

that site was almost 4,000 m to the southeast). Explosives processed at the site included ammonium picrate, 

trinitrotoluene (TNT), and ammonium nitrate.

This paper describes principal results of that study 

and of an earlier one by the USGS immediately downgradient from disposal pits at the same site. Discussed 

herein are the hydrogeologic setting of the overall 

2-km2 study area, chemical character of the shallow 

ground water and host sedimentary deposits within 

and adjacent to the contaminant plume, differing 

mobilities of the several dissolved explosives-waste 

components, and mechanisms by which those components were altered and depleted as the shallow ground 

water moved toward Walker Lake. These discussions 

are based on data collected from about 40 shallow 

exploratory wells (depths, 7-14 m) drilled adjacent to 

and downgradient from the northernmost pits during 

1975-78. Most of the water-chemistry analyses were

made on samples collected from the wells in 1977 or 

1979. At the time of study, results were documented in 

unpublished administrative reports; more recently, 

selected aspects of the investigations were described 

briefly by Van Denburgh and Goerlitz (1987) and by 

Goerlitz (1992, p. 325-330). The more extensive 

discussion herein was inspired by the possible transfer 

value of the Hawthorne results at a time of increasing 

USGS involvement in studies of waste-disposal 

contamination at military facilities. Details regarding 

procedures used at Hawthorne to drill the exploratory 

wells and to collect and analyze the sediment cores 

and water samples therefrom are available in files of 

the USGS (333 W. Nye Lane, Carson City, NV 

89706).

We are particularly grateful to Don Cavin, Louie 

Dellamonica, Bill Glenzer, Don Gruwell, Floyd Justus, 

Dell Mortensen, and Joe Viani of Hawthorne (all but 

Mr. Viani are current or retired employees of the 

ammunition plant), who provided a blend of historical 

perspective, material assistance, and genuine hospitality that made our task so much easier.

896
GEOGRAPHY, HYDROLOGY, AND 

WASTE DISPOSAL

The study area (fig. 2) is on the gently northwestward-sloping floor of Walker Lake Valley about 1,220 m 

above sea level. Adjacent mountains to the west and 

east are 1,000 to 2,000 m higher. Precipitation on the 

valley floor averages only about 100 mm/yr; atop the 

highest adjacent range, the average may not be much 

greater than 300 mm/yr [on the basis of an estimated 

long-term average at the briefly operated Mount Grant 

storage gage, 2,700-m altitude (Van Denburgh and 

Glancy, 1970, p. 10)]. Consequently, all streams in this 

arid to semiarid area are ephemeral downgradient from 

the mountain fronts.

Sedimentary deposits beneath the study area are 

unconsolidated to partly consolidated aggregations of 

lacustrine and alluvial clay, silt, sand, and gravel; their 

cumulative thickness is 800 to 1,000 m (D.H. Schaefer, 

U.S. Geological Survey, written commun., 1985). 

Shallow exploratory wells drilled during the study 

penetrate generally heterogeneous deposits in which 

clay- to gravel-size particles predominate. A distinctive 

impermeable clay unit underlies these deposits; the top 

of the clay ranges from 7-1/2 to at least 11 m below 

land surface.

The water table, which areally varied from 1-1/2 to 

6-1/2 m below land surface during the period of study, 

slopes generally northwestward (gradient, about 5.5 m/ 

1,000 m; see fig. 2). The average horizontal hydraulic 

conductivity of saturated deposits overlying the clay 

unit is on the order of 7 m/d, as estimated from results 

of slug tests. The slug-test data were analyzed by R.L. 

Carman (U.S. Geological Survey, written commun., 

1979) using the procedure described by Cooper and 

others (1967). The average effective porosity of the 

deposits within the plume may be 10 to 12 percent 

based on moisture-content measurements of sediment 

samples from below the water table. These values 

indicate that the shallow ground water moves downgradient at an average linear rate of 120 to 140 m/yr, 

on the basis of the equation given by Lohman 

and others (1972, p. 14).

The then-northernmost demilitarization site at the 

ammunition plant was used beginning in about 1952, 

and large but unrecorded volumes of liquid waste 

were generated (historical information herein is based 

mostly on written and oral communications from 

William K. Glenzer and Dell L. Mortensen during 

1976-79). Explosives at the site were ammonium

picrate (formal name, ammonium 2,4,6-trinitrophenolate) from projectile casings emptied during about 

1952-58 and TNT (2,4,6-trinitrotoluene) plus ammonium nitrate from bomb casings emptied during 1964- 

68. The solubility of these toxic compounds in water 

ranges widely, from 130 mg/L at 20°C for TNT (Gordon and Hartley, 1992, p. 327) to 11,000 mg/L and 

1,920,000 mg/L at 20°C for ammonium picrate and 

ammonium nitrate, respectively (Sax and Lewis, 1987, 

p. 165 and 163).

CHEMICAL CHARACTER OF SHALLOW 

GROUND WATER AND SATURATED 

SEDIMENTARY DEPOSITS

Although sodium and sulfate were the dominant 

ions in all ground water sampled at the study area, three 

chemical traits the areal distributions of dissolvednitrogen species (fig. 2), dissolved-solids concentrations, and major ions delineate the extent of a plume 

of contaminated ground water downgradient from the 

demilitarization site. The contrasting concentrations 

listed in table 1 provide a "chemical signature" for 

water in the plume, relative to the adjacent, presumably 

uncontaminated ground water.

Downgradient from the northernmost pits, the 

relative amount of sodium (Na+) in the plume increased 

in comparison to the more conservative chloride (CP) 

and sulfate (SO4~^) ions: at mid-plume wells C and D 

(fig. 2), the proportion of Na+ relative to Cl~ and SO.,"2 

was about 40 percent greater than it was at upgradient 

well B; at well E, it was 80 percent greater than at well 

B. The proportion of bicarbonate (HCO,~) in the plume 

also increased downgradient, relative to CP and SO,,"2 : 

at well E, it was more than 100 percent greater than at 

upgradient wells C and D.

Concentrations of dissolved explosives-waste 

components also changed downgradient from the pits. 

TNT decreased from about 500 ^ig/L at well A to less 

than the detection limit (2 ng/L) at well D. Picrate concentrations were less than the 2-|ag/L detection limit at 

and downgradient from well C (unfortunately, picrate 

analyses were not made at earlier-sampled wells closer 

to the disposal pits). Down-plume changes in the concentrations of dissolved-nitrogen species and TNT are 

shown in figure 3. At the center of the plume, less 

than about 300 m from the six northernmost pits, 

ammonium-nitrogen concentrations exceeded those 

of nitrate-nitrogen; elsewhere in the plume, nitrate 

consistently dominated.

897
AREA SHOWN IN 

FIGURE 2B

e 0.5

<*

18*

100 200 

_|___________I

300 METERS

900 FEET

®c

56-19

®B 

130

 0.2

X 

X

:3.1 ^

"0 <30? "X

X X

X X

X .30 

<3? X

 0.3

<70? ®A 

91

<50®

Figure 2. Water-table contours and concentrations of dissolved nitrogen and trinitrotoluene (TNT) in 

exploratory wells downgradient from waste-disposal pits.

898
At well C, the total concentration of dissolvednitrogen species decreased with increasing depth 

below the water table (fig. 3); measured TNT concentrations at the site decreased similarly. Downward dispersion of the explosives wastes is limited by the 

impermeable clay that underlies most or all of the 

plume 10 to 11 m below land surface.

Calculated rates of dissolved-nitrogen transport as 

of the late 1970's, in Mg/yr, were about 0.7 for a transverse section across the plume at well B (fig. 2), about 

0.4 at a section through well C, and about 0.1 at a section through well D. The dissolved-nitrogen content of 

the entire ground-water plume downgradient from the 

northernmost pits is estimated to have been on the 

order of 3 Mg.

Saturated sedimentary deposits within the plume 

generally had somewhat greater to much greater concentrations of nitrogen species and organic carbon than 

did adjacent sediments, as indicated in table 2. All 

analyses for TNT and picrate yielded concentrations 

below the 25 ng/kg detection limit, but such analyses 

are not available for deposits upgradient from well C, 

in the area where concentrations of dissolved-nitrogen 

species including TNT were highest (figs. 2 and 3). 

Comparison of the available nitrogen data for sediments and companion ground water within the plume 

indicates that almost all of the organic nitrogen and 

ammonium were components of the sedimentary 

deposits, whereas the nitrate was a component of the 

interstitial fluids.

DEPLETION OF DISSOLVED EXPLOSIVES 

WASTES

The mobilities of individual explosives-waste 

components differed considerably, as indicated in 

table 3. Nitrate, the most mobile, had a linear velocity 

more than three times that of picrate, the least mobile. 

For any dissolved component of the contaminant 

plume, a linear velocity appreciably slower than the 

linear rate of ground-water flow indicates that the 

migration of that component is being retarded (attenuated) by depletion mechanisms. Thus, table 3 indicates 

that dissolved TNT, ammonium, and especially picrate, 

were depleted downgradient from the disposal pits.

Trinitrotoluene

Microbial transformations probably were the key 

mechanisms for removal of TNT from the groundwater plume. Investigations at well A (fig. 2) revealed 

that aerobic bacteria capable of altering TNT to aminotoluene compounds (Won and others, 1974, p. 515-516; 

McCormick and others, 1976, p. 956) were present. In 

addition, a large, strictly anaerobic microbe population 

(104 to 105 organisms per 100 mL), including sulfatereducing bacteria, was found at well A, and 4-amino2,6-dinitrotoluene was detected along with TNT at well 

C. The amine-substituted intermediate products are 

known to be co-metabolized under anaerobic conditions 

(Roberts and others, 1993; Cheng and others, 1993),

EXPLANATION 

I I Plume of contaminated ground water

 1212 Water-table contour-Shows altitude of water table, April 1979. Contour interval, 2 meters. Datum is sea level

 ^^~ Direction of ground-water flow

 40 Line of equal dissolved-nitrogen concentration-Approximate^ located. Interval, in milligrams per liter, is variable

®B Exploratory well where water contained 2 micrograms per liter of TNT or more-Number is dissolved-nitrogen 

130 concentration, in milligrams per liter. Symbol (<, less than) indicates approximate value from earlier study 

terminated before final sample was collected from well. Letter identifies key well. Range of concentrations at 

well C in part B is for three screened intervals of differing depth (see hydrogeologic section in fig. 3)

D* 

6.0

Exploratory well where water contained less than 2 micrograms per liter of TNT-Number is dissolved-nitrogen 

concentration, in milligrams per liter.- Symbol (<) as defined above. Letter identifies key well

O Unsampled exploratory well-Depth-to-water data used to delineate water-table contours

0 Feature of demilitarization facility-1 , empty-and-washout building; 2, large disposal pits, about 900 meters 

upgradient from well C (ammonium-picrate wastes during 1952-58, TNT and ammonium-nitrate wastes during 

1964-68); 3, concrete and metal troughs; 4, six northernmost pits, about 450 meters upgradient from well C 

(ammonium picrate during 1957-58, TNT and ammonium nitrate during 1964-68); 5, three additional pits (used 

little, if any)

Figure 2. Continued.

899
Table 1 . Chemical character of ground water within and adjacent to the contaminant plume 

[Concentrations in milligrams per liter, except as noted; ng/L, micrograms per liter; <, less than]

Range of measured concentrations

Constituent or property

Contaminant plume

Maximum Minimum

Adjacent ground 

water

Maximum Minimum

Dissolved solids 2,800 2,200 1,700 1,100

Selected major ions:

Sodium

Sulfate

Chloride

Nitrogen (N) species:

Recoverable organic N

Nitrate as N

Nitrite as N

Ammonium as N

Total recoverable N

Carbon:

Organic 1 Inorganic

Explosives (jig/L) :

Trinitrotoluene

Picrate

950

1,400

180

17

44

3

75

130

4.9

160

500

<2

580

1,100

130

0

2

3

1

60

<2

<2

.3

.7

.1

.0

.1

.9

500

760

120

0

<

<

2

<80

<2

<2

.4

.1

.1

.1

.6

.6

370

510

89

<0.

< .

< .

1.

<80

<2

<2

.1

.1

.1

.0

.1

.1

'Carbon content of bicarbonate and carbonate.

presumably to ammonium, carbon dioxide, and other 

products, depending on which potential microbial electron acceptors~for example, sulfate, or bicarbonate, or 

ferric iron are available.

Sedimentary deposits cored at wells C and D contained no detectable TNT but had, on average, 2-1/2 

times as much organic nitrogen as cored deposits 

adjacent to the plume about 180 mg/kg compared to 

70 mg/kg. The sorbed organic nitrogen may have consisted of amino-substituted intermediate products, such 

as 4-amino-2,6-dinitrotoluene and 2,4-diamino-6- 

nitrotoluene, or dimerized or polymerized azoxy 

compounds (Roberts and others, 1993, p. 142).

Picrate

Microbial alteration probably was responsible 

for the depletion of dissolved picrate near the disposal 

pits. Microorganisms can effectively alter picrate to 

picramic acid (Won, 1977). This alteration product is, 

in turn, vulnerable to metabolism by the resident

900

microbial population, yielding dissolved inorganic 

nitrogen species and carbon dioxide. These decomposition products may have contributed to the elevated 

nitrate concentration and bicarbonate enrichment, 

respectively, in ground water at downgradient well E.

Ammonium

Sorption, and probably microbial oxidation 

(nitrification), depleted ammonium in the groundwater plume. Sediment cores from well C contained an 

average ammonium-nitrogen concentration as much as 

14 times that of cored deposits adjacent to, and farther 

downgradient within, the plume (about 33 mg/kg compared to 2.4 mg/kg). Thus, sediments at and upgradient 

from well C appear to have sorbed ammonium from the 

ammonium picrate and ammonium nitrate wastes. 

Clay-mineral sorption of ammonium in exchange for 

sodium would at least partly explain the sodium enrichment detected in the plume downgradient from well B.
1,230 -

LU 

LU

LU 

CO 

LU

O 

CD 

<

CO 

DC 

LLJ

LU 

Q

1,220

1,210-

1,200

EXPLANATION

10 - - Line of equal dissolved-nitrogen concentration-Approximately 

located. Interval, in milligrams per liter, is variable

I Exploration well, with screened interval-Number is 

6 dissolved-nitrogen concentration, in milligrams per liter. 

Letter identifies key well {fig. 2)

Disposal 

pit

Vertical

exageration

X33

DC 

LU

DC 

LLI 

CL

CO

DC 

O

LU 

O

O 

DC

Q 

LU

O

CO 

CO

DC 

LU

LU -I 

LU £

°-

n

40

20

80

- NITRATE AND NITRITE

Nitrite-N

60 -

40 -

20 -

20

1,000

500

200

100

50

20

10

5

TNT

'A

B /

<2

1.500 1,000 500 C 

DISTANCE FROM DISPOSAL PITS, IN METERS

Figure 3. Concentrations of dissolved nitrogen species at exploratory wells A-E within the contaminant 

plume. At well C, range of concentrations in graphs is for three screened intervals of differing depth.

 1 

- AMMONIUM

E 

 1

' 1 B 0 1 '

t \ 

\

1 \ 

/ \

/ \A 

/ \ 

/ * 

C f//

_ ORGANIC NITROGEN

E

 1 

D 

. ».!-----

1

--------i

c I-'*'* 

-I---'*

901
Table 2. Chemical and physical character of saturated sedimentary deposits within and adjacent to the contaminant plume1 

[mg/kg, milligrams per kilogram, dry weight; ng/kg, micrograms per kilogram, dry weight; <, less than]

Range of measured values

Contaminant plume Adjacent deposits

Constituent or property Maximum Minimum Maximum Minimum

Clay + silt (percent by weight)

Water (percent by weight)

Nitrogen (N) species (mg/kg) :

Recoverable organic N

Nitrate plus nitrite as N

Ammonium as N

Total recoverable N

Carbon (mg/kg) :

Recoverable organic

Inorganic

Explosives (Jj.g/kg) :

Trinitrotoluene

Picrate

29

16

440

6.4

80

440

800

3,800

<25

<25

4

8

11

.2

1.3

73

100

<50

<25

<25

17

14

100

1.1

4.9

110

100

6,300

<25

<25

3

11

42

46

<50

<50

<25

<25

.0

.7

'Data are for cores collected below water table at, adjacent to, and downgradient from well C (fig. 2; data are not available for deposits between 

well C and disposal pits). Chemical analyses indicate combined concentrations for paniculate components and interstitial ground water. 

2Maximum or minimum concentrations for each nitrogen species were not necessarily measured in same sample.

At well E, principal nitrogen-bearing contaminants are thought to have been related to ammonium 

picrate wastes, yet nitrate is the dominant dissolvednitrogen species (fig. 3). The nitrate may have 

resulted from oxidation of ammonium by nitrifying 

bacteria (the dissolved-oxygen content of ground 

water was not measured during the study) as well as 

from decomposition of picramic acid following 

microbial alteration of picrate.

IMPLICATIONS OF DOCUMENTED 

DEPLETION

Most dissolved components of explosivesdemilitarization wastes that percolated to the shallow 

ground-water system at the arid study area were at 

least partly depleted within a relatively short distance 

(on the order of hundreds of meters) downgradient

from the disposal pits. This was particularly true for 

the nitrogen-bearing organic constituents (TNT, 

picrate, and their organic degradation products). 

The overall effectiveness of depletion is suggested by 

the downgradient decrease in the total amount of 

dissolved nitrogen: vertical, transverse segments of 

the entire plume one longitudinal meter thick at wells 

B, C, and D (figs. 2 and 3) would have contained 5, 

3, and 1 kg of dissolved nitrogen, respectively. (Other 

factors that undoubtedly contributed to this 80-percent 

decrease were longitudinal dispersion and fluctuations 

in contaminant input with time.)

Regardless of the depletion mechanisms involved, 

the Hawthorne study shows that natural microbiological and geochemical processes can degrade or retard 

the transport of large quantities of toxic, chemically 

complex, nitrogen-rich explosives wastes, leaving 

nitrate as the only residual nitrogen-bearing component in the ground-water plume.

902
Table 3. Summary of contrasting mobilities and depletion mechanisms for dissolved explosives-waste components of

ground-water plume

[>, greater than; <, less than]

Component

(in order of

decreasing

mobility)

Travel Distance Linear

time as traveled as velocity

of 1979 1 of 19792 (meters

(years) (meters) per year)

Proposed depletion 

mechanism(s)

Nitrate4 22 >1,700 >77 Presumably little if any 

depletion.

Trinitrotoluene

Ammonium

15

22

>450 >30 Microbial transformation to 

<1,100 <73 inorganic products;

probable sorption of 

organic transformation 

products.

>450 >20 Sorption; probable 

<1,100 <50 nitrification.

Picrate 22 <450 <20 Microbial alteration to

picramic acid, then 

metabolic degradation to 

inorganic products.

'Based on reported first year of disposal at six northernmost disposal pits (ammonium picrate, 1957; TNT and ammonium nitrate, 1964). (See 

figure 2 and text.)

2Linear distance between northernmost pits and wells C (about 450 meters), D (about 1,100 meters), and E (about 1,700 meters). (See 

figure 2.)

3 Linear velocity of ion or molecule at leading edge of "cloud" for each constituent within plume. ("Linear velocity" refers to net migration 

distance along straight line, as opposed to circuitous flow path travelled by interstitial water molecules and solutes.) In comparison, estimated average linear rate of ground-water flow is thought to be 120 to 140 meters per year. Velocities listed here for explosives-waste components are slightly 

refined from those given by Van Denburgh and Goerlitz (1987).

4Nitrate at well E is assumed to have originated from microbial degradation of ammonium picrate wastes; if, instead, it is from ammonium 

nitrate, linear velocity would exceed 110 meters per year.

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903
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Abstracts of the Annual Meeting: no. Q89, p. 276.

Won, W.D., Heckly, R.J., Glover, D.J., and Hoffsommer, 

J.C., 1974, Metabolic disposition of 2,4,6-trinitrotoluene: Applied Microbiology, v. 27, no. 3, p. 513-516.

904
Reaction of Arsenic Minerals and Gold Ores with 

Alkaline Cyanide Solutions

By Walter H. Ficklin1 , Geoffrey S. Plumlee1 , and John B. McHugh1

Abstract

Gold ores often contain significant quantities 

of arsenic (As). The As may occur in several 

different mineral forms and/or oxidation states. 

The most common As minerals are arsenopyrite 

(FeAsS), orpiment (As2S3), and realgar (AsS) 

but some deposits may contain oxidized forms of 

arsenic (arsenates) and more exotic minerals, such 

as enargite (Cu3AsS4). During heap leaching with 

alkaline cyanide solutions, some As is dissolved 

from these minerals. At pH values equivalent to 

those found in heap-leach solutions, the presence 

of cyanide increases the concentration of arsenic 

that dissolves by a factor of 4 to 100 for various 

forms of As. Formation of thiocyanate is one 

factor that enhances the dissolution of arsenic 

from sulfide minerals. Orpiment reacts more 

strongly with cyanide solutions than realgar and 

much more strongly than arsenopyrite.

INTRODUCTION

Many gold ores contain significant quantities of 

arsenic (As). The As may be present in several different 

mineral forms and oxidation states. The most common 

As minerals are arsenopyrite (FeAsS), orpiment 

(As2S3 ), and realgar (AsS), but some deposits may 

contain oxidized forms of arsenic (arsenates) and more 

exotic minerals, such as enargite (Cu3AsS4). During 

heap leaching with alkaline cyanide solutions, As is 

extracted from ores, sometimes in significant quantities. 

Concentrations of As in various heap-leach solutions 

collected from operations in Colorado and Nevada are 

presented in table 1. Some of these concentrations

Table 1 . Typical arsenic concentrations 

determined in heap-leach solutions from 

selected operations in Nevada and Colorado 

[Concentrations are in milligrams per liter; number 

following State denotes different samples from 

each State]

Site Concentration

Nevada 1 

Nevada 2 

Colorado 1 

Colorado 2 

Colorado 3

30

.5 

.3 

.01 

.07

'U.S. Geological Survey, Denver, Colo.

exceed U.S. Environmental Protection Agencyand state 

drinking water regulations.

The agent responsible for dissolution of As by 

alkaline cyanide solutions is not known with certainty. 

The pH of heap-leach solutions ranges from 10 to 12, 

and the minerals may dissolve because of the action of 

alkali and (or) the cyanide in solution.

As is an important element to be considered in 

heap leaching for gold recovery because its dissolution becomes a disposal problem when the heap is no 

longer used for gold recovery. Furthermore, the 

dissolution of As minerals may account for excessive 

cyanide consumption during the heap-leaching 

process. If leaks develop in heap-leach pads, the As 

could become an environmental contaminant in soil, 

alluvium, and ground water.

In this paper we describe the effect of cyanide 

and alkaline solutions on the dissolution of orpiment, 

realgar, arsenopyrite and some poorly characterized 

forms of As-bearing tailings.which may be remined 

to recover gold that was not recovered in the first 

processing of the ore.

905
MATERIALS AND METHODS

Orpiment and realgar samples were obtained from 

the Getchell Mine in Nevada. Arsenopyrite was supplied by Steven Sutley of the U.S. Geological Survey 

(USGS). Oxidized and unoxidized samples of tailings 

from Whitewood Creek, South Dakota, were obtained 

from Kim Goddard of the USGS. The gold reference 

standard sample was obtained from David Grimes of 

the USGS.

Reagent-grade sodium hydroxide and sodium 

cyanide were used to prepare the necessary solution for 

this study. Analysis of As in samples was performed by 

graphite furnace atomic-absorption spectrophotometry 

(Ficklin, 1983). Analysis of thiocyanate (SCN ) in samples was performed by the method presented in Standard 

Methods for the Examination of Water and Wastewater 

(American Public Health Association, 1985, p. 348). 

Analysis of As species (As(IH)and As(V)) in samples 

was performed by the method of Ficklin, (1983). Iron 

concentration was determined by flame atomic-absorption spectrophotometry (Aruscavage and Crock, 1987).

EXPERIMENTAL 

Experiment 1

Experiment 1 was carried out to determine 

whether the extraction of arsenic is enhanced by 

cyanide solution. In order to do this, 0.5 g each of 

oxidized (3,000 ppm As) and unoxidized (4,000 ppm 

As) Whitewood Creek tailings and the gold reference material, which was a sample of gold ore from 

the Pinson Mine, Nev., (DGPM1 containing 180 

ppm As), were extracted by four solutions. The solutions used were 20 mL of sodium hydroxide solution 

of pH 10.85, 20 mL of 500 ppm cyanide solution 

(pH-10.85), 20 mL of 0.1N sodium hydroxide, and 

20 mL of 0.1 N sodium hydroxide that contained 

500 ppm cyanide.. The samples were placed in 

capped glass centrifuge tubes, and allowed to react 

for 1 week with occasional shaking. A duplicate set 

was reacted for 2 weeks. The resulting extracts were 

examined for total As concentration. 

Experiment 2

Samples of pulverized orpiment (0.1 g) were 

placed in eight centrifuge tubes. Forty mL of 500 ppm 

cyanide solution was added to each tube. One of the 

samples was allowed to react for 30 minutes with 

occasional shaking. Another was allowed to react for

24 hours with occasional shaking. The remaining six 

samples were opened, each in succession at approximately 48-hour intervals until about 500 hours had 

passed. Each solution was measured for total As, 

As(III), As(V), and SCN~. Similar experiments were 

conducted with realgar (0.1 g) and arsenopyrite (0.2 g). 

Iron concentration in the solutions generated by the 

reaction of arsenopyrite also was measured.

RESULTS AND DISCUSSION

The dissolution of As from gold-bearing materials 

in alkaline cyanide solutions may be due to the extractive nature of alkaline solutions or to reaction of the 

substrate with cyanide. In order to determine which of 

these processes occurs, arsenic-bearing samples were 

treated, as described in experiment 1. Results of these 

experiments are presented in figure 1.

In solutions generated by experiment 1 having a 

pH of 10.85, but without cyanide, only a small 

amount of As was extracted. In the cyanide solution 

of the same pH, approximately 100 times the amount 

of As was extracted from the oxidized tailings and 

unoxidized tailings. The amount of As extracted 

from the gold ore sample (DGPM1) was approximately four times greater with cyanide solution (fig. 

1 A). Data for samples extracted with alkali solutions 

having a pH of 13 are presented in figure IB. At this 

pH, the amount of As that was extracted is not 

enhanced by the presence of cyanide solutions. 

Strong alkaline solutions will extract As regardless 

of the concentration of cyanide; but, with the pH 

values that are likely to prevail in heap-leach pads, 

cyanide has a significant effect on the extraction of 

As.

The significant increase in the dissolution of As 

from the unoxidized tailings at a pH of 10.85 suggests 

that the dissolution of sulfide-bearing As minerals by 

cyanide may be due to the reaction of cyanide with the 

sulfide part and not the As part of the mineral.

The percentage of As extracted from each mineral 

type, described above in experiment 2, is presented in 

figure 2. In similar alkaline cyanide solutions, orpiment 

is much more soluble than realgar, which, in turn, is 

much more soluble than arsenopyrite. The analytical 

results for the concentrations of As species, total As, 

SCN~, and Fe (arsenopyrite only) for timed extractions 

of orpiment, realgar, and arsenopyrite are presented in 

figures 3A,B,C. The extraction of As with 500 ppm

906
180-

160-

a 140-

CONC., MILLIGRAMS PERKILC IN b. 

8 Sg3

i

i

I

|

1 WEEK

\

2 WEEKS

WWCOX WWCRED DGPM1 WWCOX 

\

WWCRED 

I DGPM1

EXPLANATION

ss

WITHCN 

WITHOUT CN

2000

1800-

2 1600-

O 1400-

^ 1200-

W

g 1000-

2

^ 600- 

J 600^ 

8 4 

200-

1 WEEK

2 WEEKS

WWCOX WWCRED DGPM1 WWCOX WWCRED DGPM1

EXPLANATION 

ES

WITH CN 

WITHOUT CN

Figure 1 . Extraction of arsenic from three different arsenic-bearing 

materials for 1 and 2 weeks at (A) pH of 10.85 and (B) pH of 13. 

WWCOX is oxidized tailings from Whitewood Creek S.D., WWCRED is 

unoxidized tailings from Whitewood Creek, S.D., and DGPM1 is a sample from the Pinson Mine, Nev.

cyanide solution produced As(III), As(V), and an 

unidentified As compound from orpiment and realgar, 

but only As(V) from arsenopyrite. The concentration 

of the unidentified As compound can be calculated as 

the difference of the total As minus the sum of As(III) 

and As(V). As2S3 reacts with alkaline solutions 

according to the following reaction (Hogness and 

Johnson, 1957):

As2S3 + 4OH- = AsS3-3 + AsO2- + H2O; (1)

in water As2S3 reacts according to the following 

equation:

2HAs02 + 3H2S = As2S3 + 4H2O. (2)

The solubility product constant for reaction 2 is 

2.1 x 10 22 . Based on equation 1, the unidentified As 

compound found in the orpiment and realgar experiments is most likely AsS3 3 .

SCN forms very rapidly from the reaction of 

polysulfides and thionates with cyanide (Leuthy and 

Bruce, 1979). McGill and others (1985) studied the 

reaction of cyanide with several metallic minerals and 

found that SCN forms very slowly and uses only a 

small part of the cyanide. According to reaction (2), a 

small amount of H2S forms from the dissolution of 

As2S3 . The oxidation of sulfides to sulfate proceeds 

through either polysulfides and (or) thionates

907
ARSENOPYRITE

100 200 300 400 500

HOURS

Figure 2. Percentage of arsenic extracted from orpiment, realgar and arsenopyrite in a 500 ppm cyanide solution.

600

(Boulegue, 1981). When the polysulfides or thionates 

form, the reaction with cyanide is very rapid. Consequently, additional orpiment or realgar dissolves. The 

data presented in figure 3 suggests that formation of 

SCN~ from cyanide and the sulfide part of the arsenic 

minerals is an important factor in the increasing solubility 

of orpiment and realgar over time. As long as free 

cyanide is available to react with the S released in the 

dissolution of the As-sulfide minerals, more and more 

As will eventually be dissolved.

The mechanism for dissolution of As from arsenopyrite is apparently different from that of the simple 

As-sulfide minerals. Both the Fe and As concentrations 

were uniform after about 100 hours. However, the 

formation of SCN~ continued to increase over time. 

A possible explanation is that the mineral continues to 

react with the cyanide solution to supply S for the

formation of SCN~, but that precipitation of an As-Fe 

mineral of unknown composition occurs.

SUMMARY

As is dissolved from As-bearing minerals and 

gold ores by alkaline and alkaline cyanide solutions. 

At pH values equivalent to those found in heap-leach 

solutions, cyanide increases the concentration of 

arsenic that dissolves by a factor of 4 to 100 for 

various forms of As. Formation of SCN~ is likely one 

of the controls that affects the amount of dissolution 

of the As-bearing materials examined. Based on data 

presented above, cyanide reacts most strongly with 

orpiment, moderately with realgar, and weakly with 

arsenopyrite.

908
160

140

W 120

100

EXPLANATION

As

As(lll) 

As(V) 

SCN

100 200 300 400 

HOURS

500 600

EXPLANATION

As

As(lll) 

As(V) 

SCN

REFERENCES

American Public Health Association, 1985, Standard methods for examination of water and wastewater: Washington, D.C., p. 348.

Aruscavage, P.J., and Crock, J.G., 1987, Atomic absorption 

methods, in Baedecker, P. A., cd., Methods of geochemical analysis, U.S. Geological Survey Bulletin 1770, 

chap. C,p. C1-C5.

Boulegue, Jacques, 1981, Simultaneous determination of 

sulfide, polysulfides, and thiosulfate as an aid to ore 

exploration: Journal of Geochemical Exploration, v. 15, 

p. 21-36.

Ficklin, W.H., 1983, Separation of arsenic(III) and arsenic(V) 

in ground waters by ion exchange: Talanta, v. 30, no. 5, 

p. 371-373.

Hogness, T.R., and Johnson, W.C., 1957, An introduction to 

qualitative analysis: New York, Holt, Rinehart, and 

Winston, 376 p.

Leuthy, R.G., and Bruce, S.G., Jr., 1979, Kinetics of reaction of cyanide and reduced sulfur species in aqueous 

solution: Environmental Science and Technology, v. 13, 

no. 12, p. 1481-1487.

McGill, S.L., Hendrix, J.L., and Nelson, J.H., 1985, Cyanide/ 

thiocyanate reactions in tailings, in Van Zyl, Dirk, ed., 

Cyanide and the environment Proceedings of a conference, December 11-14, 1984: Tucson, Ariz., GeoTechnical Engineering Program, Department of Civil 

Engineering, Fort Collins, Colorado State University, 

p. 143-159.

100 200 300 400 

HOURS

500 600

2 3-

z

z 

O 2' I

H

100 200 300 400 

HOURS

500

EXPLANATION

As

As(V) 

SCN 

FE

600

Figure 3. The speciation of arsenic and the formation 

of thiocyanate from (A) orpiment, (B) realgar, and 

(C) arsenopyrite as a function of time.

909
910
Anaerobic Growth of Bacterial Strain Ses3 with Selenate

as the Electron Acceptor

SyJodi Switzer Blum1 , Charles W. Culbertson 1 , and Ronald S. Oremland1

Abstract

The bacterial strain SES3 is capable of growth 

with selenate, nitrate, or Fe(III) as electron acceptors, and lactate as an electron donor. Selenate is 

reduced through selenite to elemental selenium, 

whereas nitrate was reduced through nitrite to 

ammonia, and Fe(III) was reduced to Fe(II). 

Lactate was oxidized to CO2 and acetate in an 

approximately 1:1 ratio. Results of washed-cell 

experiments comparing nitrate and selenate reduction in either nitrate-grown or selenate-grown cells 

indicated that separate enzyme systems performed 

the reductions. Washcloth experiments using 

uniformly labelled, [U-I4C] lactate, showed that 

fumarate, arsenate, and O2 also functioned as electron acceptors. The respiratory inhibitors azide, 

carbonyl cyanide m-chlorophenyl-hydrazone 

(CCCP), and dinitrophenol (DNP) inhibited both 

selenate and nitrate reduction in washed-cell 

suspensions.

INTRODUCTION

Selenium, an essential micronutrient in nature, 

becomes toxic to plants (Aslam and others, 1990; 

Cappon, 1991) and animals (Nebbia and others, 1991; 

Van Vleet and Ferrans, 1992), and inhibitory to a 

variety of microorganisms (Oremland, 1993) when 

present at high (micromolar) concentrations. Bacterial 

reduction of dissolved selenium oxyanions to insoluble 

elemental selenium represents a potential bioremediation process (Gerhardt and others, 1991; Oremland and 

others, 1991; Oremland, 1993); moreover, selenate 

reduction in nature is a widespread phenomenon 

(Steinberg and Oremland, 1990). Several studies have

U.S. Geological Survey, Menlo Park, Calif.

shown that bacteria are capable of reducing selenate to 

elemental selenium (Oremland and others, 1989; Lortie 

and others, 1992; Rech and Macy, 1992; Steinberg and 

others, 1992; Tomei and others, 1992). A bacterium, 

designated strain SES3, was selectively isolated to 

reduce selenate. SES3 is an anaerobic, Gram negative, 

motile, vibroid organism that is capable of growth with 

selenate as the terminal electron acceptor. Strain SES3 

will reduce nitrate preferentially over selenate if both 

are present in the growth medium. (Steinberg and 

others, 1992). In this paper, strain SES3 is characterized with emphasis on the relation between selenate 

and nitrate respiration.

EXPERIMENT

Strain SES3 grew by oxidizing lactate to form 

acetate and CO2 and was coupled to the reduction of 

either selenate (fig. 1A, B) or nitrate (fig. 2A, B). Elemental selenium was observed as a bright red precipitate in 

selenate reducing cultures (Steinberg and Oremland, 

1990). The precipitate formation was concurrent with 

selenate reduction. SES3 converted approximately 6 

millimoles of Fe(III) to Fe(II) concurrent with an 

approximately 10-fold increase in cell density (D.R. 

Lovely, and E.J.P. Phillips, U.S. Geological Survey, 

written communication, 1993). Pyruvate, or acetate 

+H2 also acted as electron donors, but not acetate or H2 

alone. Experiments were performed with washed-cell 

suspensions to contrast the metabolism of nitrategrown cells to selenate-grown cells. Cells of strain 

SES3 grown in nitrate medium were unable to use 

selenate unless 1 mM NO3 was present in the cell suspension, and then only a small amount was reduced 

after prolonged incubation (140 hours). Selenate reduction was inhibited by 200, ug/mL chloramphenicol (50 

percent inhibition). Nitrate reduction by nitrate-grown 

cells was rapid (2 hours) and was unaffected by 

chloramphenicol. Cells grown in selenate medium

911
60 90 120 

ELAPSED TIME IN HOURS

O 

u

A Selenate

 Selentte

A Cells/milliliter

le+010

1 le+009

1 le+008

on

- le+007 w

le+006

0 30 (0 90 120 150 180 

ELAPSED TIME IN HOURS

Figure 1. Growth and metabolism of strain SES3 with selenate. (A) Oxidation of lactate; and 

(B) Reduction of selenate; cell mass.

912
25 50 75 

ELAPSED TIME IN HOURS

100

H

O

3

7

A Nitrite

 Nitrate

t Ammonia

v Absorbaace 680 (nm)

0.2

0.4

0.3

25 50 75 

ELAPSED TIME IN HOURS

Figure 2. Growth and metabolism of strain SES3 with nitrate. (A) Oxidation of lactate; and 

(B) Reduction of nitrate; cell mass.

I

O 

on

3

913
reduced nitrate only after prolonged incubation (160 

hours) in washed-cell suspensions, and the reduction 

was inhibited by chloramphenicol (71 percent inhibition). The addition of 5 mM selenate to the nitrate-cell 

suspension accelerated nitrate reduction by selenategrown cells (25 hours), and this reduction was inhibited 

by chloramphenicol (100 percent inhibition). Selenate 

reduction was nearly complete after 50 hours in 

washed-cell suspensions of selenate-grown cells with 

slight inhibition by chloramphenicol (13 percent inhibition). Another set of washed-cell experiments contrasted the ability of nitrate-grown and selenate-grown 

cells to convert [U- 14C] lactate to 14CO2 . Nitrategrown cells converted 53 percent of the [U- 14C] lactate 

to 14CO2 if NC>3 was the electron acceptor, and none if 

SeO4 was the electron acceptor. Conversely, selenate 

grown cells converted 41 percent of the [U- 14C] lactate 

to 14C02 if SeO4 was the electron acceptor, and only 6 

percent if NO3 was the electron acceptor. Both nitrategrown and selenate-grown cells were able to couple 

oxidation of [U-I4C] lactate to 14CO2 with reduction of 

arsenate, fumarate, Fe(III), and O2 (O2 was tested for 

NO3 -grown cells only). The respiratory inhibitors 

azide, CCCP, and DNP were strong inhibitors of lactate 

oxidation for nitrate-grown, and selenate-grown cells 

(table 1).

Table 1. Respiratory inhibitors; SeO4 and NO3 grown cells

[Based on mean of two replicates. Incubation times: SeO4 grown 

cells = 3.5 hours; NOs grown cells = 2.0 hours. Cell densities: SeO4 

grown cells = 5.9x108 cells per milliliter; NOs grown cells = 3.7x109 

cells per milliliter]

Inhibitor

Percent inhibition1

SeO4 NO3 

grown cells grown cells

SeO4 or NO3 control

Rotenone (1 00 micromolar)

Cyanide (500 micromolar)

Azide (10 millimolar)

HOQNO (50 micromolar)

Antimycin A (1 00 micromolar)

CCCP (1 00 micromolar)

DNP (1 00 micromolar)

Acetone (control)

ETOH (control)

None

0

2.4

1 5.6

71. 1

18.9

11.0

78.7

76.7

3.5

5.1

80.3

0

8.2

1.2

42.0

2.0

20.1

60.5

32.4

24.4

21.6

85.2

1 Percent I4C recovery from added label in controls: 

SeO4 grown cells = 87.8 

NO3 grown cells = 89.6

Rotenone, cyanide, 2-heptyl 4-hydroxyquinoline 

N-oxide (HOQNO), and antimycin A showed little 

inhibitory effect. Rotenone, antimycin A, CCCP, and 

dinitrophenol were dissolved in acetone before addition to the washed-cell suspension, and HOQNO was 

dissolved in ethanol. No solvent effects were observed 

for selenate-grown cells, but a slight solvent effect was 

observed for nitrate-grown cells.

CONCLUSION

Bacterial strain SES3 is potentially useful for the 

bioremediation of toxic selenium oxyanions. SES3 can 

remove selenate from solution by precipitation as elemental selenium. The presence of nitrate interferes 

with this process (Steinberg and others, 1992); however, washedcell data (above) indicate that the enzyme 

systems for nitrate and selenate reduction function separately. Rech and Macy (1992) made similar conclusions about nitrate and selenate reduction in Thauera 

selenatis. Experiments with respiratory inhibitors indicate that nitrate and selenate reduction in strain SES3 is 

due to respiration.

REFERENCES

Aslam, M, Harbit, K.B., and Huffaker, R.C., 1990, Comparative effects of selenite and selenate on nitrate 

assimilation in barley seedlings: Plant, Cell and 

Environment, v. 13, p. 773-782.

Cappon, C.J., 1991, Sewage sludge as a source of environmental selenium: The Science of the Total Environment, v. 100, p. 177-205.

Gerhardt, M.B., Green, F.B., Newman, R.D., Lundquist, 

T.J., Tresan, R.B., and Oswald, W.J., 1991, Removal of 

selenium using a novel algal-bacterial process: 

Research Journal of the Water Pollution Control Federation, v. 63, p. 799-805.

Lortie, L., Gould, W.D., Rajan, S., McCready, R.G.L., and 

Cheng, K.-J., 1992, Reduction of selenate and selenite 

to elemental selenium by a Pseudomonas stutzeri isolate: Applied and Environmental Microbiology, v. 58, 

p. 4042-4044.

Nebbia, C, Soffietti, M.G., Zittlau, E., and Fink-Gremmels, 

J., 1991, pathogenesis of sodium selenite and dimethylselenide acute toxicosis in pigs cardiovascular 

changes: Research in Veterinary Science, v. 50, 

p. 269-272.

914
Oremland, R.S., in press, Biogeochemical transformations 

of selenium in anoxic environments, in Frankenberger, 

W.T., Jr., Selenium in the Environment: New York, 

Marcell Dekker.

Oremland, R.S., Hollibaugh, J.T., Maest, A.S., Presser, T.S., 

Miller, L.G., and Culbertson, C.W., 1989, Selenate 

reduction to elemental selenium by anaerobic bacteria 

in sediments and culture Biogeochemical significance 

of a novel, sulfate-independent respiration: Applied 

and Environmental Microbiology, v. 55, p. 2333-2343.

Oremland, R.S., Steinberg, N.A., Presser, T.S., and Miller, 

L.G., 1991, In situ bacterial selenate reduction in the 

agricultural drainage systems of western Nevada: 

Applied and Environmental Microbiology, v. 57, 

p. 615-617.

Rech, S.A., and Macy, J.M., 1992, The terminal reductases 

for selenate and nitrate respiration in Thauera selenatis 

are two distinct enzymes: Journal of Bacteriology, 

v. 174, p. 7316-7320.

Steinberg, N.A., Oremland, R.S., 1990, Dissimilatory selenate reduction potentials in a diversity of sediment types: 

Applied and Environmental Microbiology, v. 56, p.

3550-3557.

Steinberg, N.A., Blum, J.S., Hochstein, L., Oremland, R.S., 

1992, Nitrate is a preferred electron acceptor for 

growth of freshwater selenate-respiring bacteria: 

Applied and Environmental Microbiology, v. 58, p. 

426-428.

Tomei, F.A., Barton, L.L., Lemanski, C.L., and Zocco, T.G., 

1992, Reduction of selenate and selenite to elemental 

selenium by Wolinella succinogenes'. Canadian Journal 

of Microbiology, v. 38, p. 1328-1332.

Van Vleet, J.F., and Ferrans, V.J., 1992, Etiologic factors 

and pathologic alterations in seleniumvitamin E deficiency and excess in animals and humans: Biological 

Trace Element Research, v. 33, p. 1-21.

915
916
Chemical, Isotopic, and Microbiological Evidence for 

Denitrification During Transport of Domestic Wastewater 

Through a Thick Unsaturated Zone in the Mojave Desert

By Roy A. Schroeder1 , Peter Martin 1 , anc/J.K. Bohlke2

Abstract

Septic-tank wastewater disposed in 30-footdeep seepage pits (dry wells) at 46,000 residences 

is estimated to equal about 18 percent of natural 

recharge to the sole-source aquifer beneath the 

upper Mojave River Basin, which is rapidly 

becoming urbanized, in the high desert northeast 

of Los Angeles. Nitrogen in the downwardinfiltrating wastewaters represents a significant 

potential source of nitrate contamination to underlying ground water, but increases in nitrate 

concentration in ground water have not yet been 

observed. The low nitrate concentration in the 

ground water may be the result of lateral dispersion in the unsaturated zone, dilution below the 

water table, or denitrification of wastewater nitrate 

in the unsaturated zone.

Measured vertical rates of movement of wastewater fronts through the unsaturated zone at three 

newly occupied residences ranged from 0.07 to 

1 foot per day. Those measurements, along with 

moisture-content profiles at older residences, indicate that some wastewater has reached the water 

table beneath communities that are older than 5 to 

10 years. As wastewater percolates from seepage 

pits into the unsaturated zone, reduced nitrogen is 

converted rapidly to nitrate at shallow depths. 

Analyses of water extracts of soil cores and of soil 

moisture from suction lysimeters deep beneath 

seepage pits at several residences indicate that 

nitrate concentrations commonly decrease with 

depth. The largest nitrate decreases seem to coincide with increased content of fine-grained sediments or proximity to the water table. Nitrate-

'U.S. Geological Survey, San Diego, Calif. 

2 U.S. Geological Survey, Reston, Va.

reducing bacteria were found in soil cores 

collected from two residences. Between lysimeters at 160 and 199 feet at one residence, the 

decrease in nitrate concentration coincided with a 

large increase in sulfate, decrease in alkalinity, and 

increase in 5 I5N in nitrate. Those data are 

consistent with denitrification by oxidation of iron 

sulfide to produce ferric oxides; but if such a reaction occurs, it must be in domains that are small in 

comparison with the sampled volumes because the 

waters also contain substantial quantities of 

dissolved oxygen. The predominantly low nitrate 

concentrations in the area's ground water are 

consistent with the operation of a nitrogenremoval mechanism, possibly denitrification, as 

wastewater moves through an unsaturated zone 

that averages 150 feet in thickness. However, the 

reducing capacity of the sediments to maintain 

denitrification is not known.

INTRODUCTION

Septic-tank wastewater disposed in 30-foot-deep 

seepage pits (dry wells) at 46,000 residences in the 

communities of Victorville, Apple Valley, and 

Hesperia, is estimated to equal about 18 percent of 

natural recharge to the sole-source aquifer in the upper 

Mojave River Basin, which is rapidly becoming urbanized, in the high desert 80 miles (mi) northeast of Los 

Angeles (fig. 1).

The use of seepage pits in this area, instead of nearsurface leach fields that are more commonly used elsewhere, is necessitated by the widespread existence of 

an impermeable caliche layer near the land surface. 

This wastewater-disposal practice has caused concern 

about possible contamination of the underlying aquifer, 

especially by nitrate. In response to that concern, the 

U.S. Geological Survey began a study in 1988 to

917
117° IT R 4 W R 3 W 117 o 12

34° 32' 

CHOCTAWX CHEYENNE

Valley ^ APPLE

KIOWA<> RANCHOS 

LITTLE 

BEAVER

5 MILES

TOLTEC

CAJON 

RINCONL

5 KILOMETERS

EXPLANATION

MONITORING SITE AND NAMELysimeter hole and 

neutron logging hole

Lysimeter hole(s) 

Multiple well

Los 

Angeles

Figure 1. Location of study area and monitoring sites in the upper Mojave River Basin.

1. Measure vertical rates of wastewater movement 

through the unsaturated zone,

2. Determine which chemical constituents of the 

wastewater reach the water table, and

3. Evaluate the current and potential effects of wastewater discharge on future ground-water quality.

This paper presents a summary of results of that 

study, which was funded jointly by the California 

Regional Water Quality Control Board, Lahontan 

Region, and the U.S. Geological Survey (Southern 

California Regional Aquifer-System Analysis program 

and other funds).

918
Instrumentation at Eight Residences

Monitoring was done to follow the movement of 

wastewater and to detect chemical changes in the water 

in the unsaturated zone at eight residences (locations 

shown in fig. 1) chosen to represent a range in geohydrologic conditions, residential density, and years of 

seepage-pit operation. Test holes were drilled next to 

the seepage pits at five previously occupied residences 

and through the bottoms of the seepage pits at three 

residences that were occupied soon after completion of 

drilling and instrument installation in 1988 (see table 1). 

Cores were collected for analysis of various physical, 

chemical, and microbiological properties. Multipledepth suction ly si meters were installed for waterquality sampling at all eight residences. Galvanizedsteel access tubes were installed to measure changes in 

soil moisture by neutron logging at the three newly 

occupied residences and at one previously occupied 

residence (Cheyenne-I site).

Possible Sources of Nitrate

Downward-infiltrating domestic wastewater is 

considered to be the dominant potential source of nitrate 

contamination to ground water in the study area because

1. There are few domestic wells at which faulty well 

construction could allow wastewater to "short circuit" a thick unsaturated zone that ranges from less 

than 50 feet (ft) beneath the Mojave River to more 

than 400 ft in parts of Apple Valley and Hesperia;

2. Landscaping, and its accompanying use of nitrogen 

fertilizers, is minimal in this desert environment;

3. What little agriculture exists has always been 

confined to a narrow strip adjacent to the Mojave 

River; and

4. Although natural nitrate-rich soils are present in 

parts of the Mojave and Death Valley deserts, these 

soils generally are small in areal extent and highly 

localized, and none have been reported near the 

study area.

Observed Nitrate Concentrations in 

Ground Water

Nitrogen concentrations in wastewater samples 

from several residences averaged almost 50 milligrams 

per liter (mg/L), or 3,600 micromolar ((iMX and the 

nitrogen in the wastewater was present mostly as 

ammonium and organic nitrogen (see table 1). Nitrogen concentrations in infiltrating wastewater sampled 

below seepage pits were similar (typically 20-60 mg/L), 

but the nitrogen in the infiltrating wastewater was 

present mostly as nitrate, indicating that nitrification of 

wastewater occurs rapidly during infiltration. Despite 

this large potential source of downward-moving nitrate 

contamination, the nitrogen concentrations (mostly as 

nitrate) averaged only 1.2 mg/L (90 \iM) in recent samples from 20 municipal wells in the study area. Ground 

water sampled just below the water table at three 

residences and between 5 and 100 ft below the water 

table in multiple-depth wells at the Apple Ranches and 

Rincon sites (locations shown in fig. I) also contains 

little nitrate.

Possible reasons for the absence of widespread 

nitrate contamination in the area's ground water 

include the following:

1. Wastewater has reached the water table but is 

diluted by large-scale vertical mixing in a saturated 

zone that is as thick as 1,000 ft.

2. Wastewater has been retained within the unsaturated zone and has not yet reached the water table.

Table 1. Characteristics of monitoring sites at eight residences in the upper Mojave River Basin 

[gal/d; gallon per day; mg/L, milligram per liter; ft, foot; --, not determined]

Site 

of 

residence

Toltec 

Cheyenne-I 

Vasquez 

Choctaw 

Little Beaver 

Kiowa 

Cheyenne-ll 

Cajon

Year 

seepage pit 

constructed

1 988 

1 983 

1 988 

1 978 

1 988 

1 979 

1 987 

1 977

Winter water 

usage, 1989 

(gal/d)

388 

750 

250 

625 

875 

8I2 

338 

260

Wastewater 

nitrogen 

(mg/L)

53 

26 

46

22 

60 

40

Depth to water 

table, 1989-90 

(ft)

'220 

H2 

85 

1 24 

 ISO 

'ISO 

H7 

252

'Water table not reached by drilling; depth based on estimates from Subsurface Surveys, Inc. (I990, p. 4).

919
3. Nitrogen in the wastewater has been attenuated by 

processes such as denitrification as the wastewater 

moves through the unsaturated zone.

Each of these explanations was evaluated in this study.

"PREDICTED" NITRATE CONCENTRATIONS IN GROUND WATER

Differences in major-ion characteristics at the 

Rincon multiple-depth wells indicate that vertical 

mixing with wastewater probably will occur within the 

upper 50 ft of the saturated zone. Results of simulations 

made using a single-cell mixing model similar to that 

described by Bauman and Schafer (1985) that allows 

nitrate from wastewater to be diluted by instantaneous 

vertical mixing indicate that readily measurable 

increases in nitrate concentration would be expected in 

shallow ground water within 10 years after the wastewater reaches the water table (fig. 2). The model simulates conditions that would be expected in the area 

between the Apple Ranchos and Rincon multiple-depth 

wells a part of Apple Valley in which most of the 

houses were constructed in the early 1960's and where 

depth to ground water is sufficiently small that wastewater should have reached the water table long ago. 

The absence of high nitrate concentrations in the

multiple-depth wells suggests that either nitrogen is 

being removed from the septic-tank wastewater or the 

vertical mixing depth must be substantially greater than 

50ft.

RATE OF WASTEWATER MOVEMENT

The arrival of the wastewater wetting front in the 

unsaturated zone was recognized from (1) temporal 

changes in neutron-log profiles, (2) the recovery 

of large sample volumes from lysimeters, and 

(3) increased soil moisture in cores. Downward movement of the front is indicated by the differences 

between successive neutron logs for a recently occupied residence at Toltec (fig. 3). Lysimeters at each 

depth at that site yielded recoverable water only after 

the neutron log indicated that the wastewater wetting 

front had arrived. Soil-moisture contents are relatively 

high throughout the unsaturated zone at the established 

sites and indicate that the wastewater wetting front has 

reached the water table at most older sites.

Measured rates (from neutron-log profiles and the 

suction lysimeters) of vertical movement of the wastewater wetting front in the upper 100 ft ranged from 

0.07 to 1.0 foot per day (fl/d) at the three newly occupied residences. These measured vertical rates correspond to traveltimes of several months to about 6 years

LL\ 

O

UJ

UJ 

Q_

00

s: 

<

CD

I 

50

40

10

Wastewater loading: 

Mixing depth: 

Housing density: 

 Hydraul ic conducti vity 

Hydraulic gradient: 

Porosi ty: 

Initial nitrate175 gallons per day per residence

50 feet

1 per acre

20 feet per day

0.0012

0.3

2 milligrams pe r 1 i t e r

Concentration of nitrate-N in wastewater, 

in milligrams per liter40 60 

TIME, IN YEARS

Figure 2. Model-computed nitrate concentrations in ground water in the study area for selected 

nitrate-N concentrations in wastewater.

920
10 20 30 40 50 0 10 20 

MOISTURE CONTENT, IN PERCENT VOLUME

30 40 50

2-2-89

EXPLANATION

Date of measurement

Change in moisture

content since previous 

measurement

 Bottom of neutron 

access tube

Figure 3. Temporal variation in soil-moisture content in the unsaturated zone at the Toltec site. (Shaded areas 

indicate moisture additions in succeeding time intervals and record the downward movement of the wastewater 

wetting front.

for wastewater to reach the water table at an average 

depth of about 150 ft (table 1). However, the vertical 

rates decrease with depth, possibly owing to lateral 

movement of the wastewater induced by stratigraphic 

barriers or increasing consolidation (or cementation) of 

the soils. Evidence for lateral movement was found in 

test holes drilled 25 ft from the seepage pit at the 

Cheyenne-I site and 50 ft from the seepage pit at the 

Choctaw site.

Further evidence for lateral movement can be seen 

from calculations of water volume potentially held

within the unsaturated zone and from chemical data at 

the Cajon site. If the wastewater input were 260 gallons 

per day (gal/d) (table 1), porosity were 0.3, and depth 

to ground water were 250 ft, then wastewater would 

saturate the soil in a 5-ft cylinder directly underlying 

the seepage pit in less than 2 months if there were no 

lateral movement. However, chemical data (table 2, 

mainly chloride concentrations) indicate that wastewater at the Cajon site has moved to a depth of approximately 200 ft in about 15 years. If uniform lateral 

movement were assumed, then wastewater would satu921
Table 2. Water-quality data from lysimeters at the Cajon residence, Hesperia, California

(o [Specific conductance, in microsiemens per centimeter at 25° Celsius, is given as a range of values from as 

10 November 1991-March 1993 at each depth. Isotope ratios are in permil deviation from standards (V-SMOW for O and H, air 

All other constituent concentrations except pH are in /*M (micromole per liter), and most are from single samples. If more th

median value is given. , no data; <, actual value is less than value shown]

Depth below land

surface, in feet . . .

Constituent or property

Specific conductance ....

Delta deuterium .......

Delta inorganic carbon- 13 .

Delta nitrate nitrogen- 15 . .

Delta oxygen- 18 .......

Molecular oxygen ......

pH (standard units) .....

Hydrogen ion .........

Calcium ............

Magnesium ..........

Sodium ............

Potassium ...........

Alkalinity (as CaCO3 )' . . .

Sulfate .............

Chloride ............

Fluoride ............

Bromide ............

Silica ..............

Nitrite .............

Ammonia ...........

Nitrite plus nitrate ......

Organic nitrogen .......

Orthophosphate .......

Boron .............

Chromium ...........

Copper .............

Iron ..............

Lead ..............

Manganese ..........

Nickel .............

Zinc ..............

Organic carbon ........

35

448-553

-12.75

10.5

120

7.2

.06

820

260

2,700

280

700

110

1,600

20

.6

1,300

2

7

1,900

40

110

18

.16

.02

240

70

435-525

-9.50

9.5

140

7.4

.04

720

300

2,100

40

550

120

1,200

10

.5

1,300

<1

11

1,900

20

74

10

«

~

<.05

~

.13

~

300

97.5

463-544

-59.5

-11.35

8.2

-8.75

140

7.1

.08

770

360

2,200

30

350

120

1,300

10

.8

800

<1

4

2,100

<10

100

7

<.05

<.02

230

130

470-590

-58.0

7.8

-8.65

7.4

.04

560

~

1

24

2,400

40

2

~

350

160

493-567

-61.0

-12.40

7.8

-8.85

190

7.4

.04

320

490

1,600

50

450

150

1,000

10

.8

730

<1

31

2,500

20

5

<1

~

.1

<.05

.02

.04

.02

14

290

199

391-535

-59.0

14.8

-8.85

160

4.0

100

340

370

1,800

90

-34

620

890

10

.6

1,500

<1

64

1,600

10

<1

2

<.02

5

9

.14

12

.5

7

170

'Negative alkalinity calculated relative to a titration endpoint of pH 4.5.
rate the soil within a cylinder to a radial distance of 

about 30 ft after 15 years. Actual moisture content of 

soils in the wastewater zone is somewhat less than saturation, and some preexisting moisture is present in the 

soils; therefore, wastewater movement would be proportionately faster and lateral movement a little farther 

(it varies inversely with the square root of moisture 

content) than indicated by the calculations, but still 

much less than the average distance between seepage 

pits at adjacent residences.

Although wastewater has not yet reached the water 

table at the Cajon site, abundant soil moisture is present 

beneath the wastewater zone as evidenced by large 

sample volumes from the lysimeter at 243 ft. This 

might be the result of an accumulation of natural soil 

moisture in advance of the wastewater wetting front in 

a manner similar to that described by Young and others 

(1992). Specific-conductance measurements at the 

Cajon site between November 1991 and March 1993 

indicate that no systematic change occurred in the 

chemical compositions in either the five lysimeters 

within the wastewater zone (above 199 ft) or the lysimeters at 243 and 257 ft that are below the wastewater 

zone. During the same period, the specific conductance 

at 199 ft increased monotonically in 11 successive

samples by a total of nearly 40 percent, presumably 

because the wastewater was displacing some residual 

low-conductance and low-chloride soil moisture. The 

lysimeter at 257 ft has yielded very little water since 

1991 when the water table dropped below this depth in 

response to nearby pumping and the recent drought in 

California.

EVIDENCE FOR NITRATE REDUCTION IN 

THE UNSATURATED ZONE

The possibility that nitrate is being removed from 

wastewater in the unsaturated zone is indicated by 

decreases in nitrate concentrations with depth in 

deionized-water core extracts and in lysimeter samples 

from the Cheyenne-I site (fig. 4) and in lysimeter samples from the Cajon site (table 2). Additional evidence 

for nitrate reduction comes from the observation that 

bacteria capable of nitrate reduction are present in 

some of the soils. Higher numbers of nitrate-reducing 

bacteria and reduced nitrate concentrations seem to be 

associated with finer grained sediments and higher 

moisture contents at the Cheyenne-I site (fig. 4). The 

nitrate-concentration profile from lysimeter samples at 

the Cheyenne-I site is based on monitoring for a period 

of only about 1 year in 1988-89, after which the

0 50 100 0 5 10 15 20 25 0

GRAIN-SIZE MOISTURE CONTENT, 

DISTRIBUTION, IN PERCENT OF 

IN PERCENT WET WEIGHT

10 20 30 40 50 60 70 80

NITRATE-NITROGEN CONCENTRATION, 

IN MILLIGRAMS PER LITER

90 10 1 102 io 4 10 5

NITRATE-REDUCING 

BACTERIA, IN NUMBER 

PER GRAM SOIL (WET)

EXPLANATION

GRAIN SIZE-- 

Within interval

Gravel 

Coarse sand 

Fine sand 

Silt and clay

OTHER DATA--

Soil core 

Lysimeter (median)

Less than bacteria 

detection 1imit

Figure 4. Grain-size distribution, moisture content, nitrate concentration, and nitrate-reducing bacteria counts at the 

Cheyenne-I site.

923
lysimeters were covered by a concrete driveway. The 

concrete was removed in the spring of 1993 and 

monitoring recently was resumed to ascertain if the 

decrease in nitrate concentration at depth that is shown 

in figure 4 represents transient or more permanent conditions, and to further evaluate possible reasons for the 

concentration profile that was observed.

If the reduced nitrate concentration at 199 ft at the 

Cajon site were solely the result of mixing of highnitrate wastewater (represented by data at 160 ft) with 

low-nitrate residual soil moisture (represented by data 

at 243 ft), then the chloride data indicate that the lysimeter at 199 ft contains 88 percent wastewater. That 

percentage predicts that nitrate concentration should be 

2,200 (iM- However, data in table 2 show that the 

actual nitrate concentration is only 1,600 |oM.

Isotopic evidence for denitrification is provided by 

the nitrogen-15 enrichment of the nitrate in a sample 

from the lysimeter at 199 ft in comparison with 

samples from the shallower lysimeters (fig. 4; table 2). 

The relation between the isotope ratios and nitrate concentrations is consistent with kinetic isotope fractionation during nitrate-consuming reactions with an 

isotope-enrichment factor of about -10 permit. That 

enrichment factor is within the wide range of values 

(approximately -30 to -5 permil) previously associated 

with denitrification in a variety of ground-water 

environments (Hubner, 1986; Mariotti and others, 

1988; Bottcher and others, 1990; Schroeder and others, 

1991; Smith and others, 1991). Isotope-enrichment 

factors at the lower end of this reported range, in this

and in other studies, could indicate the influence of 

specific physicochemical factors on the reaction, or 

mixtures of unreacted and reacted waters from differing physicochemical environments (Hubner, 1986; 

Mariotti and others, 1988).

If waters entering the lysimeters at 160 and 199 ft 

were dominated by wastewater that had the same initial 

composition (both having high chloride concentrations, for example), then the differences between their 

measured compositions could be treated as coefficients 

in a net chemical reaction that occurs between those 

depths as the wastewater moves downward. If that 

assumption were correct, then the data in table 2 indicate that the wastewater gained approximately 470 |oM 

of sulfate and 970 |oM of hydrogen ion (including alkalinity lost to carbonic acid) while losing 900 |oM of 

nitrate. If those quantities were normalized to hold 

chloride constant (assuming minor dilution of wastewater by low-salinity soil water, represented by data 

from the lysimeter at 243 ft), then the corresponding 

gains and losses would be approximately 550 joM sulfate and 980 |u,M hydrogen ion for 700 |oM nitrate. 

Those coefficients and underlying assumptions have 

large uncertainties; nevertheless, they can be interpreted as evidence for a reaction involving denitrification by oxidation of iron sulfide (table 3, reaction 1). 

Iron sulfide minerals have not been analyzed in the 

unsaturated-zone sediments, but they could be present 

in small quantities.

The data are not supportive of major reactions 

involving oxidation of organic carbon (reaction 4) or

Table 3. Stoichiometries of some possible redox and precipitation reactions at the Cajon 

residence, Hesperia, California

Geochemical reaction

1 . 6NO, + 2FeS2 + 2H2O -> 3N2 + 2FeOOH + 4SO4^ + 2H+

2. 14NO,- + 5FeS2 + 4H+ -> 7N2 + 10SO42 + 5Fe2+ + 2H2O

3. 2NO,- + lOFeO + 4H2O + 2H+ -> N2 + lOFeOOH

4. 4NO3- + 5(CH2O) + 4H+ -> 2N2 + 5H2CO, + 2H2O

5. 4FeS2 + 15O2 + 10H2O -> 4FeOOH + 8SO42 + I6H+

6. 2FeS2 + 7O2 + 2H2O -> 4SO42~ + 2Fe2+ + 4H+

7. 4Fe2+ + O2 + 6H2O -> 4FeOOH + 8H+

8. Fe3+ + 2H2O -> FeOOH + 3H+

Electron 

donors

Fe,S

S

Fe

C

Fe,S

S

Fe

Electron 

acceptors

N

N

N

N

O

O

O

924
ferrous silicates or oxide minerals (reaction 3). The 

estimated ratios of sulfate gain to nitrate loss between 

160 and 199 ft at the Cajon site (0.52 and 0.79, for raw 

data and chloride-normalized data, respectively) are 

similar to the ratio required by reaction 1 (0.67). The 

estimated ratios of hydrogen-ion gain to nitrate loss 

(1.1 and 1.4) are larger than that required by reaction 

1 (0.33) and could indicate additional hydrogen-ion 

contributions from reactions such as 5 to 8 (table 3), if 

mixtures of waters from diverse soil domains were 

sampled.

Denitrification generally is inhibited in the presence 

of oxygen, whose concentration in soil-moisture samples brought to the surface by nitrogen-gas pressurization of the lysimeters ranged from 120 to 220 ^M 

(3.7-7.1 mg/L). If denitrification occurred at the Cajon 

site, the presence of oxygen in the lysimeter samples 

indicates that it could be occurring in domains (possibly 

microenvironments) that are smaller than the volumes 

sampled by the lysimeters.

Denitrification by iron sulfide oxidation has been 

reported in ground waters from saturated-zone environments (Postma and others, 1991; Korom, 1992). If 

a reaction occurred at the Cajon site, it would be 

relatively deep in the unsaturated zone near the wastewater front. It is not known whether this is a stable feature controlled by the occurrence of some unique 

lithology or mineralogy localized in sediments at that 

depth, or if it is a transient feature that will move with 

the wastewater front and eventually allow nitrate to 

reach the underlying ground water. Monitoring is continuing at the Cajon residence to identify initial arrival 

of wastewater in the lysimeters at 243 and 257 ft and to 

determine if the apparent nitrate-sulfate exchange 

reaction observed near 199 ft is a stable phenomenon 

associated with the aquifer or a more transient condition associated with the advancing wastewater front.

SUMMARY

Physical and chemical evidence indicates that 

wastewater from domestic disposal systems has 

reached the water table beneath many of the older residences in the upper Mojave River Basin. However, 

any increases in ground-water nitrate concentration 

that might have resulted from the wastewater disposal 

have not yet been detected. The absence of high 

observed nitrate concentrations in the area's ground 

water may by partly explained by the failure of recent 

wastewater discharge to have reached the water table,

by dilution through vertical mixing within the saturated 

zone, and by the scarcity of monitor wells perforated 

solely near the water table. There also is some chemical 

and isotopic evidence locally for nitrification of wastewater nitrate within the unsaturated zone; however, the 

distribution of denitrification, and the capacities of the 

soils for maintaining it, are not known. Effects of 

wastewater on the aquifer could be detected by future 

monitoring in both the saturated and unsaturated zones.

REFERENCES

Bauman, B.J., and Schafer, W.M., 1985, Estimating groundwater quality impacts from on-site sewage treatment 

systems Proceedings of the Fourth National Symposium on Individual and Small Community Sewage 

Systems, American Society of Agricultural Engineers, 

New Orleans, 1984: p. 285-294.

Bottcher, J., Strebel, O., Voerkelius, S., and Schmidt, H.-L., 

1990, Using isotope fractionation of nitrate-nitrogen 

and nitrate-oxygen for evaluation of microbial denitrification in a sandy aquifer: Journal of Hydrology, v. 114, 

p. 413-424.

Hubner, H., 1986, Isotope effects of nitrogen in the soil and 

biosphere, chapter 9, in Fritz, P., and Fontes, J.Ch., 

eds., Handbook of environmental isotope geochemistry, 

v. 2, The terrestrial environment, B: New York, 

Elsevier, p. 361-425.

Korom, S.F., 1992, Natural denitrification in the saturated 

zone: a review: Water Resources Research, v. 28, no. 6, 

p. 1657-1668.

Mariotti, A., Landreau, A., and Simon, B., 1988, 15N isotope 

biogeochemistry and natural denitrification process in 

groundwater Application to the chalk aquifer of 

northern France: Geochimica et Cosmochimica Acta, v. 

52, p. 1869-1878.

Postma, D., Boesen, C., Kristiansen, H., and Larsen, F., 

1991, Nitrate reduction in an unconfined sandy aquifer Water chemistry, reduction processes, and 

geochemical modeling: Water Resources Research, v. 

27, no. 8, p. 2027-2045.

Schroeder, R.A., Setmire, J.G., and Densmore, J.N., 1991, 

Use of stable isotopes, tritium, soluble salts, and redoxsensitive elements to distinguish ground water from 

irrigation water in the Salton Sea Basin, in Ritter, W.F., 

ed., Proceedings of National Conference, Irrigation and 

Drainage Division, American Society of Civil Engineers, Honolulu, Hawaii, July 22-26, 1991: p. 524-530.

925
Smith, R.L., Howes, B.L., and Duff, J.H., 1991, Denitrification in nitrate-contaminated groundwater Occurrence 

in steep vertical geochemical gradients: Geochimica et 

Cosmochimica Acta, v. 55, p. 1815-1825.

Subsurface Surveys, Inc., 1990, Inventory of groundwater 

stored in the Mojave River basins Final consultant 

report prepared for Mojave Water Agency, Solana 

Beach, California, 47 p.

Young, M.H., Wieranga, P.J., Hills, R.G., and Vinson, J., 

1992, Evidence of piston flow in two large-scale field 

experiments, Las Cruces trench site: EOS, Transactions, American Geophysical Union, v. 73, no. 43, 

Abstract no. H11 A-11, p. 156.

926
Potential Anhydrite Precipitation Associated With Deep

Injection of Ground-Water Brine from

Paradox Valley, Colorado

By Yousif K. Kharaka1 , William C. Evans 1 , Gil Ambats 1 , and James J. Thordsen 1

Abstract

Seepage and discharge of ground-water 

brine into the Dolores River in Paradox Valley, a 

collapsed diapiric salt anticline located in southwestern Colorado, increase the dissolved-solids 

load of the Colorado River annually by about 

2.0 x 108 kilograms. In order to abate this natural 

contamination, the U.S. Bureau of Reclamation 

plans to pump about 3,540 cubic meters per day of 

brine from 12 shallow wells located along the 

Dolores River. The brine, with a salinity of 250,000 

mg/L, will be piped to the deepest (4.88 kilometers) 

injection well in the world, and injected mainly into 

the Mississippian Leadville Limestone.

Results of geochemical modeling indicate and 

water-rock experiments confirm that a huge mass 

of anhydrite, about 10,000 kilograms per day, 

likely will precipitate from this brine when heated 

to 120 °C at 500 bar the temperature and pressure 

conditions in the Leadville aquifer. Precipitation 

of anhydrite could increase by a factor of two or 

more if the injected brine were allowed to mix 

with the highly incompatible formation water of 

the Leadville aquifer and if, as expected, the Mg in 

this brine dolomitizes the calcite of the aquifer. 

Dilution of the brine with river water, precipitation 

of its SO4 and/or addition of precipitation inhibitors prior to injection are possible remedial 

actions.

