EPA/600/R-94/162

September 1994

Symposium on Natural Attenuation of Ground Water

U.S. Environmental Protection Agency

Office of Research and Development

Washington, DC 20460

Printed on paper that contains at least 250 percent recycled fiber.
Notice

The information in this document has been reviewed in accordance with the EPA’s peer and

administrative review policies and approved for presentation and publication. Mention of trade

names or commercial products does not constitute endorsement or recommendation for use.

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Contents

Page

Executive Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

State and Federal Issues Impacting Natural Attenuation . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

State and Federal Issues Impacting Natural Attenuation: Meeting Overview . . . . . . . . . . . . . . . . . 6

Panel Discussion on Regulatory Issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

Symposium on Natural Attenuation of Ground Water: Summaries and Abstracts . . . . . . . . . . . . . 13

Regulatory Perspective on Natural Attenuation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Department of Defense Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Jim Owendoff

U.S. EPA Regulatory Perspective: Natural Attenuation as a Remedial Option for . . . . . . . . . . . 14

Contaminated Ground Water

Guy A. Tomassoni

State Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

John Shauver

Overview of Investigations Related to Natural Attenuation . . . . . . . . . . . . . . . . . . . . . . . . . . 16

USGS Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Gail E. Mallard

U.S. EPA Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Fran V. Kremer

Overview of Process and Procedures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

Site Characterization: What Should We Measure, Where (When?), and How? . . . . . . . . . . . . . 20

Michael J. Barcelona

Processes Controlling the Distribution of Oil, Air, and Water . . . . . . . . . . . . . . . . . . . . . . 26

John L. Wilson

New Tools To Locate and Characterize Oil Spills in Aquifers . . . . . . . . . . . . . . . . . . . . . . 34

Bruce J. Nielsen

Microbiological and Geochemical Degradation Processes . . . . . . . . . . . . . . . . . . . . . . . 40

E. Michael Godsy

Field and Laboratory Results: Getting the Whole Picture . . . . . . . . . . . . . . . . . . . . . . . . 44

Mary Jo Baedecker

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Case Studies: Public Service’s Site . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

In Situ Bioremediation of the Seventh Avenue Site in Denver: The Remediation of . . . . . . . . . . 48

Soils and Ground Waters

Christopher Nelson

The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through In Situ . . . . . . . . 50

Bioremediation: The Role of Mathematical Models

Tissa H. Illangasekare, David C. Szlag, and John T. Wilson

The Importance of Knowledge About Intrinsic Bioremediation for Cost-Effective . . . . . . . . . . . . 55

Site Closure: The Client’s Perspective

Harry E. Moseley

A Regulators’ Perspective of a Bioremediation Site . . . . . . . . . . . . . . . . . . . . . . . . . . . 57

Lisa C. Weers

The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through In Situ . . . . . . . . 58

Bioremediation: The Regulator’s Perspective

Mark E. Walker and Lisa C. Weers

Case Studies: Sites Where Natural Attenuation Has Been Documented . . . . . . . . . . . . . . . . . . . 59

Intrinsic Bioremediation of JP-4 Jet Fuel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

John T. Wilson, Fredrick M. Pfeffer, James W. Weaver, Don H. Kampbell, 

Todd H. Wiedemeier, Jerry E. Hansen, and Ross N. Miller

A Natural Gradient Tracer Experiment in a Heterogeneous Aquifer With Measured In Situ . . . . . . 68

Biodegradation Rates: A Case for Natural Attenuation

Thomas B. Stauffer, Christopher P. Antworth, J. Mark Boggs, and William G. MacIntyre

Traverse City: Distribution of the Avgas Spill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75

David W. Ostendorf

Traverse City: Geochemistry and Intrinsic Bioremediation of BTX Compounds . . . . . . . . . . . . . 80

Barbara H. Wilson, John T. Wilson, Don H. Kampbell, Bert E. Bledsoe, and John M. Armstrong

Mathematical Modeling of Intrinsic Bioremediation at Field Sites . . . . . . . . . . . . . . . . . . . . 85

Hanadi S. Rifai

Biogeochemical Processes in an Aquifer Contaminated by Crude Oil: An Overview of . . . . . . . . 89

Studies at the Bemidji, Minnesota, Research Site

Robert P. Eganhouse, Mary Jo Baedecker, and Isabelle M. Cozzarelli

Simulation of Flow and Transport Processes at the Bemidji, Minnesota, Crude-Oil Spill Site . . . . . 95

Hedeff I. Essaid

Contents (Continued)

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Natural Attenuation of Trichloroethylene and Similar Compounds: Case Studies . . . . . . . . . . . . . . 103

An Overview of Anaerobic Transformation of Chlorinated Solvents . . . . . . . . . . . . . . . . . . 104

Perry L. McCarty

Contamination of Ground Water With Trichloroethylene at the Building 24 Site at . . . . . . . . . . 109

Picatinny Arsenal, New Jersey

Mary Martin and Thomas E. Imbrogiotta

Intrinsic Bioremediation of TCE in Ground Water at an NPL Site in St. Joseph, Michigan . . . . . . 116

John T. Wilson, James W. Weaver, and Don H. Kampbell

Posters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

Technical Protocol for Implementing the Intrinsic Remediation with Long-Term Monitoring . . . . . . 122

Option for Natural Attenuation of Fuel Hydrocarbon Contamination in Ground Water

Todd H. Wiedemeier

Wisconsin’s Guidance on Naturally Occurring Biodegradation as a Remedial Action Option . . . . . 127

Michael J. Barden

Assessing the Efficiency of Intrinsic Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . 128

Francis H. Chapelle

A Practical Approach To Evaluating Natural Attenuation of Contaminants in Ground Water . . . . . 129

P.M. McAllister and C.Y. Chiang

The Use of Low Level Activities To Assist Intrinsic Bioremediation . . . . . . . . . . . . . . . . . . 130

Robert D. Norris, Jeffrey C. Dey, and Daniel P. Shine

Natural Attenuation of Jet Fuel in Ground Water . . . . . . . . . . . . . . . . . . . . . . . . . . . 132

G. Doyle, D. Graves, and K. Brown

Evaluation of Intrinsic Bioremediation at an Underground Storage Tank Site in Northern Utah . . . 133

R. Ryan Dupont, Darwin L. Sorensen, and Marion Kemblowski

Case Studies of Field Sites To Demonstrate Natural Attenuation of BTEX Compounds in . . . . . . 135

Ground Water

C.Y. Chiang and P.M. McAllister

Demonstrating Intrinsic Bioremediation of BTEX at a Natural Gas Plant . . . . . . . . . . . . . . . 136

Keith Piontek, Tom Sale, Jake Gallegos, Steve de Albuquerque, and John Cruze

Demonstrating the Feasibility of Intrinsic Bioremediation at a Former Manufactured Gas Plant . . . 137

Ian D. MacFarlane, Edward J. Bouwer, and Patricia J.S. Colberg

Natural and Enhanced Bioremediation of Aromatic Hydrocarbons at . . . . . . . . . . . . . . . . . 139

Seal Beach, California: Laboratory and Field Investigations

Harold A. Ball, Gary D. Hopkins, Eva Orwin, and Martin Reinhard

The Complete Dechlorination of Trichloroethene to Ethene Under Natural Conditions in a . . . . . 142

Shallow Bedrock Aquifer Located in New York State

David Major, Evan Cox, Elizabeth Edwards, and Paul W. Hare

Contents (Continued)

Page

v
Appendix A: Regulatory Meeting Participants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

Contents (Continued)

Page

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Executive Summary

Introduction

The U.S. Environmental Protection Agency’s (EPA’s)

Biosystems Technology Development Program and the

U.S. Geological Survey (USGS), with sponsorship from

the U.S. Air Force (USAF), coordinated a meeting on the

Natural Attenuation of Ground Water at the Hyatt Regency in Denver, Colorado, on August 30 through September 1, 1994. The purpose of this meeting was to

bring together interested members of the bioremediation

community, including state and federal regulators, academicians, industry, and remedial contractors, to share

regulatory and field perspectives, present laboratory

and field research, and discuss issues related to the

successful implementation and evaluation of natural attenuation. Fran V. Kremer, EPA, Cincinnati, Ohio; John

T. Wilson, EPA, Ada, Oklahoma; and Gail E. Mallard,

USGS, Reston, Virginia, served as co-organizers of the

symposium, along with support from Lieutenant Colonel

Ross N. Miller, Headquarters, USAF, Center for Environmental Excellence, Brooks Air Force Base, Texas.

In conjunction with this symposium, a smaller meeting

on State and Federal Issues Impacting Natural Attenuation was held on Monday, August 29, at the same

location. The results of the regulatory meeting were

presented at the symposium and are reported in a separate section of this proceedings. 

More than 420 participants attended the 3-day symposium which included over 25 oral presentations, an overview of Monday’s regulatory meeting and a followup

panel discussion, and a poster session with 12 presentations. The oral presentations were divided into five

topic areas: a Regulatory Perspective on Natural Attenuation, an Overview of Investigations Related to Natural Attenuation, an Overview of Processes and

Procedures, and Case Studies on the Public Service

Company’s Site, Sites Where Natural Attenuation Has

Been Documented, and Natural Attenuation of Trichloroethylene and Similar Compounds.

Regulatory Perspective on Natural

Attenuation

Day One of the symposium opened with the regulatory

perspective session, chaired by Fran Kremer, EPA’s

Biosystems Technology Development Program. Colonel

Jim Owendoff, U.S. Department of Defense (DOD), offered DOD’s perspective as a facilities owner. With over

10,000 sites where biological and nonbiological cleanup

technologies are being considered or implemented,

DOD has attempted to adopt technologies that eliminate

contaminants from media rather than transferring them

between media. Bioremediation has emerged as an

attractive remedial option. Through several activities,

DOD has sought to engender greater participation and

communication among regulatory, public, and industrial

communities in carrying out remedial efforts. 

Guy A. Tomassoni, EPA Office of Solid Waste, presented

an overview of the Superfund National Contingency

Plan’s (NCP’s) decision criteria for natural attenuation

and discussed some of the issues and obstacles concerning the acceptance of natural attenuation as an

appropriate remediation technology, primarily the issue

of reasonable remedial time frame. A state perspective

was presented by John Shauver, Michigan Department

of Natural Resources. Mr. Shauver emphasized that

bioremediation, naturally occurring or otherwise, should

not be considered a “no action” remedial option. He

spoke of the increasing complexity and stringency of

regulations over the last 10 years and discussed the

emergence of bioremediation as a viable tool to meet

these challenges. He then outlined the steps that regulators look for in evaluating the potential efficacy of a

bioremediation proposal, specifically performing adequate site characterization, providing detailed description of the remedial process, and implementing an

extensive site monitoring program. 

Overview of Investigations Related to

Natural Attenuation

Gail E. Mallard, USGS, presented her organization’s

perspective in the context of its unique role as an earth

science information center without regulatory or management responsibilities. In relation to research conducted at a petroleum spill in New Jersey, Ms. Mallard

illustrated three points that USGS stresses regarding

bioremediation: performing site characterization, presenting multiple lines of evidence, and gathering as

much information as possible. By covering all three

issues thoroughly, USGS was able to evaluate and work

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towards developing an optimal engineered solution for

the site.

Fran Kremer’s presentation on research within EPA covered the Office of Research and Development’s (ORD’s)

mission to provide technical assistance to states and

regions while fulfilling longer term environmental and

regulatory needs. Dr. Kremer described the reorganization of ORD to emphasize health risk management and

outlined EPA Administrator Carol Browner’s 5-year strategic plan for the Agency stressing collaboration, longterm research, and expansion of competitive grants.

Finally, she reviewed some recent EPA natural attenuation projects including collaborative efforts with the

U.S. Air Force on developing a protocol for evaluating

the fate of dissolved hydrocarbons and with the Public

Service Company and Colorado Department of Health

to establish a methodology for evaluating possible risks

associated with benzene, toluene, ethylbenzene, and

xylene (BTEX) in soils following remediation.

Overview of Process and Procedures 

Session Two: Overview of Process and Procedures was

chaired by Ross N. Miller, USAF, Brooks Air Force Base.

The opening speaker, Michael J. Barcelona, University

of Michigan, Ann Arbor, presented an overview of site

characterization, which sets the stage for evaluating the

progress of the natural transformation of contaminants.

In addition to providing evidence of parent compound

disappearance, active microbial populations with

biotransformation capabilities, and the appearance or

disappearance of organic and inorganic constituents,

site remediators need to demonstrate with quantitative

evidence the net removal of toxic compounds solely by

biological processes. To meet this need, Mr. Barcelona

described a process of identifying the mass of contaminants and total reactive carbon and estimating the net

removal/transformation of reactive compounds over

time.

John L. Wilson, Department of Geoscience, New Mexico Institute of Mining and Technology, Socorro, New

Mexico, described the implications of the processes

controlling the distribution of oil, air, and water on natural

attenuation. Capillary, viscous, and buoyancy forces

control the movement and distribution of nonaqueous

phase liquids (NAPLs), a major source of aquifer contamination. Natural attenuation at a NAPL contaminated

site is most effective in the downgradient aqueous

phase of the dissolved plume, because of mass transfer

limitations, toxicity concerns, and limited nutrient availability within the NAPL-contaminated zone. Dr. Wilson

contended that to evaluate the effectiveness of bioremediation within the plume, however, it is necessary to

determine the distribution of the plume’s source or

whether that source is moving. 