INTRODUCTION

A 4.88-km-deep well, the deepest injection well in 

the world, was completed in 1987 by the U.S. Bureau 

of Reclamation for its Paradox Valley Unit. The well,

'U.S. Geological Survey; Menlo Park, Calif.

located in southwestern Colorado (fig. 1), is intended 

for the disposal of 3,540 m3/d of ground water brine for 

a period of 100 years. The brine will be pumped from 

12 shallow wells (<40 m deep) located strategically 

along the Dolores River as it traverses the topographically lowest part of the Paradox Valley. The cone of 

depression created by this pumping is expected to stop 

the natural seepages and discharges of brine into the 

Dolores River, which is a tributary of the Colorado 

River. The dissolved salt load of the Colorado River 

will be reduced annually by as much as 2.0 x 108 kg 

from this operation (Flak and Brown, 1988).

Water and gas samples were obtained for chemical 

and isotope analyses from several brine wells, from two 

freshwater wells in Paradox Valley, from the Dolores 

River and from the Leadville Limestone, the primary 

injection zone in the deep well. In this summary, we 

report the results of geochemical modeling using 

SOLMINEQ (Kharaka and others, 1988) to simulate 

injecting the brine into the Leadville aquifer. These 

results are in general agreement with those obtained 

from our water-rock interaction experiments (Rosenbauer and others, 1992) and indicate that large amounts 

of anhydrite could precipitate in the Leadville aquifer.

REGIONAL SETTING

Paradox Valley is one of five northwestwardtrending major salt anticlines located in the Canyon 

Lands of the Colorado Plateau in southwestern Colorado 

and southeastern Utah (fig. 1). Paradox Valley is about 

40-km long and 3- to 5-km wide and has a relatively 

smooth floor that slopes gradually from both ends 

toward the Dolores River, which enters at about the 

midpoint of the valley and crosses perpendicular to its 

axis. The Dolores River has cut canyons about 400-m 

deep on both the northern and the southern flanks of 

Paradox Valley. The river canyons and valley flanks are

927
109e30' 109° 108°30'

39° -

Bedrock

Injection 

well

20 MILES I

38°30' -

Figure 1 . Study area showing the location of the injection well relative to Paradox 

Valley and the path of Dolores River flowing through the brine field before empting 

into the Colorado River.

nearly vertical, exposing brilliantly colored strata of 

sandstones and shales of pre-Cretaceous age. The diapiric anticline underlying Paradox Valley has about 

4,000 m of relatively pure evaporites, mainly halite, 

and evaporite-rich shales (Flak and Brown, 1988). The 

halite has been leached from the upper 200 to 300 m to 

form the residual cap composed of highly brecciated 

gypsiferous shale. The evaporites are of Pennsylvanian 

age and belong to the Paradox Member of the Hermosa 

Formation (Cater, 1970). Quaternary alluvium deposits 

of sand, silt and gravel, less than 40-m thick, are 

present along Dolores River above the residual 

gypsiferous-shale cap.

The injection well, located about 1 km south of the 

Paradox Valley salt diapir and 5 km south of the brine 

wells, penetrates rocks that range in age from 

Precambrian to Triassic. The Leadville Limestone of 

Mississippian age, composed of 127 m of limestone 

and dolomite, is the primary injection zone. The dolomitized sections of the Leadville Limestone immediately underlying the karsted top interval have the 

greatest porosity (>5 percent) and permeability. The 

Precambrian basement rocks, composed of fractured 

granitoids and chlorite schist, are a secondary injection 

zone, although no significant amounts of fluid were 

recovered from this zone during the production tests.

928
The regional hydrology of this area is described in 

several reports, including those by Hanshaw and Hill 

(1969), U.S. Bureau of Reclamation (1978), and 

Ackerman and Rush (1984). Flow in the Dolores River 

is mainly from snow melt in the San Juan and La Plata 

Mountains and, consequently, the highest flow rates 

(as much as 142 m3/s) are from April through June; the 

minimum flow rates have been increased since 1977 to 

about 1 to 3 m3/s by discharges from the McPhee 

Reservoir (Stan Powers, U.S. Bureau of Reclamation, 

oral communication, 1993). The ground-water brine 

probably underlies the entire length of the Paradox 

Valley at varying depths. The brine surfaces along the 

Dolores River floodplain in two general areas from 

about the middle of the valley to the exit point of the 

river. A relatively freshwater layer (salinity < 1,500 mg/L) 

floats on the brine, especially on the western side of the 

river where the thickness of this layer increases to 

about 30 m at a distance of 1 km from the river.

CHEMICAL COMPOSITION OF WATER

The chemical composition of selected water and 

gas samples collected from several Paradox Valley 

brine wells using submersible water pumps are given in 

tables 1 and 2, respectively. Large quantities of fluids 

were pumped from the wells prior to sampling and, in 

several wells, water samples were collected from two 

or three depth intervals. Gas compositions are included 

in this report because loss of gases, especially CO2 and 

H2S, from the brine prior to injection will modify its 

interactions in the aquifer.

The samples from the Leadville Limestone (table 1) 

were collected by use of a down-hole sampler at the 

conclusion of a long production test. The sampler, in 

the case of sample 88PVB- 123, was placed 60 m 

above the top of the Leadville aquifer; the sample 

likely is a composite sample from all the three fractured 

zones of this aquifer. This conclusion is supported by 

the fact that the chemical composition of water in this 

sample is virtually the same as that in samples collected 

at the well head using a N2 gas lift. Because the sampler 

was placed between the two lowest fractured zones of 

the Leadville Limestone, sample 88PVB-124 likely 

represents water from the deepest fracture zone of this 

aquifer. The field and laboratory methods used for 

sample treatment, storage, and analysis are generally 

those reported by Lico and others (1982).

Subsurface temperatures calculated from the 

chemical composition of water from the Leadville

Limestone using Mg-Li, Si and Mg-corrected Na-K-Ca 

geothermometry (Kharaka and Mariner, 1989) give 

concordant results of 115 ± 5 °C. This value is in close 

agreement with measured temperatures (113-116 °C) 

in the Leadville aquifer, confirming that the samples 

were obtained from this aquifer and that the chemical 

and temperature data are reliable.

The waters from Paradox Valley wells are hypersaline brines that have salinities (250,000 mg/L) more 

than seven times that of seawater and a relatively 

uniform chemical composition. The brines are Na-Cl 

type water with high concentrations of SO4, H2S, K, 

Mg, and Ca. The waters from Leadville Limestone are 

also hypersaline brines of the Na-Cl type, but their 

salinities and SO4 concentrations are lower and the Ca 

and Sr concentrations are much higher than those of 

Paradox Valley brines (table 1). The Br:Cl ratios for all 

these waters are very low and typical of brines that 

acquire their chemical composition from dissolution of 

halite and associated evaporites (Kharaka and others, 

1987). The stable isotopes of water indicate that all the 

waters are of meteoric origin; tritium concentrations 

indicate that the brine contains modern water.

GEOCHEMICAL MODELING

Computer simulations were carried out using the 

latest version of the comprehensive geochemical code 

SOLMINEQ (Kharaka and others, 1988). This code is 

suitable for this study because it has options to compute 

the activity coefficients of solutes in brines using the 

Pitzer ion-interaction formulations (Pitzer, 1981), to 

perform water-rock interactions at temperatures and 

pressures of interest, and to investigate the consequences of mixing of two waters.

Results of these simulations show that the Paradox 

Valley brine is at saturation (±0.2 kcal/mol) with 

gypsum and calcite; and is slightly undersaturated 

(-0.5 kcal/mol) with halite at the aquifer temperature of 

about 15 °C. Because the solubilities of anhydrite (the 

stable form of calcium sulfate in the injection zone) and 

calcite decrease with increasing temperature, the brine 

becomes highly supersaturated with anhydrite and 

slightly supersaturated with calcite at 120 °C and 

500 bar, the temperature and pressure conditions in 

the Leadville aquifer.

Precipitation of a large mass of anhydrite appears 

to be the main problem associated with injecting 

3,540 m3/d of Paradox Valley brine into the Leadville 

aquifer (fig. 2). There is a general agreement between

929
Table 1. Chemical composition of Paradox Valley waters from selected brine wells, from one freshwater well, from the Dolor

the injection well

[Concentrations in milligrams per liter; mV, millivolts;TDS, total dissolved solids]

Well No. 2E 

Sample No. 88PVB-1

Solute

Li

Na

K

NH4

Mg

Ca

Sr

Ba

Mn

Fe

Cl

Br

HC03

S04

H2S

SiO2

B

pH

Eh(mV)

TDS

0.42

91,700

4,200

10.9

1,480

1,390

26.1

.12

.57

.35

150,000

85.9

313

5,940

91

7.6

8.8

6.32

255,000

9E 

88PVB-2

0.43

89,700

4,340

11.5

1,500

1,390

27.1

.07

.52

<.25

148,000

83.7

285

5,840

84

5.9

8.4

6.25

-415

251,000

12E 

88PVB-3

0.4 1

89,900

4,470

1 0.9

1,520

1,330

25.6

.05

.12

.67

147,000

86.2

258

5,920

41

6.0

8.4

6.56

-450

250.000

4E 

88PVB-7

0.43

88,500

4,200

9.2

1,500

1,380

26.4

.07

.52

<.25

148,000

83.7

295

5,920

94

8.6

8.6

6.18

250,000

8E 

88PVB-8

0.42

88,900

4,430

8.3

1,440

1,320

25.8

.05

.57

<.25

150,000

85.0

285

5,830

80

5.5

8.5

6.26

-445

253,000

10E 

88PVB-9

0.42

85,100

4,270

12.1

1,560

1,260

24.6

.07

.32

1.47

145,000

83.7

226

5,860

11

5.6

7.6

6.67

243,000

.,,,_. Freshwater 

13E well

88PVB-11 88PVB-14

0.42

94,600

4,770

13.4

1,730

1,280

24.3

.05

.32

<.25

154,000

93.5

257

5,970

98

9.3

8.2

6.32

263,000

0.01

16.4

2.36

13.0

40.3

59.3

.82

.12

<.002

.01

17

.06

334

47

15.8

.03

7.49

550

Delore

River

88PVB-1

23

2.0

36.4

52.0

<.02

<.25

12.0

122

112

7.6

7.8

340
Table 2. Chemical composition of dissolved gas and total gas pressure obtained 

from selected samples of the Paradox Valley brine and from formation water in 

the Leadville Limestone 

[Concentration in micromoles per kilogram of water; pressure in bars]

Well No. 

Sample No.

Gas

N2

O2

Ar

CH4

C2H6

CO2

H2

He

H2S

Total gas pressure

8E 

88PVB-8

115

.08

5.7

324

12.7

853

.7

1.47

10E 

88VB-9

1 43

<.05

5.7

350

9.2

223

< .5

1. 66

13E 

88VB-11

203

<.05

7.9

430

16. 1

719

67

2,154

2.49

Injection 

well

88PVB-123

3,230

223

44.2

12,800

69

7,640

42

157

<1

31.9

Injection 

well

88PVB-124

2,390

148

27.8

1 1,600

59

2,780

135

122

<1

32.3

100

60 80 100 120 140 

TEMPERATURE, IN DEGREES CELSIUS

160

Figure 2. Mass of anhydrite that may precipitate from subjecting 3,540 cubic meters per day of the 

Paradox Valley brine to difference temperatures and pressures. (Results from SOLMINEQ and 

experiments (Rosenbauer and others, 1992) are for a pressure of 600 bar. Data connected with solid lines 

were calculated from solubility parameters in Moller (1988) for a NaCI solution with an ionic strength 

equivalent to that of the Paradox brine.)

931
2.0

20 40 

PERCENT RIVER WATER

60 80 

INJECTED BRINE

Figure 3. The percent amount of Dolores River water in the injection brine required to lower the 

saturation state of anhydrite to a value below 0 kilocalories per mole in order to prevent its 

precipitation in the injection aquifer. (Results are from computer simulations at 500 bar and 

assume no interaction with minerals in the injection zone.)

the experimental results (Rosenbauer and others, 1992) 

and those computed with SOLMINEQ assuming equilibrium with anhydrite (fig. 2); at 120 °C and 500 bar, 

both indicate that about 10,000 kg/d of anhydrite may 

precipitate in the aquifer if there is no interaction with 

aquifer minerals.

Because calcite is the main mineral in the Leadville 

Formation, the concentrations of Mg in the injection 

brine are relatively high, and the thermodynamic constraints are favorable, dissolved Mg in the injection 

brine can react with this calcite to form dolomite. The 

reaction is

2CaCO3 (s) + Mg++ -> CaMg(CO3 )2 ( s) + Ca' (1)

and results in release of Ca to the water in the aquifer. 

The added Ca will react with the aqueous SO4 to form 

more anhydrite; the net or total reaction is

2CaCO3 (s) + Mg+++ SO4" -> CaMg(CO3 )2 ( s) + CaSO4(S). (2)

Assuming brine equilibrium with anhydrite, calcite 

and dolomite, computations with SOLMINEQ show 

that ultimately a total of about 25,000 kg/d of anhydrite 

could precipitate in the Leadville aquifer. Our experiments (Rosenbauer and others, 1992) show dolomitization of calcite but a somewhat lower (20,000 kg/d) total

anhydrite precipitated. This discrepancy likely is due to 

the continued disequilibrium between the solution and 

minerals in the experiments even after 80 days of reaction time (Rosenbauer and others, 1992, fig. 8), 

because of the slow kinetics of dolomitization (Morrow 

andRicketts, 1988).

Computer simulations further indicate that there 

could be as much as a 50 percent increase in anhydrite 

precipitation if the injected brine were allowed to mix 

with the formation water. This increase could result 

because the two waters are highly incompatible chemically, especially with the very high SO4 concentration 

in the brine and the very high Ca in formation water 

(table 1). Precipitation due to mixing, however, might 

be minimized by initially injecting Dolores River water 

or a mixture of brine and river water to obtain a buffer 

of compatible water between the formation water and 

the injected brine.

Computer simulation and water-rock interaction 

experiments are useful and cost-effective tools to indicate the novel chemical and hydrologic approaches 

necessary for the successful injection of these highsalinity brines. These tools, for example, show that a 

mixture of about 35 percent river water with the brine 

will not precipitate anhydrite in the absence of dolomitization (fig. 3); dilution with 65 to 70 percent river 

water would promote dolomitization (which could lead

932
to increased porosity) without anhydrite precipitation. 

Simulations and experiments further indicate that other 

remedial actions to prevent precipitation of anhydrite 

could include precipitation of SO4 as gypsum or barite 

at ground level and (or) addition of precipitation inhibitors, such as phosphonates, to the brine prior to injection (Matty and Tomson, 1988).

REFERENCES

Ackerman, D.J., and Rush, F.E., 1984, Hydrogeologic 

reconnissance of the San Miguel River basin, southwestern Colorado: U.S. Geological Survey WaterResources Investigation Report 84- 4133, 25 p.

Cater, F.W., 1970, Geology of the salt anticline region in 

southwestern Colorado: U.S. Geological Survey 

Professional Paper 637, 80 p.

Flak, L.H., and Brown, John, 1988, Case history of ultra 

deep disposal well in western Colorado. IADC/SPE 

1988 Drilling Conference, Houston: SPE Paper 17222, 

24 p.

Hanshaw, B.B., and Hill G.A., 1969, Geochemistry and 

hydrodynamics of the Paradox Basin region, Utah, 

Colorado and New Mexico: Chemical Geology, v. 4, 

p. 263-294.

Kharaka, Y.K, and Mariner R.H., 1989, Chemical geothermometers and their application to formation waters 

from sedimentary basins, in Naeser, N.D., and 

McCulloh, T.H., eds., Thermal history of sedimentary 

basins methods and case histories: Springer-Verlag, 

no. 6, p. 99-117.

Kharaka, Y.K., Gunter, W.D., Aggarwal, P.K., Perkins, 

E.H., and DeBraal, J.D., 1988, SOLMINEQ.88 

A computer program for geochemical modeling of

water-rock interactions: U.S. Geological Survey Water 

Resources Investigation Report 88- 4227,420 p.

Kharaka, Y.K., Maest, A.S., Carothers, W.W., Law, L.M., 

Lamothe, P.J., and Fries, T.L., 1987, Geochemistry of 

metal-rich brines from central Mississippi Salt Dome 

basin, U.S.A.: Applied Geochemistry, v. 2, p. 543-561.

Lico, M.S., Kharaka, Y.K., Carothers, W.W., and Wright, 

V.A., 1982, Methods for collection and analysis of 

geopressured geothermal and oil-field waters: U.S. 

Geological Survey Water-Supply Paper 2194, 21 p.

Matty, J.M., and Tomson, M.B., 1988, Effect of multiple 

precipitation inhibitors on calcium carbonate nucleation: Applied Geochemistry, v. 3, p. 549-556.

Moller, Nancy, 1988, The prediction of mineral solubilities 

is natural waters: a chemical equilibrium model for the 

Na- Ca-Cl-SO4-H2O system, to high temperatures and 

concentrations: Geochimica Cosmochimica Acta, v. 52, 

p. 821-837.

Morrow, D.W., and Ricketts, B.D., 1988, Experimental 

investigation of sulfate inhibition of dolommite and its 

mineral analogues, in Sedimentology and geochemistry 

of dolostones: SEPM Special Publication No. 43, 

p. 25-38.

Pitzer, K.S., 1981, Characteristics of very concentrated 

aqueous solutions, in Rickard, D.T., and Wickman, 

F.E., eds.. Chemistry and geochemistry of solutions at 

high temperatures and pressures: Physics and Chemistry of the Earth, v. 13-14, p. 249-272.

Rosenbauer, R.J., Bischoff, J.L., and Kharaka, Y.K., 1992, 

Geochemical effects of deep-well injection of the Paradox Valley brine into Paleozoic carbonate rock, Colorado, U.S.A.: Applied Geochemistry, v. 7, p. 273-286.

U.S. Bureau of Reclamation, 1978, Paradox Valley Unit 

definite plan report, Appendix B, Hydrosalinity: U.S. 

Bureau of Reclamation, 173 p.

933
934
Apparent Solute Dispersion in an Alluvial/Aquifer

Stream System

By J. Jeffrey Starn 1

Abstract

In simulations of solute transport, the dispersion coefficient characterizes the spreading and 

mixing of the solute due to variations in groundwater flow velocity that are not explicitly included 

in the simulation. Although usually considered to 

be the result of variability in hydraulic conductivity, dispersion also can result from variations in the 

direction of ground-water flow. A computer simulation of ground-water flow and solute transport in 

an alluvial aquifer/stream system shows that fluctuations in river stage and subsequent changes in 

the direction of ground-water flow in the aquifer 

cause an increase in apparent dispersion. Solute 

plumes in an alluvial aquifer near a river move 

more slowly and become more dilute as a result of 

ground-water/surface-water interaction than if this 

interaction were not considered.

INTRODUCTION

Theoretical developments and carefully controlled 

field experiments have greatly increased knowledge of 

the way solutes disperse and mix with uncontaminated 

ground water. Most ground-water contamination does 

not take place under carefully controlled conditions; 

however, there is still a need to predict the distribution 

of solute concentrations at sites of ground-water contamination. Environmental regulations often mandate 

the reduction of levels of contamination in an aquifer 

by specifying a target concentration that a remedial 

action must achieve. To select an appropriate remedial 

action, the degree to which the target concentrations 

will be achieved must be predicted, commonly using 

sparse data and simplistic models.

'U.S. Geological Survey, Louisville, Ky.

The amount of spreading and mixing a solute 

undergoes during advection (a process termed dispersion) must be specified correctly to successfully 

predict solute concentrations. The process of dispersion lowers the mean concentration of a solute in an 

aquifer and increases the area affected by the solute. 

In the commonly used advection/dispersion equation, 

dispersion is often assumed to be Fickian; that is, a 

constant dispersivity parameter can be used to predict 

solute mixing. The dispersivity parameter describes 

mixing due to variations in ground-water velocity at 

the microscopic and pore scale and is used to account 

for the lack of detailed information about solute 

mixing at this scale (Wheatcraft and Tyler, 1988). 

At ground-water contamination sites, however, there 

commonly is a lack of detailed information about the 

hydraulic-conductivity distribution, and the dispersivity parameter must also account for this variability. 

Field-scale dispersivity is constant only when a solute 

has passed through a sufficient volume of aquifer such 

that the mean hydraulic conductivity is constant.

Field-scale dispersion is usually considered to be 

the result of the heterogeneous nature of geologic 

materials (Garabedian and others, 1991). Several 

recent field studies have concentrated on the collection 

of detailed information about the distribution of aquifer properties (Mackay and others, 1986; LeBlanc and 

others, 1991; Rehfeldt and Gelhar, 1992). This type of 

investigation is impractical for most instances of 

ground-water contamination. The dispersivity parameter must, in fact, account for all variability at a site 

that is not explicitly known. Although not usually 

considered, this variability includes changes in the 

direction of the hydraulic gradient over time as well as 

geologic variability. Dispersion usually connotes only 

mixing caused by hydraulic properties of the aquifer; 

therefore, the combined effects of aquifer heterogeneity and unsteady ground-water flow is termed apparent 

dispersion (Goode and Konikow, 1990).

935
For an example of apparent dispersion, consider 

an alluvial-aquifer/stream system. Areas of industrial 

development commonly are located near rivers. This 

juxtaposition has resulted in ground-water contamination in many alluvial aquifers near surface streams. 

Ground water generally flows toward major streams, 

especially where these streams fully penetrate alluvial 

aquifers. When stream stage is high, the direction of 

ground-water flow near the stream can be reversed. The 

distance from the stream that water levels are affected 

by this reversal and the magnitude and duration of the 

water-level change produced in the aquifer are a function of the hydraulic properties of the aquifer and of the 

magnitude and duration of the fluctuation in stream 

stage. The change in direction of ground-water flow 

near the stream causes an apparent dispersion.

The purpose of this paper is not to determine the 

exact nature of dispersivity: this subject is the topic of 

a large amount of research currently taking place 

throughout academia and the U.S. Geological Survey. 

The purpose of this paper is to report the results of a 

preliminary study to simulate solute transport in an 

alluvial-aquifer/stream system in which the stream 

stage fluctuates periodically. The effect of different 

assumptions used in the model on apparent dispersion 

are compared by an objective measure of solute 

mixing. The preliminary results of this study will provide investigators with a qualitative knowledge of the 

role of stream-stage fluctuation in the spreading and 

dilution of contaminants in an alluvial-aquifer/stream 

system.

METHOD

Ground-water flow and solute transport are simulated by use of a three-dimensional, finite difference, 

method-of-characteristics computer code (Goode and 

Konikow, 1991). This code links a commonly used 

ground-water flow model code (McDonald and Harbaugh, 1988) with a method-of-characteristics (MOC) 

solute-transport model. The governing equation and 

definition of parameters for the MOC model are identical to those used by Konikow and Bredehoeft (1978). 

The ground-water-flow part of the code is used to generate a velocity field. The MOC modules track particles 

through this velocity field. Each particle is assigned a 

concentration based only on initial conditions. The 

dispersion process is simulated by a finite-difference 

approximation at the end of each time step that 

redistributes particle concentration as a function of

the velocity and the dispersivity. The concentration at 

each node of the finite-difference grid is calculated 

from the particles within the cell block at the end of the 

dispersion process.

The conceptual model of ground-water flow is identical to that described by an analytical equation (Finder 

and others, 1969). The equation represents a simplified 

version of the aquifer/stream system in which the aquifer 

properties are homogeneous, the aquifer has a finite 

width, and there is an instantaneous rise in river stage 

followed by a subsequent, delayed rise in ground-water 

stage. The analytical solution was used to check that the 

finite-difference model was specified correctly.

The boundaries of the model domain and the aquifer 

properties used in the model were selected so as to be 

similar to conditions in alluvial aquifers along the Ohio 

River and some of its larger tributaries. In the hypothetical aquifer, the bedrock valley wall is 2,000 ft from the 

stream. The aquifer is 50 ft thick and is composed 

mainly of sand with varying amounts of gravel. The sand 

is overlain and confined by a fine-grained unit composed 

of lacustrine clays and silts. Rises in river stage that are 

the result of discharges from dams upstream can be 

approximated by an instantaneous change in stage.

The finite-difference grid of the mathematical 

model is composed of 50 by 50 ft cells. The groundwater flow model grid is 40 columns by 20 rows. Solutetransport simulations are performed on a subarea 18 

columns by 10 rows. The hydraulic conductivity used 

in the simulations is 60 ft/d, the storage coefficient is 

0.01, and the porosity is 20 percent. The edge of the 

bedrock valley wall at the right-hand side of the model 

grid is modeled as a no-flow boundary to be consistent 

with the analytical model. River stage at the left-hand 

edge of the model grid is modeled as a constant-head 

boundary. The top and bottom edges are no-flow 

boundaries. Stage in the river fluctuates from 0 to 25 ft 

every 5 days; 40 time steps in each of 8 stress periods 

are simulated. Solute concentrations at the start of the 

simulation are zero everywhere except for a 4 by 4 cell 

block in the which the concentration is 1.0 ppm of 

solute per cell. The solute is assumed to be conservative, 

that is, not affected by retardation, decay, biodegradation, or any of the other processes that act to reduce the 

mass of a contaminant in ground water.

In this simulation, vertical flow is not considered. 

In many places, the Ohio River fully penetrates the 

aquifer and vertical flow would occur during recharge 

from precipitation on the surface of the alluvial deposits; 

however, the amount of recharge to the aquifer can be

936
LU

Q 

<

LU

30

25

20

DC 

UJ

5 15

o

10

5

0

0

®-

500 1,000 1,500 2,000 

DISTANCE IN X DIRECTION, IN FEET

EXPLANATION 

Head in aquifer at end of stress period 7, river stage = 25 feet

 Head in aquifer at end of stress period 8, river stage = 0 feet

Figure 1. Section along model grid row 5 showing hydraulic-head fluctuation.

limited by fine-grained lacustrine deposits that have a 

much lower hydraulic conductivity than the aquifer. In 

this theoretical study, therefore, the flowfield is simulated in only two dimensions in a horizontal plane. 

Three-dimensional flow should be considered in the 

area near the stream; however, this is beyond the scope 

of this report.

The solute distribution is quantified at the end of 

each stress period by calculating its spatial moments. 

The zero, first, and second central moments describe, 

respectively, the mass of the solute in the system, the 

displacement of the center of the solute mass with 

respect to the origin, and the spatial covariance (a2) of 

the solute mass (Fischer and others, 1979). The mass of 

contaminant in each cell is found by multiplying the 

concentration of contaminant in a cell by the volume of 

water stored in that cell. This method allows for the 

effect of fluid storage transients (Goode, 1990). The 

finite-difference model uses the dispersion coefficient 

(D), which is related to the dispersivity parameter (a) 

by Z)=a|F|, where Vis the magnitude of the velocity. 

The dispersivity can be obtained from the rate of 

change of the variance with respect to time

or distance (x) by a = rr^^a2 and by a = r-r-a2 . The ^ 2\y\dt 2dx

dispersivity tensor is represented in two principal

directions: longitudinal, along the jc axis in the direction of flow; and transverse, along the y axis perpendicular to flow.

PRELIMINARY RESULTS

The hydraulic head distribution at the end of two 

stress periods is shown on figure 1. Line A is after an 

instantaneous rise in river stage and line B is after an 

instantaneous drop in river stage. The reversal of flow 

is apparent from the graph; however, the graph does not 

show that the change in river stage propagates a wave 

through the aquifer that causes a flow divide to move 

ahead of the wave. The rate of change of the variance 

of the plume over time is shown on figure 2. The final 

position and extent of the plume in relation to its origin 

is shown on figure 3.

The slopes of graphs of various x-axis and y-axis 

variants are listed in table 1. The slope is the rate of 

change of the variants. In the cases where the slope is 

determined with respect to distance, the slope is calculated from the data at every second stress period; otherwise, the oscillatory movement of the plume caused by 

the reversal of flow direction results in a slope with a 

low r2 . In this manner, the calculation of dispersivity is

937
16,000

Q 14,000

LU 

DC

§12,000 

O

10,000

LU 

U_

? 8,000

LU 

O

6,000

oc

> 4,000

2,000, 5 10 15 20 25 30 35 40 

TIME, IN DAYS

EXPLANATION

® Variance of solute plume in x direction 

© Variance of solute plume in y direction

Figure 2. Change in variance in the x and y directions.

based on the concentration distribution after each complete cycle of flow reversal. The dispersivities and 

ratios of longitudinal to transverse dispersivity calculated from the rate of change of the variance and the 

velocity are listed in table 2. The true dispersivities 

used in the simulation are 50 ft longitudinal and 5 ft 

transverse.

Table 1. Parameters used to determine dispersivity

Table 2. Dispersivity calculated from model results

Parameter Slope

distance over time

x-variance over distance

y-variance over distance

x-variance over time

y-variance over time

O.I 37

1,870

183

257

25.1

0.999

.999

.999

.999

.999

Direction Dispersivity Dispersivity 

with time with distance

longitudinal (1)

transverse (t)

1/t

937

91.5

10.2

937

91.6

10.2

CONCLUSIONS

The center of the plume has not moved in a 

constant direction because of the reversal in the direction of ground-water flow. The slope of distance with 

respect to time represents the apparent velocity of the 

center of the plume and not the actual velocities within 

the plume. There is apparent net movement toward the 

river because the gradients at the leading edge of the 

plume are steeper than at the trailing edge; this, in 

combination with the nonreversible spreading contrib938
(TJ 

T3

3 

O 

JQ

T3

CO 

0>

cd 

to

o 

O

No-flow boundary 

EXPLANATION

Model grid cell, 50 ft by 50 ft

Initial solute concentration, 1.0 milligram per liter per cell

Simulated line of equal concentration, in milligrams per liter. 

Contour interval 0.05 milligrams per liter.

Figure 3. Lines of equal concentration at the final position of the plume, 40 days after the start of simulation.

uted by the dispersion term, leads to more spreading at 

the front edge of the plume. The increased spreading 

shifts the calculated center of mass toward the stream.

As the plume moves back and forth, the volume of 

water it occupies increases and the concentrations 

within the plume decrease. Without consideration of 

the unsteady-velocity field, the plume could not be 

simulated with the true hydraulic gradient and (or) 

hydraulic conductivity, because the simulated center of 

the plume would not correspond with its actual position. If the hydraulic parameters were adjusted to 

match the center of the plume and the dispersivity were 

adjusted to match the spreading of the plume, the dispersivity would be too high. The use of such a dispersivity in a model could lead to gross errors in prediction 

if the velocity field were to change, as would be the 

case if a pump and treat system were being designed to 

collect contaminated ground water. Although the magnitude of the dispersivity calculated from the simulation is too large, the ratio of longitudinal to transverse 

dispersivity is very close to the true value.

The flow system proposed in this study is simplistic. 

The effect of the conceptual model is to exaggerate the 

apparent dispersion caused by fluctuations in river 

stage; nonetheless, the study shows the nature of the 

effect of flow reversals in an aquifer near a river. In a 

real aquifer, fluctuations in stage would probably cause 

changes in flow direction that were less than 180°, as 

was simulated in this study. Changes in direction less 

than 180° would probably increase the transverse dispersion relative to the longitudinal dispersion.

The main result of this study is to show that a 

contaminant plume in an alluvial/aquifer stream 

system near a river moves more slowly and becomes 

more dilute as a result of ground-water/surface-water 

interaction than if steady flow were assumed. The 

implication of this for the remediation of ground-water 

contamination is that simulations based on matching 

the center of the plume and the dispersion assuming a 

steady flow system can be in error. Efforts to remediate 

contaminated ground water should consider the effects 

of unsteady flow.

939
REFERENCES

Fischer, H.B., List, E.J., Koh, R.C.Y., Imberger, Jorg, and 

Brooks, N.H., 1979, Mixing in inland and coastal 

waters: New York, Academic Press, p. 30-54.

Garabedian, S.P., LeBlanc, D.R., Gelhar, L.W., and Celia, 

M.A., 1991, Large-scale natural gradient tracer test in 

sand and gravel, Cape Cod, Massachusetts, 2 Analysis of spatial moments for a non-reactive tracer: Water 

Resources Research, v. 27, no. 5, p. 911-924.

Goode, D.J., 1990, Governing equations and model approximation errors associated with the effects of fluidstorage transients on solute transport in aquifers: U.S. 

Geological Survey Water-Resources-Investigations 

Report 90-4156, 20 p.

Goode, D.J., and Konikow, L.F., 1990, Apparent dispersion 

in transient ground-water flow: Water Resources 

Research, v. 26, no. 10, p. 2339-2351.

___1991, Testing a method-of-characteristics model of 

three-dimensional solute transport in ground water: 

Proceedings of the Symposium on Ground Water, 

American Society of Civil Engineers, Hydraulics 

Division, Nashville, Tennessee: p. 21-27.

Konikow, L.F., and Bredehoeft, J.D., 1978, Computer 

model of two-dimensional solute transport and dispersion in ground water: U.S. Geological Survey Techniques of Water-Resources Investigations, book 7, 

chap. C2, 90 p.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W., 

Quadri, R.D., Stollenwerk, K.G., and Wood, W.W, 

1991, Large-scale natural gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 Experimental design and observed tracer movement: Water 

Resources Research, v. 27, no. 5, p. 895-910.

Mackay, D.M., Freyberg, D.L., Roberts, P.V., and Cherry, 

J.A., 1986, A natural gradient experiment on solute 

transport in a sand aquifer, 1 Approach and overview 

of plume movement: Water Resources Research, v. 22, 

no. 13, p. 2017-2029.

McDonald, M.G. and Harbaugh, A.W, 1988, A modular 

three-dimensional finite-difference ground-water flow 

model: U.S. Geological Survey Techniques of WaterResources Investigations, book 6, chap. Al, 528 p.

Pinder, G.F., Bredehoeft, J.D., and Cooper, H.H., Jr., 1969, 

Determination of aquifer diffusivity from aquifer 

response to fluctuations in river stage: Water Resources 

Research, v. 5, no. 4, p. 850-855.

Rehfeldt, K.R., and Gelhar, L.W., 1992, Stochastic analysis 

of dispersion in unsteady flow in heterogeneous aquifers: Water Resources Research, v. 28, no. 8, p. 2085- 

2099.

Wheatcraft, S.W, and Tyler, S.W 1988, An explanation of 

scale-dependent dispersivity in heterogeneous aquifers 

using concepts of fractal geometry: Water Resources 

Research, v. 24, no. 4, p. 566-578.

940
Measurements of Hydraulic Conductivity and Capillary 

Pressure under Unsaturated Conditions in a Laboratory

Triaxial System

By Harold W. Olsen 1

Abstract

A new laboratory technique is being developed to provide media-specific measurements of 

hydraulic conductivity and capillary pressure 

under unsaturated conditions for modeling 

ground-water flow and contaminant transport in 

geologic systems. The technique is designed to 

test undisturbed specimens in a triaxial chamber 

under simulated in situ stress conditions and to 

provide direct and concurrent measurements of 

vertical hydraulic conductivity, capillary pressure, 

and the variation of these properties with degree 

of saturation. Constant-flow hydraulic-conductivity measurements are conducted by infusing and 

withdrawing identical flow rates at opposite ends 

of a specimen with a flow pump and measuring 

the resultant induced head difference with a differential pressure transducer. Additional transducers and flow pumps are used to (1) measure the 

capillary pressure and effective stress, (2) control 

and change the water content of the specimen during and between hydraulic conductivity measurements, and (3) control the specimen volume. This 

new technique avoids fundamental limitations of 

the methods commonly used in current practice, 

including (1) the need for duplicate specimens 

and separate equipment for laboratory measurements of hydraulic conductivity under saturated 

conditions and soil water retention, (2) the lack of 

stress or volume control for simulating in situ 

conditions in the pressure plate or tempe cell

'U.S. Geological Survey, GD, Golden, CO.

equipment generally used to meas ure soil-waterretention curves, and (3) the need for pore-sizedistribution models to estimate the relationship 

between hydraulic conductivity and degree of saturation from soil-water-retention data.

INTRODUCTION

The variations of hydraulic conductivity (relative permeability) and capillary pressure (soil-water 

retention) with degree of saturation are functions 

needed to model ground-water flow and contaminant 

transport in geologic systems. In current practice, 

these functions for specific materials are commonly 

estimated from laboratory measurements of vertical 

hydraulic conductivity under saturated conditions 

together with soil-water-retention curves obtained 

with traditional pressure plates and tempe cells, and 

from theoretical pore-size-distribution models. This 

approach is very time consuming. It also has limitations that introduce uncertainties in the significance of 

the results. In a pressure plate or a tempe cell, an 

unconfined test specimen is mounted on a high-airentry porous plate inside a chamber that can be pressurized; the porous plate is connected to a flow-ratemeasuring device, such as a capillary tube or beaker, 

outside the chamber. A soil- water-retention curve is 

obtained by elevating the chamber air pressure in steps 

and measuring the increments of pore fluid expelled 

through the high-air-entry plate and into the external 

flow-rate-measuring device. The test is time consuming because a substantial time interval is needed to 

expel the pore fluid following each loading step. The 

specimen cannot be maintained under simulated in situ 

stress conditions because it is unconfined. Finally, the 

limitations of the pore-size-distribution models used to

941
estimate relative permeabilities from soil-water-rentention curves are not well understood.

This paper describes and illustrates the capabilities of a new laboratory technique for concurrent measurements of relative permeability and soil- waterretention functions on an undisturbed and unsaturated 

specimen that is maintained under simulated in situ 

stresses in a triaxial cell. This technique is an extension of recent innovations using triaxial systems for 

measuring vertical hydraulic conductivity under saturated conditions (Olsen and others, 1985, 1988,1991), 

and the variation of capillary pressure with degree of 

saturation under unsaturated conditions (Znidarcic and 

others, 1991).

EXPERIMENTAL SYSTEM

A diagram of the experimental system is shown 

in figure 1. A test specimen [S] is mounted in a triaxial 

cell and encased in a latex membrane between 5-bar, 

high-air-entry, ceramic discs in the pedestal and top 

cap (The number "5-bar" designates the capillarity 

characteristics of the ceramic disc in terms of the air 

pressure needed to expel water from the disc). The 

lower ceramic disc is sealed with an O-ring in the 

base pedestal. The upper ceramic disc has a center 

hole and is sealed with two O-rings in a doughnutshaped cavity in the top cap. A hole through the center 

of the top cap allows control of pore-air pressure in the 

specimen by an external pressure regulator [PR]. Constant-flow vertical hydraulic-conductivity measurements are performed on an unsaturated specimen by 

infusing and withdrawing identical flow rates at opposite ends of the specimen with a flow pump equipped 

with an infuse-withdraw actuator [P]. The water content of the specimen is varied by using a second flow 

pump [ W] to infuse or withdraw water at the base of 

the specimen. In situ stresses on the specimen can be 

simulated by use of a differential pressure regulator 

[DPR] and a bellowfram [B] to maintain a constant 

pressure difference between the hydraulic chamber 

pressure and the pore-air pressure. Alternatively, the 

specimen volume can be controlled by infusing or 

withdrawing water into/from the triaxial chamber with 

a third flow pump [V] (to the extent limited by compliance of the system). One side of each of three differential pressure transducers [M,N, and Q] is connected to 

one of the pore-water lines feeding into the base pedestal, where the water pressure below the 5-bar 

ceramic disc is monitored at the base of the specimen

(Pb). The other sides of these transducers monitor (I) 

the water pressure in the top cap above the 5-bar 

ceramic disc on top of the specimen [transducer M], 

(2) the air pressure in the center hole of the top cap 

that connects with the pore-air pressure at the top of 

the specimen [transducer Q], and (3) the chamber 

pressure [transducer N].

EXPERIMENTAL DATA

The capabilities of this system have been tested 

on an "undis turbed" sample of well-graded silty sand, 

14.5 percent of which is finer than a #200 sieve, from a 

ground-water-recharge research site 5 km north of 

Golden, Colo. (Kiusalaas, 1992). Prior to mounting 

the specimen, the hydraulic components of the laboratory apparatus were saturated with deaired and distilled water. After mounting, the specimen was 

saturated by first evacuating the air in its pore space 

and then allowing deaired and distilled water to flow 

upwards through the specimen. The pore-air line at the 

top of the specimen was then opened to the atmosphere, and the differential pressure regulator was set 

to maintain the hydraulic pressure in the chamber fluid 

69 kPa (10 lb/in2) greater than the pressure in the poreair line.

The time history of a test on the above specimen 

is illustrated in figures 2 and 3. APt.b is the difference 

between the water pressure above the ceramic disc in 

the top cap and the water pressure below the ceramic 

disc in the base pedestal. The response of this pressure 

difference to an externally generated flow of water 

through the specimen arises from resistance to water 

flow, both through the specimen and through the 

ceramic discs above and below the specimen. APa.b is 

the difference between the pore-air pressure at the top 

of the specimen and the water pressure below the 

porous stone in the base pedestal. When there is no 

flow through the system and hence no head loss 

through the ceramic discs, this pressure differencethe capillary pressure, Pc~equals the pressure difference between the pore-air pressure and the pore-water 

pressure within the specimen. Q is the externally 

imposed flow rate used for making hydraulic-conductivity (k) measurements; w is the water content of the 

specimen in percent dry soil weight; 8Vw designates a 

volumetric increment of water extracted from the 

specimen; and Pa is the pore air pressure within the 

specimen.

942
Air Pressure

H PR

H DPR

B

'o' 5' M Q

N

i\ x xi

IS S S\

Figure 1. Schematic diagram of a triaxial system for vertical hydraulicconductivity and capillary-pressure measurements on an undisturbed 

and unsaturated specimen (S). One flow pump equipped with an infusewithdraw actuator (P) transmits identical water flow rates to and from 

opposite ends of the specimen. A second flow pump equipped with a 

conventional syringe (W) controls the water content of the specimen. A 

third flow pump equipped with a conventional syringe (V) provides a 

means to maintain the specimen under volume control. Alternatively, 

the specimen can be maintained under stress control using a 

differential pressure regulator (DPR) and a bellowfram (B). Three 

transducers (M,N,&Q) monitor differences between (1) pore pressures 

above and below the specimen (M), (2) pore air and pore-water 

pressures above and below the specimen (Q), and (3) the chamber 

pressure and the pore-water pressure below the specimen.

943
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k = 2.6 &3.0 x10' 8 cm/s

50 100 150 200 

TIME, IN HOURS

250 300

J] V W = 18.8%

P a = 518 kPa B

AP a-b

Q = 1.48 x 1 O' 5 cm3/s Q = 1.48 x 1 O* 5 cm3 /s :

: k = 2.3 x1 0' 8 cm/s k = 1.2 x 10' 8 cm/s

350 375 400 425 

TIME, IN HOURS450

Figure 2. Data obtained on a specimen mounted in the triaxial system from 0 to 300 

hours (A) and from 350 to 450 hours (B) following the beginning of the 

test. Pa is the pore-air pressure, w is the water content in percent dry 

soil weight. APt-b is the difference between the pore-water pressures 

above and below the specimen. AP a _b is the difference between the 

pore-water pressure below, and the pore-air pressure above the 

specimen. 5V w designates an increment of pore water extracted from 

the specimen. Q is the flow rate used to conduct a hydraulicconductivity measurement, and k is the hydraulic conductivity.

944
Figure 2A shows changes in pressure differences (APt_b and APa_b) during the first 300 hours 

when the pore-air pressure (P^ was maintained at atm 

ospheric pressure. About halfway through this period, 

the water content of the specimen was reduced from 

22.4 to 20.0 percent by extracting 2.69 cm of pore 

water from the specimen. Before and after this change 

in water content, hydraulic conductivity measurements were conducted by generating a flow rate of 

1.48 x 10"5 cm3/s through the specimen with a singlecarriage flow pump equipped with an infuse-withdraw 

actuator. The hydraulic-conductivity values shown 

were calculated from the magnitudes of APt_b induced 

by the externally imposed flow rate. Because APt_b 

was measured above and below the ceramic discs in 

the top cap and base pedestal, respectively, it includes 

components of head loss in both the specimen and the 

ceramic discs. Note that the hydraulic conductivity did 

not change appreciably after reducing the water content, apparently because the magnitude of APt_b was 

dominated by resistance to flow through the ceramic 

discs. Low hydraulic conductivity of the ceramic discs 

is also indicated in the induced response of APa_b that 

parallels the induced APt_b during the hydraulicconductivity measurements. Between hydraulic-conductivity measurements, when there is no head loss in 

the ceramic discs from an externally imposed flow 

rate, APa.b equals the capillary pressure, Pc . Note that 

Pc increased from nearly zero to about 6 kPa after 

reducing the water content from 22.4 to 20.0 percent. 

Finally, note that the response times for these hydraulic conductivity measurements are substantial on the 

order of 20 to 30 hours.

Figure 2B shows changes in pressure differences 

(APt.b and APa.b) during the period from 350 to 450 

hours after the pore-air pressure was raised from 

0 to 518 kPa while the water content was maintained 

constant at 20.0 percent. The first part of this time interval shows a sequence of three changes in the pressure 

differences resulting from the same externally-imposed 

flow rate used for the hydraulic conductivity measurements between 0 and 300 hours. Comparison of figures 

2A and 2B shows that the increase in pore air pressure 

caused an increase in capillary pressure from 6 to 18 

kPa and a decrease in the hydraulic conductivity from 

3.0 x 10~8 to 2.3 x 10"8 cm/s. In addition, the increase in 

pore-air pressure caused a substantial decrease in the 

response time for hydraulic-conductivity measurem 

ents, from more than 20 hours to less than 1 hour. 

Beginning at about 375 hours, the water content was

reduced from 20 to 18.8 percent by extracting 

1.38 cm3 of pore water. Thereafter, a hydraulic-conductivity test was run with the same externallyimposed flow rate used previously. This reduction in 

water content increased the capillary pressure from 

about 18 kPa to about 30 kPa, and reduced the hydraulic conductivity from 2.0 x 10"8 to 1.2 x 10"8 cm/s.

Figure 3A shows the period from 450 to 550 

hours when the water content w as decreased from 

18.8 to 18.2 percent. This change increased the capillary pressure to about 40 kPa and reduced the hydraulic conductivity to 4.6 x 10"9 cm/s. Figure 3B shows 

the period from 500 to 900 hours when a decrease in 

water content from 18.2 to 16.8 percent increased the 

capillary pressure to about 90 kPa, and reduced the 

hydraulic conductivity to 6.1 x 10"'° cm/s. Comparison of the response times for hydraulic-conductivity 

measurements in figures 2 and 3 shows that test 

durations became substantially longer as hydraulic 

conductivities decreased.

Figures 4A and 4B summarize the variations of 

hydraulic conductivity and capillary potential that 

were obtained concurrently at successively lower 

water contents. The nearly constant hydraulic conductivity for water contents above 20 percent probably 

reflects an upper limit imposed by the hydraulic conductivity of the 5-bar ceramic discs. Figure 4C shows 

how the response time varies with pore-air pressure 

and hydraulic conductivity.

The data do not include estimates for the degree 

of saturation of the specimen because changes in the 

total volume of the specimen associated with changes 

in its water content were not measured. Such volume 

changes tend to occur in stress-controlled specimens 

because the effective stress varies with the capillary 

pressure. By assuming that these volume changes were 

negligible in the above tests, approximate degrees of 

saturation can be estimated that range from 100 percent to 75 percent; these percentages correspond to 

water contents of 22.4 to 16.8 percent, respectively.

SUMMARY AND DISCUSSION

The data presented in the paper demonstrate the 

feasibility of concurrent measurements of constantflow hydraulic-conductivity and capillary pressure. 

The results depict the variation in these properties as a 

function of the degree of saturation for undisturbed 

and unsaturated soil specimens mounted in triaxial 

cells.

945
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RESPONSE TIME, IN HOURS

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Figure 4. Variation of hydraulic conductivity with water content (A). Variation of 

capillary pressure with water content (B). Variation of response time 

with pore air pressure and hydraulic conductivity (C). Pa is pore-air 

pressure in kilopascals (kPa).

947
The lower limit for hydraulic-conductivity measurements on unsaturated specimens is very low, in the 

range of 10 cm/s. This limit is imposed by complications in the characteristics of APt.b that first arose 

between 395 and 410 hours (fig. 2B) and became more 

pronounced after 490 hours (fig. 3). During the 

hydraulic conductivity test between 395 and 410 

hours, APt_b continued creeping upward rather than 

reaching a constant steady-state value. During the 

hydraulic conductivity tests between 490 and 505 

hours, APt_b initially climbed at a rate that did not 

decrease with time and approach a steady state; however, a steady state was reached by reducing the externally-imposed flow rate by a factor of 5, from 14.8 to 

2.95 x 10"6 cm3 /s. This characteristic is depicted again 

for the hydraulic conductivity test between 630 and 

720 hours. APt_b initially climbed at a rate that did not 

decrease with time toward a steady state; however, an 

approximate steady state was reached by reducing the 

externally imposed flow rate by a factor of 5~from 

2.95 x 10"6 to 5.90 x 10"7 cm3/s. Another complication 

in the characteristic of APt_b follows the hydraulicconductivity test between 630 and 720 hours and the 

hydraulic-conductivity test between 810 and 860 

hours. APt_b does not return to zero after terminating 

the externally imposed flow rate. A pressure difference 

on the order of 10 kPa persists across the specimen.

The current upper limit for hydraulic-conductivity measurements is relatively low because of the component of head loss from the ceramic discs in the 

measurement of hydraulic gradients induced by externally imposed flow rates through the specimen. The 

following approaches for minimizing or eliminating this 

component of head loss are currently being explored:

(1) Locate and seal the sensors for measuring hydraulic 

heads within the ceramic discs, and do so as close as 

possible to the disc surfaces adjacent to the specimen;

(2) replace the l/4"-thick high-air-entry ceramic discs at 

the ends of the specimen with very thin high-air-entry 

membranes supported by high-k porous discs; and (3) 

mount narrow strips of high-air-entry porous fabric 

along the cylindrical surface of the specimen, connect 

these strips to access ports in the top cap and base pedestal, and use these strips as sensors for the hydraulic 

heads near the ends of the specimen. Because the feasibility of these approaches appears to be justified by 

somewhat similar applications of high-air-entry porous 

materials in the 1 iterature (for example, Lenhard and 

Parker, 1988; Oak and others, 1988; Black and Miller; 

1990 and Roseberg and McCoy, 1990), we anticipate 

that the upper bound for constant-flow k measurements 

can be extended to at least 10"4 or 10~3 cm/s.

To obtain reliable values for the degree of saturation, in addition to values for water content (in percent of dry soil weight), precise control is needed for 

the total volume of the specimen. In the system 

described, the specimen volume is controlled by transmitting water to or from the triaxial chamber with a 

flow pump. The adequacy of this approach is being 

examined by evaluating the compliance of the triaxial 

cell and chamber fluid in response to variations in the 

triaxial chamber pressure and the degree of saturation 

of the specimen. If needed, the system will be 

upgraded using a more rigid test cell or a dual-cell 

chamber equipped with an electro-pneumatic regulator 

on the outer chamber that nullifies the pressure difference between the inner and outer chambers.

REFERENCES CITED

Black, P.B. and Miller, R. D., 1990, Hydraulic conductivity 

and unfrozen water content of air-free frozen silt: 

Water Resources Research, v. 26, no. 2, p. 323-329.

Kiusalaas, Nickolas J., 1992, Estimation of groundwater 

recharge using neutron probe moisture readings near 

Golden, Colorado: Golden, Colorado School of 

Mines, unpublished M.S. thesis, 186 p.

Lenhard, R.J. and Parker, J.C., 1988, Experimental validation of the theory of extending two-phase saturationpressure relations to three-fluid phase systems for 

monotonic drainage paths: Water Resources Research, 

v. 24, no. 3, p. 373-380.

Oak, M.J., Baker, L.E. and Thomas, D.C., 1988, Threephase relative perme ability of Berea Sandstone: Paper 

SPE/DOE 17370 presented at the 1988 SPE /DOE 

Enhanced Oil Recovery Symposium held in Tulsa 

Oklahoma, April 17-20, 1988: p. 555-559.

Olsen, H. W., Nichols, R. W., and Rice, T. L., 1985, Lowgradient permeability measurements in a triaxial 

system: Geotechnique, v. 35, no. 2, p. 145-157.

Olsen, H. W., Morin, R. H., and Nichols, R. W., 1988, Flow 

pump applications in triaxial testing, in Robert T. 

Donaghe, Ronald C. Chancy, and Marshall L. Silver, 

eds., Advanced Triaxial Testing of Soil and Rock, 

ASTM STP 977: Philadelphia, American Society for 

Testing and Materials, p. 68-81.

Olsen, H. W., Gill, J. D., Willden, A. T, and Nelson, K. R., 

1991, Innovations in hydraulic-conductivity measurements. Geotechnical Engineering 1991: Transportation Research Board, National Research Council, 

Transportation Research Record 1309, p. 9-17.

Roseberg, R. J. and McCoy, E. L., 1990, Measurement of 

soil macropore air permeability: Soil Science Society 

of America Journal, v. 54, p 969-974.

Znidarcic, Dobroslov, Illangesekare, Tissa, and Manna, 

Marilena, 1991, Laboratory testing and parameter 

estimation for two-phase flow problems, Geotechnical 

Engineering Congress 1991: American Society of 

Civil Engineers, Geotechnical Special Publication 

No. 27, v. 2, p. 1089-1099.

948
Unsaturated Zone Properties at a 

Hazardous-Waste-Disposal Site at the 

Idaho National Engineering Laboratory

By Stephanie Shakofsky1 and John R. Nimmo1

Abstract

Field and laboratory investigations at a hazardous-waste site at the Idaho National Engineering 

Laboratory (INEL) have yielded measurements of 

the hydraulic properties that govern water and 

contaminant transport in the unsaturated zone. 

Saturated and unsaturated vertical hydraulic 

conductivity, soil-moisture retention, porosity, and 

particle-size distribution were measured in core 

samples collected from depths of 18 to 245 cm. 

In addition, the degree of soil aggregation and 

relative carbonate content with depth were 

analyzed in the field. In order to assess the effect 

of the disposal trenches and pits on hydrologic 

transport-governing properties, studies were made 

of (1) a simulated waste trench that was 

constructed as a model of the waste-containing 

trenches, and (2) a nearby undisturbed soil profile. 

Results show that the undisturbed soil profile has 

distinct layers whose properties differ significantly, whereas the soil profile in the simulated 

waste trench is, to a large degree, homogeneous. 

In general, the vertical movement of water is 

expected to be retarded in a distinctly layered 

medium, suggesting that the construction of 

trenches and pits has destroyed some natural 

impediments to downward flow.

'U.S. Geological Survey, Menlo Park, Calif.

INTRODUCTION

The major concern with waste isolation in 

shallow trenches is the infiltration of water and the 

eventual transport of contaminants to the ground 

water. Hazardous-waste sites in arid and semiarid 

regions are currently considered low risk because they 

receive little precipitation and are usually underlain by 

thick unsaturated zones. However, the potential for 

contaminant transport is always at issue and can be 

addressed only by a quantitative site characterization.

Knowledge of soil hydraulic properties is an 

essential requirement for predicting flow and contaminant transport through the unsaturated zone. A 

complete study involving unsaturated flow requires 

knowledge of the hydraulic conductivity (K) and the 

relation between water content (0) and matric potential 

(v|/), and should include an understanding of the soil 

profile and any artificial disturbances of the profile. 

Complicating factors, such as soil layering and spatial 

variability, soil structure and the presence of 

macropores, and unstable-flow phenomena, can critically affect the fluxes of water and contaminants.

The relations among K, 0, and \\i are commonly 

not measured directly but are estimated by a computational procedure, such as (1) the inference of soil 

hydraulic parameters from empirical or theoretical 

relations based on particle-size or pore-size distributions, and (2) the use of measured values from 

presumably comparable conditions. Such estimation 

procedures can involve errors of an order of magnitude or more, thereby reducing their usefulness for 

contaminant-transport applications. For this reason, it 

is important to use field and laboratory techniques that 

directly measure the hydraulic properties with a 

degree of accuracy adequate for predictive modeling.

949
Stale of 

Idaho

1 113*

Idaho Manorial 

Engineering Labotaioty

' 30'

43' 30'

Scale (m) 

Contour interval 10 feel. Datum is Sea Level

Diversion 

Reservoir 

Big Lost River

Figure 1. Map showing the Idaho National Engineering Laboratory (INEL) site, the Radioactive Waste Management 

Complex (RWMC), and the U.S. Geological Survey test trench area. The Subsurface Disposal Area (SDA) comprises 

about 60 percent of the western part of the RWMC. The simulated waste trench is a 4- by 5-meter area of the test 

trench area. (From Pittman, 1989, fig. 1; base from U.S. Geological Survey, 1:24,000,1972.)

950
We have undertaken a study at a radioactive- and 

chemical-waste landfill to characterize the soil properties that govern contaminant transport in an undisturbed soil profile and in an adjacent simulated waste 

trench. This paper describes the methods applied and 

presents preliminary results of this study.

STUDY SITE

The Idaho National Engineering Laboratory 

(INEL) is situated on the semiarid Snake River Plain in 

southeastern Idaho and comprises approximately 2300 

km2 (fig. 1). Since 1949, INEL has produced substantial quantities of hazardous chemical and radioactive 

wastes, primarily associated with the development of 

commercial nuclear energy. Until 1970, these wastes 

were buried at the Subsurface Disposal Area (SDA) in 

trenches and pits excavated to a depth of approximately 

3 to 6 m in the surficial sediments. Major constituents 

of the hazardous materials buried at the SDA include 

334,600 L of transuranic-contaminated organic wastes, 

38,600 L of acid (mostly HF and HC1), 104,500 L of 

sodium compounds, 88,600 L of oil, and 200 m3 of 

heavy metals (lead and mercury) (Laney and others, 

1988). Comprehensive studies to evaluate the potential 

for migration of the contaminants through approximately 180 m of alternating layers of basalt and sediments into the underlying Snake River Plain aquifer 

have been undertaken by the Department of Energy 

(DOE), the U.S. Geological Survey (USGS), and other 

institutions.

Because hydrologic experimentation in the waste 

trenches and pits is potentially hazardous, the USGS, in 

1986, constructed a simulated waste trench just north 

of the SDA boundary using the methods of waste burial 

historically practiced at the SDA (Davis and Pirtman, 

1990). This practice involves excavation of a trench to 

the top basalt layer (at a depth of 3-6 m), emplacement 

of the waste-containing barrels, and backfilling of the 

trench with the excavated soil. This study focuses on 

the simulated waste trench and the differences between 

properties of the trench backfill and those in the 

undisturbed soil adjacent to the trench.

METHODS

We used a sieve-analysis procedure to measure 

aggregate-size distribution and a hydrochloric acid test 

to determine relative carbonate content as a function 

of depth (Burke and others, 1986). We obtained

minimally disrupted cores by use of a thin-walled, 

hydraulically driven sampler. Samples were collected 

to a depth of 245 cm from both the simulated waste 

trench and the adjacent undisturbed region.

Particle-size distribution was measured by an 

optical-scattering particle-size analyzer. Soil-moistureretention curves were determined by use of a submersible pressure outflow cell (SPOC) (Constantz and 

Herkelrath, 1984), which is a modified pressure-plate 

assembly. Unsaturated hydraulic conductivities were 

estimated by the transient one-step outflow method 

of Gardner (1956) and Passioura (1976). Saturated 

hydraulic conductivity (K^) was measured by a 

falling-head method (Klute and Dirksen, 1986).

RESULTS AND DISCUSSIONS

The undisturbed soil profile description and the 

changes in soil properties with depth are shown in 

figure 2. The moisture contents indicated in figure 2 are 

from neutron-probe data collected at the time of sampling, by the USGS INEL Project office (John Pittman, 

unpublished data on file in the INEL Project office of 

the U.S. Geological Survey). Moisture-retention 

curves are shown in figure 3, and hydraulic conductivity curves are shown in figure 4.

Laboratory measurements of hydraulic properties, 

supported by the field data collected during sampling, 

demonstrate substantial differences in the properties of 

the trench backfill. Moisture-retention curves for the 

disturbed trench-soil cores are less heterogeneous 

vertically than the undisturbed cores; less horizontal 

heterogeneity also is apparent because the replicate soil 

samples collected from the same depths are from sites 

1 to 30 m apart. The disturbed soil cores from the simulated waste trench have a greater K^ than adjacent 

undisturbed soil cores from the same depth. Particlesize analysis revealed the destruction of a high-claycontent layer, and an increase in porosity was measured 

in all the disturbed cores.

Considering the effects of these measured differences on the downward movement of water and contaminants, the most important factor affecting this 

downward movement probably is the lack of pronounced layering in the simulated waste trench. In 

general, the presence of layers has a retarding effect on 

vertical flow (Jury and others, 1991, p. 138-140). In 

particular, at the INEL site, the high-clay-content, 

low-Ksat layer in the 130- to 170-cm depth interval in 

the undisturbed profile can have a significant impeding 

effect. If so, more water may remain above this depth

951
Undisturbed Clay Moisture

Dentil s0'' Reaction to content, % content Porosity Aggregates Ksat 

cfn ' profile 5%HCL (optical technique) (neutron probe) (calculated) >4mm,% (crn/s)

None Mod Strong 10 20 30 0 .15 .30 .4 .5 .6 0 25 501x10'4 1x10'3 1x10'2

Figure 2. Diagram of the undisturbed soil profile and changes in measured soil properties with depth in the 

simulated waste trench and the adjacent undisturbed soil. The increase in clay content below the 140-cm depth in 

the undisturbed profile correlates with increases in large aggregates and moisture content. Increases in porosity 

and Ksat with depth are apparent in the disturbed profiles. The disruption of natural layers can be seen by the 

relatively smooth profiles of the aggregate distribution, clay content, and relative carbonate content in soil from the 

disturbed area.

I I I 

INEL: Simulated waste trench

18cm

30cm

80cm

145cm

230cm

 i 

INEL: Undisturbed area

Figure 3. Moisture-retention drying curves for soil 

samples from five depths in the simulated waste 

trench and from the same depths in the adjacent 

undisturbed area. The substantial variation of the 

curves for the undisturbed area is indicative of layers 

with distinct hydraulic properties. The curves for soils 

from the disturbed area show the effects of 

homogenization of a naturally layered soil.

0.2 *

-800 -400 0 

MATRIC POTENTIAL, IN CENTIMETERS OF WATER

952
Q

Z

o 

o

UJ 

CO

cc

LJU 

Q.

CO 

CC 

LJU

z

UJ

o

z

o

Q

Z

o 

o

CC 

Q

1E-01

1E-02

1E-03

1E-04;

1E-05J

1E-06

1E-0?

1E-08

1E-09d

p 1E-10

1E-01

1E-02| 

1E-03J 

1E-04| 

1E-05J 

1E-06J 

1E-0?J 

1E-08J 

1E-09J

1E-10

INEL: Simulated waste trench

30 CM

0.2 0.3 0.4 0.5 0.6

INEL: Undisturbed area

0.2 0.3 0.4 0.5 0.6 

WATER CONTENT (VOLUME PER VOLUME)

Figure 4. The relation of saturated and unsaturated hydraulic 

conductivity to water content for soil samples from four depths in 

the simulated waste trench and from the same depths in the 

adjacent undisturbed area. Saturated hydraulic conductivities of 

the disturbed samples are higher than those of the undisturbed 

samples, typically by a factor of 3. The unsaturated hydraulic 

conductivities, over the range of the measured data, show a 

greater sensitivity to 6 in the undisturbed soil than in the 

disturbed soils; the slope of log K as a function of 0 for the 

undisturbed soil is, on the average, greater than that for the 

disturbed soil by a factor of 1.4.

interval, where evapotranspiration losses are greater, in 

the undisturbed than in the disturbed profile. Thus, the 

disturbed soil in the 3- to 6-m depth interval may be 

wetter and transmit more downward flow than the 

undisturbed soil. If these effects are substantial, then 

the soil-property-altering effect of trench construction 

is a critical element in characterizing the potential for 

contaminant transport.

REFERENCES

Burke, W., Gabriels, D., and Bouma, J., 1986, Soil structure 

assessment: Rotterdam, Netherlands, A.A. Balkema, 

p. 30-31.

Davis, L.C., and Pittman, J.R., 1990, Hydrological, meteorological, and geohydrological data for an unsaturated 

zone study near the Radioactive Waste Management 

Complex, Idaho National Engineering Laboratory, 

Idaho, 1987: U.S. Geological Survey Open-File Report 

90-114, p. 6-21.

Constantz, J., and Herkelrath, W.N., 1984, A submersible 

pressure outflow cell for measurement of soil water 

retention and diffusivity from 5 to 95°C: Soil Science 

Society of America Journal, v. 48, p. 7-10.

Gardner, W.R., 1956, Calculation of capillary conductivity 

from pressure plate outflow data: Soil Science Society 

of America Proceedings, v. 20, p. 317-320.

Jury, W.A., Gardner, W.R., and Gardner, W.H., 1991, Soil 

physics (5th ed.): John Wiley & Sons, p. 38-40.

Klute, A., and Dirksen, C, 1986, Hydraulic conductivity 

and diffusivity: Laboratory methods, in Klute, A., 

Methods of Soil Analysis: part I, p. 687-732.

Laney, P.T., Minkin, S.C., Baca, R.G., McElroy, D.L.,

Hubbell, J.M., Hull, L.C., Russell, B.F., Stormber, G.J., 

and Pittman, J.T., 1988, Subsurface investigations program at the Radioactive Waste Management Complex 

of the Idaho National Engineering Laboratory: DOE/ 

ID-10183, p. 55-60.

Pittman, J.R., 1989, Hydrological and meteorological data 

for an unsaturated zone study near the Radioactive 

Waste Management Complex, Idaho National Engineering Laboratory, Idaho: U.S. Geological Survey 

Open-file Report 89-74, p. 3.

Passioura, J.B., 1976, Determining soil water diffusivities 

from one-step outflow experiments: Australian Journal 

of Soil Research, v. 15, p. 1-8.

953
954
954
Fate and Transport of Atrazine at the Plains, Georgia,

Ground Water Study Site

By R. A. Leonard1 , LR. Marti 1 , D.W. Hicks2, anofJ.B. McConnell2

Abstract

Cooperative research between the U.S. 

Geological Survey, U.S. Department of 

Agriculture, Agricultural Research Service, and 

U.S. Environmental Protection Agency was 

initiated in 1986 near Plains, Georgia, to define 

processes that affect movement and fate of 

agricultural chemicals and to establish a common 

comprehensive data base for development and 

testing of mathematical models. Data from one 

crop cycle (mid-March 1992 to mid-March 1993) 

are summarized to illustrate the fate and present 

state of atrazine in the system after four annual 

treatments at normal application rates for corn. In 

1992, parent atrazine dissipated in the root zone 

with a half-life of about 18 days. The atrazine 

metabolites, desethylatrazine (DEA) and 

desisopropylatrazine (DIA) accumulated in the 

root zone to a maximum 27 days after application; 

DEA was the primary metabolite. The maximum 

DEA/atrazine ratio was 0.6 40 days after atrazine 

application. After crop harvest in mid-July 1992, 

atrazine, DEA and traces of DIA were detected 

throughout the unsaturated zone to depths of 

8.7 meters. Because the area had been treated 

with atrazine during the previous 3 years, the 

atrazine found could not be associated with a 

particular application. By March 1993, atrazine 

in the unsaturated zone had decreased from an 

estimated 360 grams per hectare to 100 grams per

'U.S. Department of Agriculture, Agricultural Research 

Service, Southeast Watershed Research Laboratory, Tifton, Ga.

U.S. Geological Survey, Water Resources Division, Atlanta,

Ga.

hectare. Atrazine and DEA in ground water, at 

the time of sampling, were at greater 

concentrations downgradient from the treated 

area, with the maximum concentrations about 

60 meters downgradient from the treated plot 

boundary. Traces of atrazine and DEA were 

detected 280 meters downgradient from the plot. 

DEA/atrazine ratios were slightly higher in 

ground water compared to the unsaturated zone, 

possibly indicating continued transformation 

along the ground-water flow path.

INTRODUCTION

Detection of agricultural chemicals in ground 

water in many locations has established a need for 

research to (1) assess the severity and extent of 

ground-water contamination by agricultural 

chemicals, and (2) develop economically viable 

management and policy alternatives to reduce 

concentrations of chemicals in ground water (U.S. 

Department of Agriculture, 1989). In 1986, the U.S. 

Geological Survey; U.S. Department of Agriculture, 

Agricultural Research Service; U.S. Environmental 

Protection Agency; and the University of Georgia, 

began a cooperative study of chemical transport in a 

representative agricultural management and 

groundwater recharge system near Plains, Georgia. 

The common goal was to develop a comprehensive 

data base for evaluation and/or calibration of several 

fate and transport models. Research objectives 

included evaluation of (1) processes that affect fate 

and transport of agricultural chemicals in the soil-root, 

unsaturated, and saturated zones; and (2) transformation and degradation processes of selected pesticides 

in the respective zones.

955
During the period 1989-91, major emphasis was 

placed on the movement and fate of atrazine (2- 

chloro-4-ethylamine-6-isopropylamino-s-triazine) 

and carbofuran (2,3,-dihydro-2,2-dimethyl-7- 

benzofuranyl methylcarbamate) in the root and upper 

unsaturated zones (0 - 3 m depth intervals). Results 

from the 1989 part of the study were summarized by 

Smith and Parrish (1991). Profile averaged pesticide 

half-life was about 15 days for atrazine and about 27 

days for carbofuran. On the basis of pesticide 

concentrations in the soil with depth, carbofuran was 

more mobile than atrazine, but neither compound was 

detected at depths greater than 3 m. Failure to detect 

pesticides in ground water and in soil at depths below 

3 m before mid-1991 was likely due to the fact that 

concentrations did not exceed detection limits of the 

methods in use at that time. This paper presents 

results of studies with atrazine conducted since 1991 

using ELISA (Enyzine-linked immunosorbent assay) 

methods and other low-level detection methods 

employed in the Agricultural Research Service 

laboratory at Tifton, Ga. The paper describes the fate 

of atrazine and its state in the Plains, Ga., study after 4 

years of atrazine application and examines 

DEA:atrazine ratios in the root, unsaturated, and 

saturated zones.

Atrazine is the second most widely used 

herbicide in the United States. Because of its 

widespread use, mobility, and persistence, atrazine and 

its metabolites have been detected in ground water 

more frequently than any other pesticide (Halberg, 

1989). Atrazine degrades in soil to several 

intermediates or metabolites, some of which are 

relatively nonmobile (Clay and Koskinen, 1990). 

Metabolites resulting from dealkylation of side chains, 

desethylatrazine and desisopropylatrazine have 

properties that are similar to those of the parent 

atrazine. The DBA metabolite is the primary 

metabolite found in soils as a result of microbial 

degradation and in ground water (Sirons and others, 

1973). The presence of DEA in relation to atrazine, 

particularly along paths of ground-water flow, has 

been examined to infer age of the contaminant or 

proximity to the source, assuming that DEA becomes 

dominant as atrazine continues to degrade over time 

(Leistra and Boesten, 1989; Denver and Sandstrom, 

1991).

DESCRIPTION OF STUDY AREA

The study area is in the Fall Line Hills district of 

the southern coastal physiographic province of 

southwestern Georgia in Sumter County (fig. 1 A). 

The planted and treated area, about 0.8 ha is located at 

the upslope end of a 6.8-ha field of Eustis loamy sand 

soil (sandy, silicious thermic Psammentic Paleudult) 

and closely related soils. The soil is underlain by the 

Tallahatta Formation of the Claiborne Group, the 

lower portion of which is saturated (Claiborne 

aquifer). The saturated zone is 8.5 to 10.5 m below 

land surface and is 2.5 to 4.5 m thick in the study area. 

The Claiborne aquifer is generally unconfined and 

underlain by the Tuscahoma Formation of the Wilcox 

group which confines deeper water-bearing zones.

METHODS

Permanent monitoring stations were located 

randomly at 12 sites within the treated area at the 

beginning of the study (fig. IB). At each site, three 

2-inch (5.1-cm) diameter wells were installed at three 

depth intervals in the aquifer. Results of sample 

collection from wells in the upper part of the aquifer 

are included in this report. Upgradient and 

downgradient monitor wells also were installed 

adjacent to the treated area. Vacuum lysimeters, soilwater sensors, and other instrumentation also were 

installed at the site (Hicks and others, 1991; Asmussen 

and Smith, 1991). During the period 1991-93, 

additional downgradient wells were installed on a line 

approximately parallel to the direction of the groundwater flow extending 350 m downgradient of the 

treated area.