Bruce J. Nielsen, USAF, Tyndall Air Force Base (AFB),

Florida, described new tools for locating and characterizing oil spills in aquifers. Tri-Services (the Air Force,

Army, and Navy) is executing a program to demonstrate,

test, and evaluate the application of cone penetrometers

to support natural attenuation demonstrations. This tool,

consisting of a laser spectrometer with fiber optics integrated with a cone penetrometer, can take a “snapshot”

of subsurface conditions showing the type and amount

of contamination, thereby helping to determine the best

method of remediation, monitor the effectiveness of

treatment, and assist technicians and engineers in judging when a site is clean. This technology has been field

tested at numerous Air Force, Army, and Navy installations, and in the future will undergo refinements to extend its sensitivities and expand its capabilities to other

types of contaminants.

E. Michael Godsy, USGS, discussed microbiological

and geochemical constraints to natural attenuation with

the focus on optimizing conditions. His presentation

covered the role of electron acceptors, factors affecting

bacterial-contaminant interactions in aquifer material

(e.g., bacterial nutrition, pH and redox potential, temperature, and physical deterrents to bioremediation),

and aerobic versus anaerobic metabolism for a variety

of organic contaminants. 

The final presentation in this session was given by Mary

Jo Baedecker, also of USGS. She discussed the integration of field and laboratory work in demonstrating the

efficacy of bioremediation strategies at a crude-oil spill

near Bemidji, Minnesota. At sites where the rates of

solubilization, volatilization, and biodegradation of hydrocarbons are such that the plume is either contained

or spreading at a slow rate, she concluded, natural

attenuation can be considered an important component

of a remediation program.

State and Federal Issues Impacting

Natural Attenuation 

Michael Jawson, EPA Robert S. Kerr Environmental

Research Laboratory (RSKERL), presented a summary

of Monday’s regulatory meeting at the end of Day One.

His presentation highlighted major themes and key issues, with a focus on what regulators need to know to

assess and evaluate natural attenuation. His emphasis

was on data gaps and needs for natural attenuation,

including the key areas of degradation rates, site characterization, monitoring, guidance materials, closure criteria, and risk assessment/management issues. A

general conclusion was that often characterization and

monitoring needs are even greater for natural attenuation than for other remedial alternatives due to the lack

of hydraulic controls on sites and the complexity of

establishing plume containment and degradation. The

more information that can be provided about site and

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contaminant characteristics and the more complete the

monitoring information, the more confident regulators

can be in evaluating remediation proposals and in closing sites where bioremediation was used. 

Mr. Jawson’s overview presentation was followed by a

panel discussion moderated by representatives of state

and federal regulatory agencies. Questions and answers focused on criteria regulators use in evaluating

applications for proposals to implement natural attenuation and for site closures. Day One closed with a poster

session, consisting of 12 posters.

Case Studies: Public Service’s Site

Days Two and Three of the symposium were devoted to

several sessions covering case studies, the first of

which concerned the successful application of in situ

bioremediation at the Public Service Company site in

Denver, Colorado. This session, chaired by EPA’s John

Wilson, consisted of five presentations made by the

various parties involved in the cleanup and subsequent

investigation, including the remedial contractor, a scientist at the site, a Public Service Company representative, and regulators from the state department of

health. Christopher Nelson, Groundwater Technology,

presented a short history of the site and explained the

treatment process from site assessment and treatability

studies to operation, monitoring, and results. At this site,

indigenous microorganisms, stimulated through the introduction of oxygen and inorganic nutrients, achieved

removal of nearly 94 percent of the original contaminant

mass. The bioremediation system operated from July

1989 to March 1992.

Tissa Illangasekare, University of Colorado, Boulder,

Colorado, described the role of mathematical models as

tools for the design and evaluation of ground-water remediation schemes and the application of a particular

model in developing a methodology for retrospective

evaluation. At the Public Service Company site, a methodology was developed to help determine, after closure,

whether or not the contaminant could be expected to

reappear and, if so, how long it would take and what

concentration levels it would reach in the ground water.

Harry E. Moseley, Public Service Company, presented

the steps the company took to remediate the site as

efficiently and cost-effectively as possible. He described

the remediation, monitoring, and evaluation processes,

along with precautionary measures taken and lessons

learned. Mr. Moseley also provided a fairly detailed

breakdown of cost data from the project.

This session concluded with two presentations by Lisa

Weers and Mark Walker, Colorado Department of

Health, regarding closure and postclosure issues at the

Public Service Company site. Ms. Weers and Mr. Walker

reviewed the factors they consider in evaluating a site

for closure, including whether the point of compliance

wells are meeting the state maximum contaminant levels, whether the existing monitoring program is adequate for detecting contamination from the release, and

whether the extent of remaining contamination could

adversely affect human health or the environment. Ms.

Weers and Mr. Walker emphasized that after a site is

closed, the owner/operator still remains liable for any

residual subsurface contamination, and the Department

of Health needs to be informed of any developments that

could increase the potential for adverse affects to human health or the environment.

Case Studies: Sites Where Natural

Attenuation Has Been Documented

The first part of the second case studies session,

chaired by Jim Spain, USAF, Tyndall AFB, Florida, consisted of five presentations focusing on sites where

natural attenuation has been documented. EPA’s John

Wilson described studies and findings related to characterization of JP-4 jet fuel spills at Eglin AFB, Florida, and

Hill AFB, Utah, and presented data on ground-water flow

and chemistry as well as measured and predicted contaminant concentrations. He also outlined the methodology behind distinguishing between contaminant

attenuation due to destruction from biological activity

and attenuation due to dilution or sorption and how

these methods were applied to correcting for dilution of

BTEX compounds at Eglin and Hill.

Thomas B. Stauffer, Tyndall AFB, described research at

Columbus AFB, Columbus, Michigan, that was conducted to measure in situ biodegradation rates of organic compounds. He presented the methods and

parameters of the pulse injection experiments used in

determining the amount of 14C carbonate produced by

complete degradation of 14C p-xylene and the amount

of 14C in water-soluble organic intermediate products. In

addition, Mr. Stauffer introduced the leaky reactor model

used to describe solute concentrations as functions of

time. He concluded with results of the experiment and a

discussion of the applicability of the leaking reactor

model.

In the next presentation, David W. Ostendorf, University

of Massachusetts, Amherst, summarized the theory describing capillary tension/saturation relations for immiscible continuous fluids. He discussed the theory’s

applicability to characterization of distribution of separate phase aviation gasoline (avgas) in solid core samples, specifically those taken from the U.S. Coast Guard

Air Station in Traverse City, Michigan. Mr. Ostendorf

also presented distribution profiles of total saturation,

free avgas, and residual avgas and reviewed the importance of such characterizations in the estimation of

separate phase contamination from monitoring-well observations.

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Barbara H. Wilson, EPA, Ada, Oklahoma, provided further discussion of the bioremediation studies conducted

at the U.S. Coast Guard Air Station in Traverse City,

specifically research into the geochemistry of BTX compounds. Ms. Wilson summarized the results of geochemical characterization of water samples and

presented evidence of natural attenuation of BTX compounds in aerobic and anaerobic aquifer material.

The next presentation, by Hanadi S. Rifai, Energy and

Environmental Systems Institute, Rice University, Houston, Texas, focused on mathematical modeling of natural

attenuation. Dr. Rifai presented four case studies that

combined field and laboratory investigations with modeling studies: the Conroe Superfund Site, Texas; the

Traverse City Site; a gas plant facility in Michigan; and

the Cliffs-Dow Superfund Site.

The second part of the session on documented case

studies, chaired by Gail Mallard, USGS, concluded on

Day Three with two presentations on studies conducted

at a crude-oil spill site in Bemidji, Minnesota. Robert P.

Eganhouse, USGS, Reston, opened with an overview

of the investigation at the Bemidji research site. Mr.

Eganhouse presented the methodology for water, sediment, and oil sampling; chemical analyses; and site

characterization. He further presented evidence of

change in ground-water geochemistry due to the inadvertent introduction of crude oil to the aquifer and provided data that supports the attenuation of oil-derived

contaminants due to biological, rather than physical,

processes. 

Supplementing Mr. Eganhouse’s talk, Hedeff I. Essaid,

USGS, Menlo Park, California, presented research into

the simulation and flow of transport processes at the

Bemidji crude-oil spill site. Mr. Essaid stressed the importance of using numerical models for integrating information collected in the field, for testing hypotheses, and

for studying the relative importance of simultaneously

occurring processes in complex, real-world systems. In

addition to outlining the procedure for collecting core

samples, Mr. Essaid summarized the assumptions and

results of the two-dimensional model used to simulate

multiphase flow through an oil pool at the site. He also

provided background and results of the two-dimensional

multispecies solute-transport model used to quantify the

field-scale degradation process and to identify the important factors affecting the distribution of solute species

in the field.

Natural Attenuation of Trichloroethylene

and Similar Compounds: Case Studies

The final session of the symposium focused on case

studies of natural attenuation of trichloroethylene (TCE)

and similar compounds. This session, also chaired by

Gail Mallard, opened with a presentation by Perry L.

McCarty, Department of Civil Engineering, Stanford University, on the anaerobic transformation of chlorinated

solvents. In his presentation, Mr. McCarty provided

background information on major chlorinated solvents

and described the chemical transformation of 1,1,1-

trichloroethane (TCA) and the biological transformation

of chlorinated aliphatic hydrocarbons (CAHs). He concluded his talk by presenting case studies that provide

evidence for natural attenuation of chlorinated solvents.

Mary Martin Chepiga, USGS, West Trenton, New Jersey, spoke next about TCE contamination of ground

water at the Building 24 Site at Picatinny Arsenal, New

Jersey. She described how the infiltration of wastewater

from lagoons behind the building and of chlorinated

solvents from a dry well in front of the building had

created a plume of contaminated ground water downgradient of Building 24. Ms. Chipega further described

the conceptual model of the physical, chemical, and

biological processes that affect the transport and mass

balance of TCE within the plume. She also presented a

preliminary solute mass balance of TCE in the unconfined aquifer at the site and compared those findings

with results of numerical solute-transport simulations.

James W. Weaver, EPA, Ada, Oklahoma, closed the

case studies session with a presentation of the natural

attenuation of TCE in ground water at the St. Joseph,

Michigan, National Priority List (NPL) site. Mr. Weaver

summarized the data collection, presentation, and interpretation methods used to infer and analyze TCE

concentration in the western plume near Lake Michigan. He emphasized that the contaminant mass flux

and apparent degradation constants were based on

visualization of the data, that presentation of the data

was free from artifacts of interpolation, and that extrapolation of the data beyond the measurement locations was controlled.

Fran Kremer provided closing remarks, and the symposium was adjourned.

4
State and Federal Issues

Impacting Natural Attenuation

5
State and Federal Issues Impacting Natural Attenuation:

Meeting Overview

Introduction

A meeting to discuss state and federal regulatory issues

impacting natural attenuation was held on Monday, August 29, 1994, in Denver, Colorado, in conjunction with

the U.S. Environmental Protection Agency’s (EPA’s)

Symposium on Natural Attenuation of Ground Water.

The objective of the meeting, which was attended by 28

state and federal regulators, was to discuss obstacles

and issues related to the use of natural attenuation as

an accepted cleanup option at waste sites. Critical topics addressed included:

• What regulators need in order to assess and evaluate

natural attenuation proposals.

• Data gaps that need to be filled to increase the viability of natural attenuation as a cleanup option.

• Site closure criteria.

• Long-term monitoring requirements.

Fran Kremer and John Wilson of EPA’s Office of Research and Development co-chaired the meeting. (Appendix A contains a list of the meeting participants.)

Definitions

The meeting opened with a discussion of terminology

and an attempt to develop a definition for natural attenuation that is accurate and avoids misconceptions and

negative connotations currently associated with related

terms such as natural attenuation. Participants agreed

that the perception of natural attenuation and related

processes as “no action” remedies represents a major

obstacle to the acceptance of natural attenuation. 

A working definition was crafted for purposes of the

meeting discussions, but participants concluded that

formalizing a suitable definition of natural attenuation is

a high priority, both to change public perception and to

promote regulatory consistency.

Overview

The major issues raised in the meeting included the

following:

• Obtaining degradation rates

• Improving site characterization

• Cost-effective monitoring

• Developing guidance materials

• Developing closure criteria

• Addressing risk assessment/management

Conclusions and recommendations from discussion of

these issues follow. 

Degradation Rates

There is a need for contaminant degradation rates,

which would be used to facilitate site characterization,

assess the viability of natural attenuation as a remedial

action, and predict how long cleanup would take. This

latter information would help regulators justify the use of

natural attenuation where it could be used for screening

purposes or where cleanup could be shown to be completed in a reasonable time frame. The need for regulators to identify an endpoint for cleanup via natural

attenuation was a recurring theme. Degradation rates

are especially needed for contaminants other than petroleum, for which a fair amount of data currently exist.

Participants suggested, however, that a published record of existing petroleum data would be very useful.