Atrazine at target rates of 2.24 kg/ha was 

applied each spring from 1989-92 when corn (Zea 

Maize L.) was planted. In 1992, after atrazine 

application, the soil-root zone was sampled to depths 

of 1.5 m periodically for 63 days in seven depth 

increments; one core was removed each time near the 

center of each plot quadrant. Soil cores were collected 

to the top of the saturated zone in mid-July 1992 and 

in early March 1993. Sample segments 15 cm long 

were collected at midpoints of 0.08, 0.76, 1.70, 3.20, 

4.70, 6.25 and 8.70 m below land surface. Atrazine, 

DEA, and DIA were extracted from soil samples by a 

1:1 methanoliwater solution.

Ground-water samples were collected 

periodically from monitor wells installed initially in

956
srsrscr

0 1OO 200X0 KLJOUETtRS 

0 100 20

FtBJ) KUNDAirr 

 SHfUNG TRANSECT

MMrrc* MEU. 

______I______

Figure 1A. Location of study area, planting area, and monitor sites.

957
EXPLANATION

A

(Sr)

MONITOR WELL Letter Is 

site Identification

QUADRANT AND IDENTIFIER

15

50

X METERS

I

I 

100 /TEET

M N

 

C

B

BB

Figure 1B. Location study area, planting area, and monitor sites.

958
and adjacent to the treated plot and from additional 

downgradient wells after they were installed at later 

dates. Ground-water samples were concentrated by 

means of cyclohexyl solid-phase extraction columns 

(SPE) with final elution in methanol. Soil extracts 

were also concentrated by means of SPE columns after 

methanol removal. Atrazine and its metabolites were 

determined by gas chromatography using a 

nitrogen:phosphorus detector and (or) high-pressure 

liquid chromatography. All samples initially were 

screened by use of ELISA procedures.

RESULTS AND DISCUSSION

Soil-Root Zone

Concentrations of atrazine, DEA, and DIA in 

the root zone during the first 63 days after atrazine 

application at corn planting in 1992 are given in table 

1. Values reported are mean concentrations in four 

cores, one from each plot quadrant. Based on standard 

deviations (table 1), average coefficients of variation 

were about 50 percent during the first 2 weeks after 

application, but increased to about 100 percent at later 

sampling dates as concentrations decreased. With 

time, local differences in degradation and transport 

rate contributes additional variation to that attributed 

to application nonuniformity. Also, in the lower 

concentration range, analytical precision in the 

laboratory decreases. In spite of variability in the data, 

initial results for recovered atrazine (2.3 kg/ha) 

compares favorably with the intended application rates 

of 2.2 kg/ha. On day 5, however, the 2.7 kg/ha of 

atrazine recovered reflects data variability since 

atrazine could not have actually increased.

During the 63-day period in 1992, atrazine 

degraded to DEA, DIA and other products not 

determined analytically and showed evidence of 

limited vertical transport in the root zone. However, 

the extent and timing of transport within and through 

the root zone may be partially obscured and 

confounded by several factors. Contamination of soil 

layers below the surface during sampling may account 

for some of the apparent changes in atrazine 

concentration, particularly on the first day after 

atrazine application. Atrazine transport is believed to 

occur in pulses in response to rainfall and irrigation 

and atrazine in these pulses may not be captured or 

revealed by soil sampling, particularly if transport

occurs in preferential pathways. Vertical transport 

through layered soils does not occur as a symmetrical 

plume because of changes in adsorptive properties and 

water-retention characteristics and with soil depth. 

Evidence for downward atrazine translocation is 

indicated by the relative increases in atrazine and DEA 

concentrations in the 15 to 90-cm deep zone 5 to 27 

days after application.

Fitting the data (table 1) to first-order kinetics 

reveals a half-life of about 18 days for the parent 

atrazine, similar to the half-life of 14 days found for a 

1989 application to the same plot (Smith and Parrish, 

1991). DEA was the dominant metabolite. The 

general form of the curves (fig. 2) that show the 

change in atrazine, DEA, and DIA over time is 

consistent with coupled first-order reactions (Leonard 

and others, 1990). The ratio of DEA to atrazine 

increased to a maximum of about 0.6,41 days after 

atrazine application.

Unsaturated Zone

Soil cores were collected near sites B, C, D, F, 

G, J, K, and L (fig. 1) in July 1992 after harvest of corn 

silage, approximately 120 days after atrazine 

application, and in February 1993. In addition to the 

surface 0-15 cm interval (0.08 m midpoint), 15-cm 

increments of the cores were collected with midpoints 

of 0.76, 1.70, 3.20,4.70, 6.25 and 8.70 m for atrazine 

analysis. Atrazine and DEA were detected in nearly 

all samples (table 2). An occasional trace of DIA near 

detection limits was observed but not reported. Only 

means and standard deviations are presented. Some 

differences in concentration profiles were observed 

that were apparently related to spatial variations in soil 

and unsaturated-zone characteristics.

As expected, atrazine and DEA concentrations 

were consistently higher in July 1992 than in March 

1993. Concentrations were greater in the surface 

horizon at both sampling times than those in lower 

horizons, a result of atrazine and DEA adsorption/ 

binding to soil organic matter. Results obtained 

represent the state of the system after 4 years of 

atrazine application. Therefore, atrazine and DEA 

detected can not be ascribed to a particular application. 

However, comparing the July 1992 and March 1993 

results indicate that atrazine and DEA were 

transported from the system or degraded during the 

intervening 8-month period. In contrast to the root 

zone, degradation rates in the unsaturated zone may be

959
to

s

Table 1. Concentrations of atrazone and atrazine metabolites in the soil-root zone (0-150 cm) at different times after atrazine 

at Plains, Ga., March 18,1993

[Values are mean concentrations for four cores; value in parentheses () is standard deviation; jag/kg, micrograms per kilogram; ND, not detected]

Days 

After

Application

Depth in Soil, in Centimeters

0-15 15-30 30-45 45-60 60-90 90-120 

Atrazine

-2

1

5

13

27

41

63

6.5(3.1)

690(250)

640(330)

330(110)

400(112)

150(46)

41(7.8)

2.6(2.1)

49(39)

270(200)

280(180)

201(115)

61(24)

43(22)

1.0(1.1)

29(10)

81(120)

81(55)

39(22)

12(8.0)

14(13)

0.6(1.1)

17(9.9)

42(69)

61(107)

23(23)

9.6(8.6)

2.0(3.9)

0.4(0.7)

16(6.4)

17(24)

58(110)

11(13)

19(31)

1.4(2.8)

0.2(0.4)

28(20)

6.2(3.2)

4.5(2.7)

2.0(0.68)

5.9(11)

ND

Des ethyl atrazine (PEA)

-2

1

5

13

27

41

63

ND

27(12)

44(19)

77(46)

89(39)

73(24)

28(12)

ND

3.7(3.9)

19(9.5)

24(12)

70(30)

86(41)

15(3.2)

ND

1.0(2.1)

5.5(5.8)

8.2(2.3)

14(14)

4.4(4.7)

3.3(1.4)

ND

2.2(2.8)

3.6(3.9)

4.5(5.8)

12(10)

2.7(3.5)

1.1(2.1)

ND

.58(.76)

1.1(1.6)

2.0(3.6)

5.1(7.1)

1.1(1.5)

1.2(2.4)

ND

0.28(.55)

1.3(1.5)

ND

0.64(1.3)

ND

ND

Desisopropylatrazine (DIA)

-2

1

5

13

27

41

63

ND

8.6(4.2)

18(5.8)

23(16)

23(11)

18(6.3)

7.5(2.2)

ND

1.5(3.1)

2.9(3.9)

5.5(3.6)

8.0(1.4)

7.2(8.6)

3.5(1.4)

ND

ND

0.50(1.0)

ND

1.8(2.2)

ND

.91(1.2)

ND

ND

ND

ND

ND

ND

.31(0.63)

ND

ND

ND

ND

ND

ND

ND

ND

ND

ND

ND

ND

ND

ND
x ATRAZINE

O DESETHYUTRAZINE (DEA) 

 O DESISOPROPYLATRAZINE (DIA)

 DESEJHYLATRAZINEiATRAZINE RATIO

10 20 30 40 50 

DAYS AFTER ATRAZINE APPLICATION

0.0

Figure 2. Changes in concentrational atrazone and atrazine metabolites in zoil-root zone over time after atrazine 

application on March 18,1992.

very low and ascribed mainly to chemical processes 

rather than biological processes.

Total mass of atrazine in the unsaturated zone 

was estimated using assumptions as to the depth 

intervals or zones represented by samples from the 

cores (table 2). The surface sample was assumed to 

represent the 0 to 0.3 m interval; while the other 

samples were assumed to represent equidistance 

intervals around the sampling point except for the 

8.7-m sample which was limited to an interval of 8.5 

to 9.0 m because of possible influences of the capillary 

fringe above the saturated zone. Using these 

assumptions and a value of 1.7 g/cm for soil bulk 

density, an estimate of 360 g/ha atrazine was 

calculated for July 1992 compared to 100 g/ha for

March 1993. A mass of 100 g/ha represents about 

5 percent of a single atrazine application.

Based on results given in table 2, no significant 

difference was found between atrazine by gas 

chromatography compared to atrazine by ELISA 

methods, although some cross-reactivity with DEA 

and DIA is expected in the ELISA methods. Although 

limited cross-reactivity has been observed, results 

have not been consistent. The comparisons in this 

report are offered only as confirmation of the presence 

of atrazine. The ratio of DEA to atrazine in the 

unsaturated zone ranged from 0.2 to 0.4 in July 1992 

and from 0.3 to 1.2 in March 1993. The tendency of 

increased DEA:atrazine ratio after about 8 months 

may indicate continued degradation of atrazine.

961
Table 2. Atrazine and desethylatrazine (DEA) concentrations and estimated mass in the unsaturated zone at Plains, Ga., 

July 14,1992, and March 10,1993

[Depth is distance below land surface at midpoint of 15-cm core segment; values are mean concentrations of eight cores; value in parentheses ( ), standard 

deviation; ELISA, enzyme-linked immunosorbent assay; GC/NPD gass chromatography-nitrogen/phosphorus detector; ng/kg, micrograms per kilogram; 

g/ha, grams per hectare; <, less than]

Depth in

Unsaturated

Zone

Atrazine

(ELISA)

( j*g /kg)

Atrazine

GC/NPD

(/ig/kg)

DBA

GC/NPD

(MgAg)

DBA/

Atrazine

Ratio

Estimated

atrazine mass

by depth

interval

(g/ha)

Meters

0.08

0.76

1.70

3.20

4.70

6.25

8.70

,9(2

,0(1

4(2

.4(2

4(2

7(3

2)

3)

1)

2) 

2) 

0)

2.0(2.3)

July 1992

10

1

1

1

2

2

2

6(3 

6(1 

3(1 

3(1 

,7(2 

9(3

8) 

0)

8)

9) 

1) 

9)

0(2.1)

8(1 

,6(0 

2( 

3( 

7(1 

6( 

4(

.2) 

.7) 

.4) 

.9) 

.3) 

.9) 

.8)

.2 

.4 

.2 

.3 

.3 

.2 

.2

March 1993

45

31

31

35

62

140

17

360 Total

0

0

1

3

4

6

8

.08

.76

.70

.20

.70

.25

.70

3

0

3

.1(1

.4(0

.3(

.3(

.9(1

 3(

.0(2

.6)

.4)

.4)

.3)

.4)

.4)

.5)

2

0

3

.8(1.7)

.4(0.5)

 3( .4)

.4( .3)

 6( .7)

 5( .6)

.3(2.5)

0.8

< .1

.3(0

.4(

 4(

 2(

1.4(

.4)

.6)

.5)

.5)

.8)

0.3

< .1

.9

1.2

.4

.7

.5

16

6

6

8

22

16

26

100 Total

Saturated Zone

Atrazine was first detected in ground water in 

July 1991 by ELISA methods, (data not shown). 

Concentrations were about 1 ng/L and detected only in 

wells in the lower or downgradient area of the plot. 

Selected wells were sampled beginning in 1989. 

However, detection limits of the methods used at that 

time were insufficient for quantification at or below 

1 ng/L. Also, some of the selected wells were located 

at upgradient sites where atrazine has not been 

detected to date (March, 1993). Once atrazine was 

detected in ground water, sampling of wells 

downgradient from the plot was begun. Since atrazine 

was found at higher concentrations in the off-plot 

wells, additional downgradient wells have been 

installed after February 1992; the latest installation in 

February 1993. The main line of wells extends along a

north-south line through the center of the plot 

extending about 450 m from the northern plot 

boundary (fig. 1). Selected results of sampling in 

November 1992 and February 1993 are given in table 3. 

Reported concentrations are for wells along the northsouth line and other wells within 15 m of the line. In 

November 1993, wells were not installed beyond 188 

m south of the northern boundary of the plot (about 88 

m beyond the plot boundary). As of April 1993, 40 

wells have been installed downgradient from the plot 

and sampling continues. During November 1992 and 

February 1993, atrazine was not detected in wells on 

the upgradient side of the plot as expected. In 

November 1992, an atrazine concentration of 2.7 ng/L 

was detected at site J, (fig. 1), and in March, 6.6 ng/L 

at site F. Sites J and F are 59 and 43 m south of the 

northern plot boundary. Atrazine was not detected in 

other wells in the plot adjacent to the north-south line

962
Table 3. Concentrations and ratios of atrazine and Desethylatrazine (DEA) in ground water at the Plains, Ga., Research Site

[DEA; desethylatrazine; ND, not detected; - not analyzed; ng/L, micrograms per liter]

Approximate 

distance from 

upgradient 

(northern) November 1992 March 1993

plot 

boundary 

in meters

-16

43

59

85

126

155

188

220

229

229

242

274

311

343

380

414

448

DEA: 

Atrazine DEA Atrazine Atrazine 

(//g/L) (/ig/L) ratio (/ig/L)

ND - - ND

ND - - 6.6

2.7 2.5 0.9 ND

ND ND - ND

7.9 5.3 .7 2.0

16.5 12.1 .7 12.8

0.7 0.5 .7 6.3

0.4

4.0

.1

1.6

1.2

.1

ND

.2

ND

ND

DEA: 

DEA Atrazine 

(/KJ/L) ratio

-

7.3 1.1

-

-

1.6 0.8

8.7 .7

5.1 .8

0.3 .8

2.9 .7

.6 6.0

.9 .6

.9 .8

.3 3.0

-

.8 4.0

-

-

during these two sampling periods. In both sampling 

periods, the maximum atrazine concentration 

measured in water was from wells located 155 m 

downgradient from the plot (table 3) 16.5 ng/L in 

November 1992, and 12.5 Mg/L in March 1993. The 

March 1993 sampling indicated some additional 

downgradient transport beyond the 155 m point 

decreasing to about 0.2 ng/L in the 380 m well, about 

280 m downgradient from the plot.

The ratio of DEA to atrazine in ground water 

was in the range of 0.6 to 0.9 in most wells. 

Occasionally ratios as high as 6 were observed, but 

concentrations of both atrazine and DEA were usually 

very low in these samples. The DEA:atrazine ratios

were generally less than those reported by Denver and 

Sandstrom (1991); data they interpreted as suggesting 

continued degradation of the parent atrazine in the 

ground-water system. The DEA:atrazine ratios in 

ground-water samples from this study generally were 

greater, however, than found in the root and 

unsaturated zones.

SUMMARY AND CONCLUSIONS

After 4 consecutive years of atrazine application 

to the research plot at Plains, Ga., atrazine and atrazine 

metabolites are distributed throughout the 

hydrogeologic system, appearing at all depths sampled

963
in the unsaturated zone and in ground water 

downgradient from the treated area. In the Eustis soil, 

most of the atrazine applied in 1992 dissipated from 

the root zone in 63 days with a half-life of about 

18 days for the atrazine parent. During this time 

period, the atrazine metabolites DEA and DIA 

appeared in maximum accumulations at 27 days, DIA 

being minor compared to DEA. The maximum 

DEA:atrazine ratio in the root zone was 0.6 at 41 days.

Unsaturated-zone sampling in July 1992 and 

March 1993 showed that atrazine and DEA were 

distributed throughout the zone above the water table 

after atrazine application each year for 4 years. 

Concentrations tended to decrease downward from the 

surface and increase near the water table. As 

expected, atrazine and DEA concentrations were 

higher in July 1992 about 120 days after application 

than in March, 1993. About 260 g/ha of atrazine 

dissipated from the unsaturated zone between July 

1992 and March 1993. The DEA:atrazine ratio in the 

unsaturated zone was similar to that in the root zone.

Highest concentrations of atrazine and DEA in 

ground water were downgradient from the plot with 

maximum atrazine concentrations of 12 to 16 ng/L, 

located about 60 meters below the lower plot 

boundary. The March 1993 sampling showed 

detectable atrazine as far as 280 meters downgradient 

from the plot. The DEA:atrazine ratio in ground water 

ranged from 0.6 to 6.0. The ratios tended to be greater 

in ground water than in the unsaturated zone and the 

maximum ratio observed in the root zone, possibly 

indicating continued transformations of atrazine along 

the flow path.

REFERENCES

Asmussen, L.E., and Smith, C.N., 1991, Study design to 

investigate and simulate agrochemical movement and 

fate in ground water recharge, in Nash, R. A., and 

Leslie, A.R., eds., Groundwater Residue Sampling 

Design: Washington, D.C., American Chemical 

Society, American Chemical Society Symposium 

Series No. 465, p. 150-164.

Clay, S.A., and Koskinen, W.C., 1990, Adsorption and 

desorption of atrazine, hydroxy-atrazine and 

s-glutathione atrazine on two soils: Weed Science, 

v. 28, p. 262-266.

Denver, J.M. and Sandstrom, M.W. 1991, Distribution of 

dissolved atrazine and two metabolites in the 

unconfined aquifer, Southeastern Delaware: U.S. 

Geological Survey, Water-Resources Investigations 

Report 91-4034.

Halberg, G.R. 1989, Pesticide pollution of ground water in 

the humid United States: Agriculture, Ecosystems and 

Environment, v. 26. p. 299-369.

Hicks, D.W., McConnell, J.B., Asmussen, L.E., Leonard, 

R.A., and Smith, C.N., 1991, Movement and Fate of 

Agricultural Chemicals in the surface and subsurface 

environments near Plains, Southwestern Georgia - 

Integrated work plan: U.S. Geological Survey OpenFile Report 91-73. 26 p.

Leistra, M., and Boesten, J., 1989, Pesticide contamination 

of ground water in Western Europe: Agriculture, 

Ecosystems, and Environment, v. 26, p. 369-389.

Leonard, R.A., Knisel, W.G., Davis, P.M., and Johnson, 

A.W., 1990, Validating GLEAMS with field data for 

fenamiphos and its metabolites: Journal of Irrigation 

and Drainage Engineering, v. 116, no. 1, p. 24-35.

Sirons, G.J., Frank, R., and Sawyer, T., 1973, Residues of 

atrazine, cyanazine, and their phytotoxic metabolites in 

a clay loam soil: Journal of Agricultural and Food 

Chemistry, v. 21, p. 1016-1020.

Smith, C.N. and Parrish, R.S. 1991, Preliminary results of 

a pesticide leaching study at Plains, Georgia: U. S. 

Geological Survey, Water Resources Investigation 

Report 91-4034, Reston, Virginia, p. 605-609.

U.S. Department of Agriculture, 1989, U.S. Department of 

Agriculture research plans for water quality: 

Agricultural Research Service and Cooperative State 

Research Service, 14 p.

964
Unsaturated Zone Transport at the 

Plains, Georgia Field Site

ByD.D. Bosch 1 , RA Leonard 1 , anc/R.K. Hubbard 1

Abstract

Soil-water and solute-transport data have 

been collected over three years on a 0.81 ha field 

site located near Plains, Georgia. Soil cores 

collected from the root and vadose zone show 

extreme variability in movement and transport of 

applied bromide (Br~) tracer. The Groundwater 

Loading Effects of Agricultural Management 

Systems - Finite Element Solute Transport 

(GLEAMS- FEST) root-vadose zone model was 

used to simulate Br" transport below the site. 

Comparison of simulated concentrations to 

horizontally averaged profile concentrations show 

good agreement for year one but poor agreement 

thereafter, probably due to inaccurate estimates of 

hydraulic parameters in deeper parts of the vadose 

zone. Because of extreme variability in transport 

data, profile-averaged concentrations may not be 

a realistic representation of the sample data.

INTRODUCTION

In 1988, the U.S. Geological Survey, U.S. 

Department of Agriculture, Agricultural Research 

Service, U.S. Environmental Protection Agency, and 

University of Georgia began a cooperative study of 

chemical transport at an agricultural field near Plains, 

Georgia. Three overall objectives of the study were 

set: (1) identify processes that affect movement and 

fate of agrichemicals in the root, vadose, and saturated 

zones, (2) test process-based models using collected 

field data, and (3) use the models to evaluate the effect

'U.S. Department of Agriculture, Agricultural Research 

Service, Southeast Watershed Research Laboratory, Tifton, 

Georgia 31793

of agricultural-management practices on ground-water 

quality.

The study area is 0.81 ha and is located in the 

recharge area of the Claiborne aquifer system. The 

soil is a Eustis loamy sand (sandy, silicious thermic 

Psammentic Paleudult). A soil berm was constructed 

around the perimeter of the plot to confine runoff. A 

H-flume, 0.46 m (meter) height, was installed at the 

southwestern (SW) corner of the plot to facilitate 

measurement of sediment and chemical transport in 

the surface runoff (Smith and Parrish, 1991).

Conventional-till corn (Zea Maize L.) has been 

planted on the plot since 1989 (Smith and Parrish, 

1991). A center-pivot irrigation system was installed 

to supply irrigation water to the site in order to ensure 

percolation of pesticides and to maintain proper soilmoisture for crop growth. The plot has been treated 

with the agricultural herbicides atrazine and alachlor, 

and with the insecticide carbofuran. A Br" tracer was 

applied to the plot in 1989 and again in 1991. In 

addition, movement of nitrogen and chloride supplied 

in applied fertilizer also is being monitored. All 

compounds were surface-applied.

This report examines the movement of water 

and transport through the vadose zone following the 

first Br" application. Two objectives were set for this 

research: (1) evaluate the variability of the Br" plume 

and relate it to physical properties of the unsaturated 

zone and (2) examine the feasibility of modeling Br" 

transport through the unsaturated zone using a onedimensional solute-transport model.

METHODS

In 1989, Br" was applied at a rate of 88 kg/ha on 

the same day corn was planted and pesticides were 

applied. The corn-planting date of June 13, 1989 was

965
about 70 days later than normal for southern Georgia 

due to difficulty instrumenting the plot.

Soil cores were collected six times after 

pesticide application in 1989, from 20 randomly 

located sites. In 1990, soil cores were collected eight 

times, again from 20 different sites each time. For 

three of these samplings, soil cores were collected 

down to the water table (9.1 m). Hand augers were 

used to collect soil-samples down to 2 m during the 

growing season, and a drill rig equipped with a 

continuous-coring device was used to collect soil 

cores. Continuous soil cores were collected on 

October 10, 1989, February 20, 1990, and November 

8, 1990. These dates correspond to 120, 253, and 514 

days after application, respectively. Total precipitation 

and irrigation falling on the plot between application 

and these sample dates was 80, 138, and 223 cm 

(centimeters), respectively.

The GLEAMS root-zone model (Leonard and 

others, 1987) was used to simulate the movement of 

Br" through the root zone during 1989 and 1990. 

Actual precipitation and irrigation data, along with 

collected soil data, were used as simulation input. The 

FEST model (Bosch, 1989) was used to model flow 

through the vadose zone. The FEST model uses a 

linear finite element numerical solution to coupled 

water and solute transport. The model uses output 

from the GLEAMS model to provide upper boundary 

conditions.

Simulations were made assuming a three-layer 

distribution: the root zone (0 to 0.71 m), upper portion 

of the vadose zone (0.71 to 2.35 m), and lower vadose 

zone (2.35 to 5.0 m). Only the first 5 m of the profile 

was simulated. The rooting zone was further divided 

into five horizons. Hydraulic properties of each of 

these layers are listed in table 1. These values were 

based on hydraulic data determined from soil cores 

collected from the plot area. These soil cores were 

collected when the plot was established to determine 

texture and hydraulic properties of the root and vadose 

zone.

RESULTS

Bromide Transport

Results of the analysis of samples collected on 

June 15, 1989, October 10, 1989, February 20, 1990, 

and November 8, 1990, are reported here. Mean Br"

concentrations and standard deviations of these 

observations for each of the sampling intervals on 

three of these dates are presented in figure 1. The 

largest concentrations of Br" were observed at 

approximately 1.5 m on October 10,1989, at 1.6 m on 

February 20, 1990, and at 2.0 m on November 8, 

1990. Examination of the maximum concentrations 

on the three dates indicates the Br" plume moved 

rapidly to 1.5 m, but slowly thereafter. In addition, 

maximum concentrations decrease with time. Large 

variability was observed in the Br" data.

Table 1 . Hydraulic properties assigned to the vadose zone 

for the model simulations

Upper 

boundary 

(meters)

0

.15

.30

.46

.61

.71

3.7

Lower 

boundary 

(meters)

.15

.30

.46

.61

.71

3.7

5.0

Vertical

saturated 

hydraulic 

conductivity

(cm/hour)

2.00

2.00

2.00

2.00

2.00

0.75

1.41

Porosity

0.28

.28

.30

.32

.32

.35

.35

Field 

capacity

0.19

.19

.19

.22

.22

.33

.33

Wilting 

point

0.03

.03

.03

.04

.04

.11

.07

The Br" distribution throughout the vadose zone 

on October 10,1989 and November 8, 1990 was 

examined graphically. This examination clearly 

showed the Br" tracer moved much deeper in the SW 

corner of the plot than it did in the northeastern (NE) 

corner. On November 8, 1990, the Br" plume had 

moved downward to a depth of approximately 3.0 m in 

the SW corner, compared to 2.0 m in the NE corner. 

In addition, the Br" distribution in the NE corner was 

less variable than that in the SW corner. These factors 

indicate a resistance to vertical transport in the NE 

corner between 1 and 2 m.

Soil cores collected from the plot indicate the 

area is underlain by unconsolidated sediments whose 

hydraulic characteristics are largely controlled by a 

clay layer extending from a depth of approximately 

0.9 m to 3.4 m (Hicks and others, 1991). In the NE 

corner of the plot, the minimum measured vertical 

saturated hydraulic conductivity (Ksat) of the dense 

layer was 0.03 cm/hr (centimeters per hour) at 1.6 m. 

The layer of reduced Ksat appears to extend from 1 to 

5 m in depth in the NE portion of the plot. In the SW 

corner, the minimum Ksat was 1.3 cm/hr at a depth of
0

£ 1

h LJ2 ?

z

HLUQ -4

-5

00 -1

a: '

LL> y -2

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' i ^ j-i I I, \ f I. ' | Hj- '

~ October 10. 1989

i i i i i i i i i i i I i i i

¥ 1 | | mean ( standard deviation :

~1 1

1 1

x -3

£ h

Q -4 fcj February 20. 1990

e: V

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to 1;'

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Lu «-j

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x -^

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i i i i i i t | i i i | i i i

N-H i :

 LJ_1 1 i

: 1 1

November 8, 1990 

-5 

i i i i i i i i i i i i i i i

0 40 80 120 1C 

BROMIDE CONCENTRATION. IN MILLIGRAMS PER LITER

Figure 1. Changes in the observed mean and standard 

deviation of bromide concentrations with depth in the 

vadose zone.

for the SW and NE quadrants on October 1 0, 1 989 and 

November 8, 1990 are shown in tables 2 and 3. Few 

of the observed differences are statistically significant 

at the 0.05 percent level. Data obtained from samples 

collected on June 15, 1989 and February 20, 1990 

were similar. Because a statistical difference could not 

be obtained, modeling simulations were compared to 

data collected over the entire plot area.

Table 2. Mean 1 in milligrams/liter and variance of the Br 

concentration on October, 10, 1989 in the southwestern and 

northeastern quadrants

Mid-point Southwestern Northeastern 

depth quadrant quadrant

(meters) Mean variance Mean Variance

O.I 0.6a 0.3 l.6b 1. 7 

.2 .6a .5 .3a .3 

.4 .3a .1 .la .0 

.5 .6a .5 .4a .1 

.7 l.3a .5 .8a .3 

.8 3.3a 7.0 2.2a 2.4 

1.0 26.8a 1148.0 64.6a 5639.0 

1.1 9l.9a 4473.0 67. la 2600.0 

1.3 I00.5a 4373.0 69. la 2336.0 

1.4 72.3a 1758.0 85.0a 8315.0 

>0 1.6 53.7a 2439.0 3.3b 12.7 

1.8 33.4a 2143.0 .Oa .0 

2.0 3 1.1 a 2903.0 .Oa .0 

2.3 9.4a 265.0 .Oa .0 

2.6 .Oa .0 .Oa .0

2.3 m. The dense layer in this corner appears to be 

less than 0.5 m thick. These data support the theory 

developed from the observed vertical distribution of 

Br"; the theory being there is greater resistance to 

vertical transport in the NE corner of the plot than in 

the SW corner.

Although not confirmed, lateral transport of Br" 

apparently occurs along the top of the dense layer in 

the plot. Hicks and others (1991) reported an increase 

in observed Br" mass in soil cores collected in the SW 

corner over time, but a decrease in cores collected 

from the NE corner. Examination of the second Br" 

application by Hicks and others (1996) also supports 

this observation.

Data from the four sampling dates were 

separated according to quadrants from which each 

soil-sample was collected and a statistical comparison 

was made. Mean Br" concentrations at each sampled 

depth interval in the NE and SW quadrants were 

statistically compared using the t test for independent 

sample means (Ott, 1984). Results of this comparison

Means in the same row followed by the same letter are not 

significantly different at the 5 percent confidence level.

Table 3. Mean 1 in milligrams/liter and variance of the Br" 

concentration on November 8,1990 in the southwestern and 

northeastern quadrants

Mid-point 

depth 

(meters)

0.5

1.0

1.4

1.6

1.8

2.0

2.3

2.6

2.9

3.2

3.5

Southwestern 

quadrant

Mean

O.Oa

.la

.la

1.6a

5.2a

23.8a

24.9a

4.2a

7.2a

l.la

.Oa

Variance

().()

.0

.0

10.0

80.0

372.0

814.0

38.0

262

6.7

.0

Northeastern 

quadrant

Mean

9.7a

.Oa

3.1a

9.4b

23.0b

15.9a

l.8b

.Oa

.Oa

.Oa

.Oa

Variance

370.0

.0

12.6

119.0

301.0

125.0

5.8

.0

.0

.0

.0

'Means in the same row followed by the same letter are not 

significantly different at the 5 percent confidence level.

967
0

0.00

-0.25 -

LC.

I

? -0.50 -

-0.75 -

-1.00

standard 

deviation

 I

i i I

0 25 50 75

BROMIDE MASS CONCENTRATION IN KILLIGRAMS PER HECTARE

Figure 2. Simulated (*) and observed mean and standard 

deviation of the bromide mass concentration in the root zone 

on June 15,1989.

Model Simulations

Figure 2 presents a comparison of GLEAMS 

simulated Br" data and Br" data obtained from samples 

collected on June 15, 1989, three days after Br" 

application. Between dates of application and this 

sampling, 4.2 cm of rainfall and irrigation water fell 

on the plot area. The peak simulated Br" mass slightly 

lags the observed, and the GLEAMS model predicts a 

more diffuse Br" plume than what was observed. 

However, given the variability ofthe data, this 

simulation result may be considered an acceptable 

representation ofthe Br" distribution in the root zone.

Simulated and observed Br" concentrations in 

the vadose zone on October 10,1989 and November 8, 

1990 are shown in figure 3. Simulated data closely 

match observed data for October 10, 1989, but the 

simulated plume for November 8, 1990 is much 

deeper than the observed. This is due to two factors. 

First, this was an attempt to simulate average Br" 

transport in the plot area. Because of variability ofthe 

physical and hydraulic properties of soils in the plot 

area, average transport is not a good representation of 

the actual transport process. While it appears the 

mean of the observations may represent the depth of 

transport in the NE corner, it is not representative of

[iiii|iiii|iiir

p standard

i mean deviation

rc*-^ '

CO

a:

CL LJ°4

0

<S)

o: LJ

Q_ 

LJ 

Q

i^

October 10, 1989

i i i

November 8, 1990

I T 1 | T 1 I i | i 1 I I

0 50 100 150 

BROMIDE CONCENTRATION IN MILLIGRAMS PER LITER

Figure 3. Simulated (*) and observed mean and 

standard deviation ofthe bromide concentration on 

October 10, 1989 and November 8, 1990.

transport in the SW corner. Also, in an attempt to 

simulate this mean behavior, average hydraulic 

characteristics were used. These average 

characteristics do not appear to yield good estimates of 

average Br" concentrations.

CONCLUSIONS

Initial investigation of soils at the Plains site led 

researchers to believe that chemical transport at the 

site would primarily be vertical, with little lateral 

transport. The surface soil is relatively coarse textured 

and initially appeared to be uniform in the horizontal 

direction. Soil and chemical data collected at this site 

in 1989 and 1990 indicate this is not the case. 

Considerable variability has been observed in texture 

and soil hydraulic conductivity throughout the plot. 

This variability has introduced a lateral flow 

component into the transport process. Lateral flow 

appears to be induced by a layer with low hydraulic 

conductivity which slopes from a depth of 0.9 m in the

968
NE corner of the plot to a depth of 2.3 m in the SW 

corner. Hydraulic conductivity of the layer increases 

in the downslope area.

This nonhomogeneity complicates data analysis. 

Because flow in the plot is not strictly onedimensional, the random sampling procedure used 

does not produce a clear picture of Br" transport 

throughout the plot area. Concentration of flow due to 

preferential flow patterns can skew sample results. 

Lateral flow can cause applied chemicals to move off 

the plot in flow patterns not consistent with groundwater flow. This can cause randomly collected 

samples to miss portions of the chemical plume 

moving off of the plot.

Although initial analysis has shown that 

observed differences are statistically insignificant, 

mean, values of Br" concentrations for the entire plot 

do not appear to be representative of the actual 

physical processes occurring. A more accurate 

representation may be made by examining observed 

data in light of prior knowledge of the hydraulic 

characteristics of the plot.

REFERENCES

Bosch, D.D., 1989, A numerical model for incorporating 

vadose zone transport into the GLEAMS root zone 

model. Proceedings of the CREAMS/GLEAMS 

symposium. University of Georgia, Athens, Georgia, 

Sept. 27-29, 1989: p. 136-148.

Hicks, D.W., McConnell, J.B., Pcrsinger, H.H., Scholz, 

J.D., and Hubbard, R.K., 1996, Two-dimensional 

distribution of a potassium bromide tracer in the 

unsaturatcd zone at the Plains, Georgia, research site: 

in Morganwalp, D.W., and Aronson, D.A., cds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting. 

Colorado Springs, Colorado, September 20-24, 1993, 

U.S. Geological Survey, Water Resources 

Investigations Report 94-4015.

Hicks, D.W., McConnell, J.B., and Hubbard, R.K., 1991. 

Effect of the unsaturatcd zone on the migration and 

retardation of a conservative tracer at the Plains, 

Georgia, research site: //; Mallard, G.E., and Aronson, 

D.A., cds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Montcrey, Calif., March 11-15, 1991, U.S. 

Geological Survey, Water Resources Investigations 

Report 91-4034, p. 597-600.

Leonard, R.A., Kniscl, W.G., and Still, D.A., 1987, 

GLEAMS Groundwater loading effects of 

agricultural management systems. Transactions of the 

ASAE: v. 30, no. 5, p. 1403-1418.

Ott, L., 1984., An introduction to statistical methods of data 

analysis: Boston, Massachusetts, PWS Publishers, p. 

138-175.

Smith, C.N. and Parrish, R.S., 1991, Preliminary results of a 

pesticide leaching study at Plains, Georgia: U.S. 

Geological Survey, Water Resources Investigations 

Report 91-4034, Rcston, Virginia, p. 605-609.

969
970
Two-Dimensional Distribution of a Bromide Tracer in the 

Unsaturated Zone at the Plains, Georgia, Research Site

By D.W. Hicks 1 , J.B. McConnell 1 , H.H. Persinger1 , J.D. Scholz 1 , and R.K. Hubbard2

Abstract

A cooperative research investigation was 

initiated in 1986 near Plains, Georgia, by the U.S. 

Geological Survey; and the LLS. Department of 

Agriculture, Agricultural Research Service, to 

describe the processes that affect the movement 

and fate of nitrogen fertilizers; the pesticides 

atrazine, alachlor, and carbofuran; and a potassium 

bromide tracer in the plant-root, unsaturated, and 

saturated zones. As a result of ongoing research, in 

1991 emphasis was placed on field-scale research 

to develop a better understanding of the physical 

and chemical processes that occur as water flows 

through variably saturated or unsaturated porous 

media. Data collected along a farmed, 420-footlong transect were used to define and evaluate the 

factors that control the rate of movement and 

distribution of agricultural chemicals in the 

unsaturated zone. Lateral (nonvertical) migration 

of soil water and chemicals at the interface 

between permeable and less permeable material is 

hypothesized to account for a significant part of 

the dispersion of agricultural chemicals in the 

unsaturated zone.

Preliminary results of this investigation indicate that the fluctuations of soil water and concentrations of bromide at specific depths and locations 

along the transect are a function of the relative 

vertical hydraulic conductivities of the adjacent 

soil layers. Soil heterogeneity substantially affects 

the migration rates and distribution of the bromide 

tracer in the unsaturated zone. Lateral flow along

'U.S. Geological Survey, Atlanta, Ga. 

2 U.S. Department of Agriculture, Agricultural Research 

Service, Tifton, Ga.

the interface of permeable and less permeable soil 

may account for the rapid fluctuations observed in 

matric suction and the distribution of the tracer 

in the unsaturated zone. For this reason, the 

deterministic methods that are routinely used in 

numerical simulation of fate and transport are not 

valid in the vast majority of field conditions when 

applied in sloping, structured, variably saturated 

porous media.

INTRODUCTION

Heavy use of agricultural chemicals on farmland in 

the southeastern United States has caused concern 

about the possible degradation of ground-water 

resources. Reconnaissance evaluations of groundwater resources in this area indicate that agricultural 

management practices have affected water quality; 

however, the geographic extent of the problem is not 

well known. Wells in the same hydrogeologic environment, and of similar depth and construction, commonly 

yield water of different water quality.

The understanding of the mechanisms affecting the 

fate and transport of agricultural chemicals in the environment is limited. A cooperative research investigation was initiated in 1986 near Plains, Ga., by the U.S. 

Geological Survey (USGS); and the U.S. Department 

of Agriculture, Agricultural Research Service (USDAARS) to (1) describe the processes that affect the 

movement and fate of nitrogen fertilizers and selected 

pesticides in the soil-root, unsaturated, and saturated 

zones; (2) develop linked mathematical models that 

describe and incorporate the process-oriented findings; 

and (3) evaluate the affect of agricultural management 

practices on the chemical quality of ground water using 

the mathematical models (Hicks and others, 199la).

Physical and chemical processes that occur as 

water flows through variably saturated or unsaturated 

porous media are not well understood at the field scale.

971
Zaslavsky and Sinai (198 la,b) suggest that the slope of 

land surface has a significant effect on the magnitude of 

horizontal components of flow in the shallow subsurface. Bakr and others (1978) indicate that, at the 

field scale, the spatial variability of the hydraulic 

properties exerts a large influence on the fate and transport of fluids in porous media. McCord and Stephens 

(1987) further evaluated the effect of surficial gradient 

on the movement of a conservative tracer (potassium 

bromide) in a homogeneous, highly porous media.

Preliminary research at the Plains, Ga., site 

indicates that nonvertical migration of soil water and 

chemicals along the interface between permeable and 

less permeable material could account for a part of the 

dispersion of agricultural chemicals in the unsaturated 

zone (Hicks and others, 1991b). Fluid-flow paths are 

commonly understood to deflect as the fluid passes 

from a medium of one hydraulic conductivity into a 

medium of a different conductivity. True one-dimensional transport in the unsaturated zone is rare and can 

only be observed where the surface soils do not have 

appreciable relief, and where the porous media is vertically homogeneous. For this reason, the one-dimensional deterministic methods which assume horizontal 

isotropic hydraulic properties that are routinely used in 

numerical simulation of fate and transport in sloping, 

structured, variably saturated porous media might not 

be valid where applied in the vast majority of field 

situations.

As a result of research at the Plains, Ga., site (Hicks 

and others, 1990,1991a,b), in 1991, new emphasis was 

placed on research to develop a better understanding of 

the processes and factors controlling the rate of movement and distribution of agricultural chemicals in the 

unsaturated zone. This paper presents preliminary findings from this evaluation of the temporal and spatial 

factors affecting the fate and transport of agricultural 

chemicals in the unsaturated zone.

DESCRIPTION OF STUDY AREA

The study area is in the Fall Line Hills district of 

the Coastal Plain physiographic province of southwestern Georgia in Sumter County (fig. 1). A 2-acre 

planting area (300 by 290.4 ft (feet)) was instrumented 

in an interstream area that separates two small watersheds. A drainage divide separating the watersheds 

passes through the northeastern corner of the planting 

area. Altitudes range from about 463 ft at the drainage 

divide, to about 456 ft in the south-southwestern part

of the planting area. Interior drainage is typical of the 

area as a result of the slight relief and sandy, highly 

permeable, surface soils (fig. 1 and table 1).

The unsaturated zone, about 26 to 30 ft thick, 

includes the undifferentiated overburden and the upper 

part of the Tallahatta Formation.The lithology of the 

undifferentiated overburden varies areally and vertically, and generally consists of alternating and intermittent layers of sand, clayey sand, and clay. The upper 

part of the Tallahatta Formation is less spatially 

variable than the undifferentiated overburden and consists mostly of fine to coarse quartz sand. However, in 

the southwestern part of the planting area, a relatively 

thin (less than 4 in. (inches) thick), discontinuous, clay 

layer is present in the Tallahatta Formation at a depth 

of about 26 to 28 ft.

METHOD OF STUDY

A 420-ft-long data-collection transect was 

established in the planting area, extending from the 

northeastern part of the planting area to the southwest, 

to evaluate the two-dimensional rate of movement and 

distribution of the bromide tracer in the unsaturated 

zone. Twenty-two sampling sites were selected along 

the transect at an equidistant spacing of 20 ft. Sampling 

sites were numbered sequentially beginning in the 

northeast at site 0 through site 21 in the southwest. Site 

0 and site 21 are not on the planting area. Site 0 was 

sampled as a control point and site 21 was not sampled 

during this investigation. The transect passes near 

existing sites D, F, J, G, and H, where data collection is 

ongoing and considerable data have been previously 

collected (fig. 1). Each of the five sites (D, F, J. G, and 

H) is equipped with nests of 2 in. diameter monitor 

wells tapping specific intervals of the saturated zone, 

and continuous soil-moisture monitors. Soil water is 

monitored using electrical, resistance-block moisture 

sensors that are installed from land surface to a depth 

of 32 ft.

Additional characterization data include bulk 

density, saturated vertical hydraulic conductivity, and 

borehole geophysical logs from selected sites (sites D, 

F, J. G, and H). Soil cores were collected for hydraulic 

characterization at five sites on the transect (sites 1, 5, 

9, 13, and 19) to a depth of about 27 ft at intervals of 1 

or 2 ft. Particle-size distribution analyses were conducted in the Georgia Agricultural Experiment Station, 

Soil Characterization Laboratory, Athens, Ga. (table 

1).

972
32°00'

31°59'30" '

EXPLANATION

Ae MONITOR WELL Letter Is

sit* identification

 SUCTION LYSIMETER 

_ PLOT BOUNDARY

Q1 QUADRANT ONE

M ... '

40 80 120 160 FEET

10 20 30 40METERS

STUDY AREA PLANTING AREA

Figure 1. Location of the study area, planting area, and monitor sites in the planting area.

Typical farming operations for a corn and wheat 

rotation have been conducted on the planting area since 

1989. During cropping, the product manufacturerrecommended applications of fertilizers and pesticides 

(atrazine, alachlor, and carbofuran) were incorporated. 

Potassium bromide salt solution (equivalent to 186 Ibs 

(pounds) of bromide) was applied at land surface as a 

conservative tracer during planting in 1989. A quantity 

of 300 Ibs of potassium bromide salt solution (equivalent to 201 Ibs of bromide) was applied during planting 

in 1991.

To evaluate the rate of migration and distribution 

of the applied chemicals, 950 soil samples were collected along the transect during seven sampling periods 

from April 1991 through July 1992. Samples were analyzed for soil-moisture content, chloride, nitrate, and 

bromide in the USGS laboratory, Atlanta, Ga. Replicate quality-assurance samples were analyzed in the 

Agricultural Research Service laboratory, Tifton, Ga.

RESULTS

Soil-characterization data and soil-core analyses 

indicate that the lithology and hydraulic characteristics 

of the unsaturated zone are heterogeneous and vary 

substantially over the transect. The northeastern part of 

the transect near sites 1-5, near the drainage divide, is 

underlain by unconsolidated sediments that are predominated by a 5 to 7 ft thick zone of dense, clayey soil 

that extends from a depth of about 5 to 12 ft below land 

surface. Silt and clay in this zone ranges from about 35 

to 58 percent of the soils matrix (table 1). In this zone, 

the vertical saturated hydraulic conductivity ranges 

from 0.02 ft/d (feet per day) at a depth of about 5 ft, to 

0.2 ft/d at a depth of 12 ft. The ratio of silt and clay to 

sand in the 5- to 12-ft- depth zone progressively 

decreases along the transect toward the southwest. At 

site 13, slightly west of the center of the transect, the 

silt and clay ranges from about 15 to 32 percent of the

973
Table 1. -Particle-size distribution along transect, Plains, Georgia 

[Analyses by Georgia Agricultural Experiment Station, Soil 

Characterization Laboratory, Athens, Ga.; -, not calculated!

Depth of 

sample 

(feet)

1

3

5

7

9

11

13

15

17

19

21

23

25

27

1

3

5

7

9

11

13

15

17

19

21

23

25

27

Very 

coarse sand 

(percent)

3.5

4.1

3.1

2.5

2.5

3.6

5.1

5.2

8.1

24

29

2.0

2.7

10.5

3.0

2.4

2.6

2.8

3.2

3.4

3.7

7.2

14.8

8.7

8.0

4.7

18.2

8.2

Coarse 

sand 

(percent)

232

20.6

166

14.9

15.8

173

21

20.1

31.5

33.1

34.1

27.7

20.8

35

25.5

20.8

14.7

16.7

18.7

173

17.1

19.1

24.6

11.9

192

11.9

30.8

30,4

Medium 

sand 

(percent)

32.9

29.3

22.8

20.1

20.2

21.3

23.4

24.8

27.3

9.7

12.3

31.1

47.2

38.6

34.5

29.2

22.7

22.7

24.4

26

25.3

25.9

26.4

35.8

35

28.9

21.7

42.8

Fine 

sand 

(percent)

Sltel

23.1

23.4

16.7

11.4

11.3

12.2

12.2

12.9

12.6

13.8

9.1

24.2

21

8.4

SiteS

21.3

23.5

21.4

14

14

16.4

15.8

16.3

17.2

26.9

19.5

36.4

12.7

12

Very 

fine sand 

(percent)

4.2

4.4

3.0

2.4

2.6

2.8

2.9

2.8

1.6

4.8

3.2

12

3.0

1.5

3.6

4.3

0.1

2.9

2.9

3.5

3.5

3.4

4.3

3.7

4.4

4.3

7.2

2.2

Total 

sand 

(percent)

86.9

81.8

622

513

52.4

512

64.6

65.8

81.1

85.4

87.7

97

94.7

94

87.9

802

61.5

59.1

632

66.6

65.4

71.9

873

87

86.1

862

90.6

95.6

Silt 

(percent)

10

10

5.6

4.6

3.2

3.9

3.4

-

-

-

~

--

--

--

8.3

9.2

5.1

5.1

4.6

4.1

4.5

-

--

--

-

-

--

..

Clay 

(percent)

3.1

82

32.2

44.1

44.4

38.9

32

-

-

--

-

-

-

--

3.8

10.6

33.4

35.8

322

293

30.1

--

--

--

-

--

--

..

Total silt 

and clay 

(percent)

13.1

182

37.8

48.7

47.6

42.8

35.4

342

18.9

14.6

123

3.0

53

6.0

12.1

19.8

58.5

40.9

46.8

33.4

34.6

28.1

12.7

13

13.9

13.8

9.4

4.4

974
Table 1. -Particle-size distribution along transect, Plains, Georgia -continued 

[Analyses by Georgia Agricultural Experiment Station, Soil 

Characterization Laboratory, Athens, Ga.; -, not calculated]

Depth of Very 

sample coarse sand 

(feet) (percent)

1

3

5

7

9

11

13

15

17

19

21

23

25

27

1

3

5

7

9

11

13

15

17

19

21

23

25

27

4.3

2.2

2.9

3.2

3.5

3.5

4.8

9.2

7.8

11.8

11.5

7.6

1.1

7.2

2.4

2.1

2.2

1.6

3.3

3.6

5.6

7.0

9.6

12.5

5.0

11

6.4

1.3

Coarse 

sand 

(percent)

27.7

19.4

211

20.8

21.1

216

21.7

305

203

27

333

21.7

14.4

29.8

211

19

17.6

iai

202

175

18.7

29.7

163

24.5

355

47.9

252

14.4

Medium 

sand 

(percent)

31.9

30.2

26.8

25.3

24.8

27.9

30.2

29.7

40.8

30.8

33.6

32.6

45.6

36.3

33

28.7

26.2

27.8

25.6

26

27.5

27.8

31.7

32.4

43.7

26.9

42.4

42.7

Fine 

sand 

(percent)

Site 9

17.9

23.6

15.6

14.7

16.1

17.8

19.8

23.6

23.5

15

11.9

19.4

25

7.2

Site 13

27.6

27.6

22.2

25.2

17.8

24.6

22.2

16.3

17.6

12.9

6.6

8.1

18.3

36.5

Very 

fine sand 

(percent)

3.6

4.2

2.8

2.8

3.0

3.1

3.2

2.2

2.6

9.6

3.8

7.3

6.7

1.6

0.1

0.1

0.1

0.1

1.9

0.1

2.9

1.7

7.2

2.6

1.3

2.2

1.5

1.8

Total 

sand Silt 

(percent) (percent)

85.4 9.4

79.6 8.8

702 4.0

66.8 5.5

68.5 10.3

74.9 9.3

79.7 15

852

95

942

94.1

88.6

92.8

82.1

852 9.1

77.5 9.0

683 6.3

72.8 5.6

69.9 7.6

71.8 4.7

76 J9 18

82.5

82.4

84.9

92.1

96.1

93.8

96.7

day 

(percent)

52

11.6

25.8

27.7

212

15.8

53

--

--

-

-

--

-

-

5.7

135

25.4

21.6

225

235

5.1

--

-

«

--

--

--

-

Total silt 

and clay 

(percent)

14.6

20.4

29.8

332

31.5

25.1

203

14.8

5.0

5.8

5J9

11.4

72

17.9

14.8

22.5

31.7

272

30.1

282

13.1

173

17.6

15.1

7.9

3.9

62

33

975
Table 1. -Particle-size distribution along transect, Plains, Georgia -continued 

[Analyses by Georgia Agricultural Experiment Station, Soil 

Characterization Laboratory, Athens, Ga.; -, not calculated]

Depth of Veiy 

sample coarse sand 

(feet) (percent)

1

3

5

7

9

11

13

15

17

19

21

23

25

27

3.0

3.5

3.7

3.0

12

2.6

6.8

11

7.0

8.7

14.5

8.2

6.5

7.3

Coarse 

sand 

(percent)

246

246

20.7

21.8

iai

15.5

22.8

33.6

21.5

242

28.7

35.9

25.6

143

Medium 

sand 

(percent)

33.2

31.3

25.4

29.5

25.9

26.3

30.3

27.6

31

34.9

19.3

38.5

36.2

11.5

Fine 

sand 

(percent)

Site 19

21.7

19.8

16

20.7

20.7

25.9

21.4

13.6

17.3

11.8

21.1

8.3

10.8

12

Very 

fine sand 

(percent)

3.7

3.0

16

3.3

3.4

4.2

16

1.5

4.5

7.0

5.3

1.5

13

1.8

Total 

sand Silt 

(percent) (percent)

86.2 4.0

82.2 7.7

68.4 5.2

783 4.9

703 4.5

74.5 4.7

83.9 3.7

873

813

86.6

88.9

92.4

81.4

37.1

Total silt 

Clay and clay 

(percent) (percent)

9.8 13.8

10.1 17.8

26.4 31.6

16.8 21.7

252 29.7

20.8 253

114 16.1

12.7

18.7

13.4

11.1

7.6

18.6

62.9

soils matrix through the upper 10 ft of the vertical soil 

section. The vertical hydraulic conductivity in this part 

of the planting area ranges from about 0.5 to 1.0 ft/d. 

At site 19, at the southwestern end of the transect, the 

percentage of clayey soils is about the same as it is at 

site 13 but substantially lower than it is at sites 1-5. 

Here, the silt and clay comprises about 14 to 30 percent 

of the soils matrix in the upper 13 ft of the vertical 

section, and the vertical hydraulic conductivity ranges 

from about 1.0 to 3.0 ft/d. The upper part of the Tallahatta Formation (about 13 to 27 ft below land surface), 

which comprises the lower part of the unsaturated 

zone, generally consists of clean, poorly sorted fine to 

coarse quartz sand. The silt and clay comprises from 

less than 3 to about 19 percent of the soils matrix, and 

the vertical hydraulic conductivity averages about 

5.0 ft/d throughout the planting area.

As a result of the marked differences in lithologic 

characteristics, the permeability and moisture-holding 

capacities along the transect are different. Figure 2

shows the relation between combined rainfall and 

irrigation, and the daily mean matric suction, at sites D, 

J, and H (near transect sites 3,13, and 17, respectively) 

at a depth of 5 ft below land surface for the period February 1 through July 18,1992. The matric suction is an 

evaluation of the energy status of the soils water; thus, 

as the soils become wetter the value of the matric suction decreases, and as the soils become dryer as a result 

of drainage and evapotranspiration, the matric suction 

increases.

The dense, clayey soils in the vicinity of site D, 

generally have the lowest vertical hydraulic conductivity of the soils in the study area. The relatively sandy 

soils downslope at sites J and H have higher permeability than soils in the upslope area near site D. Because 

of the soils fine texture (clayey) in the vicinity of site 

D, they would be expected to wet more slowly, hold 

moisture longer, and drain more slowly than the coarse 

textured (sandy) soils.

976
Cumulative rainfall and irrigation during February 

and March was about 18.7 in., which resulted in 

considerably different responses in the daily mean 

matric suction at sites D, J, and H. The observed abrupt 

changes in matric suction in the clayey soils near site D 

are somewhat anomalous (fig. 2). The daily mean 

matric suction at a depth of 5 ft at site D (clayey soils, 

low vertical hydraulic conductivity) ranged from about 

31.5 to 63 in. Heavy rainfall in late February produced 

a noticeable increase in matric suction at site D from a 

prerainfall level of 31.5 in., to a postrainfall level of 

about 63 in. within a period of a few days. At site J 

(moderate vertical hydraulic conductivity), the matric 

suction ranged from about 51.2 to 67 in. The rainfall 

did not produce a rapid fluctuation in the matric suction 

at site J and the soils apparently did not increase in wetness significantly throughout this 2-month period. The 

highly permeable soils at site H apparently wetted 

more rapidly than at site J and D, throughout this

period. The daily mean matric suction at site H ranged 

from 90.6 in. at the beginning of February to a minimum of about 29.5 in. at the end of March. The daily 

mean matric suction at sites H and J were affected 

slightly by the period of heavy rainfall that occurred in 

late February, and at site D it was affected significantly. 

The daily mean matric suction values are reasonable 

considering the significantly different hydraulic characteristics exhibited by the soils at these three sites. 

However, the intermittent anomalous response of the 

soil-water sensor at site D is not fully understood.

During the late spring and early summer in the 

Plains, Ga., area, the evapotranspiration rate commonly exceeds the rate of combined rainfall and irrigation. Because of the above-normal rainfall that 

occurred during May and June 1992, the rate of rainfall 

exceeded evapotranspiration during this period and the 

soils matrix in the planting area was much wetter than

APR T MAY 

1992

Figure 2. Relation of combined rainfall and irrigation, and daily mean matric suction at depth of 5 feet at sites J, H, and D during 

February 1 through July 18,1992, near Plains, Ga.

s

SA

§

I 

I

H

977
in a typical year. As a result, a cumulative rainfall of 5.3 

in. that occurred June 8-14 produced a marked increase 

in soil water (decrease in matric suction) at a depth of 

5 ft at sites D, J, and H (fig. 2). At sites J and H, the rainfall resulted in a decrease in matric suction of about 

19.7 to 23.6 in., followed by slow increase as the soils 

slowly lost moisture. The matric suction at these sites 

did not return to the prerainfall level until the end of 

June. At site D the June 8-14 rainfall resulted in a net 

decline in matric suction of about 31.5 in., as would be 

expected. However, the soil-water sensor indicated that 

the soils drained rapidly at site D and returned to the 

prerainfall condition within a few days after the rainfall 

ceased, and then wetted once again (fig. 2). The low 

vertical hydraulic conductivity of the soil matrix 

(clayey soils) in the vicinity of site D would prevent the 

rapid rate of vertical transport necessary to produce the 

apparent rapid increase and decrease in soil-water content. Macropore vertical transport and lateral flow 

along the interface of permeable and less permeable 

soil could account for the apparent rapid drainage of 

the soils above the 5-ft moisture sensor. The anomalous 

changes could also have resulted from sensor malfunctions caused by changes in electrical currents. Rainstorms generally have electrical fields associated with 

them. If the wetting front reaching the clayey soils at 

site D carried a current, the electrical potential relative 

to ground could cause a temporary malfunction of the 

moisture sensor resulting in anomalous readings (R.K. 

Hubbard, USDA-ARS, written commun., 1993).

Soil samples were collected at 11 sites along the 

transect on May 16,1991 (30 days after the second bromide application), to a depth of 10 ft below land surface to evaluate the rate of movement and the 

distribution of the tracer in the upper part of the unsaturated zone. The intent of this sampling was to evaluate 

the movement of the second bromide application (April 

1991), even though remnants of the initial application 

remained in the unsaturated zone. The samples were 

analyzed for soil water, nitrate, chloride, and bromide. 

The results of the bromide analyses are shown as lines 

of equal concentration on a cross-sectional graph of the 

transect (fig. 3). Site 1 is in the northeastern corner of 

the planting area and at the northeastern end of the 

transect. Site 19 is at the southwestern corner of the 

planting area. The vertical distribution of bromide in 

the unsaturated zone indicates the presence of a large 

mass of bromide at a depth of about 5 to 6 ft in the 

vicinity of sites 3-9. At site 13, the top of the bromide 

plume (interpreted to be the 1 -mg/L line) was deeper in

the unsaturated zone than elsewhere in the transect. 

The highest concentration of bromide at this site was at 

a depth of about 9 ft. At sites 15-19, two zones of relatively high bromide concentrations were detected in the 

unsaturated zone. The deeper zone of high bromide 

concentration was probably a remnant from the initial 

application in 1989. Soil samples collected from May 

16, 1991, through February 26, 1992, indicate a progressive decrease in the total bromide concentration in 

the unsaturated zone. The distribution of bromide concentration in the unsaturated zone on February 26, 

1992, is shown in figure 4. Total bromide concentration 

was significantly reduced at sites 6-10, but increased at 

site 11. Also, the zone of highest concentrations in the 

area of sites 1 -5 was at a depth of about 5 to 6 ft; indicating that the tracer had migrated vertically only about 

1 ft over a 9-month period. During this time period, the 

zone of highest bromide concentration at site 13 had 

migrated from a depth of about 9 ft to a depth of about 

12 to 14 ft below land surface. Therefore, at site 13 bromide from the 1991 application migrated at a faster rate 

than at the other sites and continued to be detected at 

the deepest point in the unsaturated zone along the 

transect.

SUMMARY AND CONCLUSIONS

A cooperative research investigation was initiated 

in 1986 near Plains, Ga., by the U.S. Geological Survey; and the U.S. Department of Agriculture, Agricultural Research Service to (1) describe the processes 

that affect the movement and fate of nitrogen fertilizers 

and selected pesticides in the soil-root, unsaturated, 

and saturated zones; (2) develop linked mathematical 

models that describe and incorporate the process-oriented findings; and (3) evaluate the affect of agricultural management practices on the chemical quality of 

ground water using the mathematical models. Physical 

and chemical processes that occur as water flows 

through variably saturated or unsaturated porous media 

is not well understood at the field scale. Previous 

research at Plains indicates that lateral (nonvertical) 

migration of soil water and chemicals along the interface between permeable and less permeable material 

could account for a part of the dispersion of agricultural 

chemicals in the unsaturated zone.

The study area is in the Fall Line Hills district of 

the Coastal Plain physiographic province of southwestern Georgia in Sumter County. A 2-acre planting area 

(300 by 290.4 ft) was instrumented in an interstream

978
SiteH 

19 I 17 15 13

SAMPLING SITES 

11 9

SiteD

w

O 10

11

40 80 120 160 200 240 280 320 

DISTANCE, IN FEET BETWEEN SAMPLING POINTS

EXPLANATION

__ 5 _ LINE OF EQUAL BROMIDE CONCENTRATION -- 

Interval 2 milligrams per kilogram

Figure 3. Distribution of bromide concentration in the unsaturated zone beneath transect A-A, May 16,1991.

360

area that separates two small watersheds. The unsaturated zone includes the undifferentiated overburden and 

the upper part of the Tallahatta Formation and is about 

26 to 30 ft thick. The lithology of the undifferentiated 

overburden differs areally and vertically. The upper 

part of the Tallahatta Formation is less spatially 

variable than the undifferentiated overburden and in 

general consists of fine to coarse quartz sand.

A 420-ft-long data-collection transect was established in the planting area, extending from the northeastern part of the planting area to the southwest, to 

evaluate the rate of movement and two-dimensional 

distribution of a bromide tracer in the unsaturated zone. 

Twenty-two sampling sites were selected along the 

transect at an equidistant spacing of 20 ft. Soil water 

was monitored using electrical, resistance-block 

moisture sensors that extend from land surface to a 

depth of 32 ft.

Typical farming operations for a corn and wheat 

rotation have been conducted on the planting area since 

1989. During cropping, the manufacturer recommended applications of fertilizers, pesticides (atrazine, 

alachlor, and carbofuran) and a conservative bromide 

tracer were applied at land surface. Potassium bromide 

salt solution (equivalent to 186 Ibs of bromide) was 

applied as a conservative tracer during planting in 

1989. A quantity of 300 Ibs of potassium bromide salt 

solution (equivalent to 201 Ibs of bromide) was applied 

during planting in 1991.

To evaluate the rate of migration and distribution of 

the applied chemicals, 950 soil samples were collected 

along the transect during seven sampling periods from 

April 1991 through July 1992. Soil-characterization data 

and soil-core analyses indicate that the lithology and 

hydraulic characteristics of the unsaturated zone are 

heterogeneous and differ substantially over the transect.

979
A

SiteH

19 | 17

8

10

£ n

OO

J ^

8 16

5 18

Q 20

22

24

26

28

30

31

15 13

SAMPLING SITES 

11 9

A'

 5 LINE OF EQUAL BROMIDE CONCENTRATION 

Interval 2 milligrams per kilogram

40 80 120 160 200 240 280 

DISTANCE, IN FEET BETWEEN SAMPLING POINTS

SiteD 

i 

3

320 360

Figure 4. Distribution of bromide concentration in the unsaturated zone beneath transect A-A, 

February 26,1992.

980
The northeastern part of the transect, near a drainage 

divide, is underlain by unconsolidated sediments that 

are predominated by a 5 to 7 ft thick zone of dense, 

clayey soil comprising about 35 to 58 percent silt and 

clay. In this zone, the vertical saturated hydraulic conductivity ranges from 0.02 ft/d at a depth of about 5 ft, 

to 0.2 ft/d at a depth of 12 ft. The ratio of silt and clay 

to sand in the 5- to 12-ft-depth zone progressively 

decreases along the transect toward the southwest. At 

site 13, slightly west of the center of the transect, the 

silt and clay comprises about 15 to 32 percent of the 

soil matrix, and the saturated vertical hydraulic conductivity ranges from about 0.5 to 1.0 ft/d. At site 19, at the 

southwestern end of the transect, the silt and clay comprises about 14 to 30 percent of the soil matrix, and the 

saturated vertical hydraulic conductivity ranges from 

1.0 to 3.0 ft/d in the upper 12 ft of the unsaturated zone. 

The upper part of the Tallahatta Formation (about 17 to 

27 ft below land surface), which comprises the lower 

part of the unsaturated zone, generally consists of clean, 

poorly sorted fine to coarse quartz sand, and the vertical hydraulic conductivity averages about 5.0 ft/d 

throughout the planting area.

As a result of the differences in hydraulic characteristics, the infiltration and moisture-holding capacities of each area of the transect differ. Daily mean 

matric suction, at sites D, J, and H at a depth of 5 ft 

below land surface for the period February 1 through 

July 18, 1992, indicates that the apparent energy status 

of the soil water varied substantially. Cumulative 

rainfall and irrigation during February and March was 

about 18.7 in., which resulted in considerably different 

changes in the daily mean matric suction at sites D, J, 

and H at a depth of 5 ft for the period.

Cumulative rainfall and irrigation during February 

and March was about 18.7 in. and in June was 5.3 in., 

which resulted in considerably different responses in 

the daily mean matric suction at sites D, J, and H. The 

observed abrupt changes in matric suction in the clayey 

soils near site D are somewhat anomalous. The daily 

mean matric suction values are reasonable considering 

the significantly different hydraulic characteristics 

exhibited by the soils at these three sites. However, the 

intermittent anomalous response of the soil-water sensor at site D is not fully understood. The low vertical 

hydraulic conductivity of the soil matrix (clayey soils) 

in the vicinity of site D would prevent the rapid rate of 

vertical transport necessary to produce the apparent 

rapid increase and decrease in soil-water content. 

Macropore vertical transport and lateral flow along the

interface of permeable and less permeable soil could 

account for the apparent rapid drainage of the soils 

above the 5-ft moisture sensor. The anomalous changes 

could also have resulted from sensor malfunctions 

caused by changes in electrical currents. Rainstorms 

generally have electrical fields associated with them. If 

the wetting front reaching the clayey soils at site D carried a current, the electrical potential relative to ground 

could cause a temporary malfunction of the moisture 

sensor resulting in anomalous readings

Soil samples were collected at 11 sites along the 

transect on May 16, 1991 (30 days after the second 

bromide application), to a depth of 10 ft below land 

surface. Vertical distribution of bromide in the unsaturated zone indicates that a large mass of the tracer was 

concentrated at a depth of about 5 to 6 ft near sites 6-10. 

At site 13, the upper boundary of the mappable 

bromide plume is deeper in the unsaturated zone than it 

is elsewhere in the transect. Soil samples collected 

from May 16, 1991, through February 26, 1992, indicate a progressive decrease in the total bromide mass in 

the unsaturated zone. Samples collected on February 

26, 1992, indicate that the total bromide mass significantly decreased at sites 6-10 but increased at site 11.

REFERENCES

Bakr, A.A., Gelhar, A.L., Gutjar, A.I., and MacMillan, R.J., 

1978, Stochastic analysis of spatial variability in 

subsurface flows, 1 Comparison of one- and threedimensional flows: Water Resources Research., v. 3, 

p. 512-519.

Bear, J., Braester, C., and Menier, P., 1987, Effective and 

relative permeabilities of anisotropic porous media: 

Transport in Porous Media, v. 2, p. 301-316.

Hicks, D.W., McConnell, J.B., and Asmussen, L.E., 1990, 

Preliminary geologic and hydrologic evaluation of a 

small watershed near Plains, Georgia: U.S. Geological 

Survey, Water-Resources Investigations Report 

90-4146, 30 p.

Hicks, D.W, McConnell, J.B., Asmussen, L.E., Leonard, 

R.A., and Smith, C.N., 199la, Movement and fate of 

agricultural chemicals in the surface and subsurface 

environments near Plains, southwestern Georgia-- 

Integrated work plan: U.S. Geological Survey, OpenFile Report 91-73, 26 p.

Hicks, D.W, McConnell, J.B., and Hubbard, R.K., 1991b, 

Effect of the unsaturated zone on the migration and 

retardation of a conservative tracer at the Plains, 

Georgia, research site: U.S. Geological Survey, WaterResources Investigations Report 91-4034, p. 597-600.

981
McCord, J.T., and Stephens, D.B., 1987, Analysis of soil- Zaslavsky, D., and Sinai, G., 198 Ib, Surface hydrology: III. 

water movement on a sandy hillslope in the apparent Causes of lateral flow: American Society of Civil Engiabsence of an impeding layer: Journal of Hydrologic neers, Journal of Hydraulics Division, v. 107, p 17-30. 

Processes, v. I, no. 3, p. 225-228.

Zaslavsky, D., and Sinai, G., 198la, Surface hydrology: I. 

Explanation of phenomena: Division American Society 

of Civil Engineers, Journal of Hydraulics, v. 107, 

p. 116.

982
Spatial Variability of Soil Properties Affecting Chemical 

Transport at the Plains, Georgia, Research Site

By R. K. Hubbard 1 , D. D. Bosch 1 , R. A. Leonard 1 , andD. W. Hicks2

Abstract

A study was conducted on a 0.81-ha site 

near Plains, Georgia, to determine transport rates 

of water and applied agricultural chemicals to 

ground water in a recharge area of the Coastal 

Plain. As part of the study, soil cores were 

collected from the root and vadose zone to a 

depth of 914 cm (30 ft). Cores were collected 

from 12 locations within the site to determine 

spatial variability of the soil physical properties. Measurements on core samples were made 

of particle size distribution, vertical saturated 

hydraulic conductivity, bulk density, and soil 

moisture retention. Measured soil moisture 

retention was compared with soil moisture retention predicted from equations based on particle 

size distribution and bulk density. Analyses of 

the soil samples showed that the physical properties varied both vertically and laterally. The soil 

has a sandy texture at the surface, has increased 

clay content from 91 to 366 cm (3 to 12 ft) deep, 

and then is very sandy to a depth of 914 cm 

(30 ft). Laterally, the plot varies in clay content 

in the 91 to 366 cm (3 to 12 ft) depth interval 

from the east to the west side, with the northeastern corner having the highest clay content. 

Lower vertical saturated hydraulic conductivities 

and higher soil-moisture retention and bulk densities are associated with increasing clay content.

1 U.S.Department of Agriculture, Agricultural Research 

Service, Southeast Watershed Research Laboratory, P.O. Box 946, 

Tifton,GA 31793.

2U.S. Geological Survey, 3039 Amwiler Road, Atlanta, GA 

30360-2824.

The comparisons between measured and predicted volumetric-water content show that the 

equations adequately predict the relative magnitude of the volumetric-water content, but that 

absolute predictions lack precision.

INTRODUCTION

Chemical transport to ground water is of 

concern both economically and environmentally. 

Loss of nutrients and pesticides from the root zone 

requires replacement by the farmer at additional cost. 

Agricultural chemicals in ground water are an 

environmental concern when concentrations exceed 

drinking-water standards.

The primary nutrient of concern in ground water 

is nitrate (NO3 -N). Nitrate is a soluble chemical 

which moves freely with water. Methemoglobinemia, or blue-baby disease, can occur in infants who 

drink water with NC^-N concentrations in excess of 

10 mg/L (Federal Register, 1975). Nitrate 

concentrations greater than 10 mg/L have been 

documented in ground water associated with 

agricultural activities in a number of locations 

(Hubbard and Sheridan, 1989).

Pesticides are of concern in ground water 

primarily because of possible carcinogenic effects. As 

of 1986, at least 17 pesticides had been reported in 

ground water (Cohen and others, 1986). It is now 

known that intrusion of low concentrations of 

pesticides into ground water occurs when certain 

combinations of conditions coincide, as, for example, 

when pesticides with high mobility and relatively long 

half-lives are applied in aquifer recharge areas 

(Leonard and others, 1988).