Data on regulated contaminants and mixtures are particularly needed, as most spills are not single compounds but mixtures of several contaminants. Field data

would be most useful, and better methods for relating

laboratory data to the field would also be helpful. Ideally,

the data produced should be site specific and take into

account site geology, contaminant transport, electron

acceptors at the site, and other site characteristics.

Degradation rates also should be tied to mass balances,

such that the mass balances coincide with projected

rates.

Site Characterization

Detailed site characterizations are needed for all remedial efforts, but natural attenuation requires particularly

intensive characterizations. To justify natural attenuation, regulators need to be assured that contaminants

will not migrate to sensitive receptors or pose health

6
threats during cleanup. Thus, cost-effective methods for

characterizing the location and possible migration of the

plume, particularly in relation to sensitive receptors and

other health exposure points (e.g., utilities, drinking

water wells, and discharge points at streams, ponds,

and lakes), are needed. Regulators want full three-dimensional plume characterization, including vertical distribution, along with information on physical and

hydrogeological parameters to determine the fate and

transport of contaminants. Also needed are visualization

systems that synthesize and display all of these characterization data in an easily understandable manner.

Site characterization is considered one of the more

costly components of natural attenuation. For small sites

(e.g., those in the leaking underground storage tank

(UST) program), assessment costs can run as high as

$40,000 or $50,000, a burden that many small businesses cannot afford. In Utah, for example, only about

10 percent of UST sites can afford to hire contractors to

perform thorough assessments. In addition, in some

states, such as Utah and Nevada, there are very few

environmental consultants available, even if cost were

not an issue. One response was that states could offer

reimbursement programs, similar to that in Florida. Another suggestion was to prepare guidance that specifically

addresses how to characterize sites with limited funding.

Monitoring

Bioremediation is not a “walkaway” remedy; on the contrary, it requires intensive, often long-term, monitoring.

Natural attenuation, in fact, can require more monitoring

and analysis than many other types of remediation.

Issues of concern to state and federal regulators included how frequently to monitor and at how many

points, exactly which parameters to measure to assess

the progress of biodegradation processes and cleanup,

and whether surrogate parameters would be adequate

if a project was constrained by budget. In addition to

establishing points of compliance, a primary goal of

monitoring is to ensure that the plume is not migrating

beyond the points of compliance. Monitoring should occur more frequently initially, then less frequently when

remediation rates, migration patterns, and seasonal influences are understood.

Potential monitoring parameters mentioned by participants included dissolved oxygen (DO), temperature gradients, and electron acceptor levels. Analysis of volatile

organic chemicals (VOCs) was cited as a requirement

that could greatly increase costs. Information on techniques and methods for analyzing different classes of

compounds would be useful. Participants also expressed

a need for validation of data from conventional wells,

and for mass balance calculations. Finally, participants

discussed the role of models versus field data, and the

extent to which models could be used to complement or

even replace field data for the purposes of predicting

plume behavior and contaminant degradation.

Type of Guidance Materials

The need for guidance materials was stressed. Information on natural attenuation needs to be presented in

simple terms so that it can be offered clearly and persuasively as a remedial alternative. More information is

needed on the types of situations for which natural

attenuation is suitable, considering such factors as contaminant types and levels and site characteristics. State

and federal participants called for protocols, in the form

of guidelines, on what parameters to measure and procedures to use. It was also suggested that user-friendly

computer models and decision support systems would

be invaluable to assist site coordinators with implementation.

Closure Criteria

A key need is the establishment of closure criteria to

determine at what point a site can be considered “clean.”

There was discussion as to which parameters should

serve as “endpoint parameters,” or parameters that

must be measured to prove that cleanup is complete.

Participants questioned the adequacy of current criteria,

particularly with regard to toxicological tests. Toxicological testing should be made on the weathered rather than

nonweathered material, and the material should come

from the site itself.

Participants also felt that criteria other than total petroleum hydrocarbons (TPHs) should be used and that

daughter products and by-products for the degradation

process should be carefully measured and assessed.

Resolving these issues would provide regulators with

more confidence in closing sites where natural attenuation was used. 

Summary: Needs for Effective Risk

Assessment/Management

The final critical theme, drawing on all of these issues,

was what information regulators would need in order

to perform risk assessment and management of natural attenuation sites. Regulators at the meeting

pointed out that they frequently receive proposals for

intrinsic remediation that do not provide adequate

data on physical parameters. Participants concluded

that they would need information related to plume

containment, including whether or not degradation

was occurring. For petroleum sites, it was felt that

bioremediation usually is occurring, but more data on

retardation and degradation coefficients would be

helpful. For nonpetroleum sites, assessment is more

7
difficult, and would require databases based on actual

sites showing the range of rates and parameters. A

database of case studies documenting both successes

and failures of natural attenuation in the field would be

of enormous value. Such a database is under development; it currently has information on over 450 bioremediation sites, some of which have intrinsic components.1

Regulators would also find useful specific information on

what distinguishes intrinsic from engineered bioremediation. All of this information should be presented in

clear, simple terms in guidance materials geared for

decision-makers and regulators.

Cost-effectiveness and timeliness are other key issues

in assessing the natural attenuation option. Sites need

to document that natural attenuation can be achieved in

a “reasonable time frame,” a phrase that carries some

flexibility depending on the relative containment of the

contamination and its threat to receptors. Some participants suggested that natural attenuation is most appropriate for sites where contamination is relatively low, little risk

is posed to drinking water supplies, and the contaminant

source has been removed. In many cases, natural attenuation can be shown to be more cost effective and speedier

than other potential remedies.

Key data needs are summarized in Table 1.

Table 1. Checklist of Data Gaps and Needs

A. Degradation Rates

• For more than petroleum and regulated products/mixtures

• Field data (relation to lab)

• Tied into site characteristics (geology, transport, etc.)

• Multiple electron acceptors

• Tied into mass balances

B. Site Characterization

• More intensive for natural attenuation

• Three-dimensional characterization

• Plume location in relation to sensitive receptors and health

 exposure points

• Data (physical, hydrogeological, etc.) to determine fate and

 transport

• Visualization systems

C. Monitoring

• Frequency

• Parameters to measure (use of surrogate parameters)

• Validity of data from conventional wells, need for mass balances

• Role of models vs. field data

D. Guidance Materials

• Information in simple terms

• Situations (sites, levels, and types of contamination) where

 natural attenuation is applicable

• Protocols (parameters to measure, procedures to use)

• Models (user friendly) and decision support systems

E. Closure Criteria

• Endpoint parameter data

• Adequacy of current criteria/toxicological tests

• Toxicological testing of “weathered” object

• Daughter and/or by-products

F. Risk Assessment/Management

• Plume containment, contaminant degradation

• Time frame

• Cost effectiveness

• Database of case studies

• Guidelines for decision-makers

1

The Bioremediation in the Field Search System (BFSS) 2.0 is a

PC-based bioremediation site database that allows users to search

for data electronically, view the data, and print reports. BFSS is

available on EPA’s Alternative Treatment Technology Information

Clearinghouse (ATTIC, 703-908-2138), Cleanup Information (CLU-IN,

301-589-8366), and Office of Research and Development (ORD,

513-569-7610) electronic bulletin board systems. BFSS is also available on diskette from EPA’s National Risk Management Research

Laboratory. To request a copy, call 513-569-7562.

8
Panel Discussion on Regulatory Issues

Panel Members:

Michael Barden, Wisconsin Department of Natural Resources, Madison, WI

Matthew Charsky, Hazardous Site Control Division, Office of Solid Waste and Emergency Response, U.S.

EPA, Washington, DC 

Daniel Cozza, Region 5, Superfund Program Management Branch, U.S. EPA, Chicago, IL

Fran Kremer, Office of Research and Development, U.S.

EPA, Cincinnati, OH

Gail Mallard, U.S. Geological Survey, Reston, VA

John Shauver, Michigan Department of Natural Resources, Lansing, MI

Guy Tomassoni, Office of Solid Waste, U.S. EPA, Washington, DC

John Wilson, Office of Research and Development, U.S.

EPA, Ada, OK

Question: The list of data gaps and needs presented

by Mike Jawson is very good. It would be helpful if the

list were written up and distributed so that we could work

from it.

Answer (Fran Kremer): Symposium participants will

receive a final report from this meeting, which will include the results of the meeting and panel discussion.

Question: The list presented was very comprehensive,

sort of a “wish list.” Is it necessary to address all of it

before regulatory agencies begin to accept applications

of bioremediation? Which two or three critical needs

should be met now, as opposed to all 20 or 25 mentioned?

Answer (Dan Cozza): For the most part, nothing on

the list is new. For any technology, you need essentially

the same list of criteria to close out a site or choose a

remedy. You can however, have flexibility, for example

with respect to monitoring and site characterization. But

since we don’t want natural attenuation to be labeled as

“walkaway” or “no action,” we may require even more

monitoring than for an active pump-and-treat system. So

for the most part, you will need nearly everything on the

list, but there will be flexibility. Every site is different, and

sometimes the information will already be available.

Closure criteria are the same regardless of the type of

technology.

Answer (John Shauver): I’d like to say “all of the

above,” but we live in the real world. Remember that

bioremediation is still site-specific; therefore, we can’t

paint with a broad brush. If you know what kinds of

systems operate best on the contaminant at the site, you

can limit the monitoring and the data that need to be

generated. Often responsible parties propose bioremediation without analyzing for sulfate, nitrate, iron, etc. or

knowing anything about the geochemistry, biochemistry,

or microbiology. It is important to think about what the

processes need and develop analytical data based on

what is likely to happen before submitting a proposal.

Because intrinsic and any other kind of bioremediation

is site specific, the state regulatory agency can’t provide

a cookbook.

Question (Fran Kremer): South Carolina has developed some guidelines for what they would like to see in

proposals that come to them. I would like to ask

Read Miner from South Carolina to comment on those

guidelines.

Answer (Read Miner): In 1992, we published a guidance document for completion of comprehensive assessments. We have had a lot of success with

contractors completing more comprehensive assessments to more fully characterize sites so that the best

remediation options could be determined. We are currently developing criteria for natural attenuation; we are

gathering information at this symposium to help complete the development of these criteria.

Question: What criteria will EPA use to evaluate a plan

specifically for a Superfund site (how many samples

should be collected and where, what parameters should

be tested for, etc.)?

Answer (Matthew Charsky): In the past, we have

looked at four factors for Superfund sites: practicability,

cost-effectiveness, the unlikely future use of drinking

water, and the time frame to reach the cleanup level.

Historically, we have used the term natural attenuation rather than natural attenuation. Natural attenuation has been used at approximately 57 sites almost

9
nationwide between 1986 and 1993. In order to use

natural attenuation, these sites have had to meet the

following criteria: contaminant concentrations are relatively low, drinking water is not threatened or can be

easily replaced, the source of contamination is no longer

present, source removal will result in a predictable decrease in ground-water contamination, and the active

restoration time frame is not radically different from that of

active remediation. 

Question: In a case where EPA has already agreed

that we can evaluate the process, how will the adequacy

of the plan be evaluated?

Answer (Dan Cozza): It is hard to specify what we

would accept (for example, a certain number of wells)

because every site is different. The bottom line is that

you have to show that the plan will not affect human

health and the environment and you need to do the site

characterization. Another key element in a plan is partnership among EPA, state and local agencies, and the

responsible party to work together on a proposal before

it comes in. Such a partnership among everyone involved is critical if natural attenuation is going to become

a reliable alternative.

Question: There seems to have been a somewhat conflicting message today: On the one hand, there has been

discussion of the need for site-specific evaluations; on

the other hand, it has been implied that one must do

mass balances, microcosm studies, and so on. I hope

we don’t lose sight of the fact that the proof that must

be provided on natural attenuation projects is a function

of the objectives and criteria you are trying to meet and

the site conditions. For example, at one site you might

have to show that a plume is not growing or is shrinking;

at another site you might have to show that a plume will

not reach a receptor location a certain distance from its

current extent. In the latter case, you need a more

rigorous proof than in the former case. We need to

recognize the need for flexibility in the level of proof

required.

Answer (Guy Tomassoni): Particularly in the RCRA

program, we often have years of data from a site. If we

have a good monitoring program over the years that

shows the plume reducing in size and concentration,

then we shouldn’t always require the regulated entity to

expend a tremendous amount of time and resources to

prove why good things have happened. There might be

relatively inexpensive ways to show that bioremediation

is taking place, but at some sites it may be very difficult

and costly to figure out why levels are going down and

why the plume is not growing. Resource Conservation

and Recovery Act (RCRA) regulations require remediation of ground water to achieve designated standards;

the regulation does not require proof of why levels are

being achieved.

Answer (Fran Kremer): There is a difference as well

between RCRA/Superfund sites and UST sites. There

are different levels of resources available to do the

necessary work, and state UST programs have varying

opinions about what is considered acceptable.

Answer (Michael Barden): At all the sites ranging

from Superfund to leaking USTs, we look for adequate

site characterization (often including vertical gradients)

so that we know where the plume is and where it is

going, and we look for adequate monitoring to show that

concentrations are decreasing over time. If we can get

a better understanding of how these processes work,

that will help us consider which sites are suitable for

natural attenuation. From the regulatory standpoint,

however, the fundamental point is that concentrations

are decreasing, not necessarily why this is happening

so long as the plume isn’t migrating. Mass balance is

the key indicating that we are losing the contaminant

mass over time, but cleanup standards are concentration-based, not mass-based, so we are looking at concentration changes.