The U.S. Geological Survey, U.S. Department 

of Agriculture, Agricultural Research Service, and 

U.S. Environmental Protection Agency initiated a

983
cooperative research investigation in 1986 to develop 

an effective model(s) of agrichemical transport to 

ground water, as well as to design and instrument a 

field research site to provide data necessary for the 

validation of the model(s). The major objectives of 

this investigation were to (1) describe the processes 

that affect the transport and fate of agrichemicals in 

the soil-root, unsaturated, and saturated zones; 

(2) develop and validate a linkedmodel(s) that 

mathematically describes the process-oriented 

findings; and (3) use the model(s) as a tool to evaluate 

the effect of agricultural management practices on the 

chemical quality of ground water.

This paper presents information on the physical 

properties of a research site with emphasis on the 

spatial variability of particle-size distribution and 

saturated hydraulic conductivity. Comparisons of soilmoisture retention as measured at selected sites with 

those predicted by equations from Rawls and others 

(1983) also are presented. The equations of Rawls and 

others (1983) were developed from root-zone soil 

samples, whereas the data from Plains on which the 

equations were tested included samples from depths 

up to 914 cm (30 ft). This comparison, thus, evaluated 

the effectiveness of the Rawls equations on both 

surface and subsurface horizons of a sandy Coastal 

Plain soil.

MATERIALS AND METHODS

The initial study area selected in 1986 included 

the upper Ty Ty Creek tributary watershed near 

Plains in Sumter County, Georgia. The area occupies 

2.67 km and is typical in relief and soil properties of 

watersheds in the Fall Line Hills district of the Coastal 

Plain physiographic province of southwestern Georgia 

(Hicks and others, 199la). In 1989, an area was added 

at the southern part of the initial study area and the 

scale of research was reduced. A 0.81-ha plot was 

selected for water and solute transport studies within 

this new area. The plot is located within a 6.76-ha 

field consisting of soil classified as Eustis sand 

(Psammentic Paleudult).

Twelve sites (A-L) within the plot area were 

selected for soil characterization (fig. 1). Three sites 

were located randomly within each of the four 

quadrants of the plot. Soil samples were collected at 

each site by coring with hollow-stem augering 

equipment. Samples were collected to a depth of 

914 cm (30 ft) using 6.1 cm i.d. plastic inserts inside

a soil probe contained within a hollow-stem auger. 

Cores were collected in sections of 91, 76 or 61 cm 

(3, 2.5, or 2 ft). In the laboratory, the cores were 

removed from the plastic sleeves and sectioned to 

provide subsamples for determination of vertical 

saturated hydraulic conductivity, particle-size 

distribution, soil- moisture retention, and bulk density.

A flexible-membrane permeameter (Daniel and 

others, 1984) was used to determine vertical saturated 

hydraulic conductivity on minimally disturbed core 

sections from the 12 sites. Particle-size analyses were 

done on samples from all 12 sites using the 

hydrometer method (Day, 1965). Soil-moisture 

retention from -0.01 to -1 bar was measured on 

selected subsamples using tempe cells (Klute, 1986; 

Reginato and van Bavel, 1962). Bulk-density 

measurements were made on the same samples used 

for soil-moisture retention measurements.

A comparison of measured soil-moisture 

retention and that predicted by equations from Rawls 

and others (1983) was made. Rawls presents linear 

regression coefficients for predicting volumetric soilwater contents at matric potentials ranging from -0.2 

to -15 bars. Volumetric water content is predicted as a 

function of relative percentages of sand, clay, and 

organic matter, and bulk density.

The coefficients for predicted volumetricmoisture content at -0.20, -0.33, -0.60, and -1.00 bar 

matric suction were used in the linear regression 

equations of Rawls and others (1983) to generate 

values for each of the samples where moisture 

retention was measured. Input to the equations 

included measured values for percentages of sand and 

clay, and bulk density. Organic-matter contents were 

assumed to be 1.0 percent for depths of 0 to 61 cm 

(0 to 2 ft), 0.5 percent for depths of 61 to 152 cm (2 to 

5 ft), and 0.1 percent for depths of 152 to 914 cm (5 to 

30 ft). From the relation of predicted volumetricwater content to matric suction, equations for this 

relation were generated for each sample. Because the 

relation between matric potential and volumetricwater content is log-linear, a log transformation of the 

matric potential was used prior to generating the 

equations by linear regression (SAS, 1989a, 1989b). 

From the predictive equations, volumetric water 

contents at -0.01, -0.30, and -1.00 bars were calculated 

for each sample. The predicted values were then 

compared with the measured values using linear 

regression (SAS, 1989a, 1989b).

984
35' 

33*

3V 

FALL LINE HILLS 

DISTRICT

COASTAL PLAIN 

PHYSIOGRAPHIC PROVINCE

STUDY SITE

100

t

200 

i

300

i

KILOMETERS

r 

100 200 MILES

EXPLANATION

A» SAMPLING SITE

AND IDENTIFICATION

(5l) QUADRANTS

50 100 FEET -f15 30 METERS

PLOT 

AREA

FIELD 

BOUNDARY

FIELD SITE

A PL°T

B

C ff\

@

G

H .

1

AREA

6

E

^^(02) 

F

 J

©

K

 L

Figure 1. Plains, Ga., research site.

985
RESULTS

Samples from the 12 sites showed variability in 

particle-size distribution both vertically and laterally. 

In the vertical direction the soil is sandy near the 

surface, has increased clay content from 

approximately 91 to 366 cm (3 to 12 ft) depth and then 

is very sandy to a depth of 914 cm (30 ft). At some 

sites, narrow clay bands were observed/penetrated 

within the 762 to 914 cm (25 to 30 ft) depth interval. 

Laterally, the variability of the plot is primarily related 

to the percentage of clay in the 91 to 457 cm (3 to 

15 ft) depth interval, and to the thickness of clayey 

material.

Particle-size distribution, saturated hydraulic 

conductivity, bulk density, and volumetric-water 

content at -0.01, -0.30, and -1.00 bar matric potential 

by depth for sites A and L are shown in tables 1 and 

2. The data from these two sites (A and L) are 

representative of the range in properties measured at 

the plot. At all sites, vertical variability in particlesize distribution was primarily in the relative 

percentages of sand and clay; increases in clay content 

were offset by decreases in sand content.

At site A, the soil-surface zone contains 

9.1 percent clay (table 1). From 61 to 213 cm (2 to 

7 ft) depth, the clay content increases to 18 to 

21 percent, and from 213 to 305 cm (7 to 10 ft) depth, 

the clay content ranges from 26 to 27 percent. Clay 

content decreases somewhat from a depth of 305 to 

457 cm (10 to 15 ft) (from 21 to 12 percent) and below 

a depth of 457 cm (15 ft) the soil is very sandy except 

for a narrow clay band 762 to 792 cm (25 to 26 ft) 

deep.

The vertical saturated hydraulic conductivity at 

site A ranges from 1.79 to 5.73 cm/hr. In general, the 

vertical saturated hydraulic conductivity is lower in 

the clayey zone than it is in the underlying sandy zone. 

The bulk density ranges from 1.51 to 1.81 g/cm3 and is 

greater in the zone with the highest clay content. Soilmoisture retention at matric potentials of-0.01, -0.30, 

and -1.00 bar also reflects the effect of clay content. 

Volumetric-water content is generally higher at each 

matric potential from depths of 107 to 411 cm (3.5 to 

13.5 ft) than in the sandier material deeper in the 

pedon.

Surface-soil texture at site L is similar to that at 

site A. At depths of 107 to 335 cm (3.5 to 11 ft), clay 

content averages 30 percent. In contrast, the average 

clay content is 22.1 percent at site A in the same depth 

range. Clay content decreases at site L from 335 to

457 cm (11 to 15 ft) deep and the soil is very sandy 

from 457 to 914 cm (15 to 30 ft) deep.

The vertical saturated hydraulic conductivity at 

site L is notably different from that at site A in the 

zone of higher clay content. The average vertical 

saturated hydraulic conductivities are 0.41 cm/hr at 

site L and 2.54 cm/hr at site A in the 107 to 366 cm 

(3.5 to 12 ft) depth interval. Similarly to site A, the 

bulk density at site L is greatest in the zone with the 

greatest clay content. Volumetric-water content at site 

L also is greatest where the clay content is the greatest.

The distributions of clay and vertical saturated 

hydraulic conductivities in the 30 to 61, 259 to 305, 

457 to 518, and 792 to 823 cm (1 to 2, 8.5 to 10, 15 to 

17, and 26 to 27 ft) depth intervals are shown in figure 

2. These four depth intervals represent properties in 

the surface zone, zone of greatest clay content, and 

upper and lower zones of the 457 to 914 cm (15 to 

30 ft) deep material, respectively. The greatest clay 

content in the 30 to 61 cm (1 to 2 ft) depth interval is 

in the northwestern corner of the plot. Clay content 

averages 10 to 12 percent elsewhere. Vertical 

saturated hydraulic conductivity is variable in the 

30 to 61 cm (1 to 2 ft) depth interval, being lowest in 

the central section of the plot and increasing towards 

both the south and the west.

Clay content in the 259 to 305 cm (8.5 to 10 ft) 

depth interval varies from 38 percent in the northeastern corner to 20 percent in the southwestern 

corner. Ground-penetrating radar patterns of the plot 

which were made prior to instrumentation indicated 

possible differences in clay content between the 

eastern and western sides of the plot. Soil samples 

from the 12 sites clearly show that clay content is 

greater on the eastern than on the western side, with 

the northeast corner having the greatest clay content in 

the 91 to 366 cm (3-12 ft) depth interval. Difficulties 

in obtaining soil samples from the northeastern 

quadrant by bucket augering are indicative of the high 

clay content. The vertical saturated hydraulic 

conductivity in the 259 to 305 cm (8.5-10 ft) depth 

interval is lower on the east side of the plot than on the 

west side.

Contour plots of the clay content at depths of 

457 to 518 cm (15 to 17 ft) and 792 to 823 cm (26 to 

27 ft) indicate that, at these depths, the texture is 

relatively uniform over the plot, and the soil is very 

sandy. In the 457 to 518 cm (15-17 ft) depth interval, 

the clay content ranges from 7 to 13 percent, whereas 

in the 792 to 823 cm (26 to 27 ft) depth interval, the

986
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987
Table 2 Physical properties of soils at Site L, Plains, Georgia

[cm, centimeters; ft, feet; cm/hr, centimeters per hour; g/cms , grams per cubic centimeter; cm'/cm', cubic 

measured]

Depth

(cm)

0-30

30-61

61-107

107-152

152-183

183-213

213-259

259-305

305-335

335-366

366-411

411-457

457-488

488-518

518-564

564-610

610-640

640-671

671-716

716-762

762-792

792-823

823-869

869-914

Interval

(ft)

0-1

1-2

2-3.5

3.5-5

5-6

6-7

7-8.5

8.5-10

10-11

11-12

12-13.5

13.5-15

15-16

16-17

17-18.5

18.5-20

20-21

21-22

22-23.5

23.5-25

25-26

26-27

27-28.5

28.5-30

Sand 

(percent)

88.4

76.7

75.6

61.4

66.5

64.1

65.5

65.9

70.7

76.8

79.8

84.9

87.2

91.9

92.7

91.2

97.5

97.5

98.4

93.9

87.8

96.0

97.9

97.6

Silt 

(percent)

7.8

10.0

10.4

3.6

6.1

3.9

3.5

5.9

2.8

3.1

4.5

2.1

4.5

3.7

1.3

3.8

2.0

2.3

1.6

4.6

5.4

1.7

1.1

1.0

Clay 

(percent)

3.8

13.3

14.0

35.0

27.4

32.0

31.0

28.2

26.5

20.1

15.7

13.0

8.3

4.4

6.0

5.0

0.5

0.2

1.5

6.8

2.3

1.0

1.4

Vertical 

saturated 

hydraulic 

conductivity 

(cm/hr)

4.18

3.20

2.85

0.48

.50

.24

.46

.57

.47

.13

2.19

2.23

2.61

4.02

2.26

3.16

3.37

1.84

4.97

2.44

5.49

5.41

5.10

Bulk 

density 

(g/cms )

1.65

1.65

1.40

1.73

1.65

1.70

1.74

1.80

1.46

1.74

1.61

1.61

1.74

1.69

1.48

1.48

1.63

1.59

1.59

1.69

1.60

1.73
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DEPTH DEPTH

(457 TO 518 CENTIMETERS. 15 TO 17 FEET) (457 TO 518 CENTIMETERS. 15 TO 17 FEET)

100

| 75

50

o

z

25

100

75

z

50

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0 25 50 75 100 0 25 50 75 100

DISTANCE EAST, IN METERS DISTANCE EAST, IN METERS

DEPTH DEPTH

(792 TO 823 CENTIMETERS. 26 TO 27 FEET) (792 TO 823 CENTIMETERS, 26 TO 27 FEET)

100 i i -r PC- , , 100

75

50

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25

25 50 75 

DISTANCE EAST, IN METERS

100 0 

(PART B)

-28-

EXPLANATION

UNE OF EQUAL CLAY CONTENT. 

IN PERCENT. CONTOUR INTERVAL. 

1 PERCENT.

 2.5-

25 50 75 100 

DISTANCE EAST. IN METERS

EXPLANATION

UNE OF EQUAL VERTICAL SATURATED 

HYDRAULIC CONDUCTIVITY, IN CENTIMETERS 

PER HOUR. CONTOUR INTERVAL. 0.5 

CENTIMETERS PER HOUR.

Figure 2b. Cloy content and vertical saturated hydraulic conductivity at depths 

of 457 to 518 centimeters (15 to 17 feet) and 792 to 823 

centimeters (26 to 27 feet).

990
clay content ranges from 2 to 9 percent; indicating that 

the sand content of the soil increases with depth in the 

457 to 914 cm (15 to 30 ft) depth interval. Vertical saturated hydraulic conductivity in the 457 to 518 cm 

(15to 17 ft) depth interval is lower in the northeastern 

quadrant than it is elsewhere in the plot. At depths of 

792 to 823 cm (26 to 27 ft) the vertical saturated 

hydraulic conductivity ranges from 1 to more than 

6 cm/hr, with the northwestern quadrant having the 

lowest conductivity.

Physical-property data from the plot indicate 

potential for variable rates of water and solute 

movement both vertically and laterally. Observations by Hicks and others (1991b) indicated that 

solutes were moving laterally through the plot from 

the northeast to the southwest. It would appear that 

solutes applied in the northeastern corner of the plot 

were impeded vertically by the clay layer and, thus, 

moved laterally towards the southwest in subsurface 

flow. Because the eastern side of the plot generally 

has higher clay content than the western side, it can be 

hypothesized that solutes move from the eastern side 

in a southwesterly direction along a series of parallel 

flow lines. Solutes were observed at greater depth and 

also in higher concentrations in the southwestern 

corner than in the northeastern corner of the plot.

Predicted versus measured soil-water contents 

at -.01,-.30, and -1.00 bar matric suction are shown in 

figure 3. The coefficients of determination (R2) for the 

relationship between predicted and measured water 

content at these three matric suctions were 0.63, 0.66, 

and 0.61, respectively. Separation of the data 

according to soil depth (root zone or subsoil) or 

texture (sandy or clayey) did not improve the 

coefficients of determination. The coefficients of 

determination indicate that the Rawls equation is 

generally useful as a relative predictor, but not an 

absolute predictor of volumetric-water content.

CONCLUSIONS

Analyses of soil samples collected from 

12 sites at the Plains, Georgia, research site showed 

that physical properties of the soil vary both vertically 

and laterally. The soil has a sandy texture at the 

surface, has increased clay content from 91 to 366 cm 

(3 to 12 ft) deep, and is then very sandy to a depth of 

914 cm (30 ft). Laterally, the clay content of the plot 

in the 91 to 366 cm (3 to 12 ft) depth interval is 

variable from east to west, with the northeastern

30

25

20

15

10

5

30

25

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2530 

tu

25

a: Si 

£§20

a: i- i«t

t

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5110 

1§ 5

si

5 10 15 20 25 30 35 40

MEASURED VOLUMETRIC WATER CONTENT 

IN CUBIC CENTIMETERS PER CUBIC CENTIMETER

5 10 15 20 25 30 35 40

MEASURED VOLUMETRIC WATER CONTENT 

IN CUBIC CENTIMETERS PER CUBIC CENTIMETER

45

5 10 15 20 25 30 35 40 45

MEASURED VOLUMETRIC WATER CONTENT 

IN CUBIC CENTIMETERS PER CUBIC CENTIMETER

Figure 3. Plots of predicted and measured water content at 

-0.01 (top graph), -0.30 (middle), and -1.0 (bottom) bar 

matric suction.

991
corner having the greatest clay content. Vertical 

saturated hydraulic conductivity decreases and soilmoisture retention and bulk density increase with 

increasing clay content. Comparisons of measured 

volumetric-water content with those predicted by 

the equations of Rawls and others (1983) showed 

that the predictive equations were generally useful. 

The spatial variability in soil physical properties at the 

site is consistent with observed patterns of chemical 

movement.

REFERENCES

Cohen, S.Z., Eiden, C., and Lorber, M.N., 1986, Monitoring 

ground water for pesticides in the USA, in Evaluation 

of Pesticides in Ground water, Garner, W.Y., Honeycutt, R.C., and Nigg, H.N. eds.: Washington, D.C., 

American Chemical Society Symposium Series No. 

315, p. 170-196.

Daniel, D.E., Troutwein, S.S., Boyton, S.S., and Foreman, 

D.E., 1984, Permeability testing with flexible-wall 

permeameters: Geotechnical Testing Journal, 

GTJODJ, v. 7, p. 113-122.

Day, P.R., 1965, Particle fractionation and particle-size 

analysis, pt. I o/Black, C.A. ed., Methods of soil 

analysis: Agronomy, v. 9, p. 545-567.

Federal Register, 1975, National interim primary drinking 

water standards: Federal Register, v. 40, p. 59,566-59, 

588.

Hicks, D.W., McConnell, J.B., Asmussen, L.E., Leonard, 

R.A., and Smith, C.N.,1991a, Movement and fate of 

agricultural chemicals in the surface and subsurface 

environments near Plains, Southwestern Georgia 

Integrated Work Plan: U.S. Geological Survey OpenFile Report 91-73.

Hicks, D.W., McConnell, J.B., and Hubbard, R.K., 1991b, 

Effect of the unsaturated zone on the migration and 

retardation of a conservative tracer at the Plains, 

Georgia, Research Site, in Mallard, G.E., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Monterey, Calif, March 11-15, 1991: U.S. 

Geological Survey, Water Resources Investigations 

Report 91-4034,, p. 597-600.

Hubbard, R.K., and Sheridan, J.M., 1989, Nitrate movement 

to ground water in the southeastern Coastal Plain: 

Journal of Soil and Water Conservation, v. 44, 

p. 20-27.

Klute, A., 1986, Water Retention: Laboratory Methods, 

Pt 1 o/Klute, A., Ed., Methods of Soil Analyses 

Physical and Mineralogical Methods: Agronomy 

Monograph, no. 9. Chap. 26, p. 635-662.

Leonard, R.A., Shirmohammadi, A., Johnson, A.W., and 

Marti, L.R., 1988, Pesticide transport in shallow 

ground water: Transactions of the American Society 

of Agricultural Engineers, v. 31, p. 776-788.

Rawls, W.J., Brakensiek, D.L., and Soni, B., 1983, Agricultural management effects on soil water processes, pt. I 

of Soil water retention and Green and Ampt infiltration 

parameters: Transactions of the American Society of 

Agricultural Engineers, v. 26, p. 1747-1752.

Reginato, R.J., and van Bavel, C.H.M., 1962, Pressure 

cell for soil cores: Soil Science Society of America, 

Proceedings, v. 26, p. 1-3.

SAS Institute, Inc., 1989a, SAS/STAT User's Guide, (4th 

ed.): Cary, NC, SAS Institute, Inc., version 6, v. 1, 

943 p.

SAS Institute, Inc., 1989b, SAS/STAT User's Guide, (4th 

ed.): Cary, NC., SAS Institute, Inc., version 6, v. 2, 

846 p.

992
Modeling Statistical Relations Among Shallow

Ground-Water Quality, Human Activities, Land Use,

and Thickness of the Unsaturated Zone on Long Island,

New York

By Paul E. Stackelberg1 and David A.V. Eckhardt2

Abstract

Water-quality data from 90 shallow observation wells in five areas of differing land use in 

Nassau and Suffolk Counties were used to 

develop statistical models that relate shallow 

ground-water quality to variables representing 

human activities and hydrogeologic conditions. 

The five study areas represent the full range of 

land uses in the two-county area and lie along the 

regional ground-water divide, where groundwater has a vertically downward component and 

thus recharges deep zones of the aquifer system.

Maximum-likelihood logistic-regression 

analysis of explanatory variables that describe the 

degree and type of human activities and hydrogeologic conditions at the 90 well sites was used 

to develop statistical models that can predict the 

probability of the presence of contaminants in 

shallow ground water within the two-county area. 

Variables found useful for representing human 

activities and hydrogeologic conditions include 

(1) population density within a 1/2-mile radius of 

each well, (2) percentages of land-use categories 

within the same area, and (3) thickness of the 

unsaturated zone at each well site. One model 

that relates the probability of the presence of volatile organic compounds (VOC's) at concentrations 

of at least 1 microgram per liter to commercial 

and high- and medium-density residential land

'U.S. Geological Survey, West Trenton, New Jersey 

2U.S. Geological Survey, Ithaca, N.Y.

was selected to illustrate model application and 

testing procedures. Resulting predictions were 

compared with an independent set of waterquality data from 378 wells in Nassau and Suffolk 

Counties. The comparison indicates that VOC's 

are more prevalent in areas with relatively little 

commercial and dense residential land than 

predicted by data from the five study areas.

INTRODUCTION

Long Island's aquifer system is one of 14 areas 

being studied as part of the U.S. Geological Survey's 

Toxic Substances Hydrology Program to determine 

relations among the chemical quality of shallow 

ground water and variables that represent human 

activities and local hydrogeologic conditions. Long 

Island's aquifer system was selected because (1) large 

amounts of water-quality and land-use data are available; (2) the aquifer system, which provides potable 

water for 2.6 million inhabitants of Nassau and 

Suffolk Counties, has been designated a sole-source 

aquifer by the U.S. Environmental Protection Agency; 

and (3) the aquifer is susceptible to contamination 

from chemicals introduced at or near the land surface 

because the material is highly permeable.

Recent studies on Long Island related volatile 

organic compound (VOC) concentrations in ground 

water to population density through monotonic regression of data from 903 wells in Nassau and Suffolk 

Counties (Eckhardt and others, 1989a). This work led 

to the collection, in 1987-88, of new water-quality data 

from a network of 90 shallow observation wells in five 

areas of differing land use in Nassau and Suffolk 

Counties. Each of the five areas was delineated to

993
represent a specific predominant land use and duration 

of sewering in 1984; these areas are categorized as 

(1) suburban with long-term sewering (longer than 22 

years), (2) suburban with recent sewering (less than 8 

years), (3) suburban without sewering, (4) agricultural, 

and (5) undeveloped (fig. 1). Each area lies along the 

regional ground-water divide, where ground water 

flows vertically downward and horizontally away 

from the divide (fig. 1).

The potential for ground-water contamination in 

the two-county area was predicted through statistical 

models that relate the presence or absence of certain 

contaminants in water from the 90 shallow observation wells to selected explanatory variables that represent potential contaminant sources (Eckhardt and 

others, 1989b). More recent work (D.A. Eckhardt and 

P.E. Stackelberg, U.S. Geological Survey, written 

commun., 1993) has expanded on these statistical 

modeling efforts by (1) modeling the presence of additional contaminants, (2) incorporating additional 

explanatory variables, and (3) using stepwise modelbuilding strategies.

This paper describes results of the recent 

statistical analyses by presenting an overview of the 

statistical techniques used to develop models presented herein, and selecting one model to illustrate 

application and testing procedures. The application 

presents the predicted probability of VOC presence 

throughout Nassau and Suffolk Counties by the 

selected model developed from data from 90 wells in 

the five study areas. A comparison of predicted and 

observed VOC presence in the two-county area is then 

presented.

STATISTICAL MODELS

The statistical models presented in this study 

were developed to describe relations between the quality of shallow ground water and variables representing 

human activities and local hydrogeologic conditions. 

The following discussion briefly summarizes the 

development, application, and testing procedures. 

Detailed discussion of the statistical techniques are 

given in Walker and Duncan (1967), Harrell and 

others (1980), Harrell and Lee (1985), Freeman 

(1987), Hosmer and Lemeshow (1989), and Helsel 

and Hirsch( 1992).

Results of Logistic Regressions

The basis for the development of models presented herein is the network of 90 shallow observation 

wells within the five areas of differing land use (fig. 1). 

Randomly distributed well locations within these areas 

are expected to provide representative data on water 

quality in the water-table aquifer (LeaMond and 

others, 1992). All wells were screened within 45 ft of 

the water table.

Maximum-likelihood logistic regression was 

used to explore relations between chemical concentrations of ground water and variables representing 

human activities and local hydrogeologic conditions 

because it provides a means of relating a mixture of 

continuous explanatory variables to a binary response 

variable (the presence or absence of a particular constituent at a specified concentration) (Harrell and Lee, 

1985). Explanatory variables used in these models are 

continuously distributed indicators of the degree and 

type of human activities at land surface and local 

hydrogeologic conditions and include several land-use 

categories, population density, and thickness of the 

unsaturated zone. Water-quality properties, such as 

specific conductance, also provide useful explanatory 

variables. The values of all variables were determined 

for a 1/2- mi radial area around each of the 90 wells by 

means of a geographic information system (GIS). 

A 1/2-mi radius was chosen because this radius represents the distance that water in the water-table aquifer 

would move in 7 years at an average horizontal flow 

rate of 1 ft/d; thus, 1987-88 water-quality data 

collected at each of the 90 wells should reflect the 

effects of land-use practices from the early and middle 

1980's. Land-use data used to develop the models 

presented herein represent 1981 patterns; population 

density represents 1985 conditions. The area within 

the 1/2-mile radius of each well is assumed to include 

the predominant sources of contamination near each 

well (Eckhardt and others, 1989a).

Maximum-likelihood estimates computed by 

the Newton-Raphson iterative method were used to 

define the best linear fit between a response variable 

(presence or absence of a constituent at a specified 

concentration) and explanatory variables. The 

maximum-likelihood procedure optimizes the likelihood that the observed water-quality data would be 

produced from a given set of explanatory-variable 

slopes (Helsel and Hirsch, 1992). The resulting 

regression parameters (intercept and slope coefficients) differ significantly from zero at a 95-percent

994
74°

41°

40° 

37 

30"

73° 

But from N«w York Stiti Dtpartmant of Transportation. 1987, 1:62,500

EXPLANATION

STUDY AR

Ground-water divide In 1983 Suburban with long-term sewering 

Illll Suburban with recent sewering

Figure 1. Location of the five study areas, Long Island, New York.

to

(O

tn
Table 1. Logistic regression models and measures of goodness of fit for five constituents

[Measures for goodness of fit: G, chi-square statistic of overall model; C, fraction of concordant 

pairs of predicted probabilities and responses]

Constituents____________Model equations1______________G C

Volatile organic compounds (at a concentration of 1 ug/L or greater)

1.VOC = -2.78 + 0.045(fflMEDRES) + 0.06(COMLU) 39 0.87

2. VOC = -2.80 + 0.31(POPDEN) + 0.032(MEDRES) 39 .86

Boron (at a concentration of 50 ug/L or greater)

3.BORON = -3.% + 0.0064(CONDUCT)

+ 0.37(POPDEN) + 0.04(AGLU) 51 .90

4.BORON = -3.06 + 0.052(fflMEDRES)

+ 0.053(AGLU) + 0.12(COMLU) 42 .86 

Nitrate (at a concentration of 3 mg/L as N or greater)

5.NITRATE = -2.10 + 0.35(POPDEN)+ 0.058(AGLU)

+ 0.055(MEDRES) 37 .86

6.NITRATE = -0.68 + 0.39(POPDEN) + 0.061(AGLU)

+ 0.062(MEDRES) - 0.032(UNSAT) 44 .88 

Insecticides (at a concentration of 0.01 ug/L or greater)

7.CARBAMATE =-4.09 + 0.067(AGLU) 39 .97

8.ORGANOCHLORINE = -2.24 + 0.087(AGLU) + 0.072(fflMEDRES)

- 0.028(UNSAT) 52 .89

Variable names and descriptions

POPDEN, Mean area-weighted population density per acre within 1/2-mile radius of well. 

(Source: 1985 Census Bureau block group data)

HMEDRES, Percentage of area within 1/2-mile radius of well classified as high-density or 

medium-density residential, 1 or more dwelling units per acre. (Source: Long Island 

Regional Planning Board [1982])

MEDRES, Percentage of area within 1/2-mile radius of well classified as medium-density residential, 1 to 5 dwelling units per acre. (Source: Long Island Regional Planning Board 

[1982])

COMLU, Percentage of area within 1/2-mile radius of well classified as commercial. (Source: 

Long Island Regional Planning Board [1982])

AGLU, Percetage of area within 1/2-mile radius of well classified as agricultural. (Source: Long 

Island Regional Planning Board [1982])

UNS AT, Thickness of unsaturated-zone at well, in feet

CONDUCT, Specific conductance of well water, (field measured) in microsiemens per centimeter at 25 °C.

confidence level (a = 0.05). Two measures of good- ues plus one-half the fraction of tied pairs, which is a

ness of fit were used to assess the predictive ability of nonparametric parameter. These measures of goodthe logistic regression equations: (1) G, the chi-square ness of fit were maximized through iterative stepwise

statistic for the log-liklihood ratio, which tests the joint procedures (Harrell and others, 1980) by forward

effect of explanatory variables in the model, and (2) C, selection and backward elimination of explanatory

the fraction of concordant predicted and observed val- variables. Table 1 presents several models for VOC's,

996
boron, nitrate, and carbamate and organochlorine 

pesticide presence and indications of their goodness of 

fit. Potential sources of these contaminants, and 

variables found useful in describing their source terms, 

are identified in the following discussion.

Volatile Organic Compounds

Equations 1 and 2 (table 1) relate the presence 

of VOC's at concentrations of at least 1 ng/L to variables that reflect sources of VOC contamination. 

VOC's were detected almost entirely in ground-water 

samples from the three suburban areas (LeaMond and 

others, 1992), which are characterized by industrial, 

commercial, and residential activities. Residue from 

road surfaces, and leaks and spills of industrial and 

commercial solvents, such as from dry cleaners and 

gas stations, can contaminate runoff and recharge 

before it infiltrates to the ground-water system. 

VOC's are also components of industrial and household products such as degreasers, furniture strippers, 

and paint thinners, and thus can be introduced to the 

ground-water system through cesspools, septic tanks, 

and leaking sewer lines. Explanatory variables found 

to be useful in describing VOC source terms include 

population density, commercial land use, and highand medium-density residential land use (table 1).

Boron

Equations 3 and 4 (table 1) relate the presence 

of boron at concentrations of at least 50 (ig/L to variables that reflect boron sources. Dissolved-boron concentrations were significantly elevated in agricultural 

and suburban areas (LeaMond and others, 1992). 

Primary sources of dissolved boron in Long Island's 

ground-water system include (1) fertilizers used on 

commercial crops, and (2) effluent from cesspools, 

septic tanks, and leaking sewer lines. Boron salts are 

components of certain fertilizer products, and boron is 

a component of laundry soaps and detergents. Explanatory variables found to be useful in describing boronsource terms include agricultural, commercial, and 

high- and medium-density residential land use, population density, and specific conductance as measured 

in the field at the time of sampling (table 1).

Nitrate

Equations 5 and 6 (table 1) relate the presence 

of nitrate at concentrations of at least 3 mg/L to variables that reflect sources of nitrate. Concentrations of

nitrate were significantly elevated within the agricultural and suburban areas (LeaMond and others, 1992). 

Primary sources of nitrate within these areas include 

(1) fertilizers used on commercial crops, lawns, 

gardens, and commercial recreational facilities such as 

golf courses, (2) leakage of effluent from cesspools, 

septic tanks, and sewer systems and (3) animal feces. 

Of the 90 wells in the five study areas, 83 indicated 

nitrate concentrations of at least 0.01 mg/L, the 

analytical detection limit. For logistic-regression 

analyses, the response-variable limit was set at 3 mg/L 

to delineate wells where ground water was most 

clearly affected by human activities; at 62 of the 90 

wells, nitrate concentrations were 3 mg/L or greater. 

Explanatory variables found to be useful in describing 

nitrate-source terms include agricultural and residential land use, population density, and thickness of the 

unsaturated zone (table 1).

Insecticides

Equations 7 and 8 (table 1) relate the presence 

of carbamate and organochlorine insecticide residues 

at concentrations of at least 1.0 and 0.01 jag/L respectively, to variables that reflect sources of insecticide. 

Carbamate insecticides were detected almost entirely 

in samples from the agricultural area, where they have 

been applied primarily to control the Colorado potato 

beetle and golden nematode (Soren and Stelz, 1985); 

none were detected in samples from the sewered suburban areas (LeaMond and others, 1992). Organochlorine insecticides, the most frequently detected 

pesticide class in all five study areas, have been 

applied in agricultural areas to control crop pests and 

in residential areas to control termites and other 

insects. Explanatory variables found to be useful in 

describing carbamate- and organochlorine-source 

terms include agricultural and high- and mediumdensity residential land use, and thickness of the 

unsaturated zone (table 1).

Application of the Statistical Models

Once developed, the linear logistic-regression 

models can be transformed to predict probabilities of 

contaminant presence through the logistic function

+exp(&o+

where P (X) is the probability of contaminant presence

997
given the vector X of explanatory variables and the 

vector p of slope parameters and the scalar model 

intercept, po.

Equation 1 (table 1) was selected to illustrate 

model application and testing procedures. To predict 

the probability of VOC presence in the shallow 

ground-water system of Nassau and Suffolk Counties, 

the appropriate explanatory variables were quantified 

for each 1/4-mi2 cell from a gridded overlay of the 

area. The probabilities computed for each cell through 

the logistic function were then displayed graphically 

to illustrate the spatial relation between the explanatory variables and the predicted presence of VOC's. 

The probability map of VOC presence at concentrations of at least 1 ng/L within the two-county area is 

shown in figure 2.

An independent set of VOC analyses from 378 

wells throughout Nassau and Suffolk Counties was 

spatially compared with the VOC distribution predicted by the model. This data set excluded the 90 

wells used for model development. The wells were 

screened within 45 ft of the water table and were 

sampled for VOC's during 1984-90. Each well was 

assigned either a 1 or 0, depending on VOC detection 

at a concentration of at least 1 ng/L, and the cumulative probability distribution of the observed binary 

presence was then compared with the distribution of 

predicted probability of presence.

A cumulative-distribution plot showing cumulative probability distributions of observed and predicted probabilities of VOC presence for the two sets 

of data (the 90 wells from which the predictive equations were developed, and the 378 wells used for comparison) is shown in figure 3. The predicted 

probabilities closely fit the observed VOC-presence 

probabilities for the five study areas because the 

predictive equations were fitted to these data by logistic regression. At probabilities greater than about 

0.75, however, the fitted model underpredicts VOC 

presence observed at the 90-well sites by as much as 

8 percent. The close match between predicted probabilities of VOC presence within the two-county area 

and that for the five study areas indicates that the five 

land-use areas accurately represent the full range of 

land uses within the entire two-county area. The second data set was also used to compare predictions of 

VOC presence, as derived through model 1 (table 1),

with the observed presence of VOC's in the entire 

two-county area (fig. 2). The resulting underprediction of VOC presence by as much as 30 percent in the 

low-probability regions indicates that VOC's are more 

prevalent in areas with relatively small amounts of 

commercial and high- and medium-density residential 

land than indicated by patterns observed in the five 

land-use areas. The underprediction can also result, in 

part, from the spatial overlap of some wells, which 

produces a redundancy of water-quality data.

SUMMARY AND CONCLUSIONS

Water-quality data for 90 shallow observation 

wells in five areas of differing land use on Long Island 

provided the information necessary to evaluate relations among shallow ground-water quality and variables representing human activities and local 

hydrogeologic conditions. Maximum-likelihood 

logistic-regression equations were developed to characterize the probability of VOC, boron, nitrate, and 

pesticide presence in shallow ground water through 

explanatory variables such as (1) population density 

within a 1/2-mi radius of each well, (2) amount of 

specified types of land use with a 1/2-mi radius of each 

well, and (3) the thickness of the unsaturated zone at 

each well site. The explanatory variables were related 

to the binary response variable (presence or absence of 

a contaminant at a specified concentration) through 

logistic regression.

A logistic-regression equation with statistically 

significant intercept and slope coefficients was 

selected to predict the probability of VOC presence at 

concentrations of 1 ng/L or greater within the shallow 

ground-water system of Nassau and Suffolk Counties. Model predictions were compared with an independent set of VOC data from 378 shallow wells 

throughout the two-county area. The underprediction 

of VOC presence in the two-county area by as much as 

30 percent in low-probability regions indicates that, in 

areas that have relatively small amounts of commercial and high-density and medium-density residential 

land, VOC's are more prevalent than indicated by patterns observed within the five specified land-use areas.

998
73 

40*

sr

30'

VOC = -2.78 + 0.045(HIMEDRES) + 0.06(COMLU)

But from N»w York SUt« D«pirtm«nt of Tnrwportitlon, 1987, 1:62,600

EXPLANATION

Probability of volatile organic compound detection 

at concentrations greater than 1.0 mlcrogram per liter

O.25 >0.25-0.50 >0.50-0.75 >0.75

Figure 2. Predicted probabilities of volatile organic compound presence at concentrations of 1 microgram per liter or greater.
LU 

O

1.0

0.9

0.8 LU 

0) 

LU

oc

Q_

O 0.7

§

U_

O

>- 0.6

CD

cr

0_

0.5

0.4

0.3

Probability of VOC Presence 

In Five Study Areas:

+ - Observed (n = 90)

0 -Predicted 

In Two-County Area:

0 - Observed (n = 378)

 -Predicted

-3-2-10123

RANK NORMAL SCORE

Figure 3. Cumulative probability distributions of observed and predicted probability of volatile organic 

compound presence at concentratons of 1 microgram per liter or greater.

REFERENCES

Eckhardt, D.A., Flipse, W.J., and Oaksford, E.T., 1989a, 

Relation between land use and ground-water quality 

in the upper glacial aquifer in Nassau and Suffolk 

Counties, Long Island, New York: U.S. Geological 

Survey Water-Resources Investigations Report 

86-4142, 35 p.

Eckhardt, D.A., Siwiec, S.F., and Cauller, S.J., 1989b, 

Regional appraisal of ground-water quality in five 

different land-use areas, Long Island, New York, 

in Mallard, G.E., and Ragone, S.E., eds., U.S. Geological Survey Toxic Substances Hydrology Program 

Proceedings of the technical meeting, Phoenix, 

Arizona, September 26-30, 1988: U.S. Geological 

Survey Water-Resources Investigations Report 

88-4220, p. 397-403.

Freeman, D.H., 1987, Applied categorical analysis: New 

York, Marcel Dekker, 318 p.

Harrell, F.E., Lee, K.L., and McKinnis, R.A., 1980, Procedures for large regression problems requiring maximum likelihood estimation, in Proceedings of the fifth 

annual SAS users group international conference: 

Cray, N.C., SAS Institute, p. 199-202.

Harrell, F.E., and Lee, K.L., 1985, The practical value of 

logistic regression, in Proceedings of the tenth annual 

SAS users group international conference: Cray, N.C., 

SAS Institute, p. 1031-1036.

Helsel, D.R., and Hirsch, R.M., 1992, Statistical methods in 

water resources: New York, Elsevier, 522 p.

Hosmer, D.W, and Lemeshow, S., 1989, Applied logistic 

regression: New York, John Wiley, 307 p.

LeaMond, C.E., Haefner, R.J., Cauller, S.J., and Stackelberg, P.E., 1992, Ground-water quality in five areas of 

differing land use in Nassau and Suffolk Counties, 

Long Island, New York: U.S. Geological Survey 

Open-File Report 91-180, 67 p.

Long Island Regional Planning Board, 1982, Quantification 

and analysis of land use for Nassau and Suffolk 

Counties, 1981: Hauppauge, N.Y., 47 p.

Soren, Julian, and Stelz, W.G., 1984, Aldicarb pesticide 

contamination of ground water in eastern Suffolk 

County, Long Island, New York: U.S. Geological 

Survey Water-Resources Investigations Report 

84-425 l,34p.

Walker, S.H., and Duncan, D.B., 1967, Estimation of the 

probability of an event as a function of several independent variables: Biometrika, v. 54, p. 167-179.

1000
Evaluating the Effect of Land Use and Sampling Depth 

on Ground-Water Quality, Long Island, New York

EyPaul E. Stackelberg 1

Abstract

The effect of nonpoint-source 

contamination was statistically evaluated through 

a comparison of water-quality data collected from 

two depth intervals at 153 wells in five areas of 

differing land use on Long Island. Each area was 

delineated to represent a predominant land use; 

the areas were categorized as (1) suburban with 

long-term sewering, (2) suburban with recent 

sewering, (3) suburban without sewers, (4) 

agricultural, and (5) undeveloped. The depth 

zones were delineated on the basis of estimated 

traveltime of ground water along vertical flow 

paths from the water table to each well screen's 

midpoint. Wells were classified as shallow 

(estimated traveltimes of less than 10 years) or 

intermediate (estimated traveltimes between 10 

and 100 years).

Concentrations of several inorganic 

constituents and values of field properties were 

found to differ significantly among land-use areas 

and depth zones. Median constituent 

concentrations tend to be (1) highest, and the 

concentration ranges the widest, in samples from 

the agricultural area, (2) lowest, and the 

concentration ranges the smallest, in samples 

from the undeveloped area, and (3) intermediate 

to high in samples from the suburban areas. 

Volatile organic compounds (VOC's) were 

detected only in the suburban areas. A common 

source of nonpoint-source contamination in 

agricultural and residential areas is fertilizers used 

on commercial crops, lawns, gardens, and golf

'U.S. Geological Survey, West Trenton, New Jersey 08628

courses. Other sources of inorganic contaminants 

and VOC's in residential areas include (1) effluent 

from cesspools, septic tanks, and leaking sewers, 

(2) road-deicing salts, and (3) runoff 

contaminated by road residues and by chemicals 

commonly used at industrial and commercial 

facilities. A decrease in concentrations of most 

inorganic constituents with depth is attributed to 

(1) physical and chemical reactions that remove 

constituents from solution, and (2) dilution by 

advection and hydrodynamic dispersion along 

flow paths.

INTRODUCTION

Relations between land use and ground-water 

quality in Nassau and Suffolk Counties have been 

investigated as part of the U.S. Geological Survey's 

Toxic Substances Hydrology Program since the early 

1980's (Eckhardt and others, 1988, and LeaMond and 

others, 1992). The most recent phase of the study was 

begun in 1990 and was designed to evaluate 

statistically the effects of land use and depth on 

ground-water quality beneath five areas of differing 

land use. The five areas were delineated according to 

predominant land use and duration of sewering in 

1984 and are categorized as (1) suburban with longterm sewering (longer than 22 years), (2) suburban 

with recent sewering (less than 8 years), (3) suburban 

without sewers, (4) agricultural, and (5) undeveloped 

(fig. 1). All five areas lie along the regional groundwater divide, where vertical ground-water flow 

components cause contaminants introduced at or near 

land surface to be transported deeper into the aquifer 

system than elsewhere. Water-quality data from 207 

randomly selected wells screened in shallow, 

intermediate, and deep zones beneath the five study

1001
1

74'

41

409 

37 

30"

QUEENS 

COUNTY

KINGS 

COUNTY

NASSAU 

COUNTY

73°

SUFFOLK 

COUNTY

10

Base from New York State Department of Transportation, 1987, 1:62,500

EXPLANATION

STU

Ground-water divide in 1983 ^j Suburban with long-term se

Mill Suburban with recent sewer

Figure 1. Location of the five study areas on Long Island, New York.
areas were statistically evaluated to determine (1) differences in water quality among the five areas and 

shallow and intermediate depth zones, (2) differences 

in water quality among the suburban areas and all 

three depth zones, and (3) whether sewering within the 

suburban areas has resulted in an improvement in 

ground-water quality. Depth zones were delineated on 

the basis of estimated traveltime of ground water 

along vertical flow paths from the water table to each 

well screen's midpoint. Wells were classified as shallow (estimated traveltimes of less than 10 years), intermediate (estimated traveltimes from 10 to 100 years), 

and deep (estimated traveltime longer than 100 years).

This paper summarizes (1) the data-collection 

and statistical methods, and (2) the effects of land use 

and depth on water quality in the shallow and 

intermediate-depth zones of the five study areas.

STUDY METHODS

The approach used in this study was based on 

the premise that contamination of ground water by 

human activities can be considered a function of land 

use that is, land uses determine the amounts and types 

of chemicals that are applied to the land surface and, 

thus, the amounts and types of contaminants that reach 

ground water (Helsel and Ragone, 1984). The study 

entailed three procedures: (1) collection of waterquality data from a network of 153 wells, (2) 

classification of wells as shallow or intermediate on 

the basis of estimated ground-water residence time, 

and (3) statistical analyses of water-quality data.

Data Collection

Water samples were collected from June 1987 

through September 1988 from a network of 153 wells 

within the five study areas. In all, 83 shallow wells 

and 70 intermediate-depth wells were sampled. The 

shallow zone was represented by 11 to 20 wells in 

each study area, and the intermediate zone by 10 to 

17 wells in each study area.

Wells were classified as monitoring, public 

supply, or unspecified withdrawal. At each 

monitoring well, water levels were measured before 

pumping, at least three casing volumes were 

evacuated, and drawdown, specific conductance, pH, 

and temperature were allowed to stabilize before 

sampling. At each public-supply and unspecifiedwithdrawal well, at least three casing volumes of water

were evacuated, and measurements of specific 

conductance, pH, and temperature were made before 

sampling.

The water-quality data analyzed herein, and 

additional details of well-locating procedures and 

sampling methods, are given in Stackelberg 

(unpublished data on file in Syosset, N.Y. office of 

U.S. Geological Survey) and LeaMond and others 

(1992).

Well-Depth Classification

Classification of wells as shallow or 

intermediate was based on the traveltime of ground 

water along vertical flow paths from the water table to 

the midpoint of each well screen to ensure that all 

wells within a given depth zone would provide 

samples of similar age. Vertical flow rates were 

calculated from cell-to-cell flow rates by the regional 

ground-water-flow model of Long Island (H.T. Buxton 

and D.A. Smolensky, U.S. Geological Survey, written 

commun., 1990), which simulates ground-water flow 

under steady-state conditions of 1968-75. These years 

represent a period in which precipitation rates 

paralleled long-term average rates, and ground-water 

pumpage was uniform (S.M. Feldman, U.S. 

Geological Survey, oral commun., 1990). The vertical 

flow rates were then used to calculate traveltimes 

along vertical flow paths from the water table to each 

screen's midpoint. Although this method does not 

address traveltimes through the unsaturated zone, the 

mean unsaturated-zone thickness is similar in all five 

land-use areas, except in the suburban unsewered area, 

where Tukey's test indicates it to be significantly 

greater than in the agricultural or undeveloped areas. 

(An explanation of Tukey's test is provided in the 

following section.) Shallow wells were those 

estimated to tap water with a residence time of less 

than 10 years, and intermediate wells were those 

estimated to obtain water 10 to 100 years old.

Statistical Analyses

Nonparametric statistical procedures and 

contingency-table analyses were used to evaluate the 

effect of land use and sampling depth on constituent 

concentrations. Nonparametric procedures, such as 

analysis of variance (ANOVA) on rank-transformed 

data, are robust that is, they are insensitive to 

outlying values and assumptions of equal variance or

1003
normality (Helsel and Hirsch, 1992; Iman and 

Conover, 1983). Contingency-table analyses provide 

a means of evaluating highly censored data sets. 

(Censored data sets contain values whose 

concentration is reported as "less than the analytical 

detection limit.") These statistical procedures are used 

to assess whether water-quality data differ 

significantly among land-use and depth categories. 

Two-way factorial ANOVA's on rank-transformed 

water-quality data were used to examine the effect of 

land use and depth simultaneously. The null 

hypothesis for this test states that mean concentration 

ranks are identical among all land-use and depth 

groups; thus no effects are due to land use, depth, or 

their interaction. If interaction is found to be 

significant, the effect of depth on mean concentration 

ranks differs among land-use areas. Without 

significant interaction, the effect of depth on mean 

concentration ranks is identical among all land-use 

areas (Helsel and Hirsch, 1992). If the null hypothesis 

was rejected by the ANOVA test, indicating that at 

least one mean concentration-rank value differed from 

another, Tukey's honest significant difference test 

(Tukey's test) was performed to determine which 

means differed.

Contingency-table analyses were used to 

examine differences in water quality among the five 

study areas and two depth zones for constituents 

whose censored values constituted more than 50 

percent of the total values. Contingency-table analyses 

do not evaluate the effects of multiple factors 

simultaneously; therefore, separate tests were 

performed on land-use and sampling-depth data to 

evaluate their effects on constituent-detection rates. 

The null hypotheses tested were that the proportion of 

samples with constituent detections was the same for 

each study area or depth zone; rejection of the null 

hypothesis indicates that the proportion of samples 

with constituent detections differed among land-use 

areas or depth zones. If a contingency-table analysis 

indicated that the proportion of samples in which a 

constituent was detected differed among study areas or 

depth zones, individual-cell chi-square statistics were 

examined to determine which areas or depth zones had 

more or less than the expected proportion of 

constituent detections. The positive or negative sign 

of the difference between observed and expected 

frequencies indicates more (+), or fewer (-), detections 

within that area or depth zone, and the magnitude of 

the chi-square statistic indicates cells contributing

most to the rejection of the null hypothesis (Helsel and 

Hirsch, 1992).

The level of significance (alpha) for all 

hypothesis testing reported herein is 0.05. Results 

from ANOVA and contingency table test are reported 

as p-values that indicate the strength of the evidence to 

accept or reject the null hypothesis. The null 

hypothesis is rejected for all p-values reported as less 

than 0.05; the smaller the p-value, the stronger the 

evidence for rejection.

Water-quality data for selected constituents are 

represented as box plots, which indicate the median, 

the data variation and skewness, and the absence or 

presence of unusual "outside" or "far-outside" values 

(Helsel and Hirsh, 1992). Summary statistics for 

water-quality data that contained censored values were 

calculated through a log-probability regression 

technique (Helsel and Hirsh, 1992), a robust procedure 

that minimizes rates of root mean square error. 

Results from Tukey's test are presented as letters 

above the box plots for each land-use area. Groups of 

data with at least one of the same letters do not differ 

significantly from one another.

EFFECTS OF LAND USE AND SAMPLING 

DEPTH ON GROUND-WATER QUALITY

The effect that land use, sampling depth, and 

interaction between land use and depth have on 

constituent concentrations are reported as p-values in 

tables 1 and 2. Summary statistics for samples from 

the shallow and intermediate zone of each area are 

reported in table 3, which includes results of Tukey's 

test to indicate whether the mean concentration ranks 

from the combination of shallow and intermediatedepth samples in each area are similar among study 

areas. Results from a contingency-table analysis for 

sulfate are given in table 4. Median concentrations 

from the combination of shallow- and intermediatedepth samples of a particular area are reported in the 

text.

Concentrations of inorganic constituents and 

field-measured properties, except pH, differed 

significantly among study areas and, except for 

dissolved oxygen, differed significantly between depth 

zones (tables 1 and 2). Interaction was found to be 

significant for temperature and concentrations of 

nitrate, potassium, and magnesium (table 1).

1004
Table 1. p-values from two-way factorial ANOVA's indicating the effect of land use, sampling depth, and their 

interaction on field-measured properties and inorganic-constituent concentrations from both depth zones of the five 

study areas, Long Island, New York

Properties and 

constituents

Specific conductance

pH

Dissolved oxygen

Temperature

Nitrate

Potassium

Sodium

Chloride

Calcium

Magnesium

Land use 

(all five areas)

0.000 1

0.5 1 55

0.000 1

0.000 1

0.000 1

0.000 1

0.000 1

0.0002

0.000 1

0.000 1

Depth zone 

(Shallow and intermed.)

0.000 1

0.0270

0.254 1

0.000 1

0.000 1

0.000 1

0.000 1

0.000 1

0.000 1

0.0121

Interaction

O.I 090

O.H2I

O.I 950

0.003 1

0.036 1

0.007 1

0.8 1 83

0.40 1 1

0.054 1

0.0205

Table 2. p-values from contingency-table analyses 

indicating the effect of land use and sampling depth on the 

proportion of samples from both depth zones with sulfate 

and volatile organic compound detections

[<, less than]

Constituent

Sulfate (all five study areas)

Volatile organic compounds

(suburban areas only):

l,l,l-trichloroethane

Trichloroethene

Tetrachloroethene

Land use

<0.00l

0.069

O.I 44

0.052

Sampling 

Depth

<O.OOI

O.I 35

0.660

0.758

The highest median values for specific 

conductance (325 (^S/cm) and concentrations of nitrate 

(7.48 mg/L as N), potassium (3.9 mg/L), calcium (27.9 

mg/L), and magnesium (7.1 mg/L) resulted from the 

combination of shallow and intermediate depth zones 

in the agricultural area, but Tukey's test results 

indicates that the mean concentration ranks for these 

constituents in the agricultural area do not differ 

significantly from those for one or more of the 

suburban areas, except for potassium, whose mean 

rank in the agricultural area is significantly higher than 

that from any other area (table 3). Elevated 

concentrations of these constituents in the agricultural 

and suburban areas result primarily from (1) fertilizers 

used on crops, lawns, and golf courses, and (2) leakage 

from cesspools, septic tanks, and sewer systems. 

Road-deicing salts will also elevate specific

conductance, especially in suburban areas, where road 

networks are extensive.

Results from contingency-table analysis indicate 

that sulfate detection in samples from both depth zones 

differed significantly among study areas (table 2). The 

sign of the difference between observed and expected 

frequencies and the magnitude of each cell's chisquare statistic (table 4) indicates that the two depth 

zones in the long-term sewered and agricultural areas 

together have more sulfate detections, and the 

unsewered area had fewer, than would be expected if 

all land-use areas were alike; therefore, sulfate appears 

to be more prevalent in the long-term sewered 

suburban and agricultural areas than elsewhere (table 

4). The proportion of samples from both depth zones 

in each study area with sulfate detections ranged from 

91 percent in the agricultural area to 46 percent in the 

unsewered suburban area. Although sulfate was 

detected frequently in samples from both depth zones 

of the undeveloped area (79 percent), concentrations in 

this area were lower than elsewhere. Elevated 

concentrations of sulfate in the agricultural area 

probably result from fertilizers; the highest sulfate 

concentration (920 mg/L) in all five areas was in the 

intermediate zone of the agricultural area. Sulfate 

concentrations were also elevated in the suburban 

areas, probably as a result of lawn and garden 

fertilizers and effluent from cesspools, septic tanks, 

and leaking sewer lines.

1005
Table 3. Summary statistics for samples from the shallow and intermediate zones of each study area, Long Island, New York

[Numbers in parentheses represent the number of wells sampled; IQR, interquartile range; Tukey's test, results of Tukey's Studentized test]

Study Areas (S, shallow wells; 1, intermediate wells)

Properties, 

constituents, 

and statistics1

Field-measured Properties

Long-term 

sewered 

S I 

(20) (13)

Recently 

sewered 

S I 

(20) (17)

Unsewered 

S I 

(20) (17)

Agricultural 

S 1 

(12) (10)

Undeveloped 

S 1

(11) (13)

Specific conductance (microsiemens per centimeter)

Median

IQR

Tukey's test

PH

Median

IQR

Tukey's test

280

1 05

5.65

0.50

1 89

158

A

5.65

0.65

A

299

1 73

139

I9l

AB

5.50

0.55

A

5.80

0.90

222

1 96

BC

5.60

0.80

A

75

50

5.60

0.30

340 227

216 250

A

5.65

0.35

A

5.85

0.70

1 07

1 05

C

5.50

0.40

A

80

103

6.40

1.40

Dissolved oxygen (milligrams per liter)

Median

IQR

Tukey's test

Temperature (degrees Celsius)

Median

IQR

Tukey's test

6.9

6. 1

14.3

0.9

3.3

3.6

A

1 4.4

3.5

A

5.8

4.3

A

1 4.4

1. 1

A

6.0

4.3

1 2.6

1. 8

8.0

3.4

B

1 3.0

1. 6

B

8.8

4.5

II. I

0.8

9.2

2.7

B

8.9

4.2

1 0.6

2.0

B

I2.l 1I.3 I1.8

0.4

BC

1. 3 1.0

C

10.1

3.0

10.8

0.9

Constituents (in milligrams per liter)

Nitrate

Median

IQR

Tukey's test

Potassium

Median

IQR

Tukey's test

Sodium

Median

IQR

Tukey's test

Chloride

Median

IQR

Tukey's test

Calcium

Median

IQR

Tukey's test

Magnesium

Median

IQR

Tukey's test

5.12

2.19

2.7

2.2

1 7.5

14.5

32.0

24.8

21.7

1 0.4

3.9

3.3

2.18

4.63

A

2. 1

1. 1

B

14.5

1 6.7

A

28.2

27.3

A

9.3

II. I

AB

2.9

1. 8

AB

5.74

6.86

A

3.9

2.9

B

23.0

22.0

A

30.8

36.2

AB

1 6.8

9.2

B

2.8

0.9

BC

4.47

5.30

0.7

0.7

ll.O

1 2.0

1 3.7

1 2.9

6.9

1 2.9

2.6

4.0

6.9 1

9.88

A

1. 9

2.3

BC

1 7.5

1 5.0

AB

22.0

22. 1

BC

1 2.5

1 3.9

C

4.5

3.0

BC

1. 17

2.47

0.8

0.8

7.0

12.8

7.2

11.2

2.8

3.0

1.5

1.8

9.4 1

7.20

A

4.4

1. 9

A

9.5

1 0.5

BC

23.7

1 3.0

AB

37.5

37.9

A

7.8

5.4

A

3.08

7.05

2.5

2.6

7.0

5.9

16. 3

17.9

17.

22.

4.

6.

3

3

6

3

0.09

3.64

B

0.8

0.2

C

8.0

17.0

C

11.7

21.4

C

3.2

2.4

C

1.3

2.6

C

0.07

5.24

0.6

0.9

5.0

4.8

7.4

3.7

4.6

9.0

1.2

3.9

1 Tukey's test results are presented as letters. Groups of data with at least one of the same letters do not differ significantly from one another. 

1006
Table 4. Contingency-table analysis showing individual-cell 

frequency, expected frequency, and chi-square statistics for 

sulfate detection in samples from both depth zones of the 

five study areas on Long Island, New York.

[<, less than]

Study area Detected Not 

detected

Long-term sewered

frequency 29 4 

expected frequency 24.16 8.84 

chi-square 0.97 2.65

Recently sewered

frequency 27 10 

expected frequency 27.08 9.91 

chi-square <.00 <.00

Unsewered

frequency 17 20 

expected frequency 27.08 9.91 

chi-square 0.94 10.26

Agricultural

frequency 20 2 

expected frequency 16.10 5.89 

chi-square 0.94 2.57

Undeveloped

frequency 19 5 

expected frequency 17.57 6.43 

chi-square 0.12 0.32

The highest median temperature (14.4°C), and 

the highest median concentration of sodium (17 mg/L) 

and chloride (31.2 mg/L) resulted from the 

combination of samples from shallow and 

intermediate depths from the long-term sewered area, 

but results of Tukey's test indicate that mean 

concentration ranks for these constituents in the longterm sewered area do not differ significantly from 

those for at least one other suburban area and that the 

mean concentration rank for chloride in the long-term 

sewered area does not differ significantly from that in 

the agricultural area (table 3). Elevated temperatures 

in the suburban areas are probably due to the 

absorption of heat from artificial surfaces such as 

asphalt or concrete before the infiltration of recharge. 

Also, water that stands in recharge basins before 

infiltrating could absorb solar energy and thereby 

increase in temperature. The elevated sodium and 

chloride concentrations in suburban areas is attributed

to the application of deicing salts on the extensive road 

networks. These constituents are also found in some 

household cleaning products and are a component of 

cesspool, septic-tank, and sewerline effluents.

The highest median concentration for dissolved 

oxygen (10.55 mg/L) resulted from the combination of 

shallow and intermediate samples in the undeveloped 

area; however, Tukey's test results indicate that the 

mean concentration rank for this area does not differ 

significantly from those for the agricultural or 

unsewered suburban areas, and these three areas have 

significantly higher values than the sewered suburban 

areas (table 3). Processes that could accelerate the 

depletion of dissolved oxygen within the sewered 

suburban areas include nitrification and increased 

bacterial activity. Nitrogen introduced to the groundwater system in the form of ammonium by cesspools, 

septic tanks, and sewerlines is converted to nitrogen 

and nitrate through oxidation, which also consumes 

dissolved oxygen. Additionally, dissolved organic 

carbon that is introduced from the same sources in 

sewered suburban areas could fuel bacterial activity 

that also consumes dissolved oxygen.

Examination of p-values in tables 1 and 2 

indicates that concentrations of each inorganic 

constituent, and measurements of each field property 

except dissolved oxygen, differed significantly with 

depth. The highest median concentration of these 

constituents and field-measured properties, except pH, 

were in samples from the shallow zone of all five 

study areas combined. Decreasing concentrations of 

these inorganic constituents with depth probably 

results from (1) physical and chemical reactions that 

remove these constituents from solution, and (2) 

dilution through advection and hydrodynamic 

dispersion as constituents move along flow paths. The 

highest median pH value was from samples from the 

intermediate zone of all five study areas combined. 

The increase in pH with depth is most likely due to the 

oxidation of organic matter; this is most apparent in 

the undeveloped area and least apparent in the 

agricultural and suburban areas (table 3). The smaller 

increase in pH with depth in the agricultural and 

suburban areas than in the undeveloped area could be 

due to the process of nitrification, which, in addition to 

consuming dissolved oxygen, produces hydrogen ions 

that lower the pH.

Dissolved-oxygen concentrations were not 

found to change significantly with depth; thus, 

concentrations in the intermediate zone of all five

1007
study areas combined were similar to those in the 

shallow zone. The extent to which dissolved oxygen 

concentrations can be maintained with depth depends 

in part on (1) the hydraulic properties of soils and 

aquifer materials through which ground water flows, 

and (2) the rate at which oxygen-depleting processes 

are occurring within the unsaturated and saturated 

zones. Elevated dissolved-oxygen concentrations in 

the intermediate zone indicate that (1) soils and 

aquifer materials in the five study areas are highly 

permeable, well drained, and subsequently well 

aerated, and (2) processes that deplete dissolved 

oxygen concentrations (such as nitrification, bacterial 

activity, and oxidation of organic matter), although not 

completely absent, are insufficient to exhaust 

dissolved-oxygen supplies within the depths 

examined.

Interaction between the effects of land use and 

sampling depth was significant for temperature, 

nitrate, potassium, and magnesium (table 1); 

indicating that the effect of depth differs among the 

land-use areas. This effect is illustrated in table 3 and 

figure 2, where concentrations of the inorganic 

constituents are seen to decrease from the shallow to 

the intermediate zone of each area except the 

undeveloped area, and temperatures decrease from the 

shallow to intermediate zone of each area except the 

long-term sewered suburban area.

The most commonly detected VOC's were 

1,1,1-trichloroethane (TCA), trichloroethene (TCE), 

and tetrachloroethene (PCE), whose maximum 

concentrations were 12,000, 660, and 280 }J.g/L, 

respectively. VOC's are introduced to the ground 

water of suburban areas primarily from industrial, 

commercial, and residential activities. Accidental 

leaks and spills of industrial chemicals and 

hydrocarbons from commercial facilities such as dry 

cleaners and gas stations, and residue from road 

surfaces, could contaminate runoff before it infiltrates 

to the ground-water system. VOC's are also 

components of industrial and household products such 

as degreasers, furniture strippers, and paint thinners, 

and, thus, can be introduced to the ground-water 

system as effluent from cesspools, septic tanks, and 

leaking sewer lines. These compounds were detected 

solely within the three suburban areas; thus, the 

following paragraph addresses their presence and 

spatial distribution only within those areas.

Of all samples from the shallow and 

intermediate zone of the long-term sewered area, 57

percent contained VOC's, as did 54 percent of those 

from the recently sewered area, and 27 percent of 

those from the unsewered suburban area. 

Contingency-table results indicate that the proportion 

of samples from both zones that contained these 

compounds did not differ among suburban areas 

significantly (table 2), although the evidence to accept 

the null hypothesis that the proportion of PCE 

detections were equal among all areas was not strong 

(p-value = 0.052). Detection frequencies for these 

compounds did not differ significantly between the 

two depth zones; therefore, the proportion of samples 

with VOC detection from the intermediate zone of all 

suburban areas combined did not differ significantly 

from the proportion for the shallow zones.

SUMMARY

Contamination from nonpoint sources has 

altered the quality of water in Long Island's aquifer 

system. The spatial distribution of the contaminants 

within the aquifer system can be related to their depth 

below the water table and to land-use patterns.

The concentrations of several inorganic 

constituents, and volatile organic compounds 

(VOC's), and values for field-measured properties in 

ground water, differ significantly among the five study 

areas. Statistical evaluation of constituent 

concentrations and detection frequencies indicated that 

median concentrations tend to be (1) highest, and 

concentration ranges the widest, in the agricultural 

area, (2) lowest, and the concentration ranges the 

smallest, in the undeveloped area, and (3) intermediate 

to high in suburban areas. VOC's were detected only 

in samples of ground water from suburban areas. A 

major source of ground-water contamination within 

agricultural and residential areas is probably the 

fertilizers used on commercial crops, lawns, gardens, 

and commercial recreational facilities such as golf 

courses. Application of these products results in 

elevated concentrations of nitrate, and sulfate, and 

inorganic constituents such as calcium, potassium, and 

magnesium. An additional source of these 

constituents, especially in high-density residential 

areas, is cesspool, septic tank, and leaking-sewerline 

effluent, which contains human wastes and common 

household-cleaning products. Road-deicing salts 

applied on extensive road networks within suburban

1008
TEMPERATURE NITRATE

20-

15-

1O- t; ?

POTASSIUM

f¥| f¥| |BC| FA! fcl

25-| r

20-

15-

1O5-

EXPLANATION

RESUtT Or TUKETS Sn«NT1Zn) 

(MSD) RANOC TEST

20-

10-

5 ~

MAGNESIUM

10- 

14-

12-

1O8-

6-

2-

 | AP | | Bl> | | Pl» | | M

M

-

.

.1

1 1

1 T

-

1D

1 1 ^ t 

AMI I PCKCENTLEv

w MEDIAN

VMJUES VMTHM IS 

TIMES THE OR

Figure 2. Distribution of temperature, nitrate, potassium, and 

magnesium in water from the shallow and intermediate zones 

of the five study areas, Long Island, New York.

1009

29mPERCENTLC /
areas cause elevated concentrations of sodium and 

chloride. Sources of VOC contamination within suburban areas include (1) residues from the extensive 

road networks, (2) commercial and industrial facilities 

commonly associated with residential areas such as 

dry cleaners and gas stations, and (3) effluent from 

cesspools, septic tanks, and sewerlines.

The quality of ground water differs significantly 

with depth; the concentrations of most inorganic 

constituents and values of field properties decrease 

from the shallow to the intermediate zone of the 

aquifer. This decrease is attributed to (1) physical and 

chemical reactions that remove constituents from 

solution, and (2) dilution by advection and 

hydrodynamic dispersion as constituents move along 

flow paths. Dissolved-oxygen concentrations did not 

decrease significantly with depth, indicating that (1) 

aquifer materials within the five study areas are highly 

permeable, well drained, and well aerated at depth, 

and (2) processes that deplete dissolved oxygen are 

insufficient to exhaust the dissolved oxygen at the 

depths examined.

REFERENCES

Eckhardt, D.A.V., Flipse, W.J., Jr., and Oaksford, E.T., 

1989, Relation between land use and ground-water 

quality in the upper glacial aquifer in Nassau and 

Suffolk Counties, Long Island, New York: U.S. 

Geological Survey Water-Resources Investigations 

Report 86-4142, 35 p.

Eckhardt, D.A.V., Siwiec, S.F., and Haefner, R.J., 1988, 

Ground-water quality in five land-use areas of Nassau 

and Suffolk Counties, Long Island, New York [Abs.]: 

American Geophysical Union, EOS Transactions, 

v. 69, no. 16, p. 354.

Helsel, D.R., and Hirsch, R.M., 1992, Statistical methods in 

water resources: New York, Elsevier, 522 p.

Helsel, D.R., and Ragone, S.E., 1984, Evaluation of

regional ground-water quality in relation to land use: 

U.S. Geological Survey Water-Resources 

Investigations Report 84-4217, 33 p.

Iman, R.L., and Conover, W.J., 1983, A modern approach to 

statistics: New York, John Wiley, 497 p.

LeaMond, C.E., Haefner, R.J., Cauller, S.J., and

Stackelberg, P.E., 1992, Ground-water quality in five 

areas of differing land use in Nassau and Suffolk 

Counties, Long Island, New York: U.S. Geological 

Survey Open-File Report 91-180, 67 p.

1010
Use of Flowpath Simulation to Determine Contributing 

Areas and Traveltimes of Nonpoint-Source GroundWater Contamination, Gloucester County, New Jersey

By Anthony S. Navoy1

Abstract

The area of direct surficial recharge, 

potentially affected by nonpoint-source 

contamination, to the Potomac-Raritan-Magothy 

aquifer system in Gloucester County, New Jersey, 

extends a maximum of 0.5 to 2 miles downdip 

from the outcrop area as determined by use of a 

finely discretized ground-water-flow model and 

flow-path simulation by use of particle-tracking 

analysis. The areas contributing water to three 

water-supply wells, determined by simulation, 

were found to contain chemical-productmanufacturing or petroleum-refining land use. A 

relatively short minimum travel time of 30 years 

between the petroleum-refining land use and one 

of the wells indicates that contamination by 

volatile organic compounds is likely. The 

minimum traveltimes of greater than 50 years to 

the other two wells indicate that, to date, 

contamination by volatile organic compounds is 

not likely. These interpretations are consistent 

with available water-quality data. The results of 

the analysis also indicate that the two wells 

heretofore unaffected by contamination probably 

will be affected by contamination in the future 

and that the concentrations of volatile organic 

compounds in water from the affected well are 

likely to increase. The method of analysis used in 

this investigation could be applied to other locales 

provided that a finely discretized ground-waterU.S. Geological Survey, West Trenton, N. J.

flow model and high-resolution land-use data are 

available.

INTRODUCTION

Ground-water contamination can pose a 

significant threat to potable water supplies. Large 

areas sometimes can appear to be affected by the same 

or similar types of contaminants whose sources either 

are not evident or are too numerous to evaluate 

individually. The term "nonpoint-source 

contamination" has been applied to these situations. 

The ability to mitigate the problems created by 

contamination, however, is predicated on 

understanding the origins of the contamination and 

transport of the contaminants. Evaluation and 

mitigation of nonpoint-source contamination is 

difficult as a result of the uncertainties involved in the 

delineation of sources and transport characteristics.

Previous investigators (Vowinkel and Battaglin, 

1989b; Eckhardt and others, 1989) have approached 

the study of nonpoint-source ground-water 

contamination with the use of stochastic methods to 

relate land use within circular buffer zones 

surrounding wells to the quality of water in the well in 

order to discern possible sources of the contamination. 

This approach can be particularly useful to assess large 

amounts of data and determine the likely extent of 

contamination. With the stochastic approach, however, 

local characteristics of the ground-water-flow system 

can not be incorporated into the analysis. These 

characteristics include variations in hydraulic 

properties, location and intensity of withdrawals, or 

variations in hydrostratigraphy, and are especially 

important in determining the origins and transport of 

observed contamination. The determination of the

1011
areas contributing direct surficial recharge to the 

ground-water-flow system and the time of travel from 

these sources to a well is of particular interest in this 

analysis; however, these factors cannot be established 

adequately by means of the stochastic methods.

The Potomac-Raritan-Magothy aquifer system 

in southern New Jersey is an opportune case-study 

area to develop methods for evaluating the origins and 

transport of nonpoint-source ground-water 

contamination. The aquifer system is the primary 

water-supply source for the Camden, New Jersey, 

metropolitan area, including parts of Burlington, 

Camden, and Gloucester Counties, shown in figure 1. 

The Potomac-Raritan-Magothy aquifer system in the 

Cretaceous Potomac Group, Raritan Formation, and 

Magothy Formation is part of the southeastwarddipping wedge of Coastal Plain sediments composed 

of gravels, sands, silts, and clays of Cretaceous to 

Holocene age (Zapecza, 1989), and is locally 

differentiated into three components the upper, middle, 

and lower aquifers (figs. IB and 1C). These aquifers are 

separated by intervening leaky confining units.

About 100 million gallons per day of water is 

pumped from the aquifer system in this area (Camp 

Dresser & McKee, Inc., 1984), regionally creating a 

large cone of depression in the aquifer system's 

potentiometric surface. Land use in the outcrop area, 

contributing recharge to the aquifer system, is diverse 

and includes manufacturing, trade, residences, and 

agriculture, which have contributed contaminants that 

have been observed in regional-scale ground-water 

sampling investigations (Fusillo and others, 1984, 

tables 4, 5, and 6; Vowinkel and Battaglin, 1989a; 

Ervin and others, 1994}. The Planning Department of 

Gloucester County maintains maps that contain a 

detailed accounting of the land use in the outcrop area. 

These maps can be used to facilitate an investigation 

of nonpoint-source contamination in Gloucester 

County at a high resolution with the assumption that 

land use can be used as a surrogate for possible 

contaminant sources.

This paper summarizes the preliminary findings 

of an ongoing study to develop a method for 

investigating nonpoint-source ground-water 

contamination in areas of complex hydrogeology with 

multiple sites of ground-water withdrawal by use of 

ground-water-flow simulation in particular, the 

determination of areas contributing direct recharge 

from the land surface and the time of travel from these 

sources to a well.

METHODS

A ground-water-flow model of the PotomacRaritan-Magothy aquifer system in the Camden area 

was developed by Navoy and Carleton (in press), by 

using the MODFLOW computer code (McDonald and 

Harbaugh, 1988), as a tool to evaluate withdrawalmanagement strategies. The design of this model is 

appropriate to the objectives of this study, as the model 

grid is finely discretized in the aquifer system's 

outcrop area. The model consists of five aquifer layers 

in a quasi-three-dimensional representation of the 

hydrogeologic units. The upper, middle, and lower 

aquifers are each represented by a model layer. Two 

additional layers represent strati graphically younger 

aquifers overlying the confined aquifer system to 

account for vertical flow across the model boundary. 

The horizontal boundaries of the model are indicated 

in figure 1A. The horizontal flows across the 

boundaries of the model from adjacent areas in the 

Coastal Plain were derived from the more coarsely 

discretized Regional Aquifer-Systems Analysis 

(RASA) model of the entire New Jersey Coastal Plain 

(Martin, 1990). A no-flow boundary represents the 

crystalline rocks underlying the aquifer system.

The simulated aquifers were discretized 

horizontally into a variably spaced grid having 99 

rows and 106 columns, and oriented approximately 

parallel to the Fall Line (fig. 1A) and the strike of the 

geologic units. The smallest cells represent an area 

880 by 1,650 feet and are in proximity to aquifer 

system's outcrop and the Delaware River. The largest 

cells represent an area 2,200 by 3,300 feet and are 

generally located in the downdip, southeastern part of 

the model, away from the outcrop area. The model 

was calibrated satisfactorily to historical water-level 

data. Detailed information on the design and 

calibration of the model is available in Navoy and 

Carleton (in press).

Ground-water flowpaths were examined to 

determine the recharge contributing areas and estimate 

traveltimes within the ground-water system. This 

analysis was accomplished by using a particletracking technique. The simulated paths of 

hypothetical particles as they travel through the 

ground-water-flow field are determined as a function 

of time (Pollock, 1988). Porosities of 25 percent for 

the aquifer layers and 40 percent for confining units 

were used. A steady-state flow simulation was 

performed with average withdrawals for 1983-88 and 

ground-water flowpaths were determined with the

1012
A.

SUSSU , ^v fff

I* / * /^\ //\ / ^. » ' 

r .^ \ .' '-^ ', >

A \ * x / ^iT' °c /

".'^'It A It IT AH

/ iueKs l|N A-^7>

\ "^ V .<^_ *X

N

I V >> »-v' »;; -s ^'> N ^'^-^

^ssrx -^^\ \: yA .l.»>TO»^

"C^V^ "* x

* *.'/ vV, c'» -f* .«^'"i?i. i /

/ sC ** L "0\S *S>£|P v% "r*» V

- x ^^o^\'f "~^"-< C N \> CV \ *, ^7 T *.. -jAVr"-X ^ ' --

; '^vs /

 ^

SA y «</» ~/o 10 20 so Mites

A'

Not to scale Confining unit

B.

Outcrop of the 

upper aquifer of the 

Potomac-Raritan-Magothy 

aquifer system

Outcrop of the 

middle aquifer of the 

Potomac-Raritan-Magothy 

/ N aquifer system

Outcrop of the 

lower aquifer of the 

Potomac-Raritan-Magothy 

aquifer system

Figure 1. A. Location of study area, B. outcrop areas and location of simulated downdip limit of direct surficial recharge in the 

Potomac-Raritan-Magothy aquifer system, and C. generalized hydrogeologic section through the study area.

1013
MODPATH particle tracker (Pollock, 1989). The fixed 

withdrawal rates used for the particle tracking are not 

entirely realistic, from a regional perspective, because 

rates were lower in the past. In the vicinity of the 

outcrop, however, rates have been fairly stable for the 

past several decades (Camp Dresser & McKee, Inc., 

1984). Withdrawals from the aquifer system in the 

study area are not likely to increase substantially in the 

future because the New Jersey Department of 

Environmental Protection and Energy has initiated 

measures to restrict future increases in withdrawals 

(New Jersey Department of Environmental Protection, 

1986). Thus, these rates are likely to be representative 

of withdrawals from the aquifer system in the study 

area for the foreseeable future and extrapolation of the 

results of the flowpath analysis into the future 

probably is valid.

Under natural, predevelopment conditions, 

ground water in the study area discharged to the 

Delaware River and its tributaries. The aquifer system 

within the study area became a recharge zone with the 

advent of water-supply pumpage. Conceptually, a 

volume of the aquifer system now contains water 

whose origin is, in part, direct recharge from the land 

surface at the outcrop in the study area. Recharge also 

originates from streams, including the Delaware 

River; however, the recharge from the land surface is 

the focus of this investigation. The extent of the direct 

surficial recharge was evaluated with the particle 

tracker. Hypothetical particles were placed on the top 

of all model cells within the aquifer-system outcrop 

area. These particles were tracked forward into the 

aquifer. The part of the aquifer in which the particles 

traveled within a travel time of 40 years from the 

outcrop was designated the "recharge-influenced 

zone." Wells within this zone would be expected to 

produce water that could be affected by nonpointsource contamination.

Land-use maps depicting the aquifer system's 

outcrop area are available from the Gloucester County 

Planning Department at a scale of 1:4,800 and were 

derived from aerial photographs taken in 1990. These 

maps are used for property-tax and transaction 

purposes and, at that scale, individual lots are 

resolvable. The land uses on these maps are divided 

into nine major categories: residential, manufacturing, 

transportation, trade, services, recreation, resource 

development, undeveloped, and unknown. These 

categories are further subdivided into a total of 88 

specific land uses. The land use of greatest relevance

to this study of nonpoint-source ground-water 

contamination is chemical-products manufacturing 

and petroleum refining. These industries have been 

present in the area for 40 to 50 years and are the likely 

source of contamination by organic chemicals. Many 

such industries are present in the study area, which has 

a long history of observed contamination (Fusillo and 

others, 1984). The land-use maps depicting the aquifer 

system's outcrop area were digitized and a 

Geographical Information System was used to 

facilitate the determination of the land use at the origin 

of each flowpath evaluated by particle tracking. Each 

active model-cell face was the origin for 400 particles 

(a total of 2,400 particles per cell) used to "backtrack" 

the contributing flowpaths from each well studied.

CONTRIBUTING AREAS AND 

TRAVELTIMES

The simulated downdip limit of direct surficial 

recharge to the aquifer system within the study area, 

shown in figure IB, is from 0.5 to 2 miles downdip 

from (southeast of) the upper aquifer's outcrop area. 

This limit, based on simulation results, delineates the 

recharge-influenced zone and is consistent with 

observed regional ground-water quality. Chemical 

compounds related to human activities, such as 

volatile organic compounds (VOC's), heavy metals, 

and nutrients, generally were not detected in water 

samples from wells completed in the aquifer system 

more than 1 mile downdip from the outcrop (Vowinkel 

and Battaglin, 1989a).

Many water-supply wells are within the rechargeinfluenced zone that could be vulnerable to nonpointsource contamination. The contributing areas of three 

wells-15-069 (Greenwich Water Department #3), 15- 

348 (Greenwich Water Department #6), and 15-212 

(Paulsboro Water Department #4) were evaluated. All 

three wells are within about 1 mile of each other and are 

screened in the middle aquifer. In order for ground 

water from the closest part of the outcrop area to reach 

the well, it must flow through a leaky confining unit. All 

three of the wells were "strong" sinks that is, all flow 

across the boundaries of the cell containing the well was 

into the cell.

Only part of the outcrop area contributes 

water to an individual well. Because chemicals are 

manufactured in the study area, VOC's are found in 

the ground water as nonpoint-source contaminants. 

Figure 2 shows the contributing areas of the

1014
INSET A 

Well 15-069

, , 

^ 

V

I 

1 FT

012

o Figure 2. Locations of wells, contributing areas, and areas of manufacturing land use.

en
selected wells determined by simulation of the 

ground-water-flow system, and the locations of 

manufacturing land use in the study area. The 

manufacturing land use is devoted predominantly to 

chemical production and petroleum refining. Inset 

A in figure 2 shows the simulated contributing area 

to well 15-069. Manufacturing land use is present at 

the northwestern end of this contributing area. Inset 

B in figure 2 shows the simulated contributing area 

to well 15-348. A small area of manufacturing land 

use is present within this contributing area. The 

three small, isolated areas extending into the 

Delaware River contribute to this well. Inset C in 

figure 2 shows the simulated contributing area to 

well 15-212. Manufacturing land use is present 

about 0.5 miles from the well and is within this 

contributing area.

The shapes of the contributing areas (fig. 2) 

are fairly irregular and do not closely resemble the 

conceptualized patterns, such as fixed-radii circles, 

truncated parabolas, or "keyholes," that are 

associated with some wellhead-protection schemes 

(U.S. Environmental Protection Agency, 1987). The 

proximity of these production wells to each other 

causes drawdown interference that results in the 

"nested" aspects shown in the simulation results.

The land use within the contributing areas 

was determined from the land-use maps and is 

compiled in tables 1, 2, and 3 for wells 15-069, 15- 

348, and 15-212, respectively. These tables include 

the number of particles simulated and traveltimes to 

the wells from each category of land use. Chemicalproducts manufacturing and petroleum-refining 

land uses are highlighted in each table.

The primary land uses (those contributing 

more than 100 particles) contributing water to well 

15-069 (table 1) are undeveloped, residential, and 

agricultural. The average traveltimes from these 

land uses range from 21 years to 134 years. Of the 

2,400 particles per cell used in the analysis, 50 are 

contributed from chemical-products manufacturing 

land use. The minimum travel time from the source 

of these particles is 51 years. The number of 

particles is only loosely related to the proportion of 

the flow to the well. The particles were assigned 

uniformly to each face of the model cell, but the 

contribution of flow from each face to the well is 

not uniform.

The primary land uses contributing water to 

well 15-348 (table 2) are residential, undeveloped, 

agricultural, and recreational. The average 

traveltimes from these land uses range from 41 to 

157 years. Three of the particles that flow to the 

well originate from chemical-products 

manufacturing land use, with a minimum travel 

time of 109 years.

Petroleum refining is a primary land use 

contributing water to well 15-212 (table 3); this 

land use contributes 327 particles with a minimum 

travel time of 30 years. Other primary land uses are 

undeveloped, residential, transportation, and 

services. The average traveltimes from these land 

uses range from 61 to 94 years.

If particles from contributing areas that 

contain manufacturing land use are associated with 

traveltimes of less than 50 years, VOC 

contamination of the water in the receiving well is 

possible. Well 15-212 fits this scenario; a large 

number of particles that flow to this well originate 

from petroleum-refining land-use areas, with a 

minimum travel time of 30 years. This is consistent 

with observed water-quality data. VOC's detected 

in well 15-212 are listed in table 4. Much smaller 

numbers of particles that flow to the other two 

wells, 15-069 and 15-368, originate from 

manufacturing land-use areas, and minimum 

traveltimes are longer than 50 years. It is feasible to 

assume that insufficient time has elapsed for 

contamination to reach these two wells. This 

scenario also is consistent with observed data, as no 

VOC's have been detected at these wells. Future 

contamination can be predicted as follows: (1) The 

two wells containing water in which no VOC's have 

been detected are threatened with contamination 

because manufacturing land use is present within 

the contributing area and "contaminated" particles 

are "in route;" and (2) concentrations of VOC's in 

water from well 15-212, where VOC concentrations 

have thus far been low, probably will increase in the 

future because the average travel time from the 

manufacturing land-use area is 61 years, whereas 

the minimum is 30 years, indicating that additional 

contamination can be expected to appear in this 

well.

1016
Table 1. Land uses of area contributing water to well 15-069

[yr, year]

Nun

nar |/Cli

iber 

tide

652

12

3

3

^sHSftSs ISS&wSw?:?

4

5

2

1

1

4

1

8

5

4

7

II

3

20

5

433

4

190

510

151

305

1

5

Qf Average

S i » (yr)

44

20

46

152

S«SSKSSS?SSa^:SSSSKSSa*£a*Wjgffii

^^^^S^^^^K

16

19

68

20

109

30

34

200

49

23

26

47

146

37

21

83

73

28

134

50

8

24

Maximum 

9 travel time (yr)

262

56

56

169

86

26

21

68

20

196

30

86

261

53

26

30

59

170

42

62

100

249

154

166

131

8

29

Minimum Q 

travel time (yr)

8

9

30

124

,.. __ , ____: ,: ..__

8

18

68

20

26

30

24

177

44

21

24

24

118

34

8

68

8

8

35

23

8

18

eneral land-use

caegory

Residential

Residential

Residential

Residential

:K^^Wii^iff^^^^^^'^Ky'gffctjifiaiiilljiicreiglicBFSmaa

Transportation

Transportation

Trade

Trade

Trade

Trade

Trade

Services

Services

Services

Services

Services

Services

Services

Recreation

Resources

Resources

Undeveloped

Undeveloped

Undeveloped

Undeveloped

Unknown

Unknown

Specific land use

Detached Single Family

Semi-Attached Single Family

Apartments

Group Quarters

Automobile Parking

Water Supply

Wholesale

Food

Automobile, Aircraft, Boats

Eating and Drinking

Other Trade

Personal Services

Warehousing and Storage

Automobile Repair

Other Repair

Other Professional Services

Contract Construction

Miscellaneous Services

Recreational Activities

Agriculture

Mining Activities

Unused, Undeveloped

Forest Areas

Rivers, Streams, Creeks

Wetlands

Unknown

Unknown

Table 2. Land uses of area contributing water to well 1 5-348

[yr, year]

Number of 

particles

67

1,078

3

2

1

2

6

5

1

1

4

40

108

238

14

47

352

162

266

Average 

travel time (yr)

208

41

23

23

174

DiiMMittiKmwW

141

107

144

79

191

141

72

51

121

118

102

56

157

145

Maximum 

travel time (yr)

258

541

24

23

174

124

144

79

191

141

172

339

634

192

267

539

309

357

Minimum 

travel time 

(yr)

172

19

22

22

174

57

144

79

191

141

44

30

46

106

19

20

30

43

General land-use 

category

Unknown

Residential

Residential

Residential

Residential

ss!',i»:>iig;-»Ka&-:aK:iai^sig;af:»;;»sxa<-

 RlnSBHtBIwSSBHEiHHEil*

Transportation

Transportation

Trade

Services

Services

Services

Services

Recreation

Resources

Resources

Undeveloped

Undeveloped

Undeveloped

Undeveloped

Specific land use

From Pennsylvania

Detached Single Family

Semi-Attached Single Family

Apartments

Group Quarters

Motor Vehicle Freight

Automobile Parking

Furniture, Home Furnishings

Other Professional Services

Contract Construction

Government Services

Primary, Nursery Education

Recreational Activities

Agriculture

Mining Activities

Unused, Undeveloped

Forest Areas

Rivers, Streams, Creeks

Wetlands

1017
Table 3. Land uses of area contributing water to well 15-212

[yr, year]

Number 

of particles

327

57

21

^iSSS^SSW^^WK^Sr^S

36

155

94

66

57

54

3

99

20

38

162

54

22

10

322

288

32

156

Average 

travel 

time (yr)

63

50

63

39

84

59

58

71

81

80

57

78

60

61

228

64

564

69

83

175

94

Maximu 

m travel 

time (yr)

162

90

68

fc -* *49*^*'

162

82

161

113

549

88

161

117

88

113

565

90

962

161

549

569

138

Minimum 

travel 

time (yr)

29

29

60

......... ........^... ..........

30

30

30

53

30

76

30

57

32

30

50

55

463

29

30

61

62

General landuse category

Residential

Residential

Residential

HlliiiiiSlillilili

Transportation

Transportation

Trade

Trade

Trade

Trade

Trade

Trade

Services

Services

Services

Services

Services

Recreation

Undeveloped

Undeveloped

Undeveloped

Undeveloped

Specific land use

Detached Single Family

Semi-Attached Single Family

Mobile Homes

^j^^^P^IS^^^^^^SB^^^h^fa^jll^lg^si

Motor Vehicle Freight

Automobile Parking

General Merchandise

Food

Automobile, Aircraft, Boats

Apparel and Accessories

Eating and Drinking

Other Trade

Finance, Insurance, Real Estate

Personal Services

Warehousing and Storage

Government Services

Miscellaneous Services

Recreational Activities

Unused, Undeveloped

Forest Areas

Rivers, Streams, Creeks

Wetlands

Table 4. Volatile organic compounds detected in well 15-212

[Concentrations shown are highest concentrations detected in samples 

collected on 11-30-82, 10-22-86,4-23-87, 11-5-87, 2-11-88, and 8-10-88; 

Hg/L, micrograms per liter]

Compound concentration

Bromoform (total) 0.6 |ig/L

1,2-Dichloropropane (total) 0.4 |ig/L

Chlorodibromomethane (total) .4 |ig/L

Chloroform (total) .2 ng/L

1,2-Dichloroethane (total) .9 |ig/L

Tetrachloroethylene (total) .2 |ig/L

Toluene (total) .2 |ig/L

1018
CONCLUSIONS

A ground-water-flow model that incorporates 

particle tracking with high-resolution land-use data 

has been used to investigate the possible extent of 

nonpoint-source contamination at a regional level and 

to simulate the source and timing of contamination at 

production wells. Results are consistent with field 

data, indicating that the method is useful in the 

analysis of nonpoint-source ground-water 

contamination.

The method used to perform the analyses 

outlined in this report could be applied easily to other 

locales, provided that a finely discretized groundwater-flow model and a detailed land-use data base are 

available. Although either of these resources may not 

be considered commonplace, ground-water-flow 

models produced as part of many U.S. Geological 

Survey cooperative studies, such as those of the 

Regional Aquifer-Systems Analysis (RASA) program, 

could be used in a similar manner with an appropriate 

land-use data base.

REFERENCES

Camp Dresser & McKee, Inc., 1984, Population and water 

demand projections: Camden Metro Study, Task 3 

Report, report submitted to New Jersey Department of 

Environmental Protection, Division of Water Supply 

and Watershed Management: Boston, Mass., Camp 

Dresser and McKee, Inc., 53 p.

Eckhardt, D.A.V., Siwiec, S.F., and Cauller, S.J., 1989, 

Regional appraisal of ground-water quality in five 

different land-use areas, Long Island, New York, in 

Mallard, G.E., and Ragone, S.E., eds., U.S. Geological 

Survey Toxic Substances Hydrology Program-- 

Proceedings of the technical meeting, Phoenix, 

Arizona, September 26-30, 1988: U.S. Geological 

Survey Water-Resources Investigations Report 88- 

4220, p. 397-403.

Ervin, E.M., Voronin, L.M., and Fusillo, TV., 1994, Water 

quality of the Potomac-Raritan-Magothy aquifer 

system in the Coastal Plain, west-central New Jersey: 

U.S. Geological Survey Water-Resources 

Investigations Report 94-4113, 114 p.

Fusillo, TV, Hochreiter, J.J., Jr., and Lord, D.G., 1984, 

Water-quality data for the Potomac-Raritan-Magothy 

aquifer system in southwestern New Jersey, 1923-83: 

U.S. Geological Survey Open-File Report 84-737, 

127 p.

Martin, M., 1990, Ground-water flow in the New Jersey 

Coastal Plain: U.S. Geological Survey Open-File 

Report 87-528, 182 p.

McDonald, M.G., and Harbaugh, A.W., 1988, A modular 

three-dimensional finite-difference ground-water flow 

model: U.S. Geological Survey Techniques of WaterResources Investigations, chap. Al, book 6, 528 p.

Navoy, A.S., and Carleton, G.B., in press. Ground-water 

flow under current and future conditions, PotomacRaritan-Magothy aquifer system, Camden area, New 

Jersey: New Jersey Department of Environmental 

Protection and Energy, New Jersey Geological Survey 

Bulletin.

New Jersey Department of Environmental Protection, 1986, 

Procedures for implementation of Water Supply 

Critical Area No. 2: Trenton, N.J., New Jersey 

Department of Environmental Protection, December

1986. 9 p.

Pollock, D.W., 1988, Semianalytical computation of path 

lines for finite-difference models: Groundwater, v. 26, 

no. 6, p. 743- 750.

Pollock, D.W., 1989, Documentation of computer programs 

to compute and display pathlines using results from the 

U.S. Geological Survey modular three-dimensional 

finite-difference ground-water flow model: U.S. 

Geological Survey Open-File Report 89-381, 188 p.

U.S. Environmental Protection Agency, 1987, Guidelines 

for delineation of wellhead protection areas: 

Washington, D.C., U.S. Environmental Protection 

Agency, Office of Ground-Water Protection, June 22,

1987. 189 p.

Vowinkel, E.F., and Battaglin, W.A., 1989a,

Hydrogeological, well construction, and land-use 

factors that can affect the evaluation of regional 

ground-water quality, in Ragone, S.E., ed., 

Proceedings of the International Association of 

Hydrological Sciences, IAHS publication no. 182, 

p. 69-77.

Vowinkel, E.F., and Battaglin, W.A., 1989b, Methods of 

evaluating the relation of ground-water quality to land 

use in a New Jersey Coastal Plain aquifer system, in 

Mallard, G.E., and Ragone, S.E., eds., U.S. Geological 

Survey Toxic Substances Hydrology Program-- 

Proceedings of the technical meeting, Phoenix, 

Arizona, September 26-30, 1988: U.S. Geological 

Survey Water-Resources Investigations Report 88- 

4220, p. 405-410.

Zapecza, O. S., 1989, Hydrogeologic framework of the New 

Jersey Coastal Plain: U.S. Geological Survey 

Professional Paper 1404-B, 49 p., 24 pi.

1019
1020
Vulnerability of Water from Public Supply Wells to

Contamination by Pesticides, Potomac-RaritanMagothy Aquifer System, New Jersey Coastal Plain

By Eric F. Vowinkel 1

Abstract

A numerical rating model used in 

conjunction with hydrogeologic, land-use, and 

water-quality data stored in a geographic 

information system was developed to rank 558 

public supply wells screened in the PotomacRaritan-Magothy aquifer system in the New Jersey 

Coastal Plain into groups of high, medium, and low 

vulnerability to contamination by pesticides. The 

vulnerability of water from public supply wells to 

contamination by pesticides is defined by variables 

describing the (1) sensitivity of the aquifer to 

contamination and (2) intensity of pesticide use in 

areas where the aquifer is sensitive. The results of 

the model indicated that 305 wells were of low 

vulnerability, 237 were moderately vulnerable, and 

16 were highly vulnerable to contamination. 

Because analyses for pesticides in water samples 

are scarce, the concentration of nitrate in water 

samples was used as a surrogate to develop and test 

the accuracy of the model. Results of statistical 

tests indicate that the concentration of nitrate is a 

significant indicator of aquifer sensitivity, 

pesticide-use intensity, and the presence of 

pesticides in ground water. Results of the rating 

model indicate that water from only 1 of the 

138 sampled wells that were ranked as low 

vulnerability contained nitrate as nitrogen in 

concentrations larger than 0.5 milligrams/L. 

Concentrations of nitrate in water samples were 

significantly larger in the medium- and high1 U.S. Geological Survey, West Trenton, N.J. 08628

vulnerability groups than in the low-vulnerability 

group.

INTRODUCTION

Regulations established by the U.S. 

Environmental Protection Agency (USEPA) will 

require water purveyors to monitor samples from 

public supply wells for pesticides on a quarterly basis 

beginning in 1994. Waivers for monitoring of 

pesticides in water samples can be granted in two 

instances: (1) if the part of the aquifer from which the 

water is withdrawn is insensitive to contamination by 

pesticides, or (2) if pesticides are not used in the area 

near the wellhead where the aquifer is sensitive to 

contamination (J.B. Louis, New Jersey Department of 

Environmental Protection and Energy, written 

commun., 1993). The U.S. Geological Survey (USGS) 

is developing a geographic information system (GIS) 

data base in cooperation with the New Jersey 

Department of Environmental Protection and Energy 

(NJDEPE) to provide data on the vulnerability of 

water from public supply wells to contamination by 

pesticides. This paper demonstrates the use of the GIS 

data base and a numerical rating model to rank the 558 

known public supply wells screened in the PotomacRaritan-Magothy aquifer system, the most widely used 

aquifer system for public supply withdrawals in New 

Jersey, into groups of low, medium, and high 

vulnerability to pesticide contamination.

The hydrogeologic framework of (Zapecza,

1989) and regional ground-water flow in (Martin,

1990) the New Jersey Coastal Plain are well 

documented. The sediments of the Potomac Group and 

Raritan and Magothy Formations of Cretaceous age 

underlie the entire New Jersey Coastal Plain and crop 

out in a narrow 5- to-16-km-wide strip of land covering

1021
about 1,000 km2 from Raritan Bay to Delaware Bay 

(fig.l). The Potomac-Raritan-Magothy aquifer system 

consists of five hydrogeologic units three aquifers, 

designated as lower, middle, and upper; and two 

confining units. The Merchantville Formation and 

Woodbury Clay, which overlies the aquifer system, is 

the most extensive confining unit in the Coastal Plain. 

A long-term water budget for the Coastal Plain under 

unstressed conditions consists of about 112 cm/yr of 

precipitation, 51 cm/yr of streamflow, and 61 cm/yr of 

evapotranspiration. In 1985, withdrawals from public 

supply wells (fig. 1) screened in the aquifer system 

were about 6.9 x 106 m3/d.

Types of land uses and land cover in the outcrop 

area of the aquifer system are diverse. In the early 

1970's, the area consisted of about 49 percent urban 

land (25 percent residential and 24 percent urbannonresidential), 19 percent agricultural land, and 32 

percent undeveloped forests and wetlands. Some areas 

that were agricultural and undeveloped during the 

1970's are now urban areas. Because ground water 

moves slowly, however, the effects of land use on 

ground-water quality probably are more closely 

related to previous rather than current land uses.

Pesticide use in various land-use areas in was 

estimated from available data collected as part of the 

NJDEPE's Pesticide Control Program (PCP). Most of 

the agricultural land is cropland, .with smaller amounts 

of pasture and orchards. The amount of active 

pesticidal ingredients applied in 1991 in agricultural 

areas in or near the outcrop area of the aquifer system 

is estimated to be about 70,000 kg at an average rate of 

about 2.5 kg/km2 . The pesticides used in the largest 

amounts were sulfur, metolachlor, chlorothalonil, 

atrazine, captan, dichloropropene, and linuron. Golf 

courses also are a potential source of pesticide 

contamination to ground water. In 1990, reported 

pesticide use at nine golf courses located in or near the 

outcrop area was about 6,000 kg at an average rate of 

about 11.2 kg/km2 . The pesticides used in the largest 

amounts on golf courses were thiram, chlorothalonil, 

propamocarb, and iprodione. Pesticide use in 

residential and urban-nonresidential areas is reported 

to the PCP by commercial lawn-care companies, but 

estimation of application rates is difficult because data 

on applications by homeowners are unavailable. 

Pesticide-use rates may be larger jn some residential 

areas than those in agricultural areas (Roy Myer, New 

Jersey Department of Environmental Protection and

75°30' 75° 74°30' 74°

40°

39° 

30'

EXPLANATION

EXTENT OF OUTCROP AREA OF 

THE POTOMAC-RARITAN-MAGOTHY 

AQUIFER SYSTEM

VULNERABILITY RATING

Low

Medium

High

I 

NEW JERSEY

Delaware 

_ Bay

PENNSYLVANIA

Figure 1 . Locations of public supply wells screened in the Potomac-Raritan-Magothy aquifer system grouped by vulnerability 

to contamination by pesticides as determined by use of the rating model.

1022
Energy, oral commun., 1993). The most commonly 

used pesticides for lawn care in New Jersey in 1990 

were 2,4-D, MCPP and pendimethalin.

METHODS OF INVESTIGATION AND 

DATA-BASE DEVELOPMENT

The vulnerability of an aquifer to contamination 

by pesticides is defined by (1) the sensitivity of the 

aquifer to contaminants from the land surface and (2) 

the intensity of pesticide-use activities in areas where 

the aquifer is sensitive to contamination. The 

sensitivity of an aquifer is defined by its physical 

properties that determine whether contaminants can 

enter the ground water, whether they can be 

transported along a flow path, and whether the ground 

water is older than when pesticides were first applied 

at the land surface in the recharge area. For example, 

relatively young ground water in the outcrop area near 

the surface-water divide that is beneath soils 

containing little organic matter and less than 30 m 

from the land surface is more sensitive to potential 

sources of pesticides than ground water 2 km downdip 

from the outcrop area where the ground water 

recharged the aquifer system more than 50 years ago. 

Similarly, the intensity of pesticide use is not uniform 

among land-use areas; it is assumed that the potential 

for contamination of ground water is greater in 

recharge areas that are agricultural, golf courses, or 

residential than in those that are undeveloped.

Hydrogeologic, land-use, and water-quality data 

were stored in a GIS data base (table 1) to allow for 

spatial analysis of the different types of data. The 

numerical vulnerability model ranks wells on the basis 

of selected variables to describe aquifer sensitivity and 

intensity of pesticide use in areas where the aquifer is 

sensitive to contamination. The sensitivity and 

intensity variables were coded so that the scores for 

the sensitivity of the aquifer and intensity of pesticide 

increased with the vulnerability to contamination by 

pesticides. Each sensitivity and intensity variable was 

coded from a low of 0 to a high of 5, with 3 being 

moderate.

The data base and rating model serve as a 

means for a generalized, regional assessment of 

vulnerability for each well. Development of a groundwater flow model to determine the zone of influence of 

each well or to locate point sources of pesticide use 

that potentially could affect the quality of water in 

each well is beyond the scope of this project. Some of 

the variables used in the model may double-account

for certain factors; in recognition of this problem, the 

model is being refined as the project progresses.

Statistical procedures used to compare 

distributions of variables among groups include the 

Kruskal-Wallis test, Tukey's multiple comparisons 

test, and discriminant analysis. Discriminant analysis 

with ranked variables was used to determine those 

variables that best describe the presence of a 

contaminant in a concentration greater than or less 

than a given level. Relations among independent and 

dependent variables were considered significant at an 

alpha equal to or less than 0.05. Boxplots are used to 

compare the entire spread of data and percentiles by 

groups. Statistics were not computed for groups 

containing fewer than five cases.

Because analyses for pesticides were available 

for water samples from less than 5 percent of the 

public supply wells screened in the aquifer system, a 

surrogate was needed to test the vulnerability of a well 

to contamination by pesticides. The surrogate had to 

be (1) available in sufficient numbers to allow 

meaningful predictions and (2) an indicator of the 

presence of pesticides, the sensitivity of the aquifer 

system, and the intensity of pesticide use. Nitrate 

concentrations in water samples satisfied both these 

requirements. Water-quality analyses for nitrate in 

samples collected by the USGS from sampling ports 

near the wellhead were available for statistical 

analysis for 267 of the 558 public supply wells 

screened in the aquifer system.

A significant relation exists among 

concentration of nitrate, presence of pesticides, and 

well depth in water samples from 81 shallow wells 

(depth to top of the screened interval less than 30 m 

below land surface) in agricultural areas of the Coastal 

Plain (Louis and Vowinkel, 1989). The concentration 

of nitrate as nitrogen (as N) in water samples was 

significantly larger (p=0.003) in wells in which 

pesticides were present than in wells in which 

pesticides were absent (fig. 2A); the 25th-percentile 

nitrate concentration in samples containing a pesticide 

was about 6 mg/L. Results of discriminant analysis 

indicate that concentrations of nitrate and the 

frequency of pesticide detection decreased with well 

depth. Pesticide concentrations ranged from 0.1 to 13 

ug/L, and water from 1 of 81 wells contained a 

pesticide in a concentration that exceeded a USEPA 

maximum contaminant level or health advisory. The 

pesticides most commonly detected were carbamate 

(aldicarb and carbofuran) insecticides and triazine 

(atrazine and simazine) and acetanilide (alachlor and 

metolachlor) herbicides. These compounds are more

1023
Table 1. Description of sensitivity and intensity variables, assumptions, and numerical scores used in the numerical ranking 

model to determine the vulnerability of water from public supply wells screened in the Potomac-Raritan-Magothy aquifer 

system to contamination by pesticides

[USGS, U.S. Geological Survey; NWIS, National Water Information System; SCS, U.S. Soil Conservation Service; NJDEPE, New Jersey Department of 

Environmental Protection and Energy; PCP, Pesticide Control Program; UNO, undeveloped: RES, urban-residential: URB, urban-nonresidential; AGR, 

agricultural; <, less than; <, equal to or less than; -, no numerical score]

Numerical score

Variable and source of data Assumptions used for ranking score 0

Aquifer-sensitivity variables (Variables 3.4. and 5 applied only to wells in the outcrop area):

I. Location of well relative 

to the outcrop area (USGS/ 

NWIS)

Effects of land uses on ground water Downquality are greater in the outcrop dip 

area than downdip from the outcrop 

area (see fig. 2B)

Effects of land uses decrease as > 1,600 

the distance from the outcrop or 

recharge area increases

Ground-water flow paths are shortest >l ,600 

and water is youngest near the 

surface-water divide and the age 

of the ground water increases with 

distance from the divide 

Sorption of some pesticides >2.0 

increases as the amount of organic 

matter in the soil increases 

The depth of the well is a measure of >l 50 

the vertical distance a contaminant 

has traveled; concentration decreases 

with well depth 

Pesticide-use-intensity variables (Variables applied only to wells in the outcrop area):

In

2. Distance of the well from 

the outcrop area, in meters 

(Zapecza, 1989)

3. Distance of the well from 

the surface-water divide, 

in meters (USGS)

4. Lower extreme of soilorganic-matter content, 

in percent (SCS)

5. Depth to the bottom of the 

well, in meters (USGS/ 

NWIS)

I. Predominant land use 

within an 800-meter-radius 

buffer zone of the wellhead 

(USGS; Fegeas and others, 

1983)

2. Distance of the well from 

nearest agricultural land, 

in meters

3. Distance of the well from 

the nearest golf course, 

in meters (NJDEPE/PCP)

The land use with the largest areal UND

extent surrounding the well has the

greatest effect on the quality of water

in the well. Agricultural areas have a

greater effect on ground-water quality

than undeveloped, urban-residential,

and urban-nonresidential areas (see

fig. 2C)

The effect of agricultural land on > 1,600

ground-water quality is greatest in

areas where agricultural land is

near the wellhead

The effect of pesticide and nitrate use > 1,600

on water quality decreases as the

distance of the well from the golf

course increases

1,200

to

1,599

1,200

to

1,599

800

to

1,199

800

to

1,199

400

to

799

400

to

799

1

to

399

1

to

399

"

1 20

to

1 49

"

90

to

119

l.O 

to

1. 9

60

to

89

*

30

to

59

0 

to

0.9

0

to

29

URB

1,200

to

1,599

1,200

to 

1,599

800

to

1,199

800

to

1,199

RES

400

to

799

400

to

799

to 

399

1

to 

399

AGR

1024
soluble in water than the organochlorine insecticides 

that are measured in samples and reported by water 

purveyors to Federal and State agencies. The model 

assumes a similar relation among the presence of pesticides, nitrate concentration, and well depth in other 

land-use areas.

Variables used to describe the sensitivity 

(table 2) of the aquifer system to contamination were 

(1) location of the well relative to the outcrop area, (2) 

distance of the well from the outcrop area if the well is 

screened downdip from the outcrop, (3) percentage of 

organic matter in the soils at the wellhead, (4) distance 

of the well from the surface-water divide, and (5) 

depth of the well screen below land surface. The best 

predictor of nitrate concentration in ground water is

z

LU O ' 

O

rr

5

ten 

zS

O=j

30

Absent Present 

PRESENCE OF PESTICIDES IN THE SAMPLE

the location of the well relative to the outcrop area; the 

nitrate concentrations in water samples from wells are 

significantly larger (p=0.003) in the outcrop area than 

downdip from the outcrop area, where the 

concentrations rarely are greater than 0.1 mg/L 

(fig. 2B). All but one of the wells downdip from the 

outcrop area containing water with "outlier" nitrate 

concentrations are less than 800 m from the outcrop 

area. For wells in the outcrop area, negative relations 

between nitrate concentration and the percentage of 

soil organic matter and between nitrate concentration 

and distance from the surface-water divide were 

expected. Results of discriminant analysis indicate that 

nitrate concentration is significantly related to well 

depth but not to percentage of organic matter or 

distance of the well from the surface-water divide.

tw

B

Downdip from In

LOCATION OF WELL 

RELATIVE TO THE OUTCROP AREA

30

20

fee

g=i

10

Residential Nonresidential Agricultural Undeveloped 

PREDOMINANT LAND USE NEAR WELL

EXPLANATION 

* Outlier

99th percentile

75th percentile

Median

25th percentile

1st percentile

Figure 2. Distribution of concentrations of nitrate in water samples from wells screened in the Potomac-Raritan-Magothy 

aquifer system by the (A) presence or absence of pesticides in the water sample in agricultural areas (modified from Louis 

and Vowinkel, 1989, fig. 3); (B) location of the well relative to the outcrop area; and (C) predominant land use within an 800- 

meter-radius buffer zone of shallow wells (modified from Vowinkel, 1991, fig. 5).

1025
Table 2. Number of public supply wells and number of wells 

sampled for nitrate in the Potomac-Raritan-Magothy aquifer 

system, by groups of aquifer sensitivity and pesticide-use 

intensity

Aquifer sensitivity

Pesticide-use intensity Low Medium High

Low

Medium

High

305/138

0/0

0/0

1/0

49/30

6/3

41/18

110/69

16/9

Variables used to describe the intensity of 

pesticide use (table 2) near a well in an area that is 

sensitive to contamination were (1) predominant land 

use within an 800-m-radius buffer zone of the well, (2) 

distance of the well from the nearest parcel of 

agricultural land, and (3) distance of the well from the 

nearest golf course. Results of discriminant analysis 

indicate that the first two variables are significantly 

associated with the concentration of nitrate in water 

samples from wells in the outcrop area. A significant 

relation exists between the concentration of nitrate in 

shallow ground water and the predominant land use 

(fig. 2C) within an 800-m-radius buffer zone of the 

wellhead (Vowinkel, 1991; Vowinkel and Battaglin, 

1989). Nitrate was rarely detected in ground water in 

undeveloped areas, where use of nitrate and pesticide 

is low. Concentrations of nitrate in shallow ground 

water in agricultural areas exceeded 10 mg/L in more 

than 33 percent of the wells in agricultural areas. 

Concentrations of nitrate in shallow ground water in 

residential and nonresidential urban areas typically 

were between those in shallow ground water in 

undeveloped and urban areas.

In some cases, outlier concentrations of nitrate 

were present in water from wells in predominantly 

undeveloped and residential areas. A closer 

examination of the results from well sites associated 

with outlier nitrate concentrations indicates that the 

wells are near agricultural areas but agriculture is not 

predominant. A significant negative relation exists, 

however, between nitrate concentration and the 

distance of the well from agricultural areas. As a result 

of these findings, agricultural areas were assigned the 

highest pesticide-use intensity, followed by golf 

courses, residential areas, urban-nonresidential areas, 

and undeveloped areas. Also, it was assumed that as 

the distance from a golf course increases, the potential 

effects of pesticide use on ground-water quality 

decrease.

RESULTS OF VULNERABILITY ANALYSIS

The model was used to group 558 public supply 

wells into 9 subgroups on the basis of sensitivity of the 

aquifer system and the intensity of pesticide use where 

the aquifer is sensitive to contamination (table 2). The 

results of the model indicated that 305 wells (55 

percent) were of low vulnerability to contamination by 

pesticides because both the aquifer sensitivity and 

pesticide-use intensity are low. In this group, all the 

wells are more than 0.8 km downdip from the outcrop 

area. Only 16 wells (3 percent) were in areas where 

both the aquifer sensitivity and pesticide-use intensity 

are high; these wells typically are less than 30 m deep 

and in predominantly agricultural areas. The 

remaining 237 wells were identified as moderately 

vulnerable. The locations of the public supply wells 

are shown in figure 1, in which the wells are identified 

as having low, moderate, or high vulnerability to 

contamination by pesticides.

Concentrations of nitrate in water from wells 

were compared among the 9 subgroups (fig. 3); the 

number of water samples in each group is listed in 

table 2. The median concentration of nitrate (as N) in 

water samples from wells in the low-vulnerability 

group (low sensitivity and low intensity) was less than 

0.1 mg/L and the concentration was larger than 0.5 

mg/L in 1 of 138 public supply wells (0.07 percent). 

The spread of nitrate concentrations was greatest in 

wells grouped as moderate or high sensitivity and 

moderate or high intensity. The median nitrate 

concentration in water samples was largest in the highvulnerability group (high sensitivity and high 

intensity); however, results of a Tukey's multiple 

comparisons test indicate no significant difference in 

concentrations between the moderate- and highvulnerability groups.

Assignment of the 41 wells identified as having 

high sensitivity and low intensity to a vulnerability 

group was subjective. These were shallow wells in 

predominantly undeveloped areas that were not near 

an agricultural area or a golf course. Because the 

pesticide-use intensity near these wells is low, the 

wells could be ranked as low vulnerability; however, 

because the data base used to quantify land use within 

an 800-m-radius buffer zone of a well contained data 

from the early 1970's, land-use changes since the 

1970's potentially could affect the quality of water in 

the wells. Although nitrate concentrations in water 

from these wells were low, the wells were rated

1026
EXPLANATION 

Outlier

Wl52

<

O=f

g

i

LLL MMM HLM HMM HHH 

SENSITIVITY. INTENSITY. VULNERABILITY

99th percentile

75th percentile

Median

25th percentile

1st percentile

L=Low, M = Moderate, H = High

Example:

HMM = High aquifer sensitivity, 

moderate pesticide-use intensity, 

and moderate vulnerability

Sample size shown in table 2; 

must be at least 5 cases to 

generate boxplot.

Figure 3. Distribution of concentrations of nitrate in water samples from public supply wells screened in the PotomacRaritan-Magothy aquifer system, by groups of aquifer sensitivity and intensity of pesticide use.

conservatively as moderately vulnerable to 

contamination by pesticides.

The accuracy of the numerical rating model to 

assign nitrate concentrations to the low-, medium-, 

and high-vulnerability groups was good. Because the 

presence of pesticides and concentrations of nitrate in 

ground water are related, the numerical rating model 

also is assumed to be applicable to the determination 

of vulnerability of water from wells to contamination 

by pesticides. The best predictors of nitrate 

concentration were the location of the well relative to 

the outcrop area, the predominant land use 

surrounding the well, and the distance of the well from 

agricultural land. The percentage of the organic-matter 

content of the soil may be a good predictor of potential 

pesticide contamination but was not effective for 

classifying nitrate concentration.

Model limitations result from the following 

assumptions made in the analysis: (1) the solubility of 

all pesticides in water is similar to that of nitrate, (2) 

pesticides are applied uniformly in agricultural and 

residential land-use areas, (3) the relation among 

concentration of nitrate, presence of pesticides, and 

well depth in agricultural areas is similar to that in 

other land-use areas, and (4) the various scales of the 

different data layers in the GIS do not affect the results 

of the statistical analyses. Because of these limitations, 

the rating model can be used only for a preliminary 

assessment of the potential vulnerability of water from

public supply wells to contamination by pesticides. 

The limitations are being addressed as the project 

develops. The validity of the rating model will be 

assessed in the future by sampling a subset of the wells 

in each of the three vulnerability groups for selected 

pesticides and other water-quality constituents that 

may be predictors of pesticide contamination of 

ground water; expected and observed results will be 

compared.

SUMMARY

The vulnerability of water from public supply 

wells to contamination by pesticides is defined by 

variables describing the (1) sensitivity of the aquifer to 

contamination and (2) intensity of pesticide use in 

areas where the aquifer is sensitive. The results of the 

model indicated that 305 wells were of low 

vulnerability, 237 were moderately vulnerable, and 16 

were highly vulnerable to contamination. Because 

analyses for pesticides in water samples are scarce, the 

concentration of nitrate in water samples was used as a 

surrogate to develop and test the accuracy of the 

model. Results of statistical tests indicate that the 

concentration of nitrate is a significant indicator of 

aquifer sensitivity, pesticide-use intensity, and the 

presence of pesticides in ground water. Results of the 

rating model indicate that water from only 1 of the 138

1027
sampled wells that were ranked as low vulnerability 

contained nitrate as nitrogen in concentrations larger 

than 0.5 mg/L. Concentrations of nitrate in water 

samples were significantly larger in the medium- and 

high-vulnerability groups than in the low-vulnerability 

group.

REFERENCES CITED

Fegeas, R.G., Claire, R.W., Guptil, S.C., Anderson, K.E., 

and Hallam, C.A., 1983, Land use and land cover 

digital data: U.S. Geological Survey Circular 895-E, 

21 p.

Louis, J.B., and Vowinkel, E.F., 1989, Effect of agricultural 

chemicals on ground-water quality in the New Jersey 

Coastal Plain, in Weigman, D.L., ed., Pesticides in 

terrestrial and aquatic environments Proceedings of a 

National Research Conference, May 11-12, 1989: 

Blacksburg, Virginia, Water Resources Research 

Center, p. 80-89.

Martin, Mary, 1990, Ground-water flow in the New Jersey 

Coastal Plain: U.S. Geological Survey Open-File 

Report 87-528, 182 p.

Vowinkel, E.F., 1991, Comparison of relations between 

shallow-ground-water quality and land use in two New 

Jersey Coastal Plain aquifer systems, in Mallard, G.E., 

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 307-313.

Vowinkel, E.F., and Battaglin, W.A., 1989, Methods of 

evaluating the relation of ground-water quality to land 

use in a New Jersey Coastal Plain aquifer system, in 

Mallard, G.E., and Ragone, S.E., eds., U.S. Geological 

Survey Toxic Substances Hydrology Program-- 

Proceedings of the technical meeting, Phoenix, Ariz., 

September 26-30, 1988: U.S. Geological Survey 

Water-Resources Investigations Report 88-4220, 

p. 405-410.

Zapecza, O.S., 1989, Hydrogeologic framework of the New 

Jersey Coastal Plain: U.S. Geological Survey 

Professional Paper 1404-B, 49 p., 24 pi.

1028
A Review of Methods for Assessing the Potential for 

Regional Ground-Water Contamination

By David R. Seller1

Abstract

The potential for ground-water contamination, 

or the sensitivity of an area to ground-water contamination, can be estimated by a quantitative, 

deterministic model that predicts fate and transport of water and contaminants locally, or by a 

broad, regional qualitative model. Although 

regional models do not predict ground-water 

contamination at specific sites as accurately as do 

the local models, the former are useful tools for 

regional planning and research. All regional 

models depict relative contamination potential. 

Some models focus on contamination potential of 

ground water in aquifers, either at the land surface 

or confined beneath low-permeability sediments, 

whereas other models focus on ground water at the 

unconfined water table. For a given area, the focus 

of ground-water protection efforts will determine 

which models are most appropriate. Some 

regional models use dimensionless numbers and 

weighting factors to derive a relative measure of 

contamination potential, whereas others assemble 

source data into a composite map and rank the 

resulting map units for contamination potential 

according to a set of rules. Both approaches have 

certain advantages. However, because most 

contamination potential maps in the area considered (Central United States) show roughly similar 

patterns of contamination potential, some models 

may be more practical than others. For example, 

if simple models give results similar to models 

requiring cumbersome numeric manipulation of 

data, there is little justification for the effort 

needed to execute the more complex model.

U.S. Geological Survey, Reston, Va.

Availability of information required by a model 

also should determine its utility. Regional, stochastic models that provide actual, not relative, 

predictions of contamination potential are needed. 

Those models should consider the uncertainty in 

source information in order to estimate contamination potential with specified levels of uncertainty.

INTRODUCTION

Contamination of ground-water resources has 

become an issue of national concern. Many models for 

assessing the potential for ground-water contamination 

have been proposed and used. Two general approaches 

exist: farm- or field-scale deterministic models that 

predict rate of transport and fate of specific chemicals, 

and regional models that show general trends of the 

potential for ground-water contamination. Here, 

regional refers to areas larger than a local community 

or farm (for example, a county or a State). These two 

categories of models have little in common.

Deterministic models that require many quantitatively defined variables are developed to estimate rates 

and patterns of contaminant migration as precisely as 

possible (see Ertel, 1990, for a summary of models), at 

least for the locality where the model was developed. 

Given the large data requirements of some deterministic 

models, their applicability to a wide variety of localities 

can be limited. Also, these models have not been, and 

perhaps cannot be, applied to areas larger than farmscale with notable success, because they require variables that are not available regionally and because it is 

more difficult to characterize precisely the variability of 

geologic materials and resulting ground-water flow 

regionally than it is locally.

Although valuable data concerning water and 

contaminant movement are provided by site studies, 

a regional perspective is required by planners and 

researchers who must deal with geologic and hydrologic features not limited to specific localities or sites.

1029
Detailed, local information is necessary for precise 

characterization of an area or for remedial engineering 

but is not particularly helpful or efficient for identifying 

larger areas with potential contamination problems. 

Regional models allow for rapid contamination potential mapping of large areas by use of existing data. 

For practical reasons, such as limited availability of 

data, regional models must rely on a few generalized, 

qualitative variables.

Regional models are known to be inaccurate 

predictors of local ground-water contamination, 

because only the broad, regional nature of the geologic 

and hydrologic framework can be characterized in 

these models. However, there is at present no alternative that will predict contamination with the reliability 

of highly detailed deterministic models. Regulators 

and policymakers must, by necessity, contend with the 

shortcomings of regional models. Potential advances 

in quantifying regional information and in developing 

stochastic models that more accurately predict contamination could significantly reduce the uncertainty of the 

information upon which regulatory and policy decisions are based. Development of new approaches that 

reduce this uncertainty would be a significant scientific 

advance as well as an economic benefit to society.

In this paper, selected regional models are 

described and compared, and their predictive abilities 

are addressed. This discussion is limited to models that 

have been applied to a specific area the glaciated central United States to simplify the model comparisons. 

The importance of geologic framework is emphasized 

here, because geologic materials dictate hydrogeologic 

properties upon which interpretations of the availability of ground-water and potential for its contamination 

are based.

REGIONAL MODELS

In order to address the potential for regional 

ground-water contamination, a model may focus, for 

example, on characteristics of the soil, the geologic 

materials at and above the unconfined water table, or 

on the larger part of the geologic framework containing 

buried aquifers. A model's focus significantly affects 

the utility of maps produced by the model. A model 

focusing on the soil or vadose zone cannot adequately 

characterize the contamination potential of confined 

aquifers.

Models address the qualitative manipulation of 

data and relative ranking of contamination potential 

map units in one of two general ways: Some translate 

the data into dimensionless numbers for manipulation 

into a relative contamination potential index value, 

whereas others classify the data into map units of relative contamination potential without representing the 

data with numbers. The most commonly used model is 

"DRASTIC," an acronym for the model's elements, or 

factors (depth to water, net recharge, aquifer media, 

soil media, topography, impact of the vadose zone 

media, and hydraulic conductivity of the aquifer). 

DRASTIC maps have been generated for individual 

counties and states across the United States. It was 

developed by a committee using the Delphi approach 

to specify relative significance for each factor. A user's 

manual (Aller and others, 1987) provides descriptions 

of common hydrogeologic settings across the United 

States and expected values and the specified "weight," 

or multiplier, for each factor. With this information, or 

with actual information for a local area, DRASTIC 

scores can be calculated for each hydrogeologic setting 

and a DRASTIC map generated. Hydrogeologic settings can be defined to include buried aquifers, 

although the model does not address them as effectively as do other models (Berg and Kempton, 1984). 

Soller (1992) notes deficiencies in the model and, particularly, in the inconsistent manner in which it has 

been applied by different modelers. As with all 

regional models, DRASTIC was designed primarily as 

a regional tool for prioritization, or screening, to indicate areas that are relatively sensitive to contamination 

and, therefore, in need of more detailed mapping and 

evaluation or monitoring.

A model developed for the Wisconsin Department 

of Natural Resources (WDNR) was used to generate a 

statewide contamination potential map (Schmidt, 

1987). In this model, all unconfined ground water is 

considered, whether in permeable, sandy surface aquifers or in low-permeability till. Ground water in confined aquifers is not addressed. Data for the factors 

(soil texture, type of surficial deposits, depth to the 

water table, type of bedrock, and thickness of glacial 

drift) are assigned an arbitrary numbers based on perceived importance. Like DRASTIC, these numbers are 

weighted and summed to produce the contamination 

potential score. Unlike DRASTIC, hydrogeologic settings are not delineated; rather, maps of each factor are 

intersected in a Geographic Information System (GIS)

1030
to produce a composite map. This map does not have 

distinct map units but has color gradations representing 

changes in relative contamination potential.

For Minnesota, a similar approach was used to map 

the potential for ground-water contamination to depths 

including the uppermost aquifer (Porcher, 1989). Like 

the WDNR approach, factors such as aquifer materials, 

recharge potential, soil materials, and vadose-zone 

materials were ranked for their effect on the potential 

for ground-water contamination and combined in a GIS 

to yield a composite relative contamination-potential 

index.

In contrast to the numeric approaches for ranking 

map areas, a method to map areas of potential groundwater contamination developed by Berg and Kempton 

(1984) of the Illinois State Geological Survey (ISGS) 

orders map units in a hierarchy from relatively low to 

high contamination potential. Also in contrast to the 

numeric models, the ISGS model relies solely on the 

textural character of the three-dimensional geologic 

framework to a specified depth. The ISGS maps 

consider aquifers. This approach has been modified in 

cooperation with the U.S. Geological Survey (USGS) 

to address contamination potential of multi-State areas 

based on information less detailed than was available 

in Illinois (Seller and Berg, 1992).

Like the ISGS-USGS maps, a CIS-based map 

produced for Michigan (Lusch and others, 1992) also 

addresses ground water in aquifers. Relative soil permeability, presence of fine-grained materials within the 

upper 40 ft, and soluble (and therefore, presumably, 

permeable) bedrock within 50 ft of land surface are 

considered. Documentation on the model is limited, 

but it appears that aquifers deeper than 50 ft are not 

considered.

Pettyjohn and others (1991) offer the only national 

appraisal of the potential for ground-water contamination. They emphasize contamination from shallow injection wells and, therefore, address aquifers. The report 

provides only individual page-size maps of each State, 

so each State must be viewed in isolation. Regional 

integration of map information, across all States and 

with a common map scale, would be invaluable.

COMPARISON OF CONTAMINATION 

POTENTIAL MAPS

Lack of adequate information is a restraint to the 

development of any model. Modelers are compelled 

to use only the more commonly available types of data

and the models are thereby limited in their ability to 

predict contamination. Some data significant in predicting movement of water and contaminants (such as 

the presence of macropores, or organic carbon content 

of soils) are less readily available, and, therefore, not 

used in regional models. Data on depth to the water 

table are not commonly available yet are required by 

some models (the DRASTIC and the WDNR model, 

for example). Schmidt (1987), in a summary of the 

WDNR map, notes that depth to water table had the 

least influence on the model and was the least reliable 

component of information because it had to be estimated (not measured) for most of the state. Models 

that address buried aquifers must consider subsurface 

geologic units. This consideration may require detailed 

data that are scarce or difficult to acquire. For areas 

where such data are unavailable, the model must be 

able to make certain inferences on subsurface units on 

the basis of other criteria (overall thickness of deposits, 

for example) or to note on the map the lack of data for 

those areas (Lusch and others, 1992; Seller and Berg, 

1992).

Regional maps can depict contamination potential 

by assigning a rank to an area with mapped boundaries, 

such as an alluvial valley, or by combining model factors in a GIS to produce map units unrelated to any 

single factor mappable at land surface (such as soil). 

The DRASTIC model represents the first case, and the 

WDNR and Minnesota models represent the second. 

The ISGS and USGS-ISGS maps are somewhat intermediate: Map units represent a combination of several 

factors but retain the general configuration of surfacemappable units. The WDNR approach shows that contamination potential is a function of many factors and 

limits the effect of any one factor by producing a map 

that almost represents a raster image of contaminationpotential scores across the State. This approach limits 

the tendency to use contamination-potential scores literally for identifiable areas, such as a specific alluvial 

valley. On the other hand, retention of map boundaries, 

such as alluvial valleys, from the source data can allow 

a user to reinterpret the map for related purposes. 

Because the three-dimensional geologic framework 

can be derived from the ISGS and USGS-ISGS maps, 

they can be used to identify, for example, aquifer 

recharge areas or thick areas of till overlying shale that 

could serve as a waste repository. DRASTIC maps 

cannot be reinterpreted as readily, because source 

information is reduced to a single, dimensionless 

number.

1031
There are general similarities among the contamination-potential maps. The most important similarity 

is also the most important information on these maps 

the differentiation of relatively high- and low-contamination-potential areas. Minor differences in the 

ranked contamination potential for any area among the 

maps are commonly meaningless because the models 

that generated the maps address more or less the same 

issue by using generalized data and assumptions based 

on expert opinion, not statistical verification. There 

are, of course, mapped areas where significant differences in ranked contamination potential occur. For 

areas covered by till and underlain by a buried aquifer, 

the WDNR model would likely assign a low contamination potential to the area because of the till cover. In 

contrast, the ISGS model would assign a relatively high 

or a moderate rating, depending on depth of the buried 

aquifer. A DRASTIC map might consider the buried 

aquifer and assign a higher rating; however, Soller 

(1992) noted a serious inconsistency with which data, 

and especially buried aquifers, were treated by persons 

producing DRASTIC maps.

The choice of an appropriate model may be based on 

many criteria, including the following. First, what 

resource is to be addressed ground water in aquifers or 

unconfined, shallow ground water in both surface aquifers and low-permeability materials? Second, what is 

the distribution and quality of information shown on the 

model-generated map? The simpler models (for example, ISGS) yield maps similar to those from more intricate models, such as DRASTIC (which involve numeric 

manipulation of many types of data). In many cases, 

input data for DRASTIC are not even available, and 

estimated or standard values must be substituted. It is 

logical to suggest that simpler models, with appropriate 

factors selected, in part, by the distribution and quality of 

available data, would be preferable.

FUTURE EVOLUTION OF REGIONAL 

MODELS

Maps of regional ground-water contamination 

potential are best used only as qualitative tools, in part 

because of the sparseness of regional data or because of 

concern about the reliability (the uncertainty) of the 

data. With some exceptions (Halliday and Wolfe, 

1991), these models do not consider the areal distribution of contaminant sources or other potentially 

relevant factors, such as precipitation. If these factors

are considered, qualitative maps of ground water 

"vulnerability" to contamination from known sources 

can be produced. Some studies (Curry, 1987; Nielsen 

and Lee, 1987; Banton and Villeneuve, 1989; Baker, 

1990) have assessed the ability of contaminationpotential maps to predict actual contamination incidents or to approximate the results of deterministic 

models applied to detailed data at local sites. Results 

have not been encouraging, in part because regional 

models from which the maps are derived do not consider all factors pertinent to ground-water contamination. Also, regional water-quality and contaminant 

data may not be of sufficiently high quality and density 

to permit a regional, statistically valid assessment of 

contamination.

Regional contamination-potential maps might 

prove to be reasonably good predictors of actual contamination incidents. Statistical analysis provides the 

means to test these maps to determine their predictive 

abilities. However, spatial statistical analysis of such 

maps is inherently difficult. Lack of an adequate data 

base on ground-water quality contributes to the difficulty in testing the utility of these maps. Despite these 

problems, the science must advance beyond the current 

qualitative approach to stochastic models of the probability of ground-water contamination. New models 

need to consider the uncertainty of source information 

if they are going to provide estimates with specified 

levels of uncertainty. Estimates of contamination probability qualified by a measure of uncertainty will lead 

to more informed decisionmaking and more efficient 

management and regulation of our natural resources.

REFERENCES

Aller, Linda, Bennett, Truman, Lehr, J.H., Petty, R.J., and 

Hackett, Glen, 1987, DRASTIC A standardized system for evaluating ground water pollution potential 

using hydrogeologic settings: U.S. Environmental 

Protection Agency, EPA/600/2-87/035, 622 p.

Baker, D.B., 1990, Groundwater quality assessment through 

cooperative private well testing An Ohio example: 

Journal of Soil and Water Conservation, v. 45, 

p. 230-235.

Banton, Olivier, and Villeneuve, Jean-Pierre, 1989, Evaluation of groundwater vulnerability to pesticides: a comparison between the pesticide DRASTIC index and the 

PRZM leaching quantities: Journal of Contaminant 

Hydrology, v. 4, no. 3, p. 285-296.

1032
Berg, R.C., and Kempton, J.P., 1984, Potential for contamination of shallow aquifers from land burial of municipal : Champaign, Illinois State Geological Survey map, 

scale 1:500,000.

Curry, D.S., 1987, Assessment of empirical methodologies 

for predicting ground water pollution from agricultural 

chemicals, in Fairchild, D.M., Ground water quality 

and agricultural practices: Chelsea, Mich., Lewis 

Publishers, p. 227-245.

Ertel, Madge (compiler), 1990, Information exchange on 

models and data needs relating to the impact of agricultural practices on water quality, Workshop proceedings, 

U.S. Geological Survey Office of Water Data Coordination, cosponsored by U.S. Geological Survey, Agricultural Research Service, and Soil Conservation Service, 

Reston, Va., February 5-8, 1990: 217 p.

Halliday, S.L., and Wolfe, M.L., 1991, Assessing ground 

water pollution potential from nitrogen fertilizer using 

a geographic information system: Water Resources 

Bulletin, v. 27, p. 237-245.

Lusch, D.P., Rader, C.P., Barrett, L.R., and Rader, N.K., 

1992, Aquifer vulnerability to surface contamination in 

Michigan: East Lansing, Center for Remote Sensing, 

Michigan State University, scale 1:1,500,000.

Nielsen, E.G., and Lee, L.K., 1987, The magnitude and 

costs of groundwater contamination from agricultural

chemicals: U.S. Department of Agriculture Economic 

Research Service, Agricultural Economic Report #576, 

38 p.

Pettyjohn, W.A., Savoca, Mark, and Self, Dale, 1991, 

Regional assessment of aquifer vulnerability and 

sensitivity in the conterminous United States: U.S. 

Environmental Protection Agency, Robert S. KenEnvironmental Research Laboratory, EPA/600/2- 

91/043, 319 p.

Porcher, Eric, 1989, Ground water contamination susceptibility in Minnesota: St. Paul, Minn., Minnesota 

Pollution Control Agency, Accompanying map with 

same title, scale 1:1,000,000.

Schmidt, R.R., 1987, Groundwater contamination susceptibility in Wisconsin: Wisconsin Department of Natural 

Resources, Wisconsin's Groundwater Management 

Plan, Report No. 5, 27 p. map scale 1:1,000,000.

Seller, D.R., 1992, Applying the DRASTIC model a 

review of county-scale maps: U.S. Geological Survey 

Open-File Report 92-297, 36 p.

Seller, D.R., and Berg, R.C., 1992, Using regional geologic 

information to assess relative aquifer contamination 

potential an example from the Central United States: 

U.S. Geological Survey Open-File Report 92-694, 

scale 1:1,000,000.

1033
1034
Hydrologic and Geochemical Factors

affecting Metal Contaminant Transport

in Final Creek Basin near Globe, Arizona

By James G. Brown 1 and Judson W. Harvey2

Abstract

Acidic-mine drainage from copper mining 

in the Final Creek basin has contaminated the 

regional aquifer. Chemical reactions neutralize 

the contaminant plume as it moves downgradient. 

Neutralized contamination subsequently surfaces 

in a perennial reach of Final Creek and 

equilibrates with the atmosphere, causing pH to 

rise and manganese minerals to precipitate on the 

streambed. Ongoing research at the site is focused 

on ground-water hydrology and geochemistry, 

metal transport, and interactions between shallow 

ground water and streamflow.

Field and laboratory experiments will 

attempt to identify the reduction(s) that balance 

the oxidation of iron in the aquifer. Reactivetransport simulations of plume evolution and 

remediation will identify the important processes 

and reactions that produced the observed 

conditions in the plume. Measurements of gravity 

change and ground-water level change at project 

wells may provide refined estimates of aquifer 

specific yield for use in reactive-transport 

simulations. Estimates of ground-water ages by 

chlorofluorocarbon age-dating techniques may 

improve estimates of traveltimes and reaction 

rates of contaminants in ground water.

Mechanisms of trace-metal scavenging by 

manganese carbonates, manganese oxides, and 

iron oxides in the perennial stream are being

'U.S. Geological Survey, Tucson, Ariz. 

2U.S. Geological Survey, Menlo Park, Calif.

investigated, and partitioning coefficients are 

being estimated to constrain hydrogeochemical 

modeling. A coupled aquifer-stream massbalance approach has been used to constrain 

effective rates of solid-aqueous partitioning of 

manganese, iron, nickel, copper, and cobalt. 

Tracer experiments are being undertaken to 

distinguish reactions occurring in near-stream 

ground water from reactions in streamflow.

INTRODUCTION

The Final Creek basin, in central Arizona 

(fig. 1), has been an area of large-scale copper mining 

since the late 1880's. Acidic drainage from mining and 

related activities has contaminated the regional 

alluvial aquifer. The U.S. Geological Survey began a 

study of the contamination with the following 

objectives: (1) Define the nature and extent of groundwater and surface-water contamination, (2) identify 

the reactions and processes occurring in the 

contaminant plume, and (3) develop and test the 

ability of geochemical and transport models to 

simulate observed conditions. The purpose of this 

paper is to provide an overview of the site and a 

summary of past and ongoing research at Final Creek.

GEOHYDROLOGY

The drainage area of Final Creek above ") 9 Inspiration Dam is 516 km . About 170 km of this

area is underlain by a basin-fill conglomerate and a 

permeable stream alluvium, which together compose 

the regional aquifer. The aquifer is bounded laterally 

and at depth in most areas by relatively impermeable

1035
110°55' 50' 110°45'

33°35'

30'

25'

33°20'

EXPLANATION 

 INTERMITTENT STREAMS

DATA SITES Number is 

site identifier

70 Well(s)

09498380 A Streamflow-gaging 

station

1 Precipitation gage

3 MILES

3 KILOMETERS

Base from U.S. Geological Survey, 

1:24,000; Globe, 1945, Rockinstraw 

Mtn., 1988, Salt River Peak, 1988, 

and Inspiration, 1945

Figure 1. Area of study.

1036
METERS 

1,150

Webster 

Lake 

(dry)

1,100-

1,050 -

1,000 -

950-

900-

850-

EXPLANATION

SITE IDENTIFIER AND WELL 

NUMBER Circle denotes 

well screen and depth

STREAMFLOW-SAMPLE SITE 

AND NUMBER

FAULT Arrows show direction 

of movement

800

DATUM IS SEA LEVEL VERTICAL SCALE GREATLY EXAGGERATED

3 MILES

3 KILOMETERS

Figure 2. Generalized hydrogeologic section of the aquifer. Line of section approximates the principal ground-water 

flow line from Webster Lake to Inspiration Dam along the channels of Webster Gulch, Bloody Tanks Wash, Miami Wash, 

and Final Creek.

rocks. The aquifer thins and becomes narrower from 

well site 500 (fig. 2) north to Inspiration Dam. This 

thinning and narrowing decreases the volume of water 

that can move and be stored in the subsurface, and sustains perennial flow into Final Creek. Streamflow in 

Final Creek is perennial in a 5-to 6 kilometer reach 

above Inspiration Dam.

Basin fill ranges from unsorted and 

unconsolidated boulders to well-stratified deposits of 

firmly cemented sand, silt, and gravel (Peterson, 1962, 

p. 41). Carbonate content of the basin fill is about 

1.5 percent (Eychaner, 1989, p. 570). Horizontal 

hydraulic conductivity of basin fill is estimated to 

range from 0.1 to 0.2 in/day (C.G. Taylor, Magma 

Copper Corporation., written commun., 1987). Stream 

alluvium consists mainly of sand and gravel but 

includes stratified deposits of clay-sized material as 

much as 12 m thick locally. Horizontal hydraulic 

conductivity of alluvium between well sites 400 and 

500 was estimated to be 260 m/d. Alluvium at well site 

600 (fig. 2), through which neutralized, contaminated

water has passed, contained from 0.3- to 0.7- percent 

carbonate material.

Ground-water levels in stream alluvium and 

adjacent basin fill respond rapidly to variations in 

recharge and pumpage. Water levels declined steadily 

from 1986 to 1990 in response to less-than-average 

rainfall in the basin and increased pumping between 

Miami Wash and Final Creek for the purposes of 

remediation. The amount of ground-water decline 

generally decreased in a downstream direction. 

Ground-water levels rose as much as 16 m (fig. 3) 

adjacent to Miami Wash by February 1993 because of 

wetter than normal winters in 1991 and 1993.

NATURE AND EXTENT OF 

CONTAMINATION

Leakage of mine wastewater and process water 

from unlined impoundments into stream alluvium 

have contributed to a 15-kilometer-long acidic 

contaminant plume within the alluvium along Miami

1037
-cc

£

Q.

Z DC 

OUJ

3,000

2,000

1,000

0

12,000

10,000

8,000

6,000

4,000

2,000

A . A 

AAA ,

A WELL 51 

+ WELL 101

A + 

A +

- A A A A

I I I I

A WELL 51 

+ WELL 101

A A

A A 

A

A 

+ A

1984 1986 1988 1990 1992

Figure 3. Ground-water levels and concentrations of 

iron and sulfate adjacent to Miami Wash.

Wash and Lower Final Creek (fig. 4). Webster Lake, the 

single largest potential source of contamination, was used 

to store waste and process water from 1941 until 1988, 

when it was drained by the order of the U.S. 

Environmental Protection Agency. Contaminated ground 

water removed by remedial pumping is either spread on 

mine tailings to evaporate or is used at the mines.

Water from well 51, which is completed in the 

acidic core of the plume, contained more than 

500 mg/L iron (Fe), 32 mg/L copper (Cu), and 

2,700 mg/L sulfate (SO4) at a pH of 3.7 in 1992 

(table 1). Between site 450 and the perennial reach of 

Final Creek, neutralization reactions caused pH of 

ground water to increase from between 4 and 5 to 

about 6 in 1992. In February 1993, neutralized ground 

water discharged to a perennial reach that began less 

than 1 km above well site 500. As the neutralized

ground water discharges into the stream, gases 

dissolved in water reach equilibrium with the 

atmosphere. In the process, carbon dioxide 

degases, and pH rises to about 8 over a distance of 

about 5 km. Neutralized water from Final Creek at 

Inspiration Dam (Streamflow-gaging station 

09498400) contained 560 mg/L dissolved calcium 

(Ca), 2,800 mg/L dissolved SO4, and 39 mg/L 

dissolved manganese (Mn) in May 1992. Dissolved Fe 

and Cu concentrations were 0.01 and 0.006 M-g/L 

(micrograms per liter), respectively. From 1984 to 

1992, concentrations of Fe in water from well 51 

decreased from 3,100 mg/L to 500 mg/L, and 

concentrations of SO4 decreased from 10,000 mg/L to 

2,700 mg/L (fig. 3). These decreases were caused by

(1) removal of a significant source of contamination,

(2) recharge from winter runoff in 1991 -92, and

(3) chemical reactions in the subsurface.