Question: Regarding closure criteria, many sites with

natural attenuation will not reach maximum contaminant

levels (MCLs) or other levels in the entire plume, and

decisions must be made to leave some contamination

in place. How do you make those decisions, and how

do you use the concept of alternate compliance points,

which fit in well with the natural attenuation approach?

Answer (John Shauver): One option is to clean up to

background levels, get a clean bill of health, and have

no future liability. But remember, risk-based numbers

are generated by an algorithm using conservative assumptions, based on complete freedom of use of the

site. With a risk-based number, the potential exists for

the regulatory community to say that the number is too

high and the site must be cleaned up again. Therefore,

in the case of natural attenuation, you must look at

cleanup not only from the standpoint of environmental

protection but also long-term liability. These sites may

be looked at for 25 or 100 years. As long as you can

show that public health and the environment are protected, we don’t care how long the site is monitored, but

you will be held to a cleanup standard based on future

use of that resource. We may be willing to write off the

site for 50 or 100 years while we watch the microorganisms degrade the contaminants, but there must be some

activity on the site to make sure that the microorganisms

are not being affected in some other way and that degradation is really taking place. We like bioremediation

because it is not temporary storage and does not transfer contaminants from one medium to another, but there

are liability risks and social risks, and the regulator must

take those risks into account in determining how much

contamination can be left on a site. Much of that determination is not based on science.

10
Answer (Michael Barden): The enabling statutes in

the states vary widely. For example, in Wisconsin,

ground-water standards must be met for all ground

water in the state, whether you drink it or not. We do

have flexibility with respect to how long you have to meet

those standards. That is where we can consider risk

management, plume migration, and the potential impact

on receptors. If there is no potential impact, we don’t

care how long it takes, but the standards must ultimately

be met.

Question: Has there been much movement toward

combining the technology of natural attenuation with

other methods of attenuation? For example, natural attenuation might be effective at very small sites, but not

at large sites, where additional types of remediation

might be necessary. But if you resort to pump-and-treat,

you may upset the microbiological environment and no

longer have the natural attenuation. Has there been

much work on combining natural attenuation with other

passive forms of treatment, such as a funnel and gate

system?

Answer (John Shauver): From our perspective, it

doesn’t matter what suite of technologies a responsible

party uses to address a site. When you have reasonably

low concentrations (parts per billion or low parts per

million range), water supplies are not at risk, and you

can show that degradation is taking place, then you can

use something more passive. The problem in areas with

increasing populations is that it is difficult to predict how

long that cleanup will take. Some communities don’t

want to wait 25 years or more because of lost economic

development opportunities. Therefore, you have to look

at community concerns and long-term land use as well

as regulatory needs.

Answer (Guy Tomassoni): States or areas of the

country might have varying ideas about what constitutes

a reasonable time frame. We are seeing developments

in more passive technologies, such as treatment walls,

that could possibly supplement completely natural systems. Combinations of a conventional or more active

remedy for a part of a plume with an intrinsic-based

remedy for the remaining plume may prove in many

settings to yield a cost-effective and protective remedial

strategy. Clearly, what an entity views as a reasonable

time frame is crucial. It is also important to realize that

time frames for cleanups using a conventional pumpand-treat remedy can be very long as well.

Question: Wisconsin has had clear and usable guidance for a year and a half on how to evaluate a site for

natural attenuation. At what percentage of the sites has

natural attenuation been found to be a viable option?

Answer (Michael Barden): We don’t have any hard

numbers because we don’t have the personnel to put all

the information into a database. We recently surveyed

site owners to determine what remediation technologies

are being used. In the past year, about 80 sites have

applied natural attenuation at least to soil contamination.

Fifteen or 20 are looking at this technology for groundwater contamination. Larger stable entities such as major oil companies and active industrial sites are looking

at this as a viable option. We are getting a lot of interest

in natural attenuation, but more as a followup to more

active remediation, because many responsible parties

are not willing to wait long enough.

Question: What is the real objective of natural attenuation—is it to stop the plume, to shrink the plume, or

even to attack the residual nonaqueous phase liquid

(NAPL)? How are objectives viewed by the regulators?

Answer (John Shauver): In Michigan, we have a multiple use ethic, looking at resource management and

resource use. If you are going to let a plume get bigger,

we are concerned about future users as well as existing

users. As a regulator, I want control so that future users

won’t be affected and don’t have to worry about a public

water supply, and I want to know that degradation is

taking place. When I say stop the plume, I want some

mechanism in place that can achieve that control. If

microorganisms are achieving that goal, then put the

monitoring system in and demonstrate that it is working.

You don’t have to do all the scientific studies, but we

want to know that degradation is taking place and that

the plume is not getting bigger in three dimensions.

We’re willing to watch it 5, 30, or 100 years while the

microorganisms do the job. But at a fresh spill where

you’ve removed the source and the plume is likely to get

much bigger (and there are homeowners nearby using

the water), we will probably say that you must keep the

plume that size, using some mechanical control if need

be. We will evaluate every site, but our basic concept is

not to contaminate more of the resource if there is a

reasonable, feasible way to maintain the size of the

plume. Therefore, we put limits on when natural attenuation is appropriate.

11
Symposium on Natural Attenuation of Ground Water:

Summaries and Abstracts

13
Regulatory Perspective on Natural Attenuation

Chairperson: Fran V. Kremer, U.S. Environmental Protection Agency

Department of Defense Perspective

Colonel Jim Owendoff, U.S. Department of Defense,

Washington, DC

The U.S. Department of Defense (DOD) consists of over

1,000 installations across the United States and overseas, with over 10,000 sites where biological and nonbiological cleanup technologies are being considered or

implemented. DOD has an annual budget of about $2.3

billion for cleanup at closing and nonclosing bases.

Funds for the overall defense budget have decreased

while funding for environmental programs has increased

or remained constant over the past several years.

DOD hopes to move toward using technologies that

eliminate contaminants from media rather than toward

those that transfer contaminants between media. DOD

also must address the issues of prioritization and sequencing of work, timeliness of cleanup, ease of technology implementation, and affordability of treatment. In

assessing these factors, DOD hopes to engender a

greater participation among the regulatory, public, and

industrial communities. The drafting of management action plans constitutes a step in the direction of greater

collaboration. These plans provide comprehensive descriptions of each site along with detailed information

about its contaminants and the strategy, schedule, and

initiatives for site cleanup. The establishment of a Restoration Advisory Board, a body of regulators and community members that would evaluate the relative risk of

site contamination, is a further move towards greater

communication of problems and solutions among the

regulatory and public communities. 