RESEARCH IN PROGRESS

Ongoing research falls within three general 

categories: ground-water geochemistry and transport, 

ground-water hydrology, and hydrochemical 

interactions between ground water and surface water. 

These broad categories provide the framework in 

which to discuss present research at the site.

Ground-water geochemistry and transport

Although the most important reactions in the 

aquifer are well understood, the redox reactions 

responsible for the observed decreases in dissolved Fe 

have not been adequately characterized. The results of 

geochemical modeling (Glynn and others, 1991) 

suggest that the dissolution of Mn oxide solids could 

be the principal reduction that balances Fe oxidation. 

Reactions involving Mn and Fe indicate that 2 mol of 

Fe should precipitate for each mole of Mn that 

dissolves, even though column experiments by 

Stollenwerk gave the ratio of 7 to 1 (Eychaner, 199la). 

This ratio suggests that (1) Mn-oxide solids are 

reduced incongruently to form other Mn minerals, 

(2) other minerals or aqueous species might also be 

reduced, and (3) the atmosphere may be a significant 

source of C>2 to the plume. The incongruent dissolution 

of Mn-oxide solids and potential reduction of other 

minerals or aqueous species to balance the redox 

budget will be examined by a combination of 

laboratory work and geochemical modeling.

1038
METERS 

1,150-n

1,100-

1,050-

1,000-

950-

900-

850-

800

EXPLANATION

SITE IDENTIFIER

LINE OF EQUAL pH AND pH 

VALUE

POINT OF KNOWN pH

FAULT Arrows show direction 

of movement

Gaging

Station

094898400

Welded 

tuff

DATUM IS SEA LEVEL VERTICAL SCALE IS GREATLY EXAGGERATED

3 MILES

J

I ! ' I

0 3 KILOMETERS

Figure 4. Distribution of pH in the aquifer in 1992. Line of section approximates the principal ground-water-flow line 

from Webster Lake to Inspiration Dam along the channels of Webster Gulch, Bloody Tanks Wash, Miami Wash, and 

Final Creek.

Table 1. Chemical analyses of contaminated water in Final 

Creek basin

[Concentrations in milligrams per liter, except as indicated; °C, degrees 

Celsius; (.iS/cm, micromsiemens per centimeter at 25 degrees Celsius. 

Water filtered through 0.45-micrometer filter before analysis except where 

indicated. <, less than.

Date sampled

pH

Temperature (°C)

Specific 

conductance

(uS/cm)

Calcium (Ca2+) 

Magnesium (Mg2+)

Bicarbonate

(HC03-)

Sulfate(SO42-)

Copper (Cu2+ ) 

Iron (Fe2+) 

Manganese (Mn2+)

Well 51

5-27-92

3.7

1 9.0

3,740

400

87

0

2,700

32 

500 

1 6

Well 451

5-28-92

4. 1

1 9.0

3,890

400 

1 80

0

2,500

15 

280

72

Well 503

5-29-92

5.4

1 8.5

3,620

520 

220

65

2,200

<.020

.22

no

Station 

09498400

5-20-92

7.8

20.0

3,240

560 

1 40

1 30

2,800

0.006 

.0! 

39

Glynn and others (1991) simulated chemical 

reactions and transport in the subsurface with the 

model PHREEQM and found a tendency for the 

predicted pH to be higher than pH measured in the 

field. Exsolution of CO2 will be measured at the 

leading edge of the acidic part of the plume to 

determine the contribution of this process to the pH 

and redox balance and ultimately to contaminant 

transport. Preliminary gas measurements of CC>2 in the 

unsaturated-zone at site 500 indicate a diffusioninduced fractionation of 4.4 per mil for carbon I3C 

over a distance of 2.38 m between neutralized water at 

the water table and the land surface the fractionation 

expected for a steady-state upward flux of CO2 (Glynn 

and Busenberg, 1996). Pierre D. Glynn, (U.S. 

Geological Survey, National Research Program, 

Reston, Virginia), plans to measure exsolution and 

estimate the flux of CO2 at or near the leading edge of 

the acidic part of the plume.

1039
The testing, verification, and application of 

coupled geochemical-transport models also will help 

answer some of the remaining questions about 

subsurface-contaminant transport. General testing of 

reactive-transport models and the use of sensitivity 

analyses to define the most important reactions will be 

continued by Glynn, in cooperation with Peter 

Engesgaard (Technical University, Denmark) and 

Kenneth L. Kipp (U.S. Geological Survey, Denver, 

Colorado). James G. Brown, (U.S. Geological Survey, 

Tucson, Arizona), R.L. Bassett (University of 

Arizona, Tucson, Arizona), and Glynn plan to use a 

sensitivity- analysis type of approach to identify and 

determine the relative importance of chemical 

reactions, dilution, and transport that may control the 

remediation and chemical recovery of the regional 

aquifer. Brown, Bassett, and Glynn also plan to 

simulate transient, reactive transport in two 

dimensions, accounting for redox, adsorption, 

dissolution-precipitation, complexation, and 

hydrolysis.

Ground-Water Hydrology

Adequate simulation of reactive-solute transport 

through the subsurface depends on the accuracy of 

estimating recharge to the aquifer and subsurfacehydrological transport. D.R. Pool (U.S. Geological 

Survey, written commun., 1993) estimated changes in 

storage and specific yield in the alluvium on the basis 

of observed changes in ground-water level and gravity, 

and plans similar measurements between Miami Wash 

and Pinal Creek. Average values of specific yield of 

sediments that were saturated by water-level rises are 

from 0.16 to 0.21. Larger specific-yield values were 

associated with sediments containing large 

percentages of sand and gravel. Small specific yields 

were associated with sediments finer than sand.

Pierre Glynn and Eurybiades Busenberg (U.S. 

Geological Survey, National Research Program, 

Reston, Virginia) measured chlorofluorocarbon (CFC) 

concentrations in ground water to estimate groundwater ages in the basin. Preliminary results using 

CFC-11 for age-dating indicate that ground water from 

a depth of about 18 to 27 m in alluvium upgradient 

from contamination sources was recharged in 1979 

(Glynn and Busenberg, 1996). Planned additional 

measurements of CFC's may indicate if an assumption 

of steady-state flow for reactive-transport modeling is 

valid. Travel-time information derived from these 

measurements, combined with the results from 

geochemical modeling, may provide estimates of 

chemical reaction-rates at the site.

Metal Transport in Shallow Ground Water and 

Streamflow

The acidic ground-water plume has moved 

downgradient at a rate of about 0.25 km/yr over the 

past several decades (Eychaner, 199la). Higher-thannormal rainfall in 1991-93 increased ground-water 

levels by as much as 15 m and moved the headwaters 

of perennial flow upvalley. In early 1993, acidiccontaminated ground water in Miami Wash rose 

temporarily to the ground surface and discharged 

directly to streamflow. The headwater of perennial 

streamflow has since reestablished itself 1 km 

upstream from the location shown in figures 1 and 3. 

Increased hydraulic gradients and shortened groundwater flow paths may cause the acidic core of the 

plume to reach perennial surface water sooner than 

expected, depending on the transport of 

uncontaminated recharge water and the degree of 

mixing of uncontaminated recharge water with the 

contaminated water in the plume. Chemical massbalance models and ground-water-flow modeling, 

where needed, will be used by Harvey and others to 

investigate the consequences of enhanced interaction 

between ground water and surface water on mass-flow 

rates of contaminants out of the basin.

Neutralized contaminated ground water with 

large Mn concentrations (table 1, well 503) is being 

discharged to the perennial reach of Pinal Creek. High 

pCO2 and low pO2 of contaminated ground water is 

modulated by gas exchange with the atmosphere after 

ground water discharges to the stream. Resulting 

increases in pH and pe favor Mn oxidation (Eychaner, 

1991b). Concentrations of colloidal Fe and Mn 

presently in streamflow are low compared to total 

concentrations in solution (Harvey and Fuller, 1996). 

The absence of high Mn colloid concentrations 

suggests that Mn is partitioned to solid phases mostly 

on surfaces of streambed solids. Colloidal transport of 

Fe, Mn, and other trace metals appears in general to be 

only a minor mechanism of metal transport during 

baseflow in Pinal Creek. The role of Fe colloids in 

trace-metal transport will likely increase if the acidic 

core of the contaminated plume reaches Pinal Creek.

The character and chemical composition of 

manganese precipitates on the streambed at Pinal 

Creek vary considerably (Lind and Hem, 1993). J.D. 

Hem and C.J. Lind (U.S. Geological Survey, Menlo 

Park, California) are continuing their studies of 

chemical controls on manganese mineral formation at 

the site. Mn carbonate precipitation on streambed

1040
sediments appears to be favored initially in upstream 

reaches; Mn oxide precipitation becomes more 

important as pH and availability of dissolved oxygen 

both increase farther downstream (Hem and Lind, 

1996). Fe and Mn oxides and carbonate minerals on 

the streambed scavenge other dissolved trace metals, 

including nickel (Ni), copper (Cu), and cobalt (Co), 

from surface flow. Trace metals have an decreasing 

affinity for Fe oxides, Mn oxides, and carbonate 

minerals, in decreasing order (Lind and Anderson, 

1992). The greater abundance of Mn solid phases 

compared to Fe solid phases suggests that trace metal 

partitioning may be controlled by reactions with Mn 

phases. Chris Fuller (U.S. Geological Survey, Menlo 

Park, California) is investigating the relative roles of 

trace-metal coprecipitation and adsorption on Mn 

phases on the streambed as mechanisms for 

attenuation of dissolved Ni, Cu, and Co in streamflow 

(Fuller, 1996). The goals are to determine dominant 

chemical mechanisms of trace-metal partitioning to 

solid phases and to estimate partitioning coefficients 

for improved hydrochemical modeling.

Harvey and others (1996) showed that 

significant partitioning of Mn and Fe occurs between 

solid and aqueous phases in shallow ground water 

before ground water is discharged into Final Creek. A 

gain in Fe and loss of Mn from solution occurs in 

ground water that discharges laterally though the 

stream-channel banks at Final Creek; in contrast, 

ground water that discharges vertically through the 

creek bottom is unaltered by chemical reaction. A 

coupled mass balance for 500 m of stream and the 

surrounding aquifer in November 1992 demonstrated 

the importance of reactions in shallow ground water 

on downstream water chemistry in Final Creek.

Water exchange between streamflow and 

stream water-filled flow paths in streambed sediments 

was investigated at Final Creek by a bromide-tracer 

experiment. Exchange between streamflow and 

ground water affects the contact area and contact times 

of streamflow with streambed solids. Rates of Mn 

oxidation and trace-metal sorption on streambed 

surfaces, therefore, may be sensitive to surface waterground water hydrological exchange. Preliminary data 

are being used by Harvey, Fuller, and Conklin 

(University of Arizona, Tucson, Arizona) to design 

future experiments at Final Creek. Aspects of 

identifying the structure of improved hydrochemical 

models for Final Creek and of estimating model

parameters and associated uncertainties are outlined in 

Wagner and Harvey (1996).

SUMMARY

Acidic drainage from copper mining has 

contaminated the regional alluvial aquifer in the Final 

Creek basin. The acidic contamination is neutralized 

along the flowpath by reactions with aquifer materials 

and native ground water. Neutralized water discharges 

in a perennial reach of Final Creek; as the water 

equilibrates with the atmosphere, pH increases and Mn 

minerals precipitate on the streambed.

Ongoing research at the site is focused on 

ground-water geochemistry, hydrology, and solute 

transport and interactions between ground water and 

perennial streamflow. Field and laboratory 

experiments will attempt to identify the reduction(s) 

that balance the oxidation of Fe in the aquifer. Twodimensional reactive-transport simulations of plume 

evolution and remediation will identify the important 

processes and reactions that produced the observed 

conditions in the plume.

Measurements of changes in gravity and levels 

of ground water at project wells may provide refined 

estimates of aquifer specific yield for use in reactivetransport simulations. Estimates of ground-water ages 

obtained by CFC age-dating techniques may provide 

improved estimates of ground-water traveltimes and, 

when combined with results from geochemical 

modeling, provide estimates of geochemical-reaction 

rates at the site.

Field, laboratory, and modeling studies of 

contaminant transport are being conducted to improve 

understanding of the controls on metal transport in 

near-stream ground water and in streamflow. Previous 

work showed that Mn precipitates are variable and 

complex, and that precipitation is generally controlled 

by CO2 outgassing and reareation, which affects pH 

and the solubility of Mn solids. Mechanisms of tracemetal scavenging by Mn carbonates, Mn oxides, and 

Fe oxides are being investigated, and partitioning 

coefficients are being estimated to improve 

hydrogeochemical modeling. Tracer experimentation 

and a coupled aquifer-stream mass-balance approach 

are being used to determine transport components and 

to constrain net effective rates of solid-aqueous 

partitioning of Mn, Fe, Ni, Cu and Co in the system.

1041
REFERENCES CITED

Eychaner, J.H., 1989, Movement of inorganic contaminants 

in acidic water near Globe, Arizona, in Mallard, G.E., 

and Ragone, S.E., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Phoenix, Arizona, September 

26-30, 1988: U.S. Geological Survey Water-Resources 

Investigations Report 88-4220, p. 567-575.

Eychaner, J.H., 199la, The Globe, Arizona, research site 

contaminants related to copper mining in a 

hydrologically integrated environment, in Mallard, 

G.E., and Aronson, D. A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings 

of the Technical Meeting, Monterey, California, March 

11-15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 439-447.

Eychaner, J.H., 199lb, Solute transport in perennial

streamflow at Pinal Creek, Arizona, in Mallard, G.E., 

and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Monterey, California, March 11- 

15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 481-485.

Fuller, C.C., 1996, Expermental studies of trace-metal 

partitioning in perennial reaches of Pinal Creek, 

Arizona, in Morganwalp, D. W, and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Glynn, P.O., 1991, Effect of impurities in gypsum on 

contaminant transport at Pinal Creek, Arizona, in 

Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the Technical Meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 466-474.

Glynn, P.O., and Busenberg, E., 1996, Unsaturated zone 

investigations and chlorofluorocarbon dating of ground 

waters in the Pinal Creek basin, Arizona, in 

Morganwalp, D. W., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the Technical Meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Glynn, P.O., Engesgaard, Peter, and Kipp, K.L., 1991, Use 

and limitations of two computer codes for simulating 

geochemical mass transport at the Pinal Creek toxicwaste site, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the Technical Meeting, 

Monterey, California, March 11-15, 1991: U.S. 

Geological Survey Water-Resources Investigations 

Report 91-4034, p. 454-460.

Harvey, J.W., and Fuller, C.C., 1996, Association of metal 

contaminants with colloidal and suspended particulate 

material in shallow ground water and surface water at 

Pinal Creek, Arizona, in Morganwalp, D. W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Harvey, J.W., and Fuller, C.C., and Wagner, B.J., 1996, 

Interactions between shallow ground water and surface 

water that affect metal contaminant transport in Pinal 

Creek basin, Arizona, in Morganwalp, D. W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

Hem, J.D., and Lind, C.J., 1996, Chemical processes in 

manganese oxide and carbonate precipitation in Pinal 

Creek, Arizona, in Morganwalp, D. W., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the Technical 

Meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Lind, C.J., and Anderson, L.D., 1992, Trace metal 

scavenging by precipitating Mn and Fe oxides, in 

Kharaka, Y.F., and Maest, A.S., eds., Proceedings of 

the 7th International Symposium on Water-Rock 

Interaction, July 13-18, 1992: WRI-7, p. 397-402.

Lind, C.J., and Hem, J.D., 1993, Manganese minerals and 

associated fine particulates in the streambed of Pinal 

Creek, Arizona A mining-related acid drainage 

problem: Applied Geochemistry, v. 8, no. 1, p. 67-80.

Longsworth, S.A., and Taylor, A.M., 1992, Chemical,

geologic, and hydrologic data from the study of acidic 

contamination in the Miami Wash Pinal Creek area, 

Arizona, water years 1990-91: U.S. Geological Survey 

Open-File Report 92-468, 59 p.

Peterson, N.P., 1962, Geology and ore deposits of the 

Globe-Miami district, Arizona: U.S. Geological 

Survey Professional Paper 342, 151 p.

Wagner, B.J., Harvey, J.W., 1996, Solute transport

parameter estimation for an injection experiment at 

Pinal creek, Arizona in Morganwalp, D. W., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

Technical Meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey 

Water-Resources Investigations Report 94-4015.

1042
Dissolved Gas and Chlorofluorocarbon Content of 

Ground Waters in the Final Creek Basin, Arizona

By Pierre D. Glynn1 and Eurybiades Busenberg1

Abstract

Dissolved nitrogen and dissolved argon 

concentrations in ground waters at the Final Creek 

toxic waste site were used to calculate recharge 

temperatures and excess air trapped during 

recharge. The average recharge temperature was 

about 12± 2C ~ 6 to 7 °C colder than the average 

ground water temperatures. Excess-air concentrations were up to 11 milliliters per liter, indicating 

very rapid recharge probably during flooding 

events in the winter and early spring. This hypothesis is supported by local precipitation and air temperature records as well as by the 10 °C 

temperature recorded in Final Creek during a large 

recharge event in February 1993. Nitrogen production by denitrification is not thought to be significant. Dissolved oxygen concentrations are high in 

all uncontaminated ground waters, a finding consistent with low organic carbon contents in the 

aquifer. High carbon dioxide partial pressures 

(pCO2) were measured in the contaminated ground 

waters, particularly in the neutralized contaminated 

ground waters. Dissolution of carbonate minerals 

by acid-contaminated ground water is thought to be 

responsible for the high pCC>2 values. Chlorofluorocarbon-12 (CFC-12) and Chlorofluorocarbon-11 

(CFC-11) concentrations decrease with depth and 

distance downgradient in the metal- and acid-contaminated ground waters. CFC-12 concentrations 

are abnormally high and are thought to result from 

historically higher-than-normal atmospheric concentrations rather than from point-source groundU.S. Geological Survey, Reston, Va.

water contamination. Ground-water ages calculated from CFC-11 concentrations are reasonable ~ 

3 to 15 years for acid-contaminated ground waters 

and 20 to 30 years for neutralized contaminated 

ground waters. Deep uncontaminated ground 

waters have no significant CFC-11 or CFC-12 concentrations, and are believed to more than 50 years 

old. The CFC-11 recharge dates are consistent with 

available tritium, deuterium, oxygen-18, and 

dissolved-gas data.

INTRODUCTION

This paper presents preliminary data gathered 

during a field trip to the Final Creek toxic waste site in 

late November 1991. The primary objective of the trip 

was to determine the possibility of obtaining reproducible measurements of Chlorofluorocarbon (CFC) concentrations, at the picogram-per-kilogram level, in the 

metal- and acid-contaminated ground waters of the 

Final Creek basin (fig. 1). A secondary objective was 

to determine whether the measured concentrations 

could be used to age date the ground waters and, therefore, refine our understanding of the flow system. 

We believe that this paper presents the first measurement of CFC concentrations in metal- and acidcontaminated ground waters.

Dissolved-gas samples also were collected and 

analyzed for nitrogen (N2), argon (Ar), oxygen (O2), 

carbon dioxide (CC^) and methane (CH4). N2 and Ar 

measurements were needed to determine recharge 

temperatures used to calculate ground-water recharge 

dates from CFC concentrations. The dissolved-gas 

analyses also provide information on reactions occurring in ground water. This paper reports results of our 

dissolved-gas study and the implications of these 

results on knowledge of the inorganic processes occurring in the ground water.

1043
110°55'

33°35'

33°20'

1100 45'

Base from U.S. Geological Survey, 

1:24,000; Globe, 1945, Rockingstraw 

Mtn., 1986, Salt River Peak, 1986, 

and Inspiration, 1945

Figure 1. Area of study.

EXPLANATION 

A A' Line of trace

^- Direction of flow 

 ... Intermittent stream reach

Data sites, number is 

site identifier

 50 Well

1+ Precipitation gage

3 MILES

3 KILOMETERS

1044
We are grateful to James Brown, Steve Longsworth, and Dorrie Gellenbeck for their invaluable 

help in collecting the field data presented in this paper. 

We also thank Rosalynd Williams, Julian Wayland and 

Ray Van Hoven, for their laboratory analyses of our 

samples.

SAMPLING PROCEDURE

Dissolved-gas sampling procedures are 

described in Hobba and others (1977) and Pearson 

and others (1978). The dissolved-gas samples were 

collected after three well-casing volumes had been 

purged, after temperature, pH, and dissolved oxygen 

concentrations had stabilized, and after samples had 

been collected for standard chemical analysis. The 

wells were usually pumped at 10 gallons per minute 

(38 liters per minute) with a Grundfoss pump.

After removing the pump and its associated 

polyvinylchloride sampling pipe, a Keck pump with 

copper tubing was lowered into the wells. After a 10 to 

30 minute purge, six water samples per well were collected at 5 to 6 minute intervals for CFC analysis. The 

samples were sealed in glass ampules following the procedure described by Busenberg and Plummer (1992).

DISSOLVED-GAS CONCENTRATIONS IN 

PINAL CREEK GROUND WATERS

The dissolved-gas samples were analyzed by 

gas chromatograph in the laboratory for dissolved N2 , 

Ar, O2, CO2 , and CH4. CH4 concentrations were 

below the detection limit (0.1 mg/L) in all samples.

The use of CFC concentrations to assign 

ground-water recharge dates requires knowledge of 

the temperature of recharge as well as the amount of 

excess air entrapped during the recharge process. Measurements of dissolved nitrogen and dissolved argon 

concentrations provide a convenient method for 

estimating these two parameters (Busenberg and 

Plummer, 1991; Busenberg and others, 1993).

Some of the factors affecting dissolved N2 and 

Ar concentrations measured in ground water are (1) 

the composition of the gas phase at the time of watergas equilibration, (2) the relative humidity of the gas 

phase, (3) the temperature at the time of equilibration,

(4) the barometric pressure at the time of equilibration,

(5) the amount of excess air (or soil gas) entrained 

during the recharge process, (6) the ionic strength of 

the water, (7) the amount of N2 produced by denitrification reactions in the ground water, (8) gas loss by 

exsolution during sampling, (9) gas-exchange processes near the water table after recharge, (10) diffusion and dispersion of dissolved gases in ground water, 

and (11) mixing of waters with distinct dissolved-gas 

contents during sampling. The assumptions made in 

calculating recharge temperatures and excess-air concentrations from dissolved N2 and Ar concentrations 

are that factors 6 through 11 can be neglected. We also 

assume that the gas phase at the time of equilibration 

has average air composition, and that the barometric 

pressure at the time of equilibration can be calculated 

from an estimate of the applicable recharge elevation 

and a standard function relating atmospheric pressure 

to altitude. N2 and Ar solubilities as a function of temperature were calculated by use of equations and constants given in Wilhelm and others (1977).

Dissolved N2 and Ar concentrations are shown 

in figure 2A. Table 1 provides a summary of the well 

characteristics associated with the well numbers given 

in the graph. Recharge temperatures and excess-air 

concentrations calculated from the N2 and Ar data are 

shown in figure 2B. Calculations were made assuming 

a minimum recharge elevation of 900 m and a maximum elevation of 1374 m. The N2 and Ar results for 

well 52 are suspect because they lie outside the generally linear trend of N2 and Ar concentrations; the Ar 

concentration is disproportionately high compared to 

the N2 concentration, and the calculated recharge 

temperature is unrealistically low.

Recharge elevation affects calculated recharge 

temperature but does not significantly affect calculated excess-air concentration, as shown in figure 2B. 

Changing recharge elevation results in a difference of 

0.12 mL/L in excess-air concentration. In comparison, 

calculated temperatures for any given water sample 

vary by about 2 °C, depending on the recharge elevation used. Measurement uncertainties in the N2 and Ar 

analyses also correspond to a 2 °C uncertainty in the 

calculated recharge temperatures. Considering these 

uncertainties, the fact that all the wells sampled for 

CFC's were not sampled for dissolved gases, and the 

lack of duplicate samples, we decided to use an average recharge temperature of 12 °C to calculate 

recharge dates from CFC concentrations. This temperature is markedly lower than the average ground-water 

temperature of 18 to 19 °C, which indicates that 

recharge must occur predominantly during winter and 

early spring. This hypothesis is supported by the 10 °C 

temperature of a water sample from Final Creek

1045
Table 1 . Characteristics of ground water wells

fy* 0.85

LLJ

K

DC

CO

fyi

0 0.75

_l

2

Z

- 0.70

O

^

I 0.65

LLJ

O

Z 

O

O 0.60

Z 

O

O

DC

< 055

1 1 1 1 ! 1 A

404 °°C A

/

/

/ " 52

/

/ 105

5°C A'' 504. S ~ 

i /

i / 

i /

c»r / "51 /

Y /

j ^402 /

10 C £ "302 /

i " 503 /

' 7(19 /

10°C / 70Z /

/«. 452

15°,' 010

A I I I I I I

sampiea in me nnai L»reeK oasin in iNovemoer iy»i .

[depths and screen lengths in meters; a, acid-contaminated water;

u, uncontaminated background water; n, neutralized contaminated

water; bf, basin fill; al, alluvium]

well

number

010

2

53

51

102

103

101

105

304

302

402

401

404

452

451

14 16 18 20 22 24 2B 28 595

NITROGEN CONCENTRATION, 50 ,

IN MILLIGRAMS PER LITER 503

EXPLANATION 502

 ^^- Excess air vector. 504

402 Data point and well number. -70 1

A - - - O Dashed line with open triangles and open circles -JM

represent equilibrium with air at 900 meters and

Well Screen Screen

depth depth length

27.1 27.1 9.1

19.8 19.7 .9

27.8 27.7 .9

33.4 33.3 .9

25.2 25.1 .9

25.3 19.0 .9

36.1 36.0 .9

48.8 48.1 .9

30.3 29.6 .9

35.8 35.7 .9

20.9 20.7 .9

34.2 34.1 .9

55.3 54.6 .9

8.2 8.2 3.0

24.4 24.4 2.9

6.7 6.7 1.5

17.0 16.3 .9

25.3 24.1 .7

38.0 37.4 .9

69.2 68.6 1.0

4.7 4.7 .9

7.3 7.3 .9

Aquifer Water

material type

bf u

al a

bf a

bf a

al a

al a

bf a

bf n

al a

al a

al a

bf a

bf u

al a-n

bf a

al n

al n

al n

bf n-u

bf u

al n

al n

Figure 2 (A). Dissolved argon and nitrogen concentrations 

measured in Final Creek ground waters. The excess-air 

vector and the line representing equilibrium with air as a 

function of temperature and elevation are given for 

reference.

(unpublished data in the files of the U.S. Geological 

Survey, Tucson district office). The sample was 

collected during a large recharge event (at a flow of 

1000 cubic feet per second in Final Creek) at Inspiration Dam in February 1993 (fig. 1).

Our calculated recharge temperatures are supported by monthly precipitation and air-temperature 

records for Miami, Arizona, provided by the National 

Weather Service. The records indicate that, on average, 21 cm of precipitation (out of a yearly total of 

50 cm) fall during the months of December through 

March and that the mean air temperature during that 

period is 8.5 °C. The other wet period of the year 

occurs in July and August and contributes on average 

14 cm of precipitation. The average air temperature 

during July and August is 27.8 °C, and evaporation is 

undoubtedly significant, although flash floods are 

common during that period.

The calculated excess-air concentrations are 

relatively high, compared to concentrations found for 

shallow ground waters in the Delmarva peninsula 

(generally less than 3 mL/L; Dunkle and others, U.S. 

Geological Survey, written commun., 1993), but are 

consistent with results from other semiarid areas, 

where rapid recharge occurs during intense precipitation and flooding events. For example, ground waters 

in the western Kalahari in South Africa contain 4 to 

16 mL/L of excess air (Heaton and others, 1983).

Ground waters from the deepest wells sampled 

(404, 504 and 105) contain the highest calculated 

excess-air concentrations up to 11 mL/L and, 

together with water from well 52, have the highest N2 

and Ar concentrations. Water from wells 404 and 504 

is uncontaminated, whereas water from well 105 is 

neutralized contaminated water. The dissolved oxygen 

data, shown in figure 2C, show that ground water at all 

three uncontaminated wells (010, 504,404) have high 

dissolved C>2 concentrations and the lowest specific 

conductance of all ground waters sampled. Water from 

well 105 has a significant O2 concentration that may 

represent a mixture of background uncontaminated

1046
18

16

C/)

LJJ 

O

LJJ 

DC

LJJ 

O

10 u

LJJ 

DC

LJJ

LJJ

B

A 

503 \ A

702: :302 30 

010;

105

451 '. , 402

A A 

i 51 i504

404 ;

52

0 2 4 6 8 10 12

EXCESS AIR CONCENTRATION, 

IN MILLILITERS PER LITER

EXPLANATION

A RECHARGE AT 900 METERS ABOVE SEA LEVEL 

 RECHARGE AT 1372 METERS ABOVE SEA LEVEL 

52 WELL NUMBER

Figure 2 (B). Excess-air concentrations and recharge 

temperatures calculated for Pinal Creek ground waters from 

the dissolved argon and nitrogen concentrations.

water and neutralized contaminated water. The calculated equilibrium partial pressures of O2 and N2 at 

the measured ground-water temperatures (generally 

18-19 °C, 15.5 °C for well 010) are shown in 

figure 2D.

The atmospheric partial pressures of N2 and O2 

are 0.69 and 0.19 atm at an elevation of 950 m (mean 

atmospheric pressure is 0.893 atm), and are even 

lower at higher elevations. Consequently, N2 partial 

pressures (pN2) in all sampled ground waters 

exceeded mean atmospheric pressure, and N2 was 

likely to degas if the ground waters were brought in 

contact with the atmosphere. Degassing of these 

ground waters could also possibly occur insitu at 

depths of up to 2 m below the water table (Glynn and 

Busenberg, 1996). Waters from wells 404, 504 and 

105 had the highest pN2 values (up to 1.3 atm). Waters 

from these wells also have high dissolved Ar concentrations in proportion to dissolved N2 concentrations,

u_ OU

LJJ

DC

LJJ

Q_ 10

0)

«<

DC

 i ^

Z

1

z

0

i

1

z

LJJ

O

z

O

O 0.1

z

LJJ

§X

f~\ nrw

i i i i

C

" 404 ~

: B 504 ;

 010

-

- -

105 

_ 

~

-

452 

302 - 

451 

m m m

\ \ R1

52 702 \ \

 \ > 

: 402 302 :

_

-

I I

1000 2000 3000 4000 5000

SPECIFIC CONDUCTANCE,

IN MICROSIEMENS PER CENTIMETER

AT 25 DEGREES CELSIUS

EXPLANATION 

404 Data point and well number.

Figure 2 (C). Dissolved oxygen concentrations in Pinal 

Creek ground waters as a function of specific 

conductance. Specific conductance is an overall measure 

of the degree of contamination .

have low dissolved organic carbon and nitrate/nitrite 

concentrations, and have relatively high dissolved O2 

concentration. Therefore, excess air trapped during 

recharge, rather than dissolved N2 production by denitrification reactions, is thought to be the cause of the 

high pN2 values observed.

The calculated equilibrium partial pressures of 

CO2 as a function of pH are shown in figure 2E. Apart 

from the uncontaminated ground waters at wells 404, 

504 and 010, and the contaminated but neutralized 

water at well 702, all other ground water samples also 

contained high pCO2 values well in excess of the 

atmospheric partial pressure (0.031 atm at an elevation 

of 950 m), as shown in figure 2E. Water from these 

wells also had pH values less than 6, with the exception of the neutralized water in well 105, which had a 

pH of 6.4. At pH values less than 6.3, dissolved CO2 is 

the dominant form of dissolved inorganic carbon, and 

acidification of an uncontaminated ground water will

1047
I

Q.

404

504

010

702

105

503

452

402

52

51

451

302

0 0.05 0.1 0.15 0.2

CARBON DIOXIDE PARTIAL PRESSURE, 

IN ATMOSPHERES

EXPLANATION 

404 Data point and well number.

CO 

LJJ 

CC 

LJJ

CO 

O

0.5

° 2

LJJ 0.05

OC

ID

CO

CO

LJJ

OC 0.02

Q.

OC 

Q.

LJJ 0.005 

O

0.002

404

504

010

105

452

451

402

503

302

702

I 52

0.7 0.8 0.9 1 1.1 1.2 1.3

NITROGEN PARTIAL PRESSURE, 

IN ATMOSPHERES

EXPLANATION 

404 Data point and well number.

1.4

Figure 2 (D). Equilibrium partial pressures of oxygen and 

nitrogen calculated from dissolved-gas concentrations for 

Final Creek ground waters at ground-water temperatures 

measured in the field.

by itself increase the equilibrium pCO2 of the water. 

As the acid-contaminated ground waters (with already 

elevated pCO2 values) become progressively neutralized (pH increases), pCO2 values increase further as a 

result of the dissolution of carbonate minerals in the 

aquifer materials, as shown in figure 2E. The high 

pCO2 values are thought to be responsible for a flux of 

approximately 9x10~3 (mol/m2 )/d of CO2 through the 

unsaturated zone at well sites 450 and 500 (Glynn and 

Busenberg, 1996).

CHLOROFLUOROCARBON CONTENTS 

AND RECHARGE DATES OF FINAL 

CREEK GROUND WATERS

CFC concentrations can be used to date ground 

waters if the concentrations measured reflect the equiFigure 2 (E). The relation between measured pH and partial 

pressures of carbon dioxide calculated for Final Creek 

ground waters (at field temperatures).

librium that was present between the atmospheric CFC 

concentrations and the ground water at the time of 

recharge. CFC-11 and CFC-12 concentrations have, 

since the 1940's and until recently, increased exponentially in the atmosphere. Therefore, recent ground 

waters should have higher CFC concentrations than 

older waters. CFC concentrations can not be used to 

age-date ground waters if (1) contamination from 

nonatmospheric sources has occurred, (2) local atmospheric concentrations differ significantly from average 

continental concentrations, (3) significant CFC 

degradation occurs in the ground-water environment 

(degradation occurs only in highly reducing environment and affects CFC-11 more strongly than CFC-12) 

and, (4) significant sorption of CFC's occurs (of concern in the presence of organic-rich aquifer materials 

and affects CFC-11 concentrations more strongly than 

CFC-12 concentrations). Busenberg and Plummer

1048
METERS A 

1,150

Former 

Acidic Lake

EXPLANATION:

Well nest showing individual 

well sampling points with 

CFC-11 concentrations in 

picograms per kilogram of 

water (standard deviations 

in parentheses). Well nest 

identifier is given above well

CONSOLIDATED 

BASIN FILL

WELDED TUFF ?

900 -

850 -

800

0246 

DATUM IS SEA LEVEL

12 14 

KILOMETERS

16 18 20 22 24 26 

VERTICAL SCALE GREATLY EXAGERATED

Figure 3. Hydrogeologic section showing chlorofluorocarbon-11 concentrations in Final Creek ground waters.

(1992), Plummer and others (1993) and Busenberg 

and others (1993) give additional information on the 

use of CFC's in age-dating ground waters.

Acid-contaminated ground waters at the Final 

Creek site have low dissolved organic carbon concentrations (4.2 mg/L), but these concentrations were 

nevertheless found to be up to eight times greater than 

the concentrations in background uncontaminated 

waters in 1988 and 1989 (Reese and Bassett, 1990). 

The presence of polycyclic aromatic hydrocarbons, 

found in the more acidic ground waters at the Final 

Creek site, indicated an anthropogenic source for the 

increase in dissolved organic carbon (Wallin and others, 1991), and raised the possibility of localized 

anthropogenic contamination of CFC's; such contamination would have precluded the use of CFC concentrations to assign ground-water recharge dates. 

Fortunately, aquifer materials at the site contain low

concentrations of organic carbon, and the ground 

waters are not strongly reducing. This knowledge minimized concerns about possible CFC biodegradation 

and sorption.

The concentrations of CFC-12 and CFC-11 are 

shown in figures 3 and 4. The CFC-12 concentrations 

commonly exceed concentrations expected for water 

at equilibrium with 1991 average continental 

atmospheric concentrations and, with a few exceptions, can not be used to date the ground waters. 

Nevertheless, CFC-12 concentrations seem to 

decrease both as a function of depth and distance 

downgradient, as shown in figure 3. The highest CFC12 concentrations near the water table are about 315 

pg/kg (wells 52,304,402), with the exception of water 

in well 103 (648 pg/kg). The concentrations in waters 

from wells 52, 304 and 402 are close to the concentration expected for a background water recharged in late

1049
METERS A 

1,150

1,100

1,050 -

1,000 -

950 -

900 -

850 -

800

Former 

/ I Acidic Lake

EXPLANATION:

Well nest snowing individual 

well sampling points with 

CFC-12 concentrations in 

picograms per kilogram of 

water (standard deviations 

in parentheses). Well nest 

identifier is given above well

CONSOLIDATED 

BASIN FILL

WELDED TUFF?

024 

DATUM IS SEA LEVEL

10 12 14 

KILOMETERS

16 18 20 22 24 26 

VERTICAL SCALE GREATLY EXAGERATED

Figure 4. Hydrogeologic section showing chlorofluorocarbon-12 concentrations in Final Creek ground waters.

1991 (approximately 270 pg/kg). Therefore, the 

possibility exists that air in the Final Creek valley may 

have had higher than normal CFC-12 concentrations 

and may be the cause for the high CFC-12 concentrations observed in the ground waters.

CFC-11 concentrations show the same overall 

trends as the CFC-12 concentrations. CFC-11 concentrations generally decrease with depth and distance 

downgradient, with the exception of water from well 

702 (well site 700), which has a CFC-11 concentration 

of 565 pg/kg, typical of waters recharged in 1987-88 

(fig- 4).

Preliminary recharge dates calculated from 

CFC-11 data (and a few CFC-12 data) are shown in 

table 2 and figure 5. The dates of recharge were calculated using a recharge temperature of 12 °C and a 

recharge altitude of 1372 m. As additional dissolvedgas data become available and an understanding of the

ground-water-flow system improves, the recharge 

temperature and recharge elevations used in the agedating calculations will be refined. Excess-air concentrations of less than 20 mL/L do not significantly affect 

calculated CFC-based recharge dates (Busenberg and 

others, 1993) and, therefore, excess-air concentrations 

are not taken into account in our calculations. It is 

important to remember that, although individual 

recharge dates are assigned to specific water samples, 

the water samples themselves will generally represent 

a mixture of waters of different ages.

Table 2 shows the good agreement between the 

calculated CFC-11- and CFC-12-based recharge dates 

for waters from well 010, illustrating the accuracy that 

can be expected for CFC-based recharge dates in 

recent uncontaminated ground waters (well 010 was 

the only well sampled that had recent uncontaminated 

ground water). The results show that younger water

1050
was being pumped from the well as sampling time 

progressed (samples were collected approximately 

6 minutes apart). The CFC-12-based recharge dates 

are a little more recent than the CFC-11-based 

recharge dates, but the difference is probably not 

significant. Unlike the other wells sampled, well 010 is 

located south of the mining and refining operation 

areas, south of the towns of Globe and Miami. Precipitation that recharged ground waters sampled at well 

010 does not appear to have been exposed to the 

higher-than-normal CFC-12 atmospheric concentrations that may have prevailed at lower elevations on 

the other side of Miami.

Table 2. Chlorofluorocarbon recharge dates for recent 

uncontaminated ground waters in the Final Creek Basin 

from well 010

[Sampling conductedon November 20, 1991; pg/kg, picogram per kilogram of water; CFC-11, chlorofluorocarbon-11; CFC-12, chlorofluorocarbon-12]

Sample CFC-11 CFC-12 CFC-11 CFC-12 

number (pg/kg) (pg/kg) date date

1

2

3

4

5

202.6

315.8

397.5

372.9

370.6

88.6

146.4

168.9

176.3

174.2

1972.5

1976.5

1980.5

1979.0

1979.0

1972.0

1978.0

1981.0

1981.5

1981.5

Ground water from the deepest uncontaminated 

wells, 404 and 504, contains no significant CFC concentrations and is, therefore, thought to have been 

recharged before 1940. Ground water from the next 

deepest well sampled, well 105, has a recharge date of 

1962. Tritium analyses of ground waters sampled in 

June 1988 and May 1989 (Brown, 1990) agree with 

these results. Water from well 404 contained a tritium 

concentration of 0.5 pCi/L (close to the detection limit 

of 0.3 pCi/L) and water from well 504 did not contain 

detectable tritium. This would indicate a pre-1953 

recharge date for these waters. Water from well 105, 

however, has a tritium concentration of 47 pCi/L, the 

second highest value measured in Final Creek ground 

waters, indicating a post-1953 recharge date consistent 

with the CFC-based recharge date of 1962. (The highest measured tritium concentration was 86 pCi/L in 

ground water from well 301; unfortunately, this well 

was not sampled in the November 1991 field trip).

Water from wells 504 and 404 is also depleted in 

oxygen-18 by 2 to 3 per mil and in deuterium by 20 

per mil relative to the contaminated ground waters 

(Longsworth and Taylor, 1992). This depletion indicates that recharge of ground waters from wells 504 

and 404 occurred at higher elevations and probably at 

colder temperatures than the recharge of contaminated 

shallow ground waters. Waters from wells 504 and 

404 are presently thought to have been recharged in 

the mountains east of the valley, whereas shallow 

contaminated ground waters originate, at least in part, 

from Webster Lake (J.G. Brown, U.S. Geological 

Survey, written commun., 1993).

CFC-based recharge dates for the acidic ground 

waters at well sites 050, 100, 300,400, and 450 range 

from 1976 to 1988, suggesting traveltimes of 3 to 15 

years. The CFC-11-based recharge dates for the contaminated ground waters are consistent with the date 

of formation of Webster lake in 1940. Tritium concentrations that generally range from 22 pCi/L to 

86 pCi/L in all contaminated ground waters (except 

for ground water at well 502, see below) at the Final 

Creek study site (Brown, 1990) also suggest post-1953 

recharge dates.

Results from particle tracking simulations 

(Brown, 1991) show shorter traveltimes than those 

calculated from CFC-based recharge dates. According 

to the particle tracking simulations, the maximum 

estimated traveltime between Webster lake and Inspiration Dam is 15 years. The ground-water-flow model 

on which Brown's (1991) particle-tracking analysis 

was based is currently considered inadequate in its 

simulation of flow in the basin fill. Nevertheless, 

traveltime estimates derived from the ground-waterflow model for the alluvium are reasonable. An average traveltime between Webster lake and Inspiration 

Dam of 8 years (4.2 minimum, 13 years maximum) 

can be calculated from Darcy's law (J.G. Brown, U.S. 

Geological Survey, written commun., 1993). The average gradient is 0.008 m/m (calculated from water table 

levels). Horizontal hydraulic conductivities range 

from 150 to 260 m/d (determined from aquifer tests) 

and porosity ranges from 0.14 to 0.26 (from measured 

specific yields and gravity-change measurements). 

The distance between Webster lake and Inspiration 

Dam is 23 km. Using similar calculations, J.G. Brown 

(U.S. Geological Survey, written commun., 1993) 

estimates minimum and maximum traveltimes of 

0.9 years and 3 years in the alluvium between Webster 

lake and well site 50. In comparison, ground waters 

sampled from the shallowest well at sit 50, well 52, 

had an average CFC-based recharge date of 1986, for a 

CFC-based age of 5 ± 2 years.

1051
METERS A 

1,150

1,100

1,050 -

1,000 -

950 -

900 -

850 -

800

EXPLANATION

Well nest showing individual 

well sampling points with 

recharge dates determined 

fromCFC-11 measurements 

Ground water samples were 

collected in November 1991. 

Well nest identifier is given 

above well.

Plnal Creek water 

at site 700: 

CFC-11:1987 

CFC-12:1991.5-1987 CONSOLIDATED 

BASIN FILL <1945-54 

CFC-12: <1940-50

WELDED TUFF ?

024 

DATUM IS SEA LEVEL

10 12 14 

KILOMETERS

16 18 20 22 24 26

VERTICAL SCALE GREATLY EXAGERATED

Figure 5. Hydrogeologic section showing dates of recharge based on chlorofluorocarbon-11 (and some chlorofluorocarbon12) concentrations in Final Creek ground waters.

Recharge dates for neutralized contaminated 

ground waters at well site 500 range from 1969 to 

1972. Ground water from well 502 has a chemical 

composition indicative of a mixture of contaminated 

neutralized ground water and old uncontaminated 

ground water; the CFC-11-based recharge date is 

1964. The tritium concentration in a water sample 

from well 502 in January 1989, however, was less than 

5.7 pCi/L (Brown, 1990), indicating either an older 

recharge date, certainly pre-1964, or a disproportionately large amount of old water in the January 1989 

water sample.

The older recharge dates of the ground waters 

from well site 500 - compared to upgradient shallow 

ground waters ~ probably result from convergent flow

caused by vertical and lateral constriction of the 

alluvium and basin fill at the northern downgradient 

end of the Final Creek valley (figs. 1 and 3). The 

convergence may have increased the mixing of old 

waters with more recent waters during sampling. For 

the same reason, neutralized contaminated ground 

waters from wells 701 and 702, downgradient from 

site 500, may have older CFC-based recharge dates 

(1960 -1967) than those determined for ground waters 

from similar depths at site 500.

The CFC-11-based recharge dates of contaminated ground waters at the Final Creek site presented 

in this paper await confirmation from additional field 

studies and ground-water age-dating.

1052
Even if confirmed, CFC-based recharge dates 

should always be used with caution. First, groundwater samples commonly represent a mixture of 

waters of differing ages, in which case the CFC-based 

age should be considered more an average, rather than 

a specific residence time. Secondly, unlike most other 

solutes in the contaminated ground waters of the Final 

Creek basin, CFC-11 and CFC-12 are presumed to be 

inert tracers. Their concentrations in ground water are 

not thought to be modified by chemical reactions. As 

a result, a CFC-based age can, at best, be considered 

only a minimum age of reactive contaminant solutes, 

because the subsurface transport of reactive solutes 

will generally be retarded with respect to the flow of 

the water and the transport of its inert solutes (Glynn 

and others, 1991).

SUMMARY

Preliminary findings of our study are 

(1) Recharge temperatures calculated on the 

basis of dissolved N2 and Ar concentrations are on 

average 6 °C cooler than the average ground-water 

temperature of 18 °C. Recharge temperatures calculated for ground waters from the deeper wells (404, 

504 and 51, especially), screened in the basin fill, are 

lower than those calculated for ground waters from 

shallow wells screened in the alluvium.

(2) Excess-air concentrations calculated from 

dissolved N2 and Ar concentrations range from 1 

to 11 mL/L of water. Ground waters from the deepest 

wells sampled (wells 404, 105, and 504) had the highest excess-air concentrations.

(3) The high excess-air concentrations and low 

recharge temperatures calculated (compared to the 

mean annual air temperature) suggest that significant 

ground-water recharge occurs as a result of flooding in 

the winter and in the early spring. The low calculated 

recharge temperatures are supported by the 10 °C 

temperature of Final Creek measured during a large 

recharge event in February 1993, and also by available 

precipitation and air-temperature records for the area.

(4) N2 equilibrium partial pressures, calculated 

at field ground-water temperatures, always exceeded 

atmospheric partial pressures by as much as a factor of 

2 in the case of water from well 404. Excess air 

entrapped during recharge, rather than N2 production 

by denitrification is considered to be the cause.

(5) Dissolved O2 was present in all uncontaminated ground waters, including deep ground waters

from wells 404 and 504. This finding is consistent with 

the low concentration of organic carbon in the 

alluvium and basin-fill aquifers, and with the low 

dissolved organic carbon concentration in the ground 

waters.

(6) Dissolution of carbonate minerals by acidcontaminated ground waters is thought to be responsible for the high CO2 equilibrium partial pressures (as 

high as 0.15 atm) calculated for the acid-contaminated 

and neutralized contaminated ground waters.

(7) CFC-12 and CFC-11 concentrations in 

ground water decrease as a function of depth and 

distance downgradient at well sites 50, 100, 300, 400, 

450, 500 and 700. Most CFC-12 concentrations 

exceed 270 pg/kg of water, the CFC-12 concentration 

expected for equilibrium with late 1991 continental 

air. Historically higher-than-normal atmospheric 

CFC-12 concentrations in the Final Creek valley are 

thought to be responsible, as opposed to direct pointsource contamination of the ground waters.

(8) Recharge dates based on CFC-11 and 

CFC-12 concentrations in uncontaminated ground 

waters at well 010 are within 2 years of each other. 

Unlike the other wells sampled, well 010 is located 

south of the mining and refining-operation areas, south 

of the towns of Globe and Miami. Precipitation that 

recharged ground waters sampled at well 010 does not 

appear to have been exposed to higher-than-normal 

CFC-12 atmospheric concentrations.

(9) Recharge dates based on CFC-11 concentrations in contaminated waters from well sites 50, 100, 

300,400,450 and 500 are reasonable. Postulated 

ground water ages range from 3 to 15 years except for 

ground waters from site 500 which are 20 years old. 

The older age of ground waters from site 500 and the 

even older age of ground waters from site 700 (determined from CFC-12 concentrations) are thought to 

result from the upward flow of old ground waters from 

the basin fill.

(10) Deep uncontaminated ground waters from 

wells 404 and 504 did not contain any significant 

CFC-11 or CFC-12 concentrations. Recharge dates for 

these ground waters must, therefore, predate 1940. 

This finding is consistent with available tritium, deuterium, oxygen-18 and dissolved-gas data. The latter 

data suggest that ground waters at wells 404 and 504 

were recharged before 1953, at a higher elevation and 

at colder temperature than that of sampled contaminated ground waters. The contaminated ground waters

1053
are thought to have been recharged, at least in part, 

from Webster lake.

CFC-based ground-water age can only at best 

give minimum ages for most of the contaminant 

solutes in ground waters of the Final Creek basin, 

because the subsurface transport of those solutes will 

generally be retarded by chemical reactions relative to 

the movement of the water and to the transport of inert 

solutes, such as CFC-11 and CFC-12.

REFERENCES

Brown, J.G., 1990, Chemical, geologic and hydrologic data 

from the study of acidic contamination in the Miami 

Wash-Final Creek area, Arizona, water years 1988-89: 

U.S. Geological Survey Open-File Report 90-395, 

75 p.

Brown, J.G., 1991, Particle tracking analysis of flow in the 

stream-aquifer system in Final Creek basin, Arizona, 

in Mallard, G.E., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, Calif, March 10-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 

91-4034, p. 448-453.

Busenberg, E., and Plummer, L.N., 1992, The use of

chlorofluorocarbons (CC13F and CC12F2 ) as hydrologic 

tracer and age-dating tools ~ The alluvium and terrace 

system of central Oklahoma: Water Resources 

Research, v. 28, p. 2257-2283.

Busenberg, E., Weeks, E.P., Plummer, L.N., and Bartholomay, R.C., 1993, Age dating ground water by use of 

chlorofluorocarbons (CC13F and CC12F2), and distribution of chlorofluorocarbons in the unsaturated zone, 

Snake River Plain aquifer, Idaho National Engineering 

Laboratory, Idaho: U.S. Geological Survey WaterResources Investigation Report 93-4054, 47 p.

Eychaner, J.H., Rehmann, M.R., and Brown, J.G., 1989, 

Chemical, geologic and hydrologic data from the study 

of acidic contamination in the Miami Wash-Pinal 

Creek area, Arizona, water years 1984-87: U.S. 

Geological Survey Open-File Report 90-395, 105 p.

Glynn, P.O., and Busenberg, E., 1996, Unsaturated zone 

diffusion of carbon dioxide and oxygen in the Pinal 

Creek Basin, Arizona, in Morganwalp, D.W., and 

Aronson, D.A. eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, September 20-24, 1991: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Glynn, P.O., Engesgaard, P., and Kipp, K.L., 1991, Use and 

limitations of two computer codes for simulating 

geochemical mass transport at the Pinal Creek toxicwaste site, in Mallard, G.E., and Aronson, D.A., eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, Calif, March 10-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 

91-4034, p. 454-460.

Heaton, T.H.E., Talma, A.S., and Vogel, J.C., 1983, Origin 

and history of nitrate in confined groundwater in the 

western Kalahari, Journal of Hydrology, v. 62, 

p. 243-262.

Hobba, W.A., Jr., Chemerys, J.C., Fisher, D.W., and Pearson, F.J., Jr., 1977, Geochemical and hydrologic data 

for wells and springs in thermal-spring areas of the 

Appalachians: U.S. Geological Survey WaterResources Investigations Report 77-25, 36 p.

Longsworth, S.A., and Taylor, A.M., 1992, Hydrologic data 

from the study of acidic contamination in the Miami 

Wash - Pinal Creek area, Arizona, water years 1990- 

91: U.S. Geological Survey Open-File Report 92-468, 

59 p.

Pearson, F.J., Jr., Fisher, D.W., and Plummer, L.N., 1978, 

Correction of ground-water chemistry and carbon 

isotopic composition for effects of CO2 degassing: 

Geochimica et Cosmochimca Acta, v. 42, p. 1799- 

1807.

Plummer, L.N., Michel, R.L., Thurman, E.M., and Glynn 

P.O., 1993, Environmental Tracers forage-dating 

young ground water, in Alley, W.M., ed., Regional 

Ground-water Quality, Van Nostrand Reinhold, 634 p.

Reese, R.S., and Bassett, R.L., 1990, Characterization of 

organic contamination of ground water in a mining 

area, Globe, Arizona, in Lehr, J.H., ed., Proceedings, 

Ground-Water Geochemistry Conference, Kansas City, 

Missouri, February 20-22, 1990: National Water Well 

Association, Association of Ground Water Scientists 

and Engineers, Ground Water Management No. 1, 

p. 221-236.

Wallin, R.W., Basett, R.L., and Eychaner, J.H., 1991,

Ground-water transport of polycyclic aromatic hydrocarbons in association with humic substances in the 

Pinal Creek basin, Globe, Arizona, in Mallard, G.E., 

and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings 

of the technical meeting, Monterey, Calif, March 10- 

15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 511-515.

Wilhelm, E., Battino, R., and Wilcock, R.J., 1977, Lowpressure solubility of gases in liquid water: Chemical 

Reviews, v. 77, p. 219-262.

1054
Unsaturated Zone Diffusion of Carbon Dioxide and 

Oxygen in the Final Creek Basin, Arizona

SyPierre D. Glynn 1 and Eurybiades Busenberg1

Abstract

Unsaturated-zone gases were sampled in 

November 1991 at two sites in the Final Creek 

basin, above the acidic to partially neutralized 

ground-water-contaminant plume. The gases were 

analyzed for carbon dioxide (CO2), oxygen (O2), 

nitrogen and argon. The isotopic composition of 

the CO2 gas also was measured. Results show a 

linear CO2 and O2 gradient. An oxygen flux 

towards the water table of approximately l.lxlO"2 

(mol/m2)/d (moles per square meter per day) is 

estimated. This oxygen flux is probably caused by 

the reaction of dissolved oxygen with reduced 

manganese and iron in the ground water. A steadystate flux of approximately 9x10~3 (mol/m2)/d of 

carbon dioxide through the unsaturated zone can 

also be calculated. The CO2 flux is thought to 

emanate from the ground-water table. The reaction 

of acidic contaminated ground water with 

carbonate minerals produces ground waters with 

high equilibrium partial pressures of CO2 relative 

to the unsaturated zone atmosphere. The high 

carbon-13/carbon-12 ( 13C/12C) ratios of the 

unsaturated zone CO2 (as high as -10.75 per mil 

8 13C with respect to the Vienna Pee Dee Belemnite

1110 standard), and the increase in the C/ C ratio 

with increasing CO2 concentrations, suggests that 

degradation of organic matter and root respiration 

in the unsaturated zone are not contributing 

significantly to the CO2 flux. Field measurements 

show a difference of close to 4.4 per mil in 8 13C 

between the 13C enriched CO2 sampled just above 

the water table and CO2 sampled 0.46 meters

'U.S. Geological Survey, Reston, Va. 22092

below the ground surface. A theoretical model 

constructed assuming (1) steady-state diffusion of 

CO2 from the water table, (2) a constant 13C/12C 

flux ratio determined by the isotopic composition 

of the CO? exsolving from the water table and (3) a 

fixed 13C/ C ratio at the ground surface, predicts a 

difference of close to 4.4 per mil 13C fractionation 

between CO2 built up just above the water table 

and the CO2 coming out of the ground surface. 

This preliminary model does not, however, 

satisfactorily explain the field observations, and 

further research and field data are needed to fully 

understand our observations.

INTRODUCTION

This paper presents preliminary data gathered al 

the Final Creek toxic-waste site in late November 

1991. One of the multiple objectives of this data 

collection effort was to determine if carbon dioxide 

(CO2) and oxygen (O2) concentration gradients 

existed in the unsaturated-zone gas phase. If these 

gradients existed, we wanted to determine (1) if the 

unsaturated zone fluxes were at steady state, (2) the 

magnitude and direction of the fluxes, and (3) the 

processes responsible for these fluxes. In numerical 

modelling of multispecies reactive transport at the 

Final Creek site, Glynn and others (1991) had made 

the assumption that the contaminated phreatic aquifer 

was open to CO2 exsolution to the unsaturated zone 

but closed to O2 ingassing from the unsaturated zone. 

These constraints were forced by the inability of the 

geochemical codes used to consider reaction kinetics 

but were conceptually unsatisfactory. Indeed, massbalance models constructed by Glynn (1991) to 

simulate the chemical evolution of ground water 

between wells 51 and 402 invoked both CO2 

exsolution and O2 ingassing.

1055
ACKNOWLEDGMENTS

We are grateful to James Brown and Steve 

Longs worth for their invaluable assistance in 

collecting the field data presented in this paper and for 

providing, together with Ken Stollenwerk, 

unpublished ground water chemistry data. Ken 

deserves special commendation for having provided 

most of the ground-water chemical analyses reported 

for the Final Creek site, all of which are the highest 

technical quality. We also thank Rosalynd Williams, 

Jessica Hopple, Julian Wayland and Ray van Hoven 

for their laboratory analyses of our samples. Finally, 

the mathematical model presented in this paper could 

not have been constructed without the extensive help 

we received from Don Thorstenson and from Edwin 

Weeks. Weeks raised several criticims of our 13C 

diffusion model. Some of his criticisms, in particular 

his belief that biological activity could be responsible 

for the low carbon-13 contents that we measured near 

the ground surface, will be addressed in a later paper, 

when more field data are available.

INSTALLATION OF UNSATURATED-ZONE 

GAS-SAMPLING POINTS

Unsaturated-zone gas-sampling points were 

installed on November 18, 1991, near the U.S. 

Geological Survey (USGS) wells at sites 500 and 450 

(fig. 1). The sampling points at both sites were 

installed in the dry Final Creek stream bed. The sand 

and gravel alluvium did not contain any discernible 

soil zone. The site 450 location floods and new 

alluvium is redeposited at least once a year. Flooding 

and sedimentation at the site 500 location probably 

occurs about once every two years.The first set of gassampling points (506U) was installed above a 

neutralized zone in the ground-water contaminant 

plume (5.6 < pH < 5.9, high dissolved manganese but 

low iron concentrations). The other set of unsaturatedzone gas-sampling points (450U) was installed 

upgradient from 506U, close to the leading edge of the 

acidic ground-water contaminant plume (4.1 < pH < 

4.9, high dissolved manganese and iron 

concentrations).

Sampling-point set 506U was placed 1.68 m 

east of ground water well 506. Sets 450UA and 

450UB (referred together as 450U) were placed about 

1.5 m from each other at the foot of the 2-m-high dry 

creek bank, about 5.49 m west of wells 451 and 452,

which, at the time, were located on the edge of the 

creek bank. The unsaturated-zone sampling points and 

wells 451, 452, and 453 were washed away by flash 

floods in the winter of 1992-93. The water table was 

6.11m below ground surface at well 451 or 4.77 m 

below ground surface at sites 450UA and 450UB 

according to a level of the stream bed taken in April 

1992 (James Brown, oral commun.). At site 506U, the 

water table was encountered 2.26 m below ground 

surface.

Two-inch diameter holes were hand-augered to 

seven-ft (2.1-m) depth (holes 506U, 450UA) and to 3- 

ft (0.91-m) depth for hole 450UB. One-eighth of an 

inch o.d. copper tubes were placed in the holes (at 

depths of 2.13, 1.80, 1.31, 0.91, and 0.46 m for hole 

506U; 2.01, 1.49, 0.82 and 0.42 m for hole 450UA; 

0.88 m for hole 450UB). The bottom end of the copper 

tubes was bent into a loop to ensure that the tube end 

pointed down but did not constitute the lowermost 

resting point of the copper tubing. A layer of gravel 

was placed around each sampling point, followed by a 

layer of sand, a layer of silica flour, a layer of sand and 

the next sampling-point gravel layer. Each layer was 

about 0.08 m deep. The top of the hole was back filled 

with natural material. The purpose of the silica flour 

was to act as a less permeable barrier between the 

various sampling points, ensuring horizontal gas flow 

from the surrounding soil to each sampling point. 

After verifying that the tubes allowed the easy flow of 

soil gas, brass Swedgelock fittings and caps were 

placed on the top ends of the copper tubes, and the 

sampling-point installations were allowed to reequilibrate to background soil conditions.

Unsaturated-zone gas sampling at 506U was 

conducted in the late afternoon on November 22 and at 

450UA and 450UB in the morning of November 23 

(morning air temperature was at least 5 °C colder than 

that of the previous afternoon). A quick screening of 

CC>2 partial pressure was conducted for each sampling 

tube. After a 300-cm3 purge, 300-700 cm3 of soil gas 

(depending on CC>2 concentration) were passed 

through a Drager CC>2 indicator tube. These field 

screening results are not presented here because they 

were much less accurate than the laboratory gasanalysis results.

A stainless-steel pump was used to sample soil 

gas. After the initial screening, and after an initial 

three-minute purge of each unsaturated zone tube 

(flow rate: 500 cm /min), two 500 cm3 gas samples 

(for carbon 13 and major gas analyses) were collected 

in glass sampling tubes.

1056
110°55'

33°35' 

33°20'

Webster J ^Tailings 

Lake (dry) 

Mine

110°45'

25' 

EXPLANATION 

~^~ Direction of flow 

 ... Intermittent stream reach

Data sites, number is 

site identifier

 50 Well 

7^ Precipitation gage

3 MILES

0 1 2 3 KILOMETERS

Base from U.S. Geological Survey, 

1:24,000; Globe, 1945, Rockingstraw 

Mtn., 1986, Salt River Peak, 1988, 

and Inspiration, 1945

Figure 1. Area of study (Site 506U is next to well group 500, site 450U is next to well group 450).

1057
CARBON DIOXIDE FLUX

CO2 concentrations varied from 0.1 to 1.1 

volume percent (fig. 2A). Although these values would 

not be particularly large for a normal soil, the absence 

of any noticeable soil zone at the sites and the 

increasing content with depth suggest an external 

source of CO2 . A linear fit of the data from 506U gave a 

slope of 0.429 ± 0.026 (standard error) volume percent 

of CO2 per meter of depth with an intercept of 0.17 ± 

0.03 percent CO2 at ground surface, signif-icantly 

higher than the atmospheric value of 0.035 percent.

Carbon-13 results are shown in figs. 2B and 2C; 

S 13C is given in standard notation in per mil with 

respect to the Vienna Pee Dee Belemnite standard, and 

was calculated as follows:

R-Rstd

R

1000 ,

std

where R and Rstd are the 13C/ 12C concentration ratios 

in the sample and in the standard, respectively.

The 8 I3C values range from -10.75 to -15.15 per 

mil at sampling-point site 506U (for a difference of 4.4 

per mil) and from -16.8 per mil to -18.45 per mil at site 

450U. Linear regression of the data gives slopes of 

2.69 ±0.13 per mil per meter of depth at site 506U, 

and 0.966 ± 0.28 at site 450U. An extrapolated 8 I3C 

value of-14.1 per mil at the water table, approximately 4.78 m below ground, was calculated for sites 

450UA and 450UB, resulting in a 8 I3C difference of 

4.6 ± 1.3 per mil between the water table and the 

ground surface. The measured S C values appear to 

change linearly as a function of depth (fig. 2B) and 

CO2 (fig. 2C). However, steady-state diffusion of 

I2CO2 and 13CO2 would imply curved 8 I3C profiles as 

a function of depth and CO2 (dashed lines in figs. 2B 

and 2C). The curved lines drawn were calculated 

assuming a linear variation in 12CO2 and 13CO2 

concentration with depth between the deepest and 

shallowest sampling points at each site.

The measured 8 I3C values and their increase 

with depth, the linear increase in CO2 concen-trations, 

the absence of vegetation cover, and the absence of a 

visible soil zone, indicate that the origin of the 

unsaturated-zone CO2 is probably not of biogenic 

origin, particularly at sampling-point site 506U. 

Indeed, Parada and others (1983) report S 13C values of 

-18 to - 20 per mil for soil CO2 in the Tucson basin (at 

about 800 m elevation). They attribute the source of 

their soil CO2 primarily to the decomposition of

organic matter and root respiration (8 I3C = -26 per mil 

for plants using the 3 photosynthetic cycle (also 

known as the Calvin cycle) and S 13C «-13 per mil for 

plants with the CAM (Crassulacean Acid Metabolism) 

photosynthetic cycle). They believe that mixing with 

atmospheric CO2 (8 I3C = -8.5 per mil), and possibly 

evaporation of water (resulting in the precipitation of 

caliche (-3.5 per mil) and the release of CO2 (-20.5 per 

mil)), also are occurring. Parada and others (1983) 

report that vegetation in their study sites was 

dominated by C3 plants, with a maximum of 20 to 30 

percent CAM plants at one site. Soil CO2 at that 

particular site had correspondingly heavier 8 C 

values (-17 to -13 per mil).

In contrast with our study, Parada and others 

(1983) observed a decrease, instead of an increase, in 

813C with depth and with increasing CO2 

concentration. Parada and others attributed this effect 

to mixing with atmospheric CO2 and to the shallower 

roots of CAM plants relative to those of C3 plants. The 

higher 8 I3C values measured in our study, compared 

to those measured by Parada and others, are another 

indication that biogenic CO2 production (from 

respiration and/or organic matter degradation) is not a 

dominant process in our study sites.

The linearity of the CO2 concentrations with 

depth indicates that a process of steady-state 

diffusion of CO2 through the unsaturated zone may 

be operative at the unsaturated-zone sites, 

particularly at 506U where we were able to auger 

down to the water table. Cerling and others (1991) 

have described the carbon-isotope fractionation 

resulting from the diffusion of CO2 , produced 

throughout a soil zone. Cerling and others measured 

a difference of 4.4 per mil in carbon-13 content 

between soil CO2 (sampled at 50-cm depth in 

montane soils with a range of soil isotopic 

compositions) and respired CO2 (sampled coming 

out of the soil). Cerling and others attributed this 

fractionation to the greater diffusivity of I2CO2 

relative to that of I3CO2 , the diffusion coefficients 

of these two molecules being related by the 

following equation (Craig, 1953; lost, 1960):

12

«'> M 13 +A/"V

= 1.0044 , (1)

where D12 and D 13 are the diffusion coefficients of 

I2CO2 and I3CO2 , M12 and M13 are the molecular 

weights of 12CO2 and 13CO2, and Mair is the average 

molecular weight of air.

1058
0.0

0.5

CO

cc 

£! 1.0

ULJ

s

Z

£1.5

LU

Q

2.0

2.5

0.0 0.2 0.4 0.6 0.8 1.0 1.2

CARBON DIOXIDE, IN PERCENT BY VOLUME

-20

-18

GC 

LU 

Q_

- -16

LU

O 

O

n -I 

z

O

COGC

-10

0 0.2 0.4 0.6 0.8 1 1.2

CARBON DIOXIDE, IN PERCENT BY VOLUME

o.o

0.5

COGC

1.0

LU

LU 

Q

2.0 -

2.5

3 Air value: -8.5 per mil

-20 -18 -16 -14 -12 -10

CARBON-13 CONTENT, IN PER MIL

o.o

0.5

CO

cc

LU

tD

1.0

z

iLU 

Q

2.0

2.5

19.8 20.0 20.2 20.4 20.6

OXYGEN, IN PERCENT BY VOLUME

20.8

EXPLANATION:

SITE506U SITE450UA SITE450UB 

A 

Figure 2. (A) Volume fraction of carbon dioxide in the unsaturated-zone gas phase as a function of depth. Linear regression 

fits are drawn for sites 506U (long-dashed lines) and 450U (short-dashed lines) in graphs A and D. (B) Carbon isotope 

composition (813C values in per mil with respect to the Vienna Pee Dee belemnite standard) of unsaturated-zone C02 gas 

as a function of depth. The dashed lines in graphs B and C represent the theoretical steady-state lines between lowermost 

and uppermost points where both C02 and 813C measurements are available. Assuming constant diffusion coefficients, if 

steady-state diffusion was the only operative process causing CO2 flux between the lower and upermost sampling points, all 

813C values measured between these points should plot on the dashed lines in graphs B and C. (C) Carbon-isotope 

composition as a function of C02 volume fraction. (D) Volume fraction of oxygen in the unsaturated-zone gas phase as a 

function of depth.

1059
Under steady-state conditions, the fluxes of 

I2CO2 and 13CO2 through the unsaturated zone must 

be constant. If the CO2 concentration at the ground 

surface is significantly lower (by a factor of 20 or 

more) than the maximum concentration in the 

unsaturated zone near the water table, we predict that 

there should be a carbon-13 fractionation of close to 

4.4 per mil between the CO2 present just above the 

water table and the CO2 exiting at the ground surface. 

This fractionation should be independent of the depth 

of the unsaturated zone and of the CO2 concentration 

gradient. In addition, under steady-state conditions, 

the COo leaving the unsaturated zone must have the 

same I3C/I2C ratio as the CO2 entering the 

unsaturated zone from the water table, if no other 

production source or sink of CO2 is present. These 

statements are described in the following 

mathematical model.

At steady-state, the diffusive transport of I2CO2 and 

I3CO2 driven purely by a concentration gradient is 

expressed by the following relations (Pick's first law of 

diffusion):

J U = D'2

= D'3

u

(2)

(3)

where J 12 and J 13 are the fluxes of I2CC>7 and I3CO2 

through the unsaturated zone, D 12 and D are the 

diffusion coefficients for these two molecular species,

1 ^ 1 ^

C w,. and C wt are their concentrations in the gas 

phase just above the water table, C' 2gs and C' 3gs are 

their concentrations at the ground surface, and 1 is the 

depth of the unsaturated-zone. The capillary fringe is 

assumed to have negligeable thickness.

Furthermore, given that the concentrations of

t "7 1 ^

CO2 and CO2 are constant and very small in air, 

there will be a positive flux of both I2CO2 and I3CO2 

from the unsaturated zone into the atmosphere, and the 

two fluxes can be related to each other by the 

boundary condition

flux (4)

where Rflux is the 13C/I2C ratio of the CO2 entering 

the unsaturated zone gas-phase. Assuming the system 

is at steady-state, the fluxes of 13CO2 and I2CO2 leaving the system must equal the fluxes entering the system and, therefore, Rflux must also be equal to the

I3C/I2C ratio of the CO2 leaving the unsaturated zone. 

We assume that CO2 production at the water table is 

the only net source of CO2 in the unsaturated zone 

and that the atmosphere provides the only net CO2 

sink.

Combining equations 1, 2, 3 and 4 and defining 

Rwt as the ratio of the CO2 to I2CO2 concentrations 

just above the water table and Rgs as the ratio at the 

ground surface leads to the following relation:

R

c12

= 1.0044 + -g

'flux

(5)

If the I2CO2 concentration at the ground surface 

is zero, equation 5 shows that there will be a carbonisotope fractionation of 4.4 per mil between the CO2 

built up just above the water table and the CO2 exiting 

at the ground surface. If the I3C/I2C ratio at the 

ground surface (Rgs)is more than 1.0044 times the 

ratio of the CO2 leaving the unsaturated zone (Rflux ) 

- that is, if the 5 I3C value at the ground surface is 

greater than the 8 I3C value of the CO2 entering or 

leaving the unsaturated zone by more than 

approximately 4.4 per mil ~ the 8 I3C value of the CO2 

just above the water table should be more than 4.4 per 

mil heavier than the 513C of the CO2 leaving (or 

entering) the unsaturated zone:

R

> 1.0044

R.

> 1.0044 (6)

flux

or approximately in terms of 8 I3C values,

(7)

The opposite relations to equations 6 and 7 also 

apply. Because the CO2 concentration at the 

ground surface is presumed to be lower than the 

concentration at the water table, Rwt/RflU x wiU 

always be positive.

Our field measurements show a difference of 

close to 4.4 per mil in S 13 C between the l3C-enriched 

CO2 sampled just above the water table and CO2 

sampled 0.46 m below ground surface. Unfortunately, 

the theoretical model constructed above predicts a 

difference of 4.4 per mil between the CO2 built up just 

above the water table and the CO2 exiting at the 

ground surface. Also, the 8 I3C value of the CO2 

sampled at 0.46 m depth (-15.15 per mil at 506U, 

-18.45 per mil at 450U) is quite different from the

1060
atmospheric value of-8.5 per mil. There must be some 

layer near the ground surface in which the CO2 fluxing 

through the unsaturated zone mixes with atmospheric 

CO2 . Given the lack of data and understanding on the 

processes occurring at this interface, further research 

is indicated before our observations can be reconciled 

with a theoretical model.

The fluxes of CO2 and O2 can be estimated from 

our measurements at sites 506U and 450U. The 

concentrations of CO2 and O2 can be calculated from 

our measured gas volume percentages by the 

following equation (Thorstenson and others, 1983), 

assuming an ideal gas:

c, = i =

RT (8)

where Cj is the concentration of gas i in moles per 

cubic centimeter; nj is the number of moles of i; V is 

the volume of gas in cubic centimeters; Xj is the mole 

fraction of gas i (for ideal gases, Xj is also equal to the 

volume fraction of i); Ptot is the total atmospheric pressure in atmospheres (atm) and, for an elevation of 900 

m (at sites 450 and 500), Ptot is equal to 0.90 atmospheres, R is the gas constant (equal to 82.05 cm3 -atm 

per mole per degree Kelvin); and T is the temperature, 

in degrees Kelvin.

Diffusion coefficients for gases in porous media 

can be estimated from their values in air by various 

relations accounting for the air-porosity of the medium 

and it's tortuous nature (Thorstenson and others, 1983; 

Troeh and others, 1982). Our estimation uses the 

empirical equations developed by Reardon and 

Moddle (1985) from CO2 and O2 diffusion 

experiments on uranium mine tailings:

DC(> = 0.155[(e-0.05 )/0.95] 1.7

D(> - 3.98x10 V/2 [(s-0.05)/0.95]' 7

(9)

(10)

Equations 9 and 10 were tested by Reardon and 

Moddle (1985) for a range of air-filled porosities, e 

(from 0.06 to 0.44). From specific yields calculated for 

the Final Creek site (Don Poole, U.S. Geological 

Survey, written commun., 1993), we estimate an 

average air-filled porosity of about 0.18 ± 0.02. Using 

equations 9 and 10, the estimated diffusivities are 5.3 

10"3 and 7.0 10~3 cm2 s" 1 (centimeters squared per 

second), respectively, for CO2 and O2 .

At site 506U, the estimated CO2 gas fraction 

just above the water table at 2.26 m is 1.14 percent 

compared to an average atmospheric CO2 content of 

0.035 percent in 1991 (Rowland, 1991). Analysis of 

dissolved gas in water from well 503, which is 

screened from 24 and 23.5 m deep, gives an 

equilibrium pCO2 of 0.103 atm (Glynn and 

Busenberg, 1996) a value 9 times greater than the 

unsaturated zone pCO2 just above the water table. The 

estimated CO2 flux at 506U at 25 °C can be calculated 

from Pick's first law of diffusion (see eq. 2) and 

equation 8:

=5.610 3(1.143-0.035)0.90/100/82.05/298

225

= 1.010 (mol/cm )/s

=8.810 3 (mol/m2 )/d

The calculation assumes a constant air-filled 

porosity throughout the unsaturated zone and, 

therefore, a constant diffusivity. A constant 

temperature of 25 °C and no temperature-driven 

diffusive transport also are assumed. The flux values 

calculated in this paper are believed to be accurate 

within a factor of 2, primarily because of the 

uncertainties in the applicable diffusion coefficients.

At site 450U, the estimated CO2 gas fraction 

just above the water table at 4.77 m is 2.6 percent and 

the estimated flux is 1.1 10" 1 ' (mol/cm2)/s or 9 10~3

 }

(mol/m )/d. Given the uncertainty in the applicable 

porosity and the applicable diffusivity, the fluxes 

calculated for sites 506U and 450U are very close to 

each other. Analysis of dissolved gas in water from 

well 452, screened from 8.23 and 5.18 m deep, gives 

an equilibrium pCO2 of 0.0705 atm CO2 (Glynn and 

Busenberg, 1996) - a value 2.7 times the estimated 

unsaturated zone pCO2 just above the water table.

The ground-water pCO2 values calculated from 

dissolved-gas analyses are greater than those 

measured in the unsaturated zone and agree with our 

hypothesis that the ground water is the primary source 

of CO2 in the unsaturated zone. The relatively high 

pCO2 values in ground water are thought to result 

from the reaction of contaminated acidic ground water 

with carbonate minerals present in the alluvial aquifer

1061
and in the underlying consolidated basin fill. CO2 is 

probably not actively bubbling out from depth. Indeed, 

every meter in depth below the water table is 

equivalent to an overpressure of 0.0968 atmospheres. 

Data on dissolved N2, CO2, O2 and Ar (Glynn and 

Busenberg, 1996) in water from wells 503 and 452 

indicate that 1.2 to 2.1 m of water depth is sufficient to 

keep gases from exsolving. The calculation assumes 

that gas-bubble nucleation is not a limiting factor. 

Also, the actual movement of exsolved gas bubbles 

towards the water table would probably be restricted 

by surface adhesion onto aquifer materials. 

Evaporation of water could be contributing CO2 to the 

unsaturated zone a possibility mentioned by Parada 

and others (1983). The importance of this process, 

however, remains uncertain because the chemical and 

isotopic composition of the water at the water table 

and in the unsaturated zone are not known. 

Evaporation is generally not thought to be important 

below 2 m depth in the unsaturated zone (W. Wood, 

U.S. Geological Survey, written commun., 1993).

The average total carbon concentration of 

groundwaters at wells 501, 506 and 452 in November 

1991 was 6 10"3 mol of carbon per liter of water. The 

average equilibrium pCO2 calculated from the pH and 

composition of the ground waters was 0.12 atm ~ a 

value close to the dissolved gas concentration 

measured in water from well 503. If the upper 1 m of 

water at the water table is capable of contributing CO2 

to the unsaturated zone, and if the water-filled porosity 

is 0.25, we calculate, given the average ground-water 

linear velocity of 5 m/d (Eychaner, 1989), that 7.5 

mol of carbon per day will flow past a 1-m x 1-m x 1- 

m unit volume below the water table and 9 mmol of 

carbon per day (or 0.1 percent of the ground-water 

carbon through-flow) will be transferred from that unit 

volume into the unsaturated zone.

OXYGEN FLUX

The O2 volume fraction in unsaturated-zone 

gases decreased as a function of depth at sites 506U 

and 450U (fig. 2D), although the decrease was not as 

linear as the CO2 increase with depth. The extrapolated O2 gas fraction just above the water table at site 

506U was 19.97 percent. The Committee on Extension 

to the Standard Atmosphere (1976) cites an 

atmospheric value of 20.95 percent. Therefore the 

estimated O2 flux from the atmosphere into the water

at 506U is 1.1 10"" (mol/cm2)/s or 1.0 10'2 

(mol/m2 )/d.

The estimated O2 volume fraction just above the 

water table at 450U was 18.28 percent. Therefore, the 

estimated O2 flux from the atmosphere into the water 

at450Uis 1.4 10' u (mol/cm2 )/s or 1.3 10'2 

(mol/m2)/d. This oxygen flux is probably caused by 

the reaction of dissolved oxygen with reduced 

manganese and iron in the ground water. Indeed, rinds 

and grain coatings of iron oxyhydroxides and 

manganese oxyhydroxides are commonly found near 

the water table (Longsworth and Taylor, 1992; Brown, 

1990; Eychaner and others, 1989). Water from wells at 

site 450 typically had high dissolved manganese and 

iron concentrations in November 1991, whereas water 

from shallow wells at site 500 had high manganese but 

low iron concentrations and, therefore, a much smaller 

capability for oxygen reduction. Given that dissolved 

manganese also oxidizes much more slowly than iron, 

it was reasonable to expect a higher flux of O2 at site 

450U relative to that calculated for site 506U.

The calculated fluxes of O2 and CO2 are of 

similar magnitude with opposite directions. This is 

consistent with the assumption of a constant total 

pressure throughout the unsaturated zone. A gradient 

in total pressure would require consideration of 

advective fluxes in addition to the diffusive fluxes 

calculated in this paper. Temperature and barometric 

pressure at the time of sampling might be partly 

responsible for differences in the CO2 and O2 

concentration-depth profiles between sites 506U and 

450U, at least at shallow depths, but this factor needs 

further investigation.

SUMMARY AND CONCLUSIONS

CO2 and O2 concentration gradients and a 

gradient in the 6 C values of the CO2 were observed 

in the unsaturated-zone gas phase at two sites in the 

Final Creek Basin in late November 1991. One 

unsaturated-zone site (450U) was installed close to the 

leading edge of the acidic ground water contaminant 

plume (pH 4.1 - 4.9 and high dissolved manganese and 

iron concentrations). The other site (506U) was 

installed above a neutralized zone of the ground water 

contaminant plume (pH 5.6 - 5.9 and high dissolved 

manganese but low iron concentrations).

The linearity of the CO2 concentration-depth 

relation suggests a steady-state flux of CO2 out of the 

unsaturated zone. The increase in 6 C values with

1062
increasing CO2 concentrations, the lack of an 

observable soil zone, and the fact that the ground 

waters below the unsaturated zone sites have pCO2 

values several times higher than the unsaturated zone 

values, indicate that the ground water, rather than 

degradation of organic matter and(or) root respiration 

throughout the unsaturated zone, may be the primary 

source of the observed CO2. Sampling during a 

different season and at a different alluvial site above 

uncontaminated ground water will help verify this 

conclusion.

Our field measurements show a difference of 

about 4.4 per mil in 6 I3C between the 13C enriched 

CC>2 sampled just above the water table and CO2 

sampled 0.46 m below the ground surface. A 

theoretical model was constructed assuming (1) 

steady-state diffusion of CO2 from the water table, (2) 

a constant 13C/ 12C flux ratio determined by the 

isotopic composition of the CO 2 exsolving from the 

water table and (3) a fixed I3C/I2C ratio at the ground 

surface. The model predicts a difference of close to 

4.4 per mil I3C fractionation between the CO2 built up 

just above the water table and the CO2 exiting at the 

ground surface (with a 5 C value different from that 

of CO2 at the ground surface). The 5 I3C value of the 

CO2 sampled at 0.46 m depth (-15.15 per mil at 506U, 

-18.45 per mil at 450U) is quite different from the 

atmospheric value of-8.5 per mil. We conclude that 

there must be some boundary zone near the ground 

surface in which the CO2 fluxing through the 

unsaturated zone is mixed with atmospheric CO2 . 

Given the lack of data and understanding on the 

processes occurring close to the ground surface, 

further research is indicated before our observations 

can be reconciled with a theoretical model.

The estimated flux of CO2 leaving the ground 

water is about 9 10"3 (mol/m2)/d at both well sites. 

This flux is reasonably small compared to the 

horizontal flux of dissolved inorganic carbon at the 

water table.

Estimates of the flux of O2 entering the ground 

water range from 1.0 10"2 (mol/m2)/day at site 506U 

and 1.3 10'2 (mol/m2 )/day at site 450U. The greater 

estimated flux at site 450U may be significant given 

the comparatively high dissolved iron concentrations 

in ground water at site 450.

REFERENCES

Brown, J.G., 1990, Chemical, geologic and hydrologic data 

from the study of acidic contamination in the Miami 

Wash-Final Creek area, Arizona, water years 1988-89: 

U.S. Geological Survey Open-File Report 90-395, 75 p.

Ceding, T.E., Solomon, D.K., Quade, J., and Bowman, J.R., 

1991, On the isotopic composition of carbon in soil 

carbon dioxide: Geochimica et Cosmochimica Acta, 

v. 55, p. 3403-3405.

Committee on Extension to the Standard Atmosphere, 1976, 

U.S. Standard Atmosphere, 1976, Washington, D.C., 

Government Printing Office, 227 p.

Craig, H., 1953, The geochemistry of the stable carbon 

isotopes: Geochimica et Cosmochimica Acta, v. 3, 

p. 53-92.

Eychaner, J.H., 1989, Movement of inorganic contaminants 

in acidic water near Globe, Arizona, in Mallard, G.E., 

and Ragone, S.E. eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Phoenix, Arizona, September 26- 

30, 1988: U.S. Geological Survey Water-Resources 

Investigations Report 88-4220, p. 567-575.

Eychaner, J.H., Rehmann, M.R., and Brown, J.G., 1989, 

Chemical, geologic and hydrologic data from the study 

of acidic contamination in the Miami Wash-Pinal 

Creek area, Arizona, water years 1984-87: U.S. 

Geological Survey Open-File Report 90-395, 105 p.

Glynn, P.D., 1991, Effect of impurities in gypsum on 

contaminant transport at Final Creek, Arizona, in 

Mallard, G.E., and Aronson, D.A. eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, Calif, March 10-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 91- 

4034, p. 466-474.

Glynn, P.D., and Busenberg, E., 1996, Dissolved gas and 

chlorofluorocarbon content of ground waters in the Final 

Creek basin, Arizona, in Morganwalp, D.W., and 

Aronson, D.A. eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colorado, 

September 20-24, 1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Glynn, P.D., Engesgaard, Peter, and Kipp, K.L., 1991, Use 

and limitations of two computer codes for simulating 

geochemical mass transport at the Pinal Creek toxicwaste site, in Mallard, G.E., and Aronson, D.A. eds., 

U.S. Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, 

Monterey, Calif, March 10-15, 1991: U.S. Geological 

Survey Water-Resources Investigations Report 91- 

4034, p. 454-460.

Jost, W., 1960, Diffusion in solids, liquids and gases 

(3d ed.): Academic Press.

1063
Longsworth, S.A., and Taylor, A.M., 1992, Hydrologic data 

from the study of acidic contamination in the Miami 

Wash-Final Creek area, Arizona, water years 1990-91: 

U.S. Geological Survey Open-File Report 92-468, 59 p.

Parada, C.B., Long, A., and Davis, S.N., 1983, Stableisotopic composition of soil carbon dioxide in the 

Tucson Basin, Arizona, USA: Isotope Geoscience, 

v. 1, p. 219-236.

Reardon, E.J. and Moddle, P.M., 1985, Gas diffusion 

coefficient measurements on uranium mill tailings: 

implications to cover layer design: Uranium, v. 2, 

p. 111-131.

Rowland, F.S., 1991, Stratospheric ozone in the 21 st 

century The chlorofluorocarbon problem: 

Evironmental Science and Technology, v.25, no. 4, 

p. 622-628.

Thorstenson, D.C., Weeks, E.P., Haas, H., and Fisher, D.W., 

1983, Distribution of gaseous I2CO2, I3CO2, and 

I4CO2 in the sub-soil unsaturated zone of the western 

US Great Plains: Radiocarbon, v. 25, no. 2, p. 315-346.

Troeh, F.R., Jabro, J.D., and Kirkham, D., 1982, Gaseous 

diffusion equations for porous materials: Geoderma, 

v. 27, p. 239-253.

1064
Interactions between Shallow Ground Water

and Surface Water that Affect Metal

Transport in Final Creek, Arizona

By Judson W. Harvey1 , Christopher C. Fuller1 , and Brian J. Wagner1

Abstract

Solute transport was investigated in shallow 

ground water and surface water along a 500-meter 

(m) reach of the perennial stream in Final Creek 

basin. A bromide tracer injection in the stream was 

conducted to quantify ground-water inflow and 

streamwater/streambed-water exchange. Respective gains and losses in iron (Fe) and manganese 

(Mn) in solution occurred in ground water 

discharging laterally through bank seeps into the 

stream, and not in ground water discharging vertically from below the active channel, or in the 

active surface channel itself. Aqueous metal 

concentrations measured in shallow ground water 

varied significantly with position across the 

narrow channel at Final Creek. Concentrations of 

Fe were twentyfold to fiftyfold higher in shallow 

ground water discharging laterally through channel banks compared to ground water discharging 

vertically from beneath the stream. Mn was 25 

percent lower in bank seepage compared to 

subchannel ground water. Cross-channel variability in metal concentrations in contaminated 

ground water could not be explained by differential dilution of ground water discharging laterally 

to the channel versus vertically through the streambed. Mass-balance calculations for the 500-m 

reach of stream and shallow aquifer verified that 

Fe and Mn dynamics in bank seepage were sufficient to affect downstream concentrations of those 

constituents. Microbial processes in channel banks 

are hypothesized to drive reductive dissolution of 

solid phase Fe in ground water discharging laterally through banks, but not in subchannel ground 

water. The mechanism of Mn loss in this channel

reach appeared to be sorption of Mn2+ on Feoxyhydroxides, which form where ground water 

emerges from bank seeps.

INTRODUCTION

Shallow ground water that surrounds the active 

channel of streams is referred to in this paper as the 

near-stream zone. Basin ground water discharges 

through the near-stream zone into the channel of gaining (effluent) streams (fig. la). In addition to an influx 

of ground water from the basin, the near-stream zone 

receives localized recharge fluxes of streamwater 

(fig. Ib). Mixing of basin ground water and streamwater in the near-stream zone and resulting chemical 

reactions are hypothesized to be an important control 

on solute chemistry of surface water in streams (Triska 

and others, 1993). Redox chemistry is commonly an 

important factor because of the tendency for a reducing

A

"U.S. Geological Survey, Menlo Park, Calif.

Figure 1. Solute transport pathways in the near-stream 

zone of a gaining stream: (A) discharge of basin groundwater solutes through near-stream zone into the active 

stream channel, and (B) localized recharge of streamwater 

solutes into streamwater-filled flow paths in the nearstream zone.

1065
zone to develop in the near-stream zone. Resulting 

chemical reactions in the near-stream zone could affect 

nutrient and contaminant inflows to streams and, thus, 

may affect downstream surface-water quality (Peterjohn and Correll, 1986). An ability to quantify and 

model near-stream hydrological and chemical processes is critical to improving an understanding of the 

timing and attenuation of nonpoint-source contaminant 

inflows through shallow aquifers into streams.

In this paper, the chemical reactions in a coupled 

stream/near-stream system in a basin where ground 

water is contaminated by acidic mine drainage are 

described. The objective of the research was to determine the importance of stream-shallow subsurfacewater interactions to metal transport in ground water 

and surface water and to downstream water quality in 

Final Creek basin. Contaminated ground water that 

presently discharges to Final Creek is a calcium 

sulfate-type water that has high Mn concentrations, 

somewhat lower concentrations of other metals (Ni, 

Co, Cu, Fe among others), and higher pH than contaminated water in the acidic core of the plume (Brown and 

Harvey, 1996). A preliminary experiment was conducted at Final Creek in November 1992 to investigate 

whether a mass-balance approach could quantify attenuation of Mn, Fe and other trace metals along a transport pathway from shallow ground water into surface 

water. Preliminary results are reported; the approach 

will ultimately be used to quantify a net "effective" rate 

of biologically and chemically driven reactions in the 

system. The mass-balance approach compliments other 

research at Final Creek aimed at identifying specific 

chemical reactions at Final Creek (Lind and Hem, 

1993; Fuller, 1996).

STUDY SITE AND METHODS

During low-flow periods without rainfall, streamflow in Final Creek basin begins 6 kilometers (km) 

above Inspiration Dam (fig. 2). The active surface 

channel meanders in an unbraided channel over a 

streambed composed of coarse sand and fine gravel; 

the average streambed slope is approximately 1 percent. We investigated solute transport during November of 1992, a month without significant rainfall at 

Final Creek basin. On November 21, a bromidelabeled-tracer solution was injected into the surface 

channel near the head of perennial surface flow at a 

constant rate for 30 hours (fig. 2). Bromide concentrations at downstream monitoring points were used to 

compute streamflow discharge and net ground-water

influx to the channel by the tracer-dilution method (Kilpatrick and Cobb, 1985). Concurrent with the tracer 

experiment, we collected and filtered water samples 

from subchannel ground water (to a depth of 1 m), surface water in the active channel, and in surface outflow 

from bank seeps and small tributaries (fig. 2). Methods 

for sample collection and analysis of metal concentrations are reported in Harvey and Fuller (1996). In addition, streamflow data and monitoring of bromide-tracer 

concentrations were used as the basis for a transient 

simulation of bromide transport to estimate stream solute-transport characteristics. A nonlinear least-squares 

optimization approach was used to determine stream 

and storage zone cross-sectional areas, a dispersion 

coefficient, and a mass-transfer coefficient between the 

stream and "storage zones" (Wagner and Harvey, 

1996).

BROMIDE TRACER INJECTION TO 

DETERMINE GROUND-WATER INFLUX 

AND STREAMWATER/STREAMBEDWATER EXCHANGE

Ground-water influx and streamwater/streambedwater exchange fluxes were estimated from the 

bromide-tracer-dilution measurements. Lateral inflow 

of ground water in the first 500 m of streamflow 

accounted for 35 percent of all ground-water inflow to 

the entire 6-km stream reach upstream of Inspiration 

Dam; this computation is based on comparison of computed ground-water inflows with streamflow measured 

at Inspiration Dam.

Bromide tracer concentration histories in the 

stream were simulated with a one-dimensional stream 

transport model with lateral ground-water inflow and 

solute exchange with storage zones (Wagner and Harvey, 1996). Characteristics of the modeled solute storage zones determined for Final Creek differed from 

values previously determined for higher-gradient, 

gravelbed streams. For example, the ratio of the crosssectional area of the storage zone to that of the active 

channel at Final Creek (between sites 1 and 2) is about 

0.2, compared to ratios which are typically 10 times 

higher in gravelbed pool and riffle streams with higher 

gradients (Broshears and others, 1993). This result suggests that the zone of transient storage of streamwater 

solutes in the subsurface (fig. Ib) might be smaller in 

low gradient sandbed streams compared to higher gradient gravelbed streams.

1066
110°55' 54' 53' 110°52'

33°34'

33'

33°32'

y

Inspiration 

Dam

 Q

EXPLANATION

 ... ... INTERMITTENT REACH

 PERENNIAL REACH

DATA SITES Number is 

site identifier

500 Observation well group

(09498400) A Streamflow-gaging station

9 A Sample site

0 Channelized inflow

+ Bank seep

1/2 1 MILE

See inset i 

500^"

0.5 1 KILOMETER

I______I______

\

Figure 2. Locations of ground-water and surface-water sites sampled during the solute-transport study, November 1992.

The simulation cannot determine where storage is 

occurring-that is, along channel margins where recirculating zones are present, or in the subsurface. However, 

the presence of bromide-labeled water at shallow depths 

(up to 6 centimeters (cm)) in the streambed indicates that 

streamwater is in contact with more sediment than the 

amount exposed on the streambed surface. Sampling in 

the subsurface supports the hypothesis that transient 

storage of streamwater solutes might be explained by 

convective flow of streamwater through the streambed.

Results of the preliminary tracer study indicated 

that the flux of streamwater through the streambed and 

contact time of that water with sediment surfaces is 

significant, and potentially could enhance Mn oxidation. Harvey and Fuller (1996) concluded that most of 

the Mn oxidation that occurs in Final Creek probably is 

catalyzed on streambed surfaces; thus streamwater/ 

streambed sediment contact times and contact area are 

likely to be critical controls on Mn loss from solution. 

At this time we have established only the potential for

1067
a stream-subsurface interaction that could enhance Mn 

oxidation. Data and observations from November 1992 

(Fuller, 1996) suggest that most Mn oxidation occurs in 

the stream well downstream of the beginning of perennial streamflow where our experiment was conducted. 

An experiment downstream from site 3 is currently 

being planned to determine the importance of streamsubsurface water exchange on Mn oxidation at Final 

Creek.

CHEMICAL PROCESSES IN GROUNDWATER INFLOW DURING PASSAGE 

THROUGH THE NEAR-STREAM ZONE

The mass balance for a solute in a coupled streamshallow aquifer system is, assuming steady-state 

conditions,

(CU QJ + (C^QgJ - (CdQdj + *' = 0 (1)

where i indicates the reach of interest, C is solute concentration (mass/volume), Q is water flow (volume/ 

time), subscripts u and d indicate upstream and downstream ends of reach /, respectively, and R is a solute 

reaction rate (mass/time). The subscript gw refers to 

reach-averaged solute concentrations and water fluxes 

in basin ground water that discharges into reach i.

Dilution of contaminated ground-water inflow to 

the stream with uncontaminated ground water was not 

expected to be important during base flow recession in 

this system; however, dilution effects were checked by 

modeling Na transport. Na in contaminated ground 

water of the basin fill aquifer is highly concentrated 

compared to uncontaminated water; the contaminated 

to uncontaminated ratio of Na concentration in ground 

water is 4:1 (Glynn, 1991). Na concentrations in 

inflowing ground water for each reach were computed 

from equation 1 by using measured water fluxes and 

concentrations of Na in streamwater. Assuming that 

transport of Na was conservative, mass-balance calculations demonstrated that contaminated ground water 

enters the stream without significant dilution by inflow 

of uncontaminated water. For example, the calculated 

Na concentrations for ground-water inflow in each 

subreach is a good approximation to measured concentrations in subchannel ground water and in bank seepage in that reach (fig. 3). The overall variation in 

measured Na concentrations at all sites is less than 5 

percent, including a comparison to Na concentrations 

measured in deep piezometers (well observation group 

500, fig. 2) that penetrate the aquifer to 40 m.

Fe concentrations in subchannel ground water, and 

in deeper piezometers (well observation group 500, 

fig. 2) were approximately equal to the Fe detection 

limit in this system. Fe concentrations in the stream 

were computed from equation 1 assuming an Fe concentration in inflowing ground water that was equal to 

the detection limit Dashed lines in figure 4a connect 

instream concentrations that were calculated from 

equation 1, assuming conservative transport. Measured 

concentrations of Fe in streamflow at sites 1 and 3 were 

larger than calculated concentrations. A positive difference between measured and calculated concentrations 

in the stream represents a gain of that metal in solution 

along the flow path. Figure 4a, therefore, indicates that 

a gain in aqueous Fe occurred somewhere along the 

flow path beginning in ground water.

The mass-balance approach cannot specify the 

exact location of the gain in Fe or the reaction that produces it. Aqueous Fe is hypothetically thought to be 

produced by reaction in shallow ground water as it discharges laterally through channel banks and into the 

stream. Detailed measurements along the 500-m reach 

indicated that Fe concentrations in bank seeps were 

twentyfold to fiftyfold higher than in subchannel 

ground water (fig. 4a). Furthermore, the Na balance 

indicated that higher measured concentrations of Fe in 

bank seeps cannot be explained by mixing of ground 

water from contaminated and uncontaminated sources; 

therefore, ground water passing through bank seeps has 

virtually the same origin as ground water discharging 

from directly beneath the active channel. Riparian vegetation is present on channel banks, and organic matter 

content of sediment appears higher. Microbial respiration in channel bank sediments appears to be lowering 

the Eh sufficiently to cause reductive dissolution of Fe 

in channel bank sediments during passage of shallow 

ground water through the channel banks.

The Mn concentration in ground water emerging 

from bank seeps was 25 percent less than in subchannel 

ground water. The mass balance for Mn assumed conservative transport; concentrations of Mn in ground 

water for each reach were computed from a best-fit 

relation between Mn and Na in subchannel ground 

water (r2 = 0.724, n = 24). Mn concentrations predicted by mass balance matched measured concentrations of Mn in the stream quite well except at site 2, 

where the measured Mn concentration was larger than 

the concentration predicted by the mass balance (fig. 

4b). The mass balance indicates loss of Mn from solution in shallow ground water or in streamflow in the

1068
0.0039

pc

yj 0.003S

Ul

o

I

o

0.0037

0.0036

0.0035

0.0034

Q 

O 

co 0.0033

0 1 Sites

-eHA

Water Ty

Streamwat

Subchann

Bank seep

Tributary In

Ground wa

0 100 200 300 400 

DOWNSTREAM DISTANCE FROM STREAM ORIGIN, IN 

Figure 3. Sodium concentrations in shallow ground water and surface water as a function of downstream distance from 

Final Creek.

o

O)

to
I-J

10-5

LU

I J

o

*

1

10-7

0 1

-f

Detection 

Limit Water Type

Streamirater, measured

----- Streamwater, calculated

 Subchannel ground water 

+ Bank seepage

0 100 200 300 400 500 

DOWNSTREAM DISTANCE FROM STREAM ORIGIN, IN METERS

0

CO 

Ul I

O

o

LU 

CO 

LU

0.0020

0.0018

0.0016

0.0014

0.0012

0.0010

0.0008

0 1 A

Sites

Wafer Type

+

A

Streamwater, measured 

Streamwater. calculated 

Subchannel ground water 

Bank seepage 

Tributary Inflow

0 100 200 300 400 500 

DOWNSTREAM DISTANCE FROM STREAM ORIGIN, IN METERS

Figure 4. Variation in iron and manganese concentrations in streamflow, subchannel ground water, ground 

water discharging from bank seeps, and small tributaries at Final Creek, (A) iron, (B) manganese.

1070
reach between sites 1 and 2. Lower Mn concentrations 

in bank seepage compared to subchannel ground water 

or the stream suggests that Mn loss occurs in shallow 

ground water as the water passes through bank seeps. 

This supposition is supported by observations that oxidation of Mn occurs slowly, if at all, in streamflow in 

the first 500 m of perennial streamflow in Final Creek 

(Eychaner, 1991; Harvey and Fuller, 1994, this report).

Mn transfer to the solid phase appears, therefore, to 

be occurring by precipitation in shallow ground water 

at some point during the passage through bank seeps. 

Some of the filterable Fe that is gained by dissolution 

in channel banks oxidizes and forms Fe floes at the 

point of discharge of shallow ground water from bank 

seeps. Fe oxides may be sorbing Mn, and (or) acting as 

a catalyst for oxidation of Mn (Davies and Morgan, 

1989), which could account for the low Mn concentrations in bank seepage compared to Mn concentrations 

in subchannel ground water, and for the loss of Mn 

from solution in the stream reach between sites 1 and 2.

SUMMARY

The solute transport in 500-m reach of stream and 

shallow aquifer immediately downstream of the beginning of perennial streamflow was studied. A 30-hour 

instream bromide-tracer experiment demonstrated that 

streamwater/streambed-water exchange increased the 

area of contact and the contact time of stream water and 

stream solutes with bed sediment. Mn oxidation potentially is enhanced by streamwater/streambed-water 

interactions at sites downstream of the study reach in 

Final Creek. In the study reach, Fe dissolution and Mn 

precipitation occur in shallow ground water as it passes 

through channel-bank sediments prior to discharge into 

the stream. The Fe concentration was twentyfold to fiftyfold higher and Mn was 25 percent lower in bank 

seepage compared to concentrations discharging from 

deeper within the aquifer and in ground water discharging directly beneath the streambed. Comparatively 

higher organic matter content and bacterial metabolism 

in bank sediments appears to drive reductive dissolution of Fe. Mn probably sorbs or coprecipitates with Fe 

oxides that precipitate as the water seeps from the 

channel sides. These reactions do not occur directly 

beneath the channel. Data on the chemistry of subchannel ground water, streamwater, and bank seepage will 

be used, along with the mass-balance model, to constrain the magnitude of reactive chemical processes in 

perennial flow.

REFERENCES

Broshears, R.E., Bencala, K.E., Kimball, B.A., and 

McKnight, D.M., 1993, Tracer-dilution experiments 

and solute transport simulations for a mountain stream, 

Saint Kevin Gulch, Colorado: U.S. Geological Survey 

Water-Resources Investigations Report 92-4081, 18 p.

Brown, J.G., and Harvey, J.W., 1996, Hydrologic and 

geochemical factors affecting metal contaminant 

transport in Final Creek Basin near Globe, Arizona, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. Geological Survey Toxic Substances Hydrology Program-- 

Proceedings of the technical meeting, Colorado 

Springs, Colorado, September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

94-4015.

Davies, S.H.R., and Morgan, J.J., 1989, Manganese (II) 

oxidation kinetics on metal oxide surfaces: Journal of 

Colloid and Interface Science, v. 129, p. 63-77.

Eagleson, PS., 1970, Dynamic Hydrology: New York, 

McGraw-Hill, 462 p.

Eychaner, J.H., 1991, Solute transport in perennial streamflow at Final Creek, Arizona, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program-Proceedings of the 

Technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, p. 481-485.

Fuller, C.C., 1996, Experimental studies of trace metal 

partitioning in perennial reaches of Final Creek, 

Arizona, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Glynn, P.O., 1991, Effect of impurities in gypsum on contaminant transport at Final Creek, Arizona, in Mallard, 

G.E., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the Technical meeting, Monterey, California, March 

11-15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 466-474.

Harvey, J.W., and Bencala, K.E., 1993, The effect of 

streambed topography on surface-subsurface water 

exchange in mountain catchments: Water Resources 

Research, v. 29, p. 89-98.

Harvey, J.W., and Fuller, C.C., 1996, Association of 

selected metals with colloidal and suspended 

particulate material in shallow ground water and 

surface water at Final Creek, Arizona, in Morganwalp, 

D.W., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program-Proceedings of 

the technical meeting, Colorado Springs, Colorado,

1071
September 20-24,1993: U.S. Geological Survey WaterResources Investigations Report 94-4015.

Kilpatrick, F.A., and Cobb, E.D., 1985, Measurement of 

discharge using tracers: U.S. Geological Survey Techniques of Water Resources Investigations, book 3, 

chap. A16, 46 p.

Lind, C.J., and Hem, J.D., 1993, Manganese minerals and 

associated fine particulates in streambed of Final 

Creek, Arizona A mining-related acid drainage problem: Applied Geochemistry, v. 8, p. 67-80.

Peterjohn, W.T., and Correll, D.L., 1986, Nutrient dynamics 

in an agricultural watershed Observations on the role 

of a riparian forest: Ecology, v. 65, p. 1466-1475.

Triska, F.J., Duff, J.H., and Avanzino, R.J., 1993, The role 

of water exchange between a stream channel and its 

hyporheic zone in nitrogen cycling at the terrestrial 

aquatic interface: Hydrobiologia, v. 251, p. 167-184.

Wagner, B.J., and Harvey, J.W, 1996, Solute parameter 

estimation for an injection experiment at Final Creek, 

Arizona, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances Hydrology Program Proceedings of the technical meeting, 

Colorado Springs, Colorado, September 20-24, 1993: 

U.S. Geological Survey Water-Resources Investigations Report 94-4015.

1072
Association of Selected Metals with Colloidal and 

Suspended Paniculate Material in Shallow Ground 

Water and Surface Water at Final Creek, Arizona

ByJudson W. Harvey1 and Christopher C. Fuller1

Abstract

Trace metals are potentially transportable 

as solid phases in ground water and surface water 

in association with colloids or suspended 

particulates. Field data that specifies the 

partitioning of metals between dissolved and 

suspended solid phases is necessary for 

developing physically based models of metal 

transport. The purpose of the present study was to 

determine colloidal and suspended-particulate 

concentrations for selected metals in shallow 

ground water and in streamflow of the perennial 

reach of Final Creek during base-flow conditions. 

Concentrations of manganese (Mn), iron (Fe), 

nickel (Ni), copper (Cu), and cobalt (Co) were 

measured in unfiltered water samples, and 

compared with 0.45 pm (micrometer) and 0.001 

m filtrates of the same water samples. Total metal 

concentrations in unfiltered water were 

partitioned into suspended particulate (>0.45 ^im), 

colloidal (0.001 - 0.45 ^im), and dissolved 

(O.001 urn) components. Median colloidal and 

suspended-particulate concentrations for Mn, Ni, 

Cu and Co were several percent or less of total 

concentrations of those metals. In contrast, the 

sum of Fe colloidal and suspended- particulate 

phases was a large percentage of total Fe 

concentrations in some cases, which indicated 

that a colloidal or particulate Fe phase was 

significant at Final Creek. At present total Fe 

concentrations in suspension are low in shallow

'U.S. Geological Survey, Menlo Park, Calif.

ground water and in streamflow at Final Creek 

during base-flow conditions. Colloids that do 

form do not appear to be significant sorbents of 

other trace metals, compared to Mn and Fe solid 

phases on the streambed. Colloid- facilitated 

metal transport will likely increase in importance 

if acidic ground water (which has Fe greater than 

50 milligrams per liter) reaches Final Creek.

INTRODUCTION

Given favorable chemical conditions, colloids 

or small-diameter particulates can move long 

distances in suspension in ground water and surface 

water (see review by McCarthy and Zachara, 1989). 

Metals that have been partitioned to the solid phase are 

still transportable under such conditions if adsorbed to 

colloidal matter (Ryan and Gscwhend, 1990; Penrose 

and others, 1990). Colloids are commonly formed by 

precipitation of Fe hydrous-metal oxides in 

contaminated water, including sewage-contaminated 

ground water (Langmuir, 1969) and streams 

contaminated by acidic mine drainage (Kimball and 

others, 1992). Fe colloids are also observed in natural 

ground waters (Ryan and Gscwhend, 1990). High 

concentrations of Mn are associated with acidic mine 

drainage (Wentz, 1974), and Mn colloids are 

occasionally observed (Hart and others, 1992). 

However, Mn colloids can be less common than Fe 

colloids in systems affected by acidic mine drainage, 

because of the kinetic limitation on oxidation of Mn2"1" 

(Hem, 1981; Davies and Morgan, 1989). Because of 

the strong ability of Fe and Mn oxides to scavenge 

other dissolved metals at Final Creek (Lind and 

Anderson, 1992), transport of many metals depends on

1073
whether He and Mn oxides form as colloids or coatings 

on streambed sediment (Kimball and others, 1992).

The history of contamination at Miami Wash 

and lower Final Creek basin, Arizona is summarized 

in Brown and Harvey (1996). Briefly, acidic mine 

waste water infiltrated from a holding pond into 

alluvium of Miami Wash and has moved 15 km 

downgradient in the subsurface into the Final Creek 

basin. Initially during transport of acidic mine waste 

water, the low pH favored dissolution and transport of 

trace metals in the dissolved phase. Acidity was 

neutralized by carbonate dissolution and mixing with 

uncontaminated water, which retarded the movement 

of the steep pH front in the contaminated groundwater plume (Eychaner, 199la). Iron oxyhydroxides 

precipitated near the pH front, and functioned as the 

principal sorbent for Co, Cu, and Ni (Stollenwerk, 

1991), which retarded movement of those metals out 

of the basin. The pH front is presently about 1 km 

(kilometer) upgradient from the head of perennial 

streamflow in lower Final Creek basin.

Colloid concentrations in ground water are 

significant near the pH front of the contaminant 

plume; oxidation of Fe2+ near the pH front apparently 

resulted in formation of Fe colloids (Puls and 

Eychaner, 1990). Colloid concentrations were much 

lower at a well site 2 km downgradient from the pH 

front (Puls and Eychaner, 1990); therefore colloids do 

not appear to be transported a great distance in ground 

water at the site. Eow colloid concentrations could 

have resulted from the effects of high ionic strength of 

contaminated water on colloid mobility. According to 

theory, high concentrations of Ca2+ and other cations 

increase nonspecific sorption, which destabilizes 

colloids (Stumm and Morgan, 1981).

Rapid adjustment of pH and pe occurs as 

shallow ground water is discharged to surface water in 

Final Creek, because of CO2 outgassing and by O2 

reareation (Eychaner, 1991b). Ficklin and others 

(1991) provided evidence that large quantities of Fe 

and Mn oxides were forming near the water table at 

"observation well group 500" in fig. 1, near the head of 

perennial streamflow. The formation and transport of 

Fe and Mn colloids at the shallow ground waterstreamflow interface was not previously investigated 

at the site. Understanding the potential importance of 

colloids to metal transport in shallow ground water 

and in streamflow is basic to the research effort in 

Final Creek basin.

METHODS

Final Creek streamflow is perennial in a 6-km 

reach where ground water is forced to the surface as 

the alluvial deposit narrows and shallows towards its 

terminus at Inspiration Dam (fig. 1). Selected metals 

were sampled in Final Creek during November 19-22, 

1992 - a period without rainfall and not preceded by 

substantial rainfall for several months.

Shallow multilevel samplers were installed 

vertically in the streambed at three locations in the 

central channel 1 month before metal sampling began. 

Eocations for multilevel samplers were sites 0, 1, 3 

and 5 (fig. 1). Sampler design followed the multilevel 

sampler design of EeBlanc and others (1991). 

Sampler ports were located at 100, 50, 25, 12, 6, and 3 

cm below the sediment surface. Ground-water 

samples were pumped through a peristaltic pump at a 

rate of approximately 200 mE/min. The pump and 

filtration system was closed from the atmosphere and 

protected from light to reduce effects of aeration and 

photoreduction on metal concentrations (Puls and 

Eychaner, 1990; Ryan and Gswhend, 1990; McKnight 

and others, 1988).

110' 54'

X

'

EXPLANATION

 ... ... INTERMITTENT REACH

 PERENNIAL REACH

DATA SITES-Number is 

site io'emilitr

500 Observation well group

(09<9S400)A. Streemflow-fejinj station

SA Sample site

0 Channelled inflow

 Bank seep

1/2 1 MILE

0.5 1 KILOMETER

Figure 1. Locations of ground-water and surface-water 

sites sampled during the solute-transport study, November 

1992.

1074
Prior to collection of raw and filtered samples 

for metal analysis temperature, specific conductance, 

dissolved oxygen, pH, and oxidation-reduction 

potential were measured in a flow cell until values of 

the measured properties had stabilized. After 

stabilization occurred, sample flow was redirected to a 

second flow cell to verify the pH measurement. A 

sample for the measurement of alkalinity was 

collected in outflow from the second flow cell by 

overfilling three bottle volumes before capping.

Following the measurement of above properties, 

the sample feed tube was disconnected and a raw 

sample was collected. The sample feed tubing was 

then connected to two tangential-flow filtration units 

that were connected in series, with the 0.45 jum pore 

size membrane preceding the 0.001 jum pore size 

membrane in the series. The initial filtrate was used to 

rinse the sample bottles (acid rinsed 125 mL 

polyethylene bottles) and then discarded. Raw and 

filtered water samples were acidified with trace-metalgrade HNC>3 to 0.05 equivalents of acid per liter.

Final Creek surface water, bank-seeps, and 

tributaries were sampled at numerous locations along 

the creek. Sampling was conducted in a manner 

similar to shallow ground water sampling, except that 

the flow cell was not used to measure water-quality 

properties. At those locations pH, temperature, 

specific conductivity and dissolved oxygen 

concentration were measured with conventional field 

meters.

Metal concentrations were measured for all 

samples using inductively coupled plasma atomic 

emission spectroscopy (ICP-AES). Total Fe 

concentration was also determined for selected 

samples using the ferrozene colorimetric method 

(Stookey, 1970). The analytical precision for 

determining concentrations of each metal was 

estimated by calculating an average standard deviation 

from triplicate determinations on each sample. An 

approximate coefficient of variation (CV) for each 

metal was calculated by dividing the mean metal 

concentration by the average standard deviation of 

triplicate determinations.

Filtration with a 0.45 |4.m pore-size membrane is 

routinely used to determine concentrations of 

dissolved constituents. If colloids are present, the 

dissolved fraction of metals is overestimated because 

of passage of colloidal phases through the filter. In 

order to test for metal-contaminant associations with 

colloids and suspended particulates in shallow

groundwater and surface water the method of 

sequential filtration of water samples was used 

(Kennedy and Zellweger, 1974; Laxen and Chandler, 

1983; Kimball and others, 1992). Paired water 

samples were collected at each measurement site: one 

sample was unfiltered and the other was sequentially 

filtered through two filter membranes of decreasing 

size (0.45 |4.m and 0.001 |4.m pore size).

Size-class fractions of particulates were 

operationally distinguished as follows; suspended 

particulate (>0.45 jum diameter), colloidal (0.001 - 

0.45 jam), and dissolved (<0.001 jum) components. 

Colloidal and suspended-particulate concentrations 

were calculated as the difference between 

concentrations in 0.45 |4.m and 0.001 jam filtrates, and 

unfiltered and 0.45 jum filtrates, respectively. The 

analytical precision of colloidal and suspendedparticulate concentrations is inflated because of the 

subtraction of two individual concentrations, each 

with associated precisions. For present purposes the 

average precision of a colloidal or suspendedparticulate concentration for a particular metal was 

estimated as,

(I)

where a co| is the average standard deviation of the 

calculated colloidal or suspended-particulate concentration for a particular metal, and a is the average standard deviation of the individual determinations for that 

metal.

Detection limits for individual determinations of 

Fe concentrations using ICP-AES were poor, 

necessitating the use of the colorimetric 

determinations for calculating Fe colloidal and 

particulate concentrations. Only five samples were 

analyzed colorimetrically for Fe, and determinations 

were done on unfiltered and 0.001 |um filtrates only. 

Thus only a sum of colloidal and particulate 

concentrations could be determined for Fe.

RESULTS

The potential importance of colloids and 

suspended particulates is apparent on bivariate plots of 

metal concentration in filtered water samples as a 

function of unfiltered samples. In figure 2, 

concentrations of Ni, Cu and Fe in filtered water 

samples are plotted as a function of concentrations in 

unfiltered water. The data plot roughly on an

1075
equivalence line with slope equal to one. The 

precision envelopes in figure 2 define a region equal to 

plus or minus a constant percentage of the metal 

concentration as determined from the coefficient of 

variation for that metal ([2CV2] 172). The precision 

envelopes consider analytical error in both filtered and 

unfiltered concentrations and are an appropriate guide 

for roughly evaluating whether data departures from 

the equivalence line are meaningful. Colloidal and 

particulate-metal concentrations are potentially 

important if concentrations plot outside and below a 

precision envelope surrounding the equivalence line; 

such a result implies that filtered concentrations are 

lower than unfiltered concentrations due to the 

removal of a paniculate metal phase during filtration.

Most plotted values in Figures 2a and 2b (for Ni 

and Cu respectively) plot within the precision 

envelope for those metals; outliers fall only slightly 

outside the precision envelope on both sides of the 

equivalence line. In contrast, all Fe values plot below 

the precision envelope and some values plot far below 

it (fig. 2c). Examination of plots for all selected metals 

investigated in this study (Mn, Ni, Cu, Co and Fe) 

indicated that only an Fe colloidal and paniculate 

phases was a significant fraction of the total unfiltered 

concentration for that metal at Final Creek.

Colloidal and suspended-paniculate 

concentrations were also examined using box plots 

(fig. 3). Box plots divide observations into quartiles 

and the 'notches' demarcate an approximate 95 percent 

confidence interval for the distribution around the 

median value. The median colloidal and suspendedpaniculate Mn concentrations were similar in 

magnitude to the average precision estimate for those 

determinations (fig. 3a). For Mn the notches overlap 

the limit of uncertainty for detecting Mn colloids and 

suspended particulates; this was also true for Ni, Cu 

and Co. This general result indicates that colloidal and 

suspended-paniculate concentrations were probably 

not large enough to be statistically distinguished from 

zero.

Too few analyses were available to construct a 

box plot for Fe (fig. 3b). Examining the data directly, 

four out of five Fe paniculate concentrations were 

above 0.02 mg/L, the analytical precision estimate for 

an Fe colloidal plus paniculate phase, and two of those 

concentrations were more than ten-fold greater than 

the analytical detection limit. Fe colloids and 

suspended particulates therefore appeared to be an 

important fraction of total Fe concentrations.

ll

LU LJ

is

^, Q.

a

1.0

0.9

0.8

o 0.45 Mm 

0.001 fj.m

0.5 0.6 0.7 0.8 0.9 1.0

N :<EL CONCENTRATION IN UNFILTERED WATER.

IN MILLIGRAMS PER LITER

1.1

(b)

0.22

0.18

0.1 4

0.10

± 0.06

002

o QA5(j.m 

0.001 nm

EQUIVALENCE LINE 

ANALYTICAL PRECISION

0.02 0.06 0.10 0.14 0.18

CC = ?ER CONCENTRATION IN UNFILTERED WATER,

IN MILLIGRAMS PER LITER

0.22

(c)

10°

0 0.45 Aim 

0.001 /im

EQUIVALENCE LINE 

ANALYTICAL PRECISION

10-1 TOO

-ON CONCENTRATION IN UNFILTERED WATER,

IN MILLIGRAMS PER LITER

Figure 2. Metal concentrations in filtered water samples as 

a function of concentrations in unfiltered samples; (a) Nickel, 

(b) Copper and (c) Iron.

1076
(a)

4 

3 

2 

I 

0

-1

-2

-3

-4

-5

MAX

p ^ r~

j^ftj...............<»W.i|f;Akt?!WSit*t->

COLLOIDAL PARTICULATE

1 2 

MANGANESE PHASES

(b)

0.5

0.0

-0.5

-1.0

eg -1.5 

o

-2.0

-2.5

- -ANALYTICAL- PRECI5K3 N -----

COLLOIDAL PLUS PARTICULATE PHASE

Figure 3. Concentrations of metals associated with colloidal 

and suspended particulate material, (a) Manganese, 

(b) Iron.

DISCUSSION

Colloidal and particulate-metal concentrations 

at Final Creek were low relative to dissolved-metal 

concentrations during base-flow conditions. Fe, Ni 

and Cu colloidal and particulate-metal concentrations 

were slightly higher than precision estimates, which 

implies that some metals are possibly being 

transported as colloids and suspended particulates. 

The precision estimates for colloidal and suspended 

particulate concentrations themselves were relatively 

high due to the inflation of error that is inherent in 

calculating colloidal and suspended-particulate 

concentrations by differences. Results presented here 

indicate that colloidal and suspended-particulate metal 

concentrations probably cannot be distinguished

statistically from zero at Final Creek during base-flow 

conditions.

Faires and Eychaner (1991) and Eychaner 

(1991b) reported that metals at Final Creek (excluding 

Fe) are lost from solution in the downstream direction 

in Final Creek. A significant decrease in dissolved 

concentrations of Mn, Ni and Co concentrations in the 

downstream direction in Final Creek was also 

observed in November 1992 (Fuller, 1996). The 

present study demonstrated that the decrease in 

concentrations of dissolved metals in the downstream 

direction probably can not be explained by the 

formation of metal colloids and suspended particulate 

phases.

0-4- Oxidation of Fe is a major source of colloids 

in other shallow ground-water and surface-water 

systems (Kimball and others, 1992; Ryan and 

Gschwend, 1990). At the Final Creek study site, Fe 

appeared to be associated with a colloidal and 

suspended-particulate phase greater than 0.00l|im in 

diameter. That material probably was colloidal Feoxyhydroxides, formed by oxidation of Fe2"1" in ground 

water discharging to the stream. The order-ofmagnitude higher Fe concentrations in seeps 

compared to those in ground water beneath the 

channel bed (Harvey and others, 1996) indicates that 

the source of Fe in Final Creek was discharge from 

bank seeps; this source is further supported by the 

observation of Fe floes at seep inflows.

Mn oxidation in natural waters is known to 

occur at rates several orders of magnitude faster than 

homogeneous oxidation in solution by dissolved 

oxygen. The oxidation rate can be enhanced by 

adsorption of Mn2"1" onto the surfaces of Fe 

oxyhydroxides, Al oxides, silica (Davies and Morgan, 

1989), and/or newly formed surfaces of Mn oxides 

(Hem, 1978). If present, Mn colloids are at 

concentrations of 1 mg/L or less compared to a 

dissolved concentration of approximately 90 mg/L. 

Such low concentrations of Mn colloids in Final Creek 

indicates that oxidation of Mn2 "1" throughout the 6-km 

reach probably occurs primarily through a surfacecatalyzed reaction on coatings of bed sediments. 

Trace metals such as Ni and Co are most likely 

partitioned to the solid phase by coprecipitation or 

sorption on Mn oxide and/or carbonate phases on the 

streambed (Lind and Anderson, 1992).

1077
SUMMARY

The fraction of total Mn, Ni, Cu and Co 

concentrations in Final Creek as colloidal and 

suspended-particulate material was low during baseflow conditions. Colloidal Fe was a significant 

fraction of total Fe concentrations in Final Creek, but 

since total concentrations of Fe in the stream are low 

at present, the importance of Fe colloids as a sorbent 

for other metals is at present probably minor compared 

to Mn and Fe solid phases on the streambed. 

Chemical conditions at Final Creek also apparently do 

not favor significant Mn-colloid formation and 

associated transport of solid-phase trace metals. Fe 

colloids will likely increase and substantially affect 

metal transport if the acidic core of the contaminant 

plume, where the Fe concentrations exceeds 50 mg/L, 

reaches Final Creek.

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geochemical factors affecting metal contaminant 

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Geological Survey Toxic Substances Hydrology 

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U.S.Geological Survey Water-Resources 

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oxidation kinetics on metal oxide surfaces: Journal of 

colloid and Interface Science, v. 129, p. 63-77.

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Eychaner, J.H., 1991b, Solute transport in perennial

streamflow at Pinal Creek, Arizona, in Mallard, G.E. 

and D.A. Aronson, eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S.Geological Survey Water-Resources 

Investigations Report 91-4034, p.481-485.

Faires, L.M., and Eychaner, J.H., 1991, Trace-element 

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D.A. Aronson, eds., U.S. Geological Survey Toxic 

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technical meeting, Monterey, California, March 11-15, 

1991: U.S.Geological Survey Water-Resources 

Investigations Report 91-4034, p. 461-465.

Ficklin, W.H., Love, A.H., and Papp, C.S.E., 1991, Solidphase variations in an aquifer as the aqueous solution 

changes, Globe, Arizona, in Mallard, G.E. and D.A. 

Aronson, eds., U.S. Geological Survey Toxic 

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technical meeting, Monterey, California, March 11-15, 

1991: U.S.Geological Survey Water-Resources 

Investigations Report 91-4034, p. 475-480.

Fuller, C.C., 1996, Experimental studies of trace metal 

partitioning in perennial reaches of Pinal Creek, 

Arizona, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program-Proceedings of the technical 

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24, 1993: U.S.Geological Survey Water-Resources 

Investigations Report 94-4015.

Hart, B.T., Noller, B.N., Legras, Christopher, and Currey, 

Nicholas, 1992, Manganese speciation in Magela 

Creek, Northern Australia: Australian Journal of 

Marine and Freshwater Research, v. 43, p. 421-441.

Hem, J.D., 1978, Redox processes at surfaces of manganese 

oxide and their effect on aqueous metal ions: Chemical 

Geology, v. 21, p. 199-218.

Hem, J.D., 1981, Rates of manganese oxidation in aqueous 

systems: Geochimica et Cosmochimica Acta, v. 45, p. 

1369-1374.

Kimball, B.A., McKnight, D.M., Wetherbee, G.A., and 

Harnish, R.A, 1992, Mechanisms of iron 

photoreduction in a metal-rich, acidic stream (St. 

Kevin Gulch, Colorado, U.S.A.): Chemical Geology, 

v. 96, p. 227-239.

Kennedy, V.C., Zellweger, G.W., and Jones, B.F., 1974, 

Filter pore-size effects on the analysis of Al, Fe, Mn, 

and Ti in water: Water Resources Research, v. 10, 

no. 4, p. 785-790.

Laxen, D.P.H., and Chandler, I.M., 1983, Size distribution 

of iron and manganese species in freshwaters: 

Geochimica et Cosmochimica Acta, v. 47, p. 731-741.

Langmuir, D., 1969, Geochemistry of iron in a coastal plain 

groundwater of the Camden, New Jersey Area: U.S. 

Geological Survey Professional Paper 650-c, p. 224- 

C235.

LeBlanc, D.R., Garabedian, S.P., Hess, K.M., Gelhar, L.W, 

Quadri, R.W., Stollenwerk, K.G., and Wood, W.W., 

1991, Large-scale natural gradient tracer test in sand 

and gravel, Cape Cod, Massachusetts, 1 

Experimental design and observed tracer movement: 

Water Resources Research, v. 27, p. 895-910.

1078
Lind, C.J., and Anderson, L.A., 1992, Trace metal 

scavenging by precipitating Mn and Fe oxides, in 

Kharaka and Maest, eds., Water-Rock Interaction: 

Rotterdam, Balkema, p. 397-402.

Lind, C.J., and Hem, J.D., 1993, Manganese minerals and 

associated fine particulates in streambed of Final 

Creek, Arizona: a mining-related acid drainage 

problem: Applied Geochemistry, v. 8, p. 67-80.

McCarthy, J.F., and Zachara, J.M., 1989, Subsurface

transport of contaminants: Environmental Science and 

Technology, v. 23, p. 496-502.

McKnight, D.M., Kimball, B.A., and Bencala, K.E., 1988, 

Iron photoreduction and oxidation in an acidic 

mountain stream: Science, v. 240, p. 637-640.

Penrose, W.R., Polzer, W.L., Essington, E.H., Nelson, D.M., 

and Orlandini, K.E., 1990, Mobility of plutonium and 

americium through a shallow aquifer in a semiarid 

region: Environmental Science and Technology, 1990, 

v. 24, p. 228-234.

Puls, R.W., and Eychaner, J.H., 1990, Sampling of

groundwater for inorganics - pumping rate, filtration 

and oxidation effects, in Fourth National Outdoor 

Action Conference on Aquifer Restoration, Ground 

Water Monitoring and Geophysical Methods, Las 

Vegas, Nevada, May 14-17, 1990, Proceedings: 

Dublin, Ohio, Water Well Journal Publishing 

Company, Ground Water Management, book 2 of a 

series, p. 313 - 327.

Puls, R.W, Powell, R.M., and Rees, T.F., 1991, Stability 

and transport of inorganic colloids through 

contaminated aquifer material, in Mallard, G.E. and 

D.A. Aronson, eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11- 

15, 1991: U.S.Geological Survey Water-Resources 

Investigations Report 91-4034, p. 507-510.

Ryan, J.N., and Gschwend, P.M., 1990, Colloid

mobilization in two Atlantic coastal plain aquifers: 

field studies: Water Resources Research, v. 26, no. 2, 

p. 307-322.

Stollenwerk, K.G., 1991, Simulation of Copper, Cobalt, and 

Nickel sorption in an alluvial aquifer near Globe, 

Arizona, in Mallard, G.E. and D.A. Aronson, eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program-Proceedings of the technical meeting, 

Monterey, California, March 11-15, 1991: 

U.S.Geological Survey Water-Resources 

Investigations Report 91-4034, p. 502-506.

Stookey, L.L., 1970, Ferrozine-a new spectrophotometric 

reagent for iron: Analytical Chemistry, v. 42, 

p. 779-781.

Stumm, W, and Morgan, J.J., 1981, Aquatic Chemistry: an 

introduction emphasizing chemical equilibria in 

natural waters: New York, John Wiley, 780 p.

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streams in Colorado, 1971-72, Colorado Water 

Resources Circular, 21, 117 p.

1079
1080
Solute-Transport Parameter Estimation for an Injection 

Experiment at Final Creek, Arizona

By Brian J. Wagner 1 anc/Judson W. Harvey 1

Abstract

Parameter estimation is an important step in 

the development of contaminant-transport 

simulation models. In this paper, we demonstrate 

an inverse modeling methodology for estimating 

the transport parameters that characterize the 

migration, attenuation, and redistribution of 

contaminants. The inverse model proceeds in two 

stages: In stage one, a finite difference solutetransport simulation model is coupled with 

nonlinear least-squares regression to identify the 

model parameter values that "best" reproduce the 

measured solute concentrations. In stage two, 

solute-transport simulation is combined with firstorder uncertainty analysis to quantify parameterestimate uncertainty. The methodology is widely 

applicable to transport models for ground-water 

and stream systems; here we demonstrate the 

application to a stream system. Using data from an 

injection experiment at Final Creek, Arizona, the 

inverse model is applied to analyze parameters for 

a one-dimensional solute-transport model with 

advection, dispersion, lateral inflow, and transient 

storage. The parameters estimated by the inverse 

model are dispersion coefficient, stream crosssectional area, storage-zone cross-sectional area, 

and stream-storage exchange coefficient. The 

parameter estimates and associated uncertainties 

support the interpretation that the transient-storage 

mechanism is active in Final Creek. A discussion 

of concentration sensitivity to these four estimated 

parameters is presented to explain how analysis of

'U.S. Geological Survey, Menlo Park, Calif.

uncertainties in parameter estimates can be used to 

identify sampling strategies for reliable parameter 

estimation.

INTRODUCTION

Toxic substances can be introduced into stream 

systems through point-source releases of contaminants, overland flow of contaminated water, and(or) 

by subsurface discharge from contaminated groundwater. Because an episode of contamination can affect 

stream-water and ground-water quality at distant 

locations, accurate prediction of the movement and 

redistribution of contaminants is essential. Solutetransport models are valuable tools for investigating 

the effect of contamination on downstream water 

quality. However, accurate prediction of the arrival 

times and distributions of contaminants at specific 

points in streams is difficult because the true values of 

the hydrologic and chemical parameters used in the 

solute transport model are unknown. Direct 

measurement of these parameters is difficult or 

impossible, and estimation by trial-and-error 

calibration does not guarantee that the estimates are 

optimal nor does it quantify the reliability of the 

parameter estimates. Our study, which follows the 

approach of Wagner and Corel ick (1986), uses a 

statistically based inverse model to estimate transport 

parameters and quantify the reliability of the 

estimates. Using data from an injection experiment at 

Final Creek, Arizona, we demonstrate here how the 

inverse model can be used to identify optimal 

parameter estimates and analyze estimate uncertainty. 

We then extend the work of Wagner and Gorelick 

(1986) to examine how parameter estimate uncertainty 

analysis can aid in evaluating the value of data for 

parameter estimation and sampling-network design.

1081
SOLUTE TRANSPORT IN STREAMS

Modeling of contaminant transport must begin 

with assumptions about the processes that control 

solute movement and redistribution. Transport 

modeling in streams typically begins with the onedimensional advective-dispersive equation. However, 

the advective-dispersive equation alone does not 

adequately describe contaminant transport in many 

small stream systems. For example, solute transport in 

small streams is commonly characterized by 

concentration histories which are attenuated and 

delayed with respect to concentration histories 

predicted by the advective-dispersive equation 

(Bencala and Walters, 1983). Solute transport in these 

streams is affected by dilution by lateral inflow and by 

exchange of solute between the main channel and a 

"storage" zone of slow-moving or immobile water at 

the channel margin or in the subsurface. In this 

conceptualization, the hydrologic regime is divided 

into two coupled systems: a system of flowing water in 

the main stream channel, and a system of storage 

zones containing immobile water. The two systems are 

coupled by a mass-transfer mechanism that exchanges 

solutes between the mobile and immobile regimes. 

The model for advective-dispersive transport with 

inflow and transient storage was formulated by 

Bencala (1984) as

dc -

A f)x A f)x

flC S A 

S = -«A (C(

fit A<, C) (2)

where

C is solute concentration in the stream (mg L ),

Q is volumetric flow rate (m3 s' 1 ),

A is cross-sectional area of stream channel

(m2),

2 .-I D is dispersion coefficient (m s" ), 

qL is lateral volumetric inflow rate (per length)

(m3 s" 1 m' 1 ), 

CL is solute concentration in the lateral inflow

(mgL- 1 ), 

C$ is solute concentration in the storage zone

As is cross-sectional area of the storage zone

(m2 ),

a is stream-storage exchange coefficient (s ), 

t is time (s), and 

jc is distance (m).

Equations 1 and 2 define a quasi-two-dimensional 

solute-transport model. In equation 1, the first two 

terms on the right-hand side represent the advection 

and dispersion mechanisms, the third term represents 

lateral inflow, and the fourth term represents transfer 

to the storage zone. Equation 2 couples the storage 

zones to the stream and assumes mass transfer to be 

the only process active in the storage zones. Although 

the model defined by equations 1 and 2 is a simplification of the complex transport processes active in a 

stream, it has been demonstrated to reproduce the transient storage observed in small stream systems (see for 

example Bencala and Walters 1983, Bencala, 1984, 

Jackman and others, 1984, Bencala and others, 1990). 

Consequently, determining the "storage capacity" of a 

stream is a necessary first step in any study aimed at 

exploring more complex solute-stream interactions.

INVERSE MODEL FOR PARAMETER 

ESTIMATION

Most of the parameters that characterize 

transport in streams are difficult or impossible to 

measure directly and are instead estimated by fitting 

the model to solute concentration data obtained in 

stream tracer experiments. Obviously, manual 

calibration cannot guarantee the parameter estimates 

are in any sense optimal. Moreover, given some 

uncertainty in the concentration data and the error 

inherent in the conceptual/mathematical model, there 

is always a degree of uncertainty associated with the 

parameter estimates. Consequently the model-based 

predictions are also uncertain. A statistically based 

simulation-optimization inverse model has the 

advantage of identifying optimal parameter estimates 

and accounting for the uncertainty inherent in the 

parameter estimates and model predictions. The 

approach taken here is to couple solute-transport 

simulation with nonlinear least-squares regression 

analysis and first-order uncertainty analysis. The goal 

of the transport simulation - nonlinear regression 

inverse model is to identify the parameter values for 

the transport model that "best" reproduce the 

measured solute concentrations. First-order 

uncertainty analysis is then applied to quantify the 

reliability of the optimal parameter estimates.

In brief, the objective of a simulation-regression 

inverse model is to identify the parameter estimates 

that minimize the weighted sum of squared deviations 

between simulated and observed concentrations.

1082
Parameter estimates that meet this objective are 

considered optimal for the given simulation model and 

concentration data. In order to apply the nonlinear 

regression methodology, it is necessary to postulate a 

statistical "model" for the random errors associated 

with the solute concentration data. These errors are 

defined as the difference between observed and 

simulated concentrations, and include measurement 

errors and errors resulting from the improper 

specification of the numerical simulation model. For 

this work, it is assumed the concentration errors are 

independent and normally distributed with mean zero, 

but it should be noted these assumptions are not 

required for application of the method. Under the 

assumption of independent and normally distributed 

errors, the nonlinear regression estimates are 

equivalent to maximum-likelihood estimates (Draper 

and Smith, 1981). A detailed discussion of nonlinear 

regression analysis for solute-transport parameter 

estimation is in Wagner and Gorelick (1986) and 

Wagner (1992).

An important component of parameter 

estimation is the analysis of parameter-estimate 

uncertainty. Asymptotically, the parameter estimates 

are normally distributed. As such, they are completely 

characterized by their mean and covariance. The 

parameter estimates serve as our estimate of the mean. 

The first-order approximation to the parameter 

estimate covariance matrix, COV, is (Yeh and Yoon, 

1981, Carrera and Neuman, 1986)

= a2 (J'J) ' (3)

where a" is the variance of the random concentration 

errors and J is the jacobian, the matrix of sensitivities 

of concentrations with respect to changes in the model 

parameters. In practice, a2 is unknown and is estimated based on the residual sum of squares (Draper 

and Smith, 1981). Together, the parameter estimates 

and the estimate covariance matrix serve to define statistically our knowledge of the parameters that determine transport. These statistical measures can be used 

to identify parameters that are poorly estimated, discriminate between competing model structures, assess 

the impact of parameter uncertainty on model 

predictions, and(or) evaluate the worth of data for 

experimental and monitoring network design.

SOLUTE TRANSPORT PARAMETER 

ESTIMATION FOR PINAL CREEK

For this study, the system of interest is Final 

Creek Basin, Arizona (fig. 1). Final Creek is a small 

meandering stream with low slope (approximately 

1 percent) and a coarse sand and gravel bed. In 

November, 1992, a tracer experiment was conducted 

to study the transport of solutes in Final Creek 

(Harvey and others, 1996). A bromide tracer was 

introduced into the stream and solute concentrations 

were monitored at five stations along a 809-m reach 

(see inset, fig. 1), resulting in five comprehensive 

concentration histories. Using solute-concentration 

data collected during this experiment, we wish to 

analyze the transport parameters that characterize 

solute migration, attenuation, and redistribution.

This report, describes the transport between the 

injection site and sampling location 2, which is located 

140 m downstream from the injection point. A CrankNicolson finite-difference approximation to the 

advective-dispersive and transient-storage transport 

model (equations 1 and 2) was developed to serve as 

the basis for the parameter-estimation study. The 

volumetric flow rate, Q, and the lateral inflow rate, qL, 

were calculated from streamflow measurements at the 

injection site and sampling site 2. The background 

bromide concentration was assumed known on the

110*55' 54' 53' UPS?

33'34'

33'

33'S?

X

Inspiration 

Dam

\

0.5

\

\

EXPLANATION

 ... ... INTERMITTENT REACH

 PERENNIAL REACH

DATA SITES-Number is 

site identifier

500 Observation well group

(09498400) A Streamflow-gaging station

9A Sample site

0 Channelized inflow

 Bank seep

0 1/2 1 MILE

1 KILOMETER

Figure 1. Site map for Final Creek Basin showing 

locations of surface and ground-water sampling stations for 

November, 1992 bromide tracer test.

1083
basis of background concentrations measured at the 

injection point and sampling site 2. These background 

concentrations serve to define the solute concentration 

of the lateral inflow, Q, and the initial concentrations 

in the stream and storage zones. Likewise, concentrations measured at sampling site 1 were used to 

define the concentration history at the upstream 

boundary of the simulation model. The resulting 

solute-transport simulation model has four unknown 

parameters - dispersion coefficient, D, stream crosssectional area, A, storage-zone cross-sectional area, As, 

and stream-storage exchange coefficient, a. It should 

be noted that lateral mixing of the tracer was enhanced 

by narrowing the stream channel just below the 

injection point to less than ten centimeters wide. This 

ensured complete cross-channel mixing, which is an 

assumption of the one-dimensional solute transport 

model used here.

By use of the inverse model described earlier, the 

four unknown transport parameters were estimated on 

the basis of 21 in-stream concentration measure-ments 

at sampling site 2 (fig. 2). For comparison, the inverse 

model was also applied to estimate the dispersion 

coefficient and stream cross-sectional area assuming 

the transient-storage mechanism was inactive (that is, 

a=0). The optimal parameter estimates and their 

uncertainties are summarized in table 1. The optimal 

fitted concentration breakthrough curves at sampling

Measured 

concentration

10.4 10.6 10.8 

TIME, IN HOURS

11

Figure 2. Optimal fitted concentration breakthrough 

curves for transport with transient storage (solid line) and 

without transient storage (dashed line). Circles represent 

concentration data used in the parameter estimation study.

location 2 are presented in figure 2. The parameter 

estimates and fitted breakthrough curves suggest that 

the transient-storage process is operative in Final 

Creek. In order to reproduce the measured 

concentration breakthrough accurately, the inverse 

model "activates" the transient-storage mechanism by

Table 1 . Results of solute transport parameter estimation for Final Creek tracer study

[D, dispersion coefficient; A, stream cross-sectional area; As, storage-zone cross-sectional area; a, stream-storage 

exchange coefficient; m /s, meters squared per second; m , meters squared; s , inverse seconds]

Parameter Parameter 

Estimate

Estimate 

Standard 

Deviation

Approximate 95 percent 

Confidence Limits

Lower Upper

Parameter Estimation with Transient Storage

D, m2/s 

A, m2 

Av , m2 

a,s-'

8.83 x IQ-2 

l.86x IO'2 

3.l2xlO'3 

6.70x lO'4

l.60xl(T2 

2.36 x I (T4 

2.64 x I O'4 

I.OSx lO'4

5.70 x IO'2 

1 .81 x IO'2 

2.60 xlO'3 

4.68 xlO'4

1.20 x 10' 1 

1.91 x 10"2 

3.64 x ID'3 

8.72 x lO'4

Parameter Estimation without Transient Storage

D, m2/s 

A,m2

7.27 x I O'2 

l.95xlO'2

3.40 x IO'2 

4.l3x IO'4

6.06 x 10'3 

1.87X10'2

1.39x10-' 

2.03 x lO'2

1084
selecting a nonzero value for the stream-storage 

exchange coefficient along with the "best fit" storagezone cross-sectional area. Furthermore, the uncertainty 

analysis indicates that these parameter estimates are 

reliable. The parameter-estimate standard deviations, 

as a percentage of the estimates, range from 

approximately 1 percent for the stream cross-sectional 

area to approximately 18 percent for the dispersion 

coefficient. The result is a fitted concentration 

breakthrough that is an excellent simulation of the 

observed data. Conversely, the advective-dispersive 

equation alone cannot adequately reproduce the 

transient storage effect, which is obvious at the 

"shoulder" of the breakthrough curve where the solute 

concentrations rise less rapidly than those predicted by 

the simple advective-dispersive model. The effect of 

this "lack-of-model-fit" is evident in the parameterestimate standard deviations, which have increased to 

approximately 2 percent of the stream cross-sectional 

area estimate and approximately 46 percent of the 

dispersion coefficient estimate.

The optimal parameter estimates and fitted 

concentration histories indicate the transient-storage 

mechanism is an important transport process in Final 

Creek. The validity of the model rests on the 

assumption that the storage parameter estimates are 

physically realistic. By comparing the estimates with

those in other studies (Bencala and Walters, 1983, 

Bencala, 1984, Bencala and others, 1990), it is evident 

that the values are at least plausible. Both the 

exchange coefficient and storage cross-sectional area 

estimates fall within the range of values in those 

studies.

SENSITIVITY ANALYSIS AND SAMPLING 

NETWORK DESIGN

Sampling-network design is, like parameter 

estimation, an optimization problem. The goal is to 

identify the sampling strategy (timing, location and 

number of measurements) that maximizes the 

information content of the data collected. The 

definition of "information content" depends on the 

sampling objective(s), but, in the case of parameter 

estimation, a logical goal would be to identify a 

sampling strategy that will yield reliable parameter 

estimates. In studies of solute-transport parameter 

estimation, the concentration sensitivity (partial 

derivative of concentration with respect to changes in 

a parameter value) plays an important role in defining 

the worth of data for parameter estimation (see 

Knopman and Voss, 1987). The concentration 

sensitivities for the parameter estimation problem 

presented in the previous section are shown in figure 3.

O

I

1

CO

HI 

CO

75

I 50

 - Q 25 zo

1000

(0

o> n O) u

2 

o

c/5 

-1000

1 1

1

1 1 1 1 1 1

,-v

, Fitted concentration ~ 

history

' \ \ i i i I

I I I

/' Sensitivity for 

/ measured 

concentration

Concentration 

sensitivity

10.5

-6000

11 10.5 

TIME, IN HOURS

Figure 3. Concentration sensitivity plots for (a) dispersion coefficient (b) stream cross-sectional area (c) storage-zone crosssectional area and (d) stream-storage exchange coefficient, sampling site 2 in Final Creek, Ariz. Solid line represents 

concentration sensitivity. Dashed line represents optimal fitted concentration breakthrough. Circles denote concentration 

sensitivities for the data used in the parameter estimation study.

1085
Shown are the sensitivities of simulated streamwater 

concentrations to dispersion (dCI dD, fig. 3a), stream 

cross-sectional area (dC/ dA, fig. 3b), storage-zone 

cross-sectional area (dCI dAs, fig. 3c) and streamstorage exchange coefficient (5C/ da, fig. 3d). 

Superimposed on the sensitivity plots is the 

concentration breakthrough curve simulated using the 

optimal parameter estimates. Highlighted by circles 

are the concentration sensitivities for the data used in 

the parameter estimation study.

From inspection of equation 3 and the 

concentration sensitivity plots in figure 3, the 

following principles of sampling network design 

emerge: (1) Information about a parameter is most 

efficiently gained by sampling at points with high 

sensitivity to the parameter. Concentration 

sensitivities define the elements of the jacobian 

matrix in equation 3. By maximizing the absolute 

sensitivities, we maximize the elements of the matrix 

J*J, which is inversely proportional to the parameter 

estimate covariance (see eq. 3). (2) Data that 

minimize the uncertainty of one parameter may not 

minimize the uncertainty of other parameters. 

Inspection of figure 3c indicates the maximum 

sensitivity for storage-zone cross sectional area 

occurs at 10.6 hours. Sensitivities for the other 

parameters, however, are far from their maximum 

absolute values at this time. (3) The concentration 

history is sensitive to the parameters over specific, 

narrow time spans. The maximum sensitivity to the 

dispersion coefficient occurs at the first arrival of 

solutes. For the stream cross-sectional area, the 

maximum sensitivity occurs at the mid-point of the 

rise of the breakthrough curve. For the storage-zone 

cross-sectional area and stream-storage exchange 

coefficient, the maximum sensitivities occur at the 

tail and shoulder, respectively, of the breakthrough 

curve. (4) The sensitivity plots presented in figure 3 

represent local derivatives of concentration with 

respect to these parameters. The range of parameter 

values for which these sensitivities are valid will 

depend on the degree of nonlinearity of the 

governing transport model.

Observation (4) above suggests that the design 

of an "optimal" sampling network will be based on 

assumptions regarding the governing transport model 

and the parameter values that, paradoxically, the 

sampling network is being designed to estimate. Two 

steps can be taken to overcome this difficulty. When 

possible, data collection should be approached as a

multi-stage process in which, after each sampling 

stage, the parameter estimates are refined and the 

sampling network is revised or supplemented. When 

multi-stage sampling is not possible, such as in the 

case of a stream-tracer study, robust sampling 

strategies should be identified to ensure the parameter 

estimates are reliable for the range of possible 

parameter values.

SUMMARY

Solute-transport simulation is a valuable tool for 

investigating the effect of contamination on water quality. 

A key step in the development of solute-transport 

simulation models is the estimation of parameters that 

characterize solute migration, attenuation, and 

redistribution. A technique for estimating the parameters 

that characterize the transport of contaminants has been 

demonstrated using data from an injection experiment at 

Final Creek, Arizona. The methodology, which couples 

solute-transport simulation with nonlinear least-squares 

regression and first-order uncertainty analysis, was 

applied to analyze parameters for a one-dimensional 

solute-transport model with advection, dispersion, lateral 

inflow, and transient storage. Results of the analysis 

suggest that accurate simulation of transport in Final 

Creek requires modeling both the advective-dispersive 

and transient-storage processes. Conversely, parameter 

estimation for the simple advective-dispersive model 

indicates that advection and dispersion alone cannot 

adequately reproduce the effects of transient storage. 

Current work focuses on extending the analysis presented 

here to study the trends in parameter estimates along an 

809-m reach of Final Creek.

A logical goal of sampling-network design for a 

tracer experiment is to identify a sampling strategy that 

will yield reliable parameter estimates. We demonstrated 

how the analysis of concentration sensitivities to the 

unknown parameters of the transport model can be used 

to evaluate the value of data for parameter estimation. 

Although this example did not address some important 

issues of parameter estimation (for example, parameter 

non-identifiability), it nonetheless demonstrates the 

quantitative relation between concentration sensitivities, 

estimation uncertainty, and sampling-network design.

1086
REFERENCES CITED

Bencala, K.E. 1984, Interactions of solutes and streambed 

sediment 2. A dynamic analysis of coupled hydrologic 

and chemical processes that determine solute transport: 

Water Resources Research, v. 19, no. 3, p. 718-724.

Bencala, K.E., McKnight, D.M., and Zellweger, G.W., 

1990, Characterization of transport in an acidic and 

metal-rich mountain stream based on a lithium tracer 

injection and simulations of transient storage: Water 

Resources Research, v. 26, no. 5, p. 718-724.

Bencala, K.E., and Walters, R.A., 1983, Simulation of

solute transport in a mountain pool-and-riffle stream 

A transient storage model: Water Resources Research, 

v. 19, no. 3, p. 718-724.

Carrera, J., and Neuman, S.P., 1986, Estimation of aquifer 

parameters under transient and steady-state conditions, 

1 Maximum likelihood method incorporating prior 

information: Water Resources Research, v. 22, no. 2, 

p. 199-210.

Draper, N.R., and Smith, H., 1981, Applied Regression 

Analysis (2d ed.): New York, John Wiley, 709 p.

Harvey, J.W., Fuller, C.C., and Wagner, B.J., 1996,

Interactions between shallow ground water and surface 

water that affect metal transport in Final Creek, 

Arizona, in Morganwalp, D.W., and Aronson, D.A., 

eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colorado, September 20- 

24, 1993: U.S. Geological Survey Open-File Report 

94-4015.

Jackman, A.P., Walters, R.A., and Kennedy, V.C., 1984, 

Transport and concentration controls for chloride, 

strontium potassium and lead in Uvas Creek, a small 

cobble-bed stream in Santa Clara County, California, 

U.S.A., 2 Mathematical modeling: Journal of 

Hydrology, v. 75, p. 111-141.

Knopman, D.S., and Voss, C.I., 1987, Behavior of 

sensitivities in the one-dimensional advectiondispersion equation Implications for parameter 

estimation and sampling design: Water Resources 

Research, v. 23, no. 2, p. 253-272.

Wagner, B.J., 1992, Simultaneous parameter estimation and 

contaminant source characterization for coupled 

groundwater flow and contaminant transport 

modelling: Journal of Hydrology, v. 135, p. 275-303.

Wagner, B.J, and Gorelick, S.M., 1986, A statistical 

methodology for estimating transport parameters: 

Theory and applications to one-dimensional advectivedispersive systems: Water Resources Research, v. 22, 

no. 8, p. 1303-1315.

Yeh, W.W-G., and Yoon, Y.S., 1981, Aquifer parameter 

identification with optimum dimension in 

parameterization: Water Resources Research, v. 17, 

no. 3, p. 664-672.

1087
1088
Alteration of Alluvium by Acidic Ground Water Resulting 

from Copper Mining at Final Creek, Arizona

By Carol J. Lind1 and Kenneth G. Stollenwerk2

Abstract

A composited core of alluvium collected 

downgradient from a metal-rich, acidic groundwater plume was eluted with ground water from 

the plume in a column experiment.

The major minerals present in the alluvium 

before and after elution were identified, and 

resulting changes in mineral content and in 

ground-water composition were correlated. Of 

the 12 identified minerals, the ground water 

dissolved two carbonate minerals and pyrolusite 

and partially dissolved the feldspars.

INTRODUCTION

More than 100 years of copper (Cu) mining has 

created a metal-rich, acidic ground-water plume in the 

Final Creek, Arizona, drainage basin (Envirologic 

Systems, Inc., 1983). The plume is formed as the 

ground water moves through the alluvium below the 

intermittent reach of Final Creek and mixes with water 

infiltrating from the Cu-mining area. This contaminated 

ground water reacts with the alluvium, dissolving some 

minerals and depositing others. The reacted ground 

water, moving downgradient from the mining area 

towards the creek mouth, contributes to the streamflow 

in the perennial reach of Final Creek (fig. 1).

This paper defines, by means of X-raydiffraction (XRD) analysis, the major minerals present 

before and after contact with the metal-rich, acidic 

ground water and correlates changes in mineralogy 

with changes in ground-water composition. This work 

is the first step in the study of the alluvium

'U.S. Geological Survey, Menlo Park, Calif. 

2U.S. Geological Survey, Denver, Colo.

composition before and after exposure to this 

contaminated water. The ultimate study objective is to 

define the location and, if possible, oxidation states of 

trace metals, especially iron (Fe) and manganese 

(Mn), within the mineral structures of the major 

minerals present and at the surfaces of these minerals.

MINERALOGIC BACKGROUND

The Final Creek shallow alluvium consists of 

sand, gravel, and fragments of mineralized schist and 

granite porphyry. The shallow alluvium is surrounded 

by basin fill that consists of interfingering lenticular 

beds of gravel, sand, and silt, many of them firmly 

indurated by calcareous cement. The sand consists 

mostly of dacitic and decomposed diabase, but grains 

of other rock types also are present. Most of the 

fragments are well rounded (apparently from being 

transported considerable distances) and consist mostly 

of rocks of the Apache Group and Troy Quartzite: 

other rock materials consist of limestone, dacite and 

diabase (Ransome, 1903; and Peterson, 1962).

MATERIALS AND METHODS

The studied alluvium was a composite of five 

samples of the <2-mm (millimeter) core material from 

well 600 (fig. 1). Well 600 is in the perennial reach of 

the creek, downgradient from the acidic, metal-rich, 

ground-water plume. The depth intervals, in feet, 

from which the five composited alluvial samples were 

collected are 11 to 13, 16 to 17,20 to 23,26 to 28, and 

36 to 39.5. The samples contained 0.26 to 

0.67 percent carbonate (as calcite) and the composited 

sample contained 0.41 percent carbonate. The ground 

water to be eluted was sampled from well 101, which 

is directly in the path of drainage from Cu-mining 

activities and adjacent to well 100 (fig. 1).

1089
ARIZONA

STUDY 

EA

INDEX MAP

EXPLANATION 

AQUIFER BOUNDARY

 400 WELL SITE AND 

NUMBER

. 9f4M400 STREAMFIOW-SAWPU 

SHE ANDNVA«£H

< GENERALIZED DIRECTION OF GROUNDWATER FLOW

Figure 1. Map showing the Final Creek drainage basin, the mines, tailings, and well sites (adapted from Eychaner (1991), 

p. 440, fig. 1).

A column, 30 cm (centimeters) long by 2.5 cm 

inside diameter, was packed with the composited 

alluvium, flushed with nitrogen, eluted with deionized 

water (also flushed with nitrogen), and then eluted 

with the ground water which had a pH of 3.86. A total 

of 13.6 pore volumes were required to change the 

groundwater effluent pH from 7.6 to 4.6. A pH of 4.6

represents pH of water well within the advancing 

ground-water plume in 1990 (Lind and Hem, 1993). 

Table 1 lists the original ground water and the final 

effluent aliquot compositions.

The measurement of magnetic susceptibility and 

grain size distribution in the alluvium before and after 

exposure to the ground water (original alluvium and

1090
eluted alluvium, respectively) provided a general 

knowledge of alluvial composition and its alteration 

by the contaminated ground water. Also, the 

concentration of certain minerals by segregation 

according to magnetic susceptibility facilitated their 

identification by XRD. The original and eluted 

alluvium were each separated by size into six fractions 

by dry sieving through an acid-washed, nonmetallic 

sieve (table 2). The more magnetic particles (MMP) 

were removed from each of these 12 fractions with a 

hand-held horseshoe magnet covered with a 

disposable glove. The alluvial particles segregated by 

hand magnet are categorized as nonmagnetic (NMP, 

not removed by contact with magnet poles), magnetic 

(MP, removed by contact with magnet poles), strongly 

magnetic (SMP, removed by waving magnet slightly 

above the sediment), and, for the 0.25- to 2-mm size 

fraction, very strongly magnetic (VSMP, removed by 

waving magnet about 1 cm above the alluvium). The 

NMP 150- to 250-(j,m size fractions were further 

segregated with a Frantz magnetic separator according 

to their slight susceptibility to a magnetic field.

Table 1. Composition of groundwater from well-101 and of 

final aliquot of eluted water

[Ground water from Well lOI, sampled May 27, 1992; mV, millivolts; uS 

@25°\ microsiemens at 25 degrees; mg/L, milligrams per liter; std, standard; C, carbon; Ca, calcium; O, oxygen;( ), essentially all iron is divalent; 

?, values unreliable; and NA, no analysis.]

_. . Units of Property or constituent 

measurement

Oxidation- mV

reduction potential

Specific Conductance ^iS/cm @25o

Dissolved Oxygen mg/L

pH (field) std. units

(laboratory)

Total hardness mg/L as

CaCO3

Major dissolved mg/L

ions I 1

Aluminum

Calcium

Copper

Iron

Magnesium

Manganese

Sodium

Chloride

Sulfate

Value or concentration

Ground water 

before elution

430

4570

0.1

3.6

3.86

1500

74.

350.

47.

740.

150.

25.

84.

140.

300.

Final eluted 

aliquot

4.63

-19.

872.

21.

(710.)

189.

59.

9

NA

NA

'Besides the metals listed, one or more mg/L of Co, Ni, Sr, and Zn 

were in the groundwater before elution. Only Co and Zn were measured in 

the eluant and the analysis precision did not warrant listing their concentrations.

With the exception of the 20- to 38-(im eluted 

alluvium and the 63- to 75-(j,m MMP of the original 

and the eluted alluvium, a representative aliquot of the 

each magnetically segregated part (including those 

separated by the Frantz magnetic separator) of every 

size fraction was removed by splitting, finely ground, 

and then examined by XRD. The results were 

analyzed using the Nicolet computer program and the 

major minerals were identified. The gain or loss of a 

mineral phase caused by elution of the ground water 

was estimated by comparison of XRD patterns for the 

original alluvium with those for the eluted alluvium. 

This gain or loss in a magnetically segregated part of a 

size fraction was based on the change in XRD relative 

intensities (RI), relative peak heights, for the mineral 

in question compared with the XRD RI of the 

predominant mineral for that segregated part. The 

predominant minerals were magnetite for the most 

magnetic part and quartz for the lesser magnetic parts 

of each size-fraction. The XRD RI for a given mineral 

are the ratios of counts-per-second of the secondary 

peaks divided by counts-per-second of the strongest 

peak multiplied by 100. Small concentrations changes 

for the most predominant mineral for a segregated part 

would not be observed using this method.

RESULTS AND DISCUSSION

General Observations

Much of the alluvium had a grain size of 0.25 to 

2 mm and the weight-percent of each size fraction 

(relative to the total sample) decreased with 

decreasing grain size. The 11 percent decrease in the 

0.25- to 2-mm size fraction and the 11 percent increase 

in the 38- to 250-(j,m size fractions after elution 

indicates that the acidic ground water dissolved and 

(or) disaggregated and reduced the size of the alluvial 

particles (table 2). In addition to dissolution of the 

alluvium by the ground water, the loss of some of the 

20- to 38-(j,m size fraction could be caused by particles 

being washed out by the eluant.

In the alluvium, MMP and NMP must be 

dissolving simultaneously because in all of the <250- 

to the >38-(j,m size fractions, the weight-percent of 

NMP was about the same after elution as before and 

there was only a one percent decrease in the 0.25- to 2- 

mm size fraction, probably not experimentally 

significant (table 2).

1091
Table 2. Distribution of particles in alluvium by size and magnetic susceptibility

[NM P, nonmagnetic part = not removed when in contact with the poles of a magnet, a horseshoe hand magnet covered by a 

disposable glove; MP, magnetic part = removed by contact with magnet poles; SMP, strongly magnetic part = removed by 

waving magnet slightly above the alluvium; VSMP, very strongly magnetic part = removed by waving magnet about 

1 centimeter above alluvium; urn, micrometer; %, weight-percent; xx, original alluvium; and [xx], eluted alluvium;]

Size 

(¥m)

250-2000 

150-250

75-150

63-75

38-63

220-38

NMP 

(% of size fraction)

87.88 

[86.90] 

92.52

[92.95] 

91.78

[91.87] 

96.16

[96.11] 

99.34

[99.03]

NMP MP SMP VSMP Total

(% of the entire sample)

57.60 

[47.29] 

18.15

[22.77] 

8.78

[13.41] 

1.92

[3.63] 

3.00

[2.64] 

.26

[.03]

5.19 

[6.48] 

1.11

[1.20] 

.49

[-74] 

.04

[.06] 

.02

[.03]

2.29 0.46 

[.37] ['.28] 

.36

[.52] 

.29

[.45] 

.04

[.09]

65.54 

[54.42] 

19.62

[14.60] 

2.00

[3.78] 

3.02

[2.67] 

.26

[.03]

1 00.00

[99.99]

A small amount of this fraction was spilled before it was weighed. 

2Mostly nonmagnetic but some magnetic particles evident.

Distribution of Identified Minerals in Original 

and Reacted Alluvium3

After elution, the decrease of XRD RI for a 

given mineral for all size fractions was considered 

evidence of dissolution of that mineral; the 

indiscriminate fluctuation of XRD IR for a mineral 

between no change, slight increase, and slight 

decrease as the particle size decreased was considered 

a result of redistribution of the mineral caused by the 

dissolution of other mineral(s) bound with it.

Twelve minerals were identified. Magnetite 

[Fe3C>4] was by far the predominant mineral in the 

most magnetic part of each size fraction and found 

only in these parts. Quartz [SiO2] was a minor 

constituent of the most magnetic part of each size 

fraction and the predominant mineral of the other 

parts. The SMP of the original alluvium contained 

pyrolusite [MnO2] and possibly prophanite [MnTiO3 ] 

and galaxite [MnAl2O4]. All MMP's contained

3The symbols for elements are aluminum (Al), carbon (C), 

calcium (Ca), fluorine (F), hydrogen (H), potassium (K), magnesium (Mg), sodium (Na), oxygen (O), silica (Si), and titanium (Ti).

hematite [Fe2O3] in all size fractions and ilmenite 

[FeTiO3 ] in the >75-jam size fractions. Muscovite 

[(KAl2(Si3Al)O 10(OH,F)2] and tremolite 

[Ca2Mg5SigO22(OH)2] were most abundant in the 

MP's of each size fraction, and chlorite 

[(Mg,Fe,Al)6(Si,Al)4O|o(OH)8] was most abundant in 

the part separated by the Frantz magnetic separator at 

0.4 amps with a slide slope of 20°. The feldspars (low 

albite [NaAlSi3O8], and labradorite 

[(Ca,Na)(Al,Si)AlSi2O8]), were present in all but the 

VSMP of the 0.25- to 2-mm size fraction. A Ca-Mg 

carbonate with an [100] intensity reflection between 

that of dolomite [CaMg(CO3)2] and huntite 

[Mg3Ca(CO3)4] and possibly kutnahorite 

[Ca(Mn,Mg)(CO3)2] were noted in the NMP of the 

150- to 250-jam size fraction (fig. 2).

The ground-water elution removed pyrolusite 

and the carbonates, partially dissolved the feldspars, 

caused a XRD IR decrease for hematite in all the 

<250-fim size fractions, and slightly decreased most 

XRD RI's for tremolite (fig.2). The partial dissolution 

of feldspar and possibly of quartz could contribute to 

redistribution of ilmenite, muscovite, and chlorite 

among magnetically segregated parts and among size 

fractions.

1092
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The XRD did not detect minor minerals or 

newly precipitated minerals in the 150- to 250-|im size 

fraction separated by the Frantz magnetic separator. 

Correlation of the precipitation model of Bigham and 

others (1992) with ground-water composition suggests 

the possible presence of ferrihydrite (~Fe5HO8 .4H2O) 

in the original alluvium and schwertmannite 

(~Fe8O8(OH)6SO4) in the eluted alluvium, but XRD 

did not detect these minerals.

Mineralogical Changes Correlated With Water 

Chemistry

At the time the elution was stopped, the groundwater/alluvium interaction was removing Al, Cu, and 

Fe from the ground water and Ca, Mg, and Mn from 

the alluvium. Many identified minerals in the 

alluvium contained Al and (or) Fe, but after elution, no 

new Al, Cu, and (or) Fe solids were discernable by 

XRD. Quite possibly, the percentage of new, wellcrystallized minerals was too low to be seen by XRD 

or, more likely, much of the Al and Fe was 

precipitating as finely divided amorphous solids that 

could also remove Cu. The observed increase in Ca, 

Mg, and Mn concentrations in the eluant could be 

explained by dissolution of feldspars, carbonates, and 

perhaps tremolite to release Ca; carbonates and 

possibly tremolite to release Mg; and pyrolusite and 

carbonate to release Mn. The data did not permit 

conclusions concerning distribution of the trace metals 

Co, Cd, Ni, Sr, and Zn (table 1).

SUMMARY

A composited core of alluvium collected 

downgradient from a metal-rich, acidic ground-water 

plume was eluted with the metal-rich, acidic ground 

water in a column experiment. Alluvium from before 

and after the elution was segregated by size and 

magnetic susceptibility and then examined by XRD. 

The alteration of minerals by the elution was estimated 

by comparison of their XRD relative peak intensities

with those of the major mineral present in the 

respective magnetically segregated part of each grainsize fraction.

Twelve minerals were identified. The 

dissolution of the carbonates (dolomitic in structure) 

pyrolusite, and perhaps tremolite, and the partial 

dissolution of feldspars could explain increases in Ca, 

Mg, and Mn concentrations in the final ground-water 

eluant; precipitation, exchange and (or) surface 

reactions could explain decreases in Al, Cu, and Fe 

concentrations in this water. The chemical formulas of 

many of the 12 initially identified minerals contain Fe 

or Mn. No new Al, Cu, or Fe minerals were identified 

after elution.

SELECTED REFERENCES

Bigham, J.M., Schwertmann, Udo., and Carlson, Lisa., 

1992, Mineralogy of precipitates formed by the 

biogeochemical oxidation of Fe(II) in mine drainage: 

Catena Supplement v. 21, 14 p.

Envirologic Systems, Inc., 1983, Mining activities and 

water-quality report: Florence, Arizona, Central 

Arizona Association of Governments, Report METF-7, 

137 p.

Eychaner, J.H., 1991, The Globe, Arizona, research site 

Contaminants related to copper mining in a 

hydrologically integrated environment, in Mallard, 

G.E., and Aronson, D.A., eds., U.S. Geological Survey 

Toxic Substances Hydrology Program Proceedings of 

the technical meeting, Monterey, California, March 11- 

15, 1991: U.S. Geological Survey Water-Resources 

Investigations Report 91-4034, p. 440, fig. 1.

Lind, C.J., and Hem, C.J., 1993, Manganese Minerals and 

associated fine particulates in the streambed of Pinal 

Creek, Arizona A mining-related acid drainage 

problem: Applied Geochemistry, v. 8, no. 1 p. 67-80.

Peterson, N.P., 1962, Geology and ore deposits of the 

Globe-Miami district, Arizona: U.S. Geological 

Survey Professional Paper 342, 151 p.

Ransome, F.L., 1903, Geology of the Globe copper district, 

Arizona: U.S. Geological Survey Professional 

Paper 12, 168 p.

1094
Chemical Processes in Manganese Oxide and 

Carbonate Precipitation in Pinal Creek, Arizona

By John D. Hem1 and Carol J. Lind1

Abstract

Two water samples from well 503 near the 

head of the perennial reach of Pinal Creek and 

one sample of surface flow at the Setka Ranch 

crossing about 2 kilometers downstream from the 

well were used as the manganese-bearing solution 

in automated pH-stat titration with a 0.10 molar 

NaOH solution while bubbling CO2-free air into 

the mixture. The three samples had manganese 

concentrations ranging from 63 to 94 mg/L and 

dissolved solids concentrations ranging from of 

3,500 to 3,570 mg/L, mostly consisting of Ca and 

804. The ground-water samples initially were 

about 5 x 10"3 molar in undissociated dissolved 

CO2 and had pH's near 6. The creek water had a 

pH near 7 and a CO2 concentration smaller than 

10"3 molar. Titration of one ground-water sample 

done without removing the dissolved CO2 

produced a mixture of hausmannite (MiijO^ and 

kutnahorite (CaMn(CO3)2). The other two 

samples were treated to decrease the dissolved 

CO2 species before titration. One of these samples 

yielded mainly hausmannite during titration that 

altered on aging to manganite (MnOOH). The other 

sample yielded a mixture of oxides that altered on 

aging to a mixed Ca + Mn species similar to 

todorokite ((Mn2+, Ca) Mn**"5 O,, 4H2O) in 

which most of the manganese is in the 4+ 

oxidation state. X-ray-diffraction studies of 

precipitates from Pinal Creek identified 

kutnahorite in carbonate cemented crusts near the 

head of perennial flow. The mixed Ca + Mn4+ 

oxides also were identified near the head of

'U.S. Geological Survey, WRD, Menlo Park, Calif.

perennial flow and in black precipitates farther 

downstream.

INTRODUCTION

The ground water in alluvium underlying Pinal 

Creek has been contaminated by seepage of waste 

solutions produced in copper mining and ore treatment 

that were impounded by piles of mill tailings in the 

tributary basin of Miami Wash. In February 1988 the 

impounded water had a pH of 2.7. Dissolved 

concentrations at that time included: sulfate, 

20,000 mg/L; iron, 5,970 mg/L; aluminum, 850 mg/L; 

copper, 210 mg/L; manganese, 100 mg/L; zinc, 

32 mg/L; and cobalt, 20 mg/L (Brown, 1990).

The contaminated water has escaped from the 

impoundment as underflow in the alluvium that 

underlies the tailings piles and moved downstream 

beneath Miami Wash into the alluvial aquifer that 

underlies the valley of Pinal Creek (fig. 1). As it 

moved downstream the acidity in this seepage plume 

was neutralized by reactions with solids (mainly 

CaCO3 ) in the alluvium, and most metals are removed. 

The neutralized seepage attains a pH between about 

5.5 and 6.5. About 8 km downstream from the 

confluence of Pinal Creek and Miami Wash, the valley 

narrows, forcing groundwater to the surface and 

creating perennial flow in the creek (fig. 1). 

Predominant solutes in the neutralized water at this 

point are calcium and sulfate at concentrations near 

saturation with respect to gypsum. Chemical analyses 

of surface flow at the Setka Ranch crossing, near the 

head of flow, are given by Brown (1990). The analyses 

of surface flow in the creek and of adjacent shallow 

ground water show that in recent years dissolved 

manganese concentrations have commonly ranged 

from 50 to 100 mg/L. Black manganese oxide crusts 

occur on rock and sand surfaces in the creek bed.

1095
SETKA 

RANCH 

CROSSING

UNITED STATES

EXPLANATION 

 ^OBSERVATION-WELL

GENERAUZED DIRECTION

OF aow

2 MILES

Figure 1. Location map of study area showing sampling sites. (Modified from Brown, 1990, fig. 1).

1096
This paper presents results obtained when 

manganese was precipitated from samples of 

contaminated water from the Final Creek site under 

carefully controlled laboratory oxidation conditions. 

Precipitates obtained from the contaminated water are 

compared with the precipitates that occur in the Final 

Creek system and with those obtained in earlier 

research.

EXPERIMENTAL

Previous laboratory studies have defined effects 

of temperature and pH on rates of oxidation and the 

form of oxide produced. The latter also is affected by 

the predominant anion in the solution (Hem, 1978, 

1980, 1981; Hem and Lind, 1983). The oxidation 

occurs preferentially at solid mineral surfaces, 

especially those of preexisting manganese oxide.

Table 1. Chemical composition of water from Well 503, near 

Final Creek, 16 km. northwest of Globe, Arizona and from 

surface flow in Final Creek at Setka Ranch, 18 km northwest 

of Globe, Arizona

[Concentrations of major constituents are milligrams per liter, except as 

noted]

Source date

Silica (Si02)

Manganese (Mn)

Calcium (Ca)

Magnesium (Mg)

Sodium (Na)

Undissociated

Diss. C02 Calc. (H2C03 )

Bicarbonate (HC03)

Sulfate (S04)

Chloride (Cl)

Dissolved solids (Sum)

Well 503 

Jan. 11, 

1989

69

86

690

1 40

91

3801

1132

2,300

120

3,550

Well 503 

Nov. 22, 

1991

67

94

655

1 70

113

3201

572

2,250

124

3,500

Final Creek 

Setka Ranch 

Jan. 26, 1989

60

63

680

1 50

91

46'

no2

2,400

120

3,570

Specific Conductance

(uS/cm, 25°C)

pH (field)

pH (lab)

Dissolved Oxygen (02 )

Temperature (°C)

Source of Data

3,800

5.94

6.44

6.44

18.0

Brown (1990)

3,610

5.72

8.763

<.l

18.0

Analyst

J.D. Hem

3,880

6.80

4.934

6.7

18.5

Brown (1990)

Based on field measurements of HC03 ~ and pH. 

2Determined in field at time of sampling. 

3 pH after treatment to lower H2C03". 

4pH after treatment to decrease HCO3".

Two samples of ground water from observation 

well 503 (fig. 1) collected in 1989 and 1991, and a 

sample of water from Final Creek at the nearby Setka 

Ranch crossing collected in 1989, were used in the 

laboratory experiments. Analyses of these samples 

are given in table 1. The manganese was oxidized by 

a continuous titration technique, in which manganesebearing solution was added to a temperaturecontrolled reaction vessel at a constant slow rate by an 

autoburet. The pH was maintained at a prescribed 

constant value by addition of a standard base (0.10 

molar NaOH) from a pH-stat, while air freed from 

CO2 was bubbled into the stirred reacting solution. A 

steady state was reached after a brief induction period 

and the composition and mineralogy of the precipitate 

and solution composition were determined by taking 

samples from the reaction vessel. After titration was 

stopped the solution with its precipitate was allowed to 

age at room temperature. Additional samples of the 

solution and precipitate were analyzed during aging.

The first precipitation experiment (503-1) used 

water from well 503, which is screened at the depth 

interval of 23.4 to 24.1 m. When the sample was 

collected, its pH and bicarbonate alkalinity were 

determined. From these data the concentration of 

undissociated carbon dioxide (CO2) given in table 1 

was calculated. Some of the CO2 had escaped by the 

time the laboratory pH measurement was made, but 

the amount present was still substantially greater than 

100 mg/L (as H2CO3). For the other two experiments, 

PC-1 and 503-2, the water samples were pretreated to 

decrease their dissolved CO2-species content. Sample 

PC-1, water from Pinal Creek at Setka Ranch, was 

acidified to pH 4.93 and bubbled with CO2-free air, 

which removed CO2 and most of the HCC>3~ before the 

titration began. The third titration used a later (Nov. 

22,1991) sample from well 503, which was pretreated 

by bubbling into it CO2-free air, without prior 

acidification. This treatment did not decrease the 

HCC>3~ concentration, but the final pH was 8.76.

The experimental conditions in 503-1 were 

intended to simulate conditions in ground water 

immediately after coming into contact with air, as in 

seepage just as it enters Pinal Creek, or at the surface 

of a rising water table. Under these conditions both 

oxygen and CO2 would be in solution. The second 

and third titrations, PC-1 and 503-2, were intended to 

simulate conditions that might be attained after 

aeration in Pinal Creek had removed most of the 

dissolved CO2 . The pH of the water samples was held 

at 9.00 in the first few hours of each titration and was

1097
lowered to 8.75 for the rest of the titration. At the 

gaging and sampling station at Inspiration Dam, about 

8 km downstream from well 503, data of Brown 

(1990) and Longsworth and Taylor (1992) show that 

the pH is generally about 8.0 ± 0.2 units. The higher 

experimental pH increases the rate of oxidation, but 

previous experiments have shown the form of oxide is 

not affected significantly.

The oxidation state of manganese in the 

precipitates was determined by a redox titration 

technique (Hem, 1980). The specific forms of 

manganese oxide that were present were determined 

by X-ray and electron diffraction. Mineralogic 

identification of precipitates recovered from Final 

Creek were given earlier by Lind and Hem (1993).

RESULTS AND DISCUSSION

Activities of solutes in experimental solutions 

during redox titration and aging are given in table 2, as 

are values for the oxidation state of precipitated 

manganese. During titration experiment 503-1 the 

measured oxidation number of precipitated manganese 

was consistently below 2.50. The lowest oxidation 

number for manganese oxides produced in our earlier 

work where dissolved HCC>3~ was absent was near 

2.67, the value to be expected for hausmannite 

(M^CXt). The lower values seen in this titration 

indicate the presence of other divalent manganese 

species in the precipitate. The X-ray-difrration data 

indicated the presence of kutnahorite, a mineral with a 

dolomite crystal structure, in pure form having the 

formula CaMn(CC>3)2 . Some magnesium can 

substitute for manganese in natural kutnahorite. This 

mineral was also found in streambed precipitates in 

Final Creek by Lind and Hem (1993). During aging, 

the oxidation number of the precipitated manganese 

increased to 2.65. Hausmannite was in the precipitate 

formed during titration, and the kutnahorite probably 

was converted to oxide during aging.

In titration experiment PC-1, the oxidation 

number of precipitated manganese was 2.66 or greater 

during titration, indicating that carbonates probably 

were absent. The measured value rose to 2.99 during 

aging. The initial precipitate appeared to be 

hausmannite, which altered during aging to the Mn 

oxide, manganite. This alteration is similar to that 

seen in earlier titration experiments in simpler 

solutions (Hem and Lind, 1983). In the first few hours 

of the titration experiment 503-2, the air bubbler was

not operated and a small amount of precipitate in 

which the manganese oxidation number was below 2.5 

was formed. During aeration the oxidation number 

approached 3.0. During aging the manganese 

oxidation number increased, finally reaching 3.62 after 

112 days. In this sample, the X-ray-diffraction data 

indicate the presence of the mixed Ca and Mn 

oxides todorokite ((Mn2+, Ca) Mn4+5 O,, 4H2O) and 

takanelite ((Mn2+, Ca) Mn4+4 O9 3H2O). These 

oxides also were identified by Lind and Hem (1993) in 

black cemented precipitates collected from Final 

Creek at the streamflow gaging station at Inspiration 

Dam. These precipitates had been in place for at least a 

year.

A thermodynamic evaluation of the 

feasibility of various precipitation and alteration 

reactions was made by calculating affinities of reaction 

(A) for the precipitation of oxide and carbonate 

species, in the experimental solutions 503-1, PC-1 and 

503-2. The affinity of reaction is calculated from the 

relation

A = -2.303 RT (log Q - log K),

where R is the gas constant; T is temperature on the 

kelvin scale; Q is the activity quotient calculated by 

entering observed reactant and product activities in the 

mass-law equation for the reaction considered, written 

as producing a precipitate; and K is the equilibrium 

constant for that reaction. Positive values for A indicate the reaction is favored to proceed to the right, as 

written.

Calculated values for A (table 3) for the 

precipitation of kutnahorite are positive during all the 

titrations, as are those for hausmannite and manganite. 

Kinetics appear to favor the oxidation reactions. The 

reaction affinities for the mixed Ca + Mn4+ oxides 

could not be calculated because standard free energies 

of formation for these species are not known.

CONCLUSIONS

Substantial amounts of dissolved undissociated 

CO2 species are present in the contaminated 

groundwater and in the perennial-flow reach of Pinal 

Creek. Most of the CO2 escapes as a gas when the 

water reaches the creek. The most favorable condition 

for carbonate precipitation probably occurs near the 

streambed surface where ground water seeps into the 

creek.

1098
Table 2. Ion activities and manganese oxidation states during titration and aging

[Thermodynamic activities in moles per liter, - = not determined]

Experiment Titration 

and sample time, 

number hours

503- 1-G 48

H

I

PC-l-E 39

F

503-2-H 22

1

J

Aging 

time, 

days

0

3

83

0

132

0

25

112

Log of activity, 

mol/L

Ca*T

-2.36

-2.38

-2.31

-2.31

-

-2.35

-

-

Mn**

-3.90

-4.20

-3.56

-4.24

-6.34

-4.24

-4.45

-4.57

H*

-8.71

-8.34

-7.92

-8.75

-8.07

-9.00

-7.68

-7.26

HC03-

-3.11

-3.09

-

-4.23

-

-4.15

-

-

Mn oxidation 

number

2.44

2.53

2.65

2.69

2.99

2.87

3.21

3.62

Table 3. Thermodynamic feasibility of chemical reactions for precipitation or alteration of solids

[Al = kutnahorite, A_2 = calcite, A* = hausmannite. As = hausmannite -»y manganite; kcal, - = not determined]

Experiment 

and sample 

number

503- 1-G

H

I

PC-l-E

F

503-2-H

I

J

Titration 

time, 

hours

48

39

22

Aging 

time, 

days

0

3

83

0

132

0

25

112

Mn 

Oxidation 

number

2.44

2.53

2.65

2.69

2.99

2.87

3.21

3.62

Affinity of reaction, kcal

Ai

5.65

4.26

-

2.31

-

3.15

-

-

A2

1.90

1.39

-

.49

-

.89

-

-

A4

9.33

7.91

7.68

8.98

4.26

9.66

5.77

4.46

As

-4.04

-2.62

-2.35

-3.68

1.04

-4.36

-.48

.83

As the pH increases downstream, and dissolved 

oxygen becomes more available, the rate of manganese 

oxidation increases. The high concentration 

O_l_ 

of calcium y| ,

in the water probably favors the mixed Ca Mn 

oxide species identified in downstream deposits and in 

aged precipitates from titration experiment 503-2. 

Comparison of experimental results with observed 

manganese precipitation in Final Creek demonstrates 

that a better understanding can be achieved by applying 

tested chemical principles.

REFERENCES

Brown, J.G., 1990, Chemical, geologic and hydrologic data 

for the study of acidic contamination in the Miami 

Wash-Final Creek area, Arizona, water years 1988-89: 

U.S. Geological Survey Open-File Report 90-395, 75 p.

Hem, J.D., 1978, Redox processes at surfaces of manganese 

oxide and their effects on aqueous metal ions: 

Chemical Geology, v. 21, p. 199-218.

 1980, Redox coprecipitation mechanisms of 

manganese oxides, in Kavanaugh, M.C., and Leckie, 

J.O., eds., Particulates in water: Advances in 

Chemistry Series No. 189. Washington, D.C., 

American Chemical Society, p. 45-72.

1981, Rates of manganese oxidation in aqueous

systems: Geochimica Cosmochimica Acta, v. 45, 

p. 1369-1374.

Hem, J.D., and Lind, C.J. 1983, Nonequilibrium models for 

predicting forms of precipitated manganese oxides: 

Geochimica Cosmochimica Acta, v. 47, p. 2037-2046.

Lind, C.J., and Hem, J.D., 1993, Manganese minerals and 

associated fine particulates in the streambed of Final 

Creek, Arizona A mining-related acid drainage 

problem: Applied Geochemistry, v. 8, no. 1, p. 67-80.

Longsworth, S.A., and Taylor, A.M., 1992, Hydrologic data 

from the study of acidic contamination in the Miami 

Wash Final Creek area, Arizona, water years 1990-91: 

U.S. Geological Survey Open-File Report 92-468, 

59 p.

1099

PA6H.*r/oo FOLLOW:-

Experimental Studies of Trace-Metal Partitioning 

in Perennial Reaches of Pinal Creek, Arizona

By Christopher C. Fuller1

Abstract

Development of solute transport models to 

predict the fate of metal contaminants in Pinal 

Creek will require quantification of the reactions 

controlling metal partitioning. Manganese oxides 

are thought to control dissolved trace-metal 

concentrations in the perennial flow of the creek. 

However, these process are not adequately characterized. Laboratory experiments under geochemical conditions relevant to Pinal Creek using fieldderived materials are proposed to evaluate metal 

uptake by coprecipitation with manganese during 

oxidation and precipitation of manganese oxides 

and by adsorption onto absorbent phases present in 

the streambed sediments. The reversibility of the 

metal sorption by these processes also will be 

investigated.

INTRODUCTION

The increase in dissolved manganese (Mn), nickel 

(Ni) and cobalt (Co) in the perennial flow of Pinal 

Creek (fig. I) from 1985 to present (Eychaner, 1991 a) 

has indicated the need to develop predictive models 

that describe the transport of these metals. Using simple conservative and reactive mixing models, Eychaner 

(I991b) suggested that the downstream decrease in 

dissolved Mn results largely from Mn oxidation and 

precipitation of Mn oxides. Downstream decreases in 

surface-water concentrations of Ni, Co, and other trace 

metals in March 1990, also were observed (Faires and 

Eychaner, 1991). Because of their abundance, 

Mn-oxide phases are believed to be largely responsible 

for trace-metal partitioning by sediments (Lind and 

Anderson, 1992). However, trace-metal partitioning

'U.S. Geological Survey, Mcnlo Park, Calif.

processes in Pinal Creek have not been sufficiently 

quantified for incorporation into reactive transport 

models. A more detailed description of the study site is 

given by Brown and Harvey (1996).

Development of solute transport models to 

adequately predict the fate of metals in Pinal Creek will 

require a quantitative description of trace-metal partitioning processes under geochemical conditions relevant to the perennial-flow regime of the stream. This 

will be of particular importance for simulating the 

potential inflow of an acidic ground-water plume, 

which contains high concentrations of dissolved copper (Cu), iron (Fe) and other metals (Eychaner, 1991 a). 

The objective of this work is to investigate trace-metal 

partitioning processes in Pinal Creek by using field 

samples in laboratory experiments. The results will be 

used to provide the appropriate trace-metal partitioning 

and rate coefficients for incorporation into reactive 

transport models in the perennial reaches of Pinal 

Creek (Harvey and others, 1996).

This paper describes laboratory experiments being 

conducted under geochemical conditions similar to 

those in the perennial reach of Pinal Creek. The 

experiments are designed to study metal adsorption and 

coprecipitation by Mn oxide and carbonate phases that 

are currently forming in the perennial reach of the 

stream. Downstream trends in concentrations of 

dissolved Mn, Ni, and Co from November 1992 are 

presented first to describe the type of initial experiments needed to model reactive transport under present 

conditions in the stream.

METHODS OF FIELD SAMPLE 

COLLECTION, ANALYSIS, AND RESULTS

Surface-water and ground-water samples were 

collected in November, 1992, under base-flow conditions along a 6-km reach of Pinal Creek from the head 

of perennial flow to just above Inspiration Dam (fig. 1).

1101
110°55' if 53' 110°52'

33°34'

33'

33°32' -

Inspiration A (09498400) 

Dam

500 

(09498400)1

EXPLANATION 

INTERMITTENT REACH 

PERENNIAL REACH

DATA SITES-Number is 

site identifier

Observation well group 

Streamflow-gaging station 

Sample site 

Channelized inflow 

Bank seep

1/2 1 MILE

See inset : 

500^

0.5 1 KILOMETER

Figure 1. Locations of surface-water sites sampled November 1992.

The methods of sample collection are described in 

Harvey and Fuller (1996). Water samples for metal 

analysis were filtered in series through 0.45 and 0.001 

jam (micron) pore-size filters, acidified, and analyzed 

by inductively coupled plasma emission spectrometry 

(ICP-AES).

Over this study reach, stream-water pH increased 

from 5.6 at the head of flow to 7.9 above Inspiration 

Dam (fig. 2). The downstream increase in pH results 

from carbon dioxide outgassing (Eychaner, 199la). 

Over the first 0.5 km downstream from the head of 

flow, the concentration of dissolved Mn increased from 

1.2 to 1.8 mM (millimolar) and was attributed to 

increased concentration of Mn in ground-water inflows 

(Harvey and others, 1996). The concentration of dissolved Mn decreased over the next 5 km by about 45 

percent to 1 mM (fig. 2). Although the relative decrease 

was less than observed by Eychaner (199 Ib) in March 

1990, it probably also is due largely to the of formation 

Mn oxide and carbonate. The Mn concentration in the 

colloidal size fraction at all sampling sites was less than

the uncertainty in the difference between the 0.45- and 

0.001-urn filtrates. The low Mn-colloid concentration 

suggests that the oxidation of dissolved Mn2+ occurs as 

a surface reaction on streambed sediments (Harvey and 

Fuller, 1996).

A similar downstream trend to manganese was 

observed for Ni (fig. 2). After an increase over the first 

0.5 km of flow, Ni concentration decreased from 18 jiM 

(micromolar) to 10 uM at 5.8 km; most of the decrease 

occurred within the first 3 km of flow. The colloidal Ni 

concentration also was not significant. The molar ratio 

of dissolved Ni to dissolved Mn (Ni/Mn ratio) 

decreased about 50 percent indicating a preferential 

loss of Ni relative to Mn within the first 1.5 km of flow 

(fig. 2) assuming a common source of dissolved Ni and 

Mn. Little subsequent change in Ni/Mn ratio is 

observed farther downstream. The preferential loss of 

Ni over Mn occurred within the reach where the rate of 

pH increase was greatest. The decrease in Ni concentration relative to Mn concentration may be due to 

sorption of Ni onto existing oxide phases in the

1102
8

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DISTANCE FROM INSPIRATION 

DAM, IN KILOMETERS

DISTANCE FROM INSPIRATION 

DAM, IN KILOMETERS

Figure 2. Dissolved constituents in Final Creek, November 1992. (Distances are in kilometers from the head of perennial flow 

located 6.2 kilometers upstream from Inspiration Dam in November 1992. Dissolved manganese is units of millimolar; nickel and 

cobalt in micromolar. Ratios of dissolved constituents are calculated in units of molar concentration.)

1103
streambed and to Ni/Mn coprecipitation during Mn 

oxidation and precipitation (Hem and others, 1989). 

The relatively constant Ni/Mn ratio farther downstream suggests that the subsequent decrease in Ni 

concentration occurs in conjunction with the decrease 

in Mn concentration. Although the observed downstream decrease in metal concentration could, in part, 

be due to dilution, the modeling results of Eychaner 

(1991b) suggest that the changes in metal concentration were largely due to reaction processes.

The Co concentration increased from 1 to 5 uM 

within a 0.5-km reach from the head of flow (fig. 2) and 

subsequently decreased over the next 1 km by about a 

factor of two. Below this point, significant changes in 

Co concentration are not detectable because of greater 

uncertainty of low Co concentrations that are near the 

detection limit of ICP-AES. The Co to Mn molar ratio 

(Co/Mn ratio) decreased downstream to 1.5 km indicating a preferential loss of Co relative to Mn. 

Although the extraction of streambed materials suggest 

Ni and Co are primarily associated with Mn-oxide 

phases (Lind and Anderson, 1992), the partitioning 

processes controlling concentrations of dissolved Ni 

and Co have not been well characterized. A series of 

laboratory experiments is proposed to evaluate sorption processes that may control Ni and Co concentrations in Final Creek.

LABORATORY SORPTION EXPERIMENTS

Trace-metal concentrations in natural waters are 

typically less than saturation with respect to pure oxide 

and hydroxide phases and are commonly thought to be 

controlled by adsorption to mineral surfaces (Davis and 

Kent, 1990). The adsorption of metals onto Mn oxides 

increases with increasing pH. The relative binding 

strength to Mn oxides decreases in the following order: 

lead>Cu>Co>zinc>Ni (McKenzie, 1980). Metal 

adsorption onto synthetic 5-MnO> has been modeled by 

a surface complexation approach (Catts and Langmuir, 

1986). In addition to adsorption onto Mn oxide 

surfaces, coprecipitation of trace metals (for example 

Ni, Cu and Co) can occur when dissolved metals are 

present during the oxidation and precipitation of the 

oxide phase; coprecipitation can result in the incorporation of trace metals into the Mn oxide structure (Hem 

and others, 1989). Uptake during coprecipitation can 

greatly exceed adsorption onto existing surfaces 

(Fuller and others, 1993), and may reduce the potential 

for desorption.

Because Mn forms a variety of oxide and carbonate 

phase in Final Creek (Lind and Hem, 1993), fieldderived solid phases will be used to study metal sorption instead of synthetic Mn-oxide phases. Trace-metal 

coprecipitation with Mn during oxidation and precipitation will be studied using surface and ground water 

collected at the field site. These experiments will be 

similar to precipitation experiments used to determine 

the mineralogy of Mn precipitates (Hem and Lind, this 

volume). The pH will be increased to facilitate Mn oxidation by controlled reaeration to lower the partial 

pressure of carbon dioxide. Dissolved metal concentrations will be monitored as a function of time and pH to 

determine the amount of Ni and Co coprecipitated with 

the Mn phases.

The adsorption of trace metals by grain-size and 

mineralogic fractions of streambed sediments will be 

studied in batch as a function of metal concentration 

and pH using an artificial stream water (ASW). In contrast to the coprecipitation experiments, Mn will not 

precipitate from the ASW in these systems. The 

relative importance of trace-metal partitioning by 

adsorption and coprecipitation can then be compared.

The reversibility of metal sorption and the extent of 

desorption caused by changes in stream-water chemistry will be investigated. These experiments will include 

shifting the sorption equilibrium to favor desorption by 

decreasing either pH or dissolved metal concentration. 

Isotopic exchange will also be used to determine the 

fraction of adsorbed metals that can desorb (Davis and 

others, 1987). These experiments will be performed on 

streambed material containing ambient sorbed-metal 

concentrations as well as on Mn precipitates and sediments from the sorption experiments outlined above. 

The reversibility studies may help identify the important sorption processes for inclusion in solute-transport 

models for Final Creek.

SUMMARY

A series of laboratory trace-metal (Ni, Co, Cu) 

adsorption, coprecipitation, and reversibility experiments are being conducted using field-derived materials under geochemical conditions similar to those in the 

perennial flow of Final Creek. The experimental results 

will help identify the importance of various partitioning processes between solution and mineral phases and 

provide sorption-desorption coefficients for use in 

reactive transport modeling of trace metals in the 

stream. For example, under present conditions in Final

1104
Creek, Ni may be removed during transport by adsorption onto existing phases in streambed sediments as 

well as by coprecipitation with Mn-oxide phases. The 

magnitude of uptake and desorption potential may 

differ significantly for these two sorption processes.

REFERENCES

Brown, J.G., and Harvey, J.W., 1996, Hydrologic and 

geochemical factors affecting metal contaminant 

transport in Final Creek Basin near Globe, Arizona, in 

Morganwalp, D.W., and Aronson, D.A., eds., U.S. 

Geological Survey Toxic Substances Hydrology 

Program Proceedings of the technical meeting, Colorado Springs, Colo., September 20-24, 1993: U.S. Geological Survey Water-Resources Investigations Report 

94-40-15.

Carts, J.G., and Langmuir, D., 1986, Adsorption of Cu, Pb 

and Zn by 5MnO2 Applicability of the site bindingsurface complexation model: Applied Geochemistry, 

v.l, p. 255-264.

Davis, J.A., Fuller, C.C., and Cook, A.D., 1987, A model 

for trace metal sorption by calcite Adsorption of Cd 

and subsequent solid solution formation: Geochimica et 

Cosmochimica Acta, v. 51, p. 1477-1490.

Davis, J.A., and Kent, D.B., 1990, Surface complexation 

modeling in aqueous geochemistry, in Hochella, M.F., 

and White A.F., eds., Mineral-Water Interface 

Geochemistry: Washington, DC., Mineralogical 

Society of America, p. 177-260.

Eychaner, J.H., 1991 a, The Globe, Arizona, research sitecontaminants related to copper mining in a hydrologically integrated environment, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: USGS Water Resources Investigation Report 91- 

4034, 1991, p. 439-447.

Eychaner, J.H., 1991b, Solute transport in perennial streamflow at Pinal Creek, arizona, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91 -4034, 1991, p. 481 -485.

Faires, L.M., and Eychaner, J.H., 1991, Trace-element 

trends at Pinal Creek, Arizona, in Mallard, G.E., and 

Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Monterey, California, March 11-15, 

1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, 1991, p. 461-465.

Fuller, C.C., Davis, J.A., Waychunas, G.A., 1993, Surface 

chemistry of ferrihydritc, II Kinetics of arsenate 

adsorption and coprecipitation, Geochimica et Cosmochimica Acta, v. 57, p. 2271-2282.

Harvey, J.W., and Fuller, C.C., 1996, Association of metal 

contaminants with colloidal and suspended particulate 

material in shallow ground water and surface water at 

Pinal Creek, Arizona, in Morganwalp, D.W., and Aronson, D.A., eds., U.S. Geological Survey Toxic 

Substances Hydrology Program Proceedings of the 

technical meeting, Colorado Springs, Colo., September 

20-24, 1993: U.S. Geological Survey Water-Resources 

Investigations Report 94-4015.

Harvey, J.W., Fuller, C.C., and Wagner, B.J., 1996, Interactions between shallow groundwater and surface water 

that affect metal contaminant transport in Pinal Creek 

basin, Arizon, in Morganwalp, D.W., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colo., September 20-24, 

1993: U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Hem, J.D., Lind, C.J., and Roberson, C.E., 1989, Coprecipitation and redox reactions of manganese oxides with 

copper and nickel: Geochimica et Cosmochimica Acta, 

v. 53, p. 2811-2822.

Hem, J.D., and Lind, C.J., 1996, Chemical processes in 

manganese oxide and carbonate precipitation in Pinal 

Creek, Arizona, in Morganwalp, D.W., and Aronson, 

D.A., eds., U.S. Geological Survey Toxic Substances 

Hydrology Program Proceedings of the technical 

meeting, Colorado Springs, Colo., September 20-24, 

1993: U.S. Geological Survey Water-Resources Investigations Report 94-4015.

Lind, C.J., and Anderson, L.D., 1992, trace-metal scavenging by precipitating Mn and Fe oxides, in Kharaka, 

Y.K., and Maest, A.S., eds., Proceedings of the 7th 

International Symposium on Water Rock Interaction, 

Balkema, Rotterdam: p. 397-402.

Lind, C.J., and Hem, J.D., 1993, Manganese minerals and 

associated fine particulates in the streambed of Pinal 

Creek, Arizona A mining-related acid drainage 

problem: Applied Geochemistry, v. 8, no. 1, p. 67-80.

McKenzie, R.M., 1980, The adsorption of lead and other 

heavy metals on oxides of manganese and iron: Australian Journal of Soil Research, v. 18, p. 61-73.

1105
1106
Relations Among Algal Species and Chemical Gradients 

in Final Creek, near Globe, Arizona

By Patrice H. Spindler1 and Milton R. Sommerfeld2

Abstract

The relations among algal species and 

water chemistry in Final Creek, near Globe, 

Arizona, were investigated. Periphyton, water, 

and sediment samples were collected from seven 

sites during 1991-92. Fourteen species of algae 

and 16 chemical properties were selected for 

correlation and cluster analyses. Three major 

algal associations in Final Creek were identified 

by cluster analysis: an acid- and metal-tolerant 

association at upstream sites, an alkalophilous 

association at downstream sites, and an 

association of ubiquitously distributed 

opportunistic species. Indicator species of the 

acid- and metal-rich environment at upstream 

sites were Microspom stagnorum, Surirella 

minuta and Ulothrix variabilis. The exclusion of 

Nitzschia linearis and Vaucheria geminata from 

upstream sites and their abundance at downstream 

sites suggest that these species are acid and metal 

intolerant.

INTRODUCTION

Algal communities are valuable in water-quality 

assessment because they are very responsive to 

physicochemical changes in habitat (Stevenson and

Represents a portion of a thesis by P.H.S. submitted in partial fulfillment of the requirements of a Master of Natural Science 

in Botany, Arizona State University. Current address: Arizona 

Department of Environmental Quality, Office of Water Quality, 

Biocriteria Program, 3033 N. Central Ave., Phoenix, Arizona 

85012.

2 Department of Botany, Arizona State University, Tempe, 

Arizona 85287.

Lowe, 1986). The taxonomic composition of the algal 

community provides an alternative to traditional 

chemical methods for assessing the effect of acid mine 

drainage on a stream ecosystem.

Specific taxa and species richness and diversity 

indices have been used to characterize streams 

affected by acidic mine drainage. English streams 

contaminated by metals are characterized by a 

Microspora community whereas streams with low 

concentrations of metals contain a Zynematales 

community of Spirogyra and Mougeotia species 

(Foster, 1982). Increasing concentrations of zinc 

shifted the algal community from diatoms to green or 

blue-green algae in an outdoor stream mesocosm in 

Virginia (Center and others, 1987). Low species 

richness (Stokes, 1983) and diversity (Hargreaves and 

others, 1975) have also been used as indicators of 

impaired water quality.

Final Creek receives drainage from an upstream 

copper mining district near Globe-Miami, Arizona. 

Significant increases in pH and significant decreases in 

metal concentrations, such as manganese and nickel 

occur with distance downstream (Spindler, 1993). The 

purpose of this paper is to characterize the algal 

community and its relation to the water quality of 

Final Creek. This research was partly funded by a 

grant from the U.S. Geological Survey (order 

#230102).

METHODS

Samples of algae, water and sediment were 

collected at seven sampling sites within a 6-km reach 

of Final Creek during May, July, and October 1991 

and January 1992. The sampling sites were located at 

distances of 0.0, 1.0, 1.5,2.0, 2.5,4.5, and 5.5 

kilometers from the headwater (fig. 1). These sites 

are hereafter referred to by sites 1-7, respectively.

1107
Gag-Ing Station

09498400

EXPLANATION

 Intermittent flow

Perennial flow 

 Sampling site

01235 KILKTHS V-.

\\\\l\\ x \\

3T2P 

Base fm U.S. Forest Service, 

Tonto National Forest, 1985

Figure 1. Area of study.

1108
Site 2 was located at Setka Ranch (station 09498380) 

and site 7 was located at Inspiration Dam (station 

094984000).

Three replicate algae samples were collected at 

each site. Each sample consisted of a composite of six 

individual samples collected along a randomly 

determined transect. A 5-cm plastic cylinder 

emplaced to a depth of 2 cm in the sediment was used 

as a benthic sampler to collect consistent-volume, 

39 cm3 periphyton-containing sediment samples. 

Algal material was separated from sediments by the 

sedimentation method (American Public Health 

Association, 1985). Diatom and nondiatom species 

were prepared and counted separately, because 

nondiatoms degrade rapidly after sample collection 

whereas diatoms can be permanently mounted. 

Nondiatom species abundances were enumerated 

from unpreserved samples by use of a SedgewickRafter chamber and an inverted microscope equipped 

with an ocular grid. Diatom species were enumerated 

from permanent bum mounts of sample material 

(Stevenson and Lowe, 1986) by use of a lighttransmission microscope equipped with an ocular 

micrometer. Taxonomic identifications were made for 

1,000 cells in each sample-500 diatoms and 500 

nondiatoms.

Four abundance categories (dominant, 

common, infrequent, rare) were used to characterize 

the frequency of occurrence of each species for each 

sample event. Only the dominant species were 

selected for cluster analysis and correlation analysis 

to identify the major associations of algae and their 

relations to water chemistry.

Algal associations were identified in a 

dendrogram obtained by cluster analysis using 

SYSTAT software (Wilkinson, 1988). Cluster analysis 

was based on Ward's minimum variance method, 

which joins successive groups using their squared 

Euclidean distances. Correlations between each of 14 

algal species and 16 chemical properties were 

analyzed by use of the Number Cruncher Statistical 

System (Hintze, 1985).

Three replicate, filtered water samples and four 

replicate, composited sediment samples were 

collected from each site. Water samples were 

analyzed for alkalinity, ortho-phosphate, and nitrate 

following standard analytical methods (U.S. 

Environmental Protection Agency, 1983). Water and 

sediment samples were analyzed for the metals 

cobalt, copper, iron, manganese, nickel and zinc

following procedures outlined in Standard Methods 

(American Public Health Association, 1985). Only 

ranges of values for each property have been 

presented in this paper. Additional detail about 

chemical methods and results can be found in 

Spindler(1993).

RESULTS

Three major algal species associations in Final 

Creek were identified by cluster analysis (fig. 2). The 

first algal association consisted of Caloneis hultenii, 

Microspora quadrata, M. stagnorum, Nitzschia 

ignorata, Surirella minuta, Tribonema bombycinum, 

Ulothrix subtilissima, and U. variabilis. The second 

algal association consisted of Cydotella 

meneghiniana, Nitzschia linearis, and Vaucheria 

geminata. The third algal association consisted of 

Nitzschia palea, Surirella angustata, and Tribonema 

minus.

The first algal association displayed positive 

correlations with metals and sulfate concentrations 

and negative correlations with pH and alkalinity 

(table 1). Several of the species in this association 

have been reported in the literature as tolerant of 

acidic mine drainage. Foster (1982) indicated that a 

grouping of Microspora species could be termed an 

indicator community for heavy-metal contamination. 

Whitton (1970) concluded that Microspora and 

Ulothrix species are resistant to copper, lead, and 

zinc. Ulothrix subtilissima was found upstream and 

far downstream of acidic mine drainage, whereas 

Ulothrix variabilis was ubiquitous in an Arizona 

stream receiving acidic mine drainage (Lampkin and 

Sommerfeld, 1982). Palmer (1977) placed Surirella 

ovata, synonymous with S. minuta, on a list of the 80 

most tolerant species, although the genus had not 

been previously listed as an indicator species for 

acidic mine drainage. Surirella minuta tolerates a 

wide range in pH in Pinal Creek.This first algal 

association was found in the greatest abundance at 

upstream sites in Pinal Creek, where pH was less than 

7.0 units and metal concentrations were highest 

(Spindler, 1993).

The second algal association composed of 

Cydotella meneghiniana, Nitzschia linearis, and 

Vaucheria geminata species displayed positive 

correlations with pH and alkalinity and negative 

correlations with concentrations of metals and sulfate.

1109
0.000

Euclidean Distance 

Vaucheria geminata 

Nitzschia Tinear is 

Cyclotella meneghiniana 

Ulothrix subtilissima 

Ulothrix variabilis 

Tribonema bombycinum 

Surirella minuta 

Nitzschia ignorata 

Microspora quadrata 

Caloneis hultenii 

Microspora stagnorum 

Nitzschia palea 

Tribonema minus 

SurirelTa angustata

Associatio

Associatio

Associatio

Figure 2. Dendrogram of algal associations in Final Creek. Mean abundances for dominant species distributed among seven 

minimum variance method. Species were grouped using their squared Euclidean distances.
Table 1. Coefficients of correlation (r) between algal species abundances and selected chemical constituents and pH, Final Cr

[Species abundances were natural log transformed means from four seasonal sampling events for each of seven sites. Chemical properties were means of four 

sites; only r values significant at p < 0.05 were listed, except where noted; Alk, alkalinity]

Species

Concentration ranges (mg/L) and correlation coeff

Alk 

45-99

PH 

5.9-7.8

SO4 

2000-2300

Co 

0.05-0.13

Cu 

0.03-0.07

Fe 

0.07-0.23

Association 1

Caloneis hultenii Peters, var. hultenii -0.71 0.78

Microspora quadrata Hazen

Microspora stagnorwn (Kuetz.) Lagerheim -.89 -0.78 .97

Nitzschia ignorata Krasske

SurirellaminutaBreb. -.92 -.96 .88

Tribonema bombycinum (C.A.Ag)Derbes,Solier -.81 -.77 .76

Ulothrix subtilissima Rabenhorst -.78

Ulothrix variabilis Kuetz. -.97 -.96 .80

0.83

0.82 

.87 

.79 

.86

0.77 

.84 

.80 

.82

Association 2

Cyclotella meneghiniana Kuetz.

Nitzschia linearis W. Smith

Vaucheria geminata (Vauch.) De Candolle

Association 3

Nitzschia palea (Kuetz.) W. Smith 

Surirella angustata Kuetz. 

Tribonema minus (Wille) Hazen

.77 

.98 

.89

-.85

.89

.77 

.88 

.77

-.78 

Significant at p< 0.100
Nitzschia linearis and Vaucheria geminata were listed 

by Palmer (1977) as clean-water species. Vaucheria 

geminata had the most limited distribution of all the 

dominant species; it was found only at downstream 

sites 4-7. Lampkin and Sommerfeld (1982) found that 

Vaucheria was not present in the reach of Lynx Creek 

that receives acidic mine drainage but was present in a 

tributary to Lynx Creek. These data and observations 

suggest that V geminata may be an indicator of the 

recovery zone in Final Creek, where pH increases to 

the range of 7.0 to 8.0 and alkalinity concentration 

increases to greater than 90 mg/L. Rushforth, and 

others (1981) showed that Cyclotella meneghiniana 

showed a preference for high concentrations of manganese and selenium and was also positively correlated with high concentrations of aluminum, copper, 

iron, and lead. Under these conditions, one might 

expect to find C meneghiniana near the beginning of 

perennial flow, where concentrations of these metals 

were highest. The fact that C. meneghiniana was 

found in part of Final Creek where acidic mine drainage was minimal suggests that other factors may play 

a role in determining its niche. This association was 

most abundant at downstream sites in Final Creek, 

where pH was higher and metal concentrations were 

lower relative to upstream sites.

The third algal group, which consists of the 

species Nitzschia palea, Surirella angustata, and 

Tribonema minus, was distributed throughout the 

study reach and was characterized by a lack of strong 

correlations with high or low pH or metal 

concentrations. Nitzschia palea has been characterized 

as tolerant of low pH (Hargreaves and others, 1975), 

as an indicator of elevated copper and chromium 

concentrations (Palmer, 1959), and as an indicator of 

elevated aluminum, arsenic, iron, manganese, and 

selenium concentrations (Rushforth and others, 1981). 

Surirella and Tribonema are among the 60 most 

pollution tolerant genera listed by Palmer (1977). This 

group of algae did not form a specific association, 

defined by increased density under a particular set of 

chemical conditions. Rather, these species shared the 

common attribute of being tolerant of a wide range of 

pH and metal concentrations and of being ubiquitously 

distributed in Final Creek. Foster (1982) concluded 

that the most abundant algae at sites where metal 

concentrations were high also were found at sites 

where metals were present in a broad range of 

concentrations. Such species have adapted to 

chemically disturbed environments but are not strictly

indicative of a particular metal. According to results 

of the present study, these three species have 

apparently adapted to a chemically disturbed 

environment having a pH range of about 2 units and a 

range in concentrations of metals.

The shift in chemical characteristics within a 6- 

km perennial reach of Final Creek is reflected by a 

corresponding shift in the algal composition. If the 

water that continues to move downstream in Final 

Creek continues to be acidic (Eychaner, 1989), a 

corresponding shift in the algal associations could be 

expected, such as the dislocation of readily visible 

Vaucheria mats from sites 4-7. The disappearance of 

this genus may provide an early indication of changing 

stream conditions associated with increasing acidity 

and metal concentrations.

SUMMARY

The algae of Final Creek are distributed in a 

pattern that reflects the response of the algae to 

changing chemical conditions. Three major algal 

associations in Final Creek were identified in this 

study: (1) an acid- and metal-tolerant association at 

upstream sites, (2) an alkalophilous association at 

downstream sites, and (3) an association of 

ubiquitously distributed opportunistic species. 

Indicator species of the acid- and metal-rich 

environment at upstream sites were Microspora 

stagnorum, Surirella minuta, and Ulothrix variabilis. 

The absence of Nitzschia linearis and Vaucheria 

geminata from upstream sites and their abundance at 

downstream sites suggests that these species are acid 

and metal intolerant. A follow-up study in five to 10 

years is recommended to determine whether algae 

distribution will continue to change in relation to 

stream conditions.

REFERENCES

American Public Health Association, 1985, Standard

methods for the examination of water and wastewater 

(16th ed.): Washington, DC., American Public Health 

Association, 1,268 p.

Eychaner, J.H., 1989, Movement of inorganic contaminants 

in acidic water near Globe, Arizona, in Mallard, G.E. 

and Ragone, E.E., eds., Proceedings of the 2nd Annual 

U.S. Geological Survey Toxic Substances Hydrology 

Technical Meeting, Phoenix, Arizona, September 26- 

30, 1988: U.S. Geological Survey Water-Resources 

Investigations Report 88-4220, p. 567-575.

1112
Foster, P.L., 1982, Species associations and metal contents 

of algae from rivers polluted by heavy metals: 

Freshwater Biology, v. 12, p. 17-39.

Center, R.B., Cherry, D.S., Smith, E.P., and Cairns, J.Jr., 

1987, Algal-periphyton population and community 

changes from zinc stress in stream mesocosms: 

Hydrobiologia, v. 153, p. 261-275.

Hargreaves, J.W., Lloyd, E.J.H., and Whitton, B.A., 1975, 

Chemistry and vegetation of highly acidic streams: 

Freshwater Biology, v. 5, p. 563-576.

Hintze, J.L., 1985, Number cruncher statistical system 

(NCSS), ver. 4.2: Kaysville, Utah, 235 p.

Lampkin, A.J. and Sommerfeld, M.R., 1982, Algal

distribution in a small intermittent stream receiving 

acid mine drainage: Journal of Phycology, v. 18, 

p. 196-199.

Palmer, C.M., 1959, Algae in water supplies. U.S. 

Department of Health, Education, and Welfare, 

Cincinnati, Ohio, 88 p.

Palmer, C.M., 1977, Algae and water pollution, an

illustrated manual on the identification, significance 

and control of algae in water supplies and in polluted 

water: Cincinnati, Ohio, U.S. Environmental 

Protection Agency-600/9-77-036, 124 p.

Rushforth, S.R., Brotherson, J.D., Fungladda, J., and

Evenson, W.E., 1981, The effects of dissolved heavy 

metals on attached diatoms in the Uintah Basin of 

Utah, USA: Hydrobiologia, v. 83, p. 313-323.

Spindler, P.H., 1993, Algal distribution and chemical 

gradients in Final Creek, Gila County, Arizona: 

Masters Thesis, Arizona State University.

Stevenson, R. J. and Lowe, R. L., 1986, Sampling and 

interpretation of algal patterns for water quality 

assessments, in Isom, E.G., ed., Rationale for sampling 

and interpretation of ecological data in assessment of 

freshwater ecosystems: American Society for Testing 

and Materials, Special Technical Publication 894, 

Philadelphia, Pa., p. 118-149.

Stokes, P.M., 1983, Responses of freshwater algae to 

metals, in Round and Chapman, eds., Progress in 

Phycological Research: New York, Elsevier Science 

Publishers Co., v. 2, p. 87-112.

U.S. Environmental Protection Agency, 1983, Methods for 

chemical analysis of water and wastes: EPA 600/4-79- 

020, Cincinnati, Ohio, 518 p.

Whitton, B.A., 1970, Toxicity of heavy metals to freshwater 

algae: a review: Phykos, v. 9, p. 117-125.

Wilkinson, L., 1988, SYSTAT The system for statistics: 

SYSTAT, Inc.

1113