In view of the large number of DOD sites which have

hydrocarbon contaminants, intrinsic bioremediation

holds substantial promise for application. For intrinsic

bioremediation to be accepted, the remedial project

manager (RPM) must perform site evaluations and present evidence that bioremediation has occurred as well

as implement monitoring systems to confirm the continuance of natural biodegradation. Utilizing these principles and guidelines, DOD will be able to implement

intrinsic bioremediation while protecting and preserving

human health and the environment. 

U.S. EPA Regulatory Perspective: Natural

Attenuation as a Remedial Option for

Contaminated Ground Water

Guy A. Tomassoni, U.S. Environmental Protection

Agency, Office of Solid Waste, Washington, DC

The two primary goals of both the RCRA and CERCLA

(Superfund) programs for contaminated ground water

are (1) protecting human health and the environment,

and (2) restoring contaminated ground water to beneficial uses, where technically practicable. EPA has recognized in policy statements and guidance that one

approach to achieve these goals in certain settings can

involve natural attenuation. Thus, EPA policy allows for

natural attenuation to be used as a component of protective and cost-effective remedies.

Natural attenuation is addressed in the preambles of

both the National Contingency Plan (NCP) (55 FR 46

March 8, 1990) and the proposed Subpart S rule (55 FR

145 July 27, 1990), which provide the regulatory framework for implementing the Superfund and RCRA Corrective Action programs, respectively. The NCP defines a

remedy based on natural attenuation as one where

biodegradation, dispersion, dilution, and adsorption will

effectively achieve remedial goals. The NCP preamble

also states that natural attenuation is generally recommended only when active restoration is not practicable,

cost effective, or warranted based on site-specific conditions (e.g., where ground water is not a likely source

of drinking water), or where natural attenuation could

achieve remedial goals in a reasonable time frame.

Similar to the NCP, the Subpart S preamble discusses

natural attenuation in the context of reasonable cleanup

times, and where the likelihood of exposure is minimal.

Furthermore, the Subpart S preamble states that natural

attenuation could play a major role in a remedy that can

be carried out over an extended period of time.

14
An example of EPA guidance that recognizes natural

attenuation was issued very recently by the Office of

Underground Storage Tanks. The guidance document

titled, Evaluating Alternative Cleanup Technologies for

Underground Storage Tanks: A Guide for Corrective

Action Plan Reviewers, covers eight technologies

including natural attenuation. The chapter on natural

attenuation includes guidance on initial screening, detailed evaluation, and monitoring. While EPA’s recognition in policy and guidance helps to determine when to

select natural attenuation as part of remedial action,

significant but not insurmountable obstacles still exist.

Two major obstacles include (1) the association of natural attenuation with a no action remedy; and (2) the

preference for rapid remediation.

Associating natural attenuation with no action often results in the conclusion that the regulators would be

letting the responsible party “off-the-hook.” Overcoming

this obstacle will involve educating all interested parties

that a remedy based on natural attenuation is intended

to achieve remedial goals and will still involve a significant cost to the responsible party. For example, a natural attenuation remedy will generally require thorough

site characterization, source control or removal, documentation (where appropriate) of attenuation processes,

and comprehensive long-term monitoring to ensure

protectiveness and contaminant reduction. Natural attenuation remedies should also generally include contingencies for more “active” remedial measures in the event

that natural processes will not achieve the goals, or where

such measures are deemed necessary for part of the

contaminant plume. Activities such as these need to be

conveyed to distinguish a remedy involving natural attenuation from a no action or walkaway decision.

The preference for rapid remediation can be an obstacle

because natural attenuation is very often associated

with prolonged remedial time frames. Overcoming this

obstacle can be achieved by (1) recognizing that remedial time frames associated with conventional pumpand-treat systems also may be prolonged; and (2)

recognizing that longer remediation time frames may be

appropriate in certain circumstances where there is no

urgent need for the resource, and human and environmental exposures can be prevented. 

Scientific evidence has documented nature’s capacity to

rectify environmental contamination in certain settings.

Where natural processes are determined capable of

achieving remedial goals for ground water, EPA affords

the flexibility to evaluate and select natural attenuation

as a component of a protective remedy. The further

acceptance of remedies involving natural attenuation will

likely be facilitated by continued advances in the scientific

understanding of the processes involved, and the ability to

address public and regulatory concern that such remedies

be protective and will achieve the intended cleanup goals. 

State Perspective

John Shauver, Michigan Department of Natural

Resources, Lansing, MI

Bioremediation is not a no action, walkaway alternative,

whether it occurs naturally or with assistance. A presentation given 10 years ago at the 7th Ground Water

Conference at Michigan State University allows one to

examine the Department of Natural Resources perspective on ground water in Michigan and assess the

changes that have occurred in the past 10 years. Although the approach to seeking legal relief at any

ground-water contamination site has remained the

same, regulations have become more complex and

stringent. Consequently, regulation prompted the research and development of cleanup technologies that

are quicker and less costly, and that provide greater

long-term liability. Bioremediation has emerged as a tool

that can meet those needs. 

Bioremediation has significant advantages over other

treatment technologies. True degradation of contaminants will limit long-term potentially responsible party

(PRP) liability. In addition, although initial startup costs

and long-term monitoring costs may be somewhat

higher than for other remedies, bioremediation allows

for in-place treatment, thereby eliminating the need for

repeated excavations over time. Bioremediation also

offers an element of environmental sanity in that it does

indeed eliminate waste by eliminating toxicity and the

need for continued storage and disposal.

Whether natural or enhanced, bioremediation is a treatment technology that must be described to regulators in

detail. Being able to describe the requirements of the

remedial process to regulators helps them to understand the treatment process, to settle any permitting

issues that may arise, and to be assured that the technology has a reasonable chance of remediating the

contaminants so that the resource is protective of public

health and the environment. 

The first step in any treatment process should consist of

an adequate site characterization. Next, it is necessary

to understand microorganism behavior and response to

the contamination so that plume shrinkage or expansion

can be described and predicted. Monitoring systems

can assist in addressing plume migration in current and

long-term site characterizations.

The biggest barriers to the implementation and success

of bioremediation in Michigan are the numerous unrealistic and uninformed proposals that are brought before

state environmental agencies. An adequate site characterization must be performed, and initial site investigations are not necessarily less costly for bioremediation

technologies than for other processes.

15
Overview of Investigations Related to Natural Attenuation

Chairperson: Fran V. Kremer, U.S. Environmental Protection Agency

USGS Perspective

Gail E. Mallard, U.S. Geological Survey, Reston, VA

Many of the investigations that the U.S. Geological Survey (USGS) conducts on the processes that affect the

fate and transport of contaminants directly relate to the

topic of intrinsic remediation. A division of the Department of Interior, USGS is an earth science information

agency that has no regulatory or management responsibilities. Thus, USGS is able to provide information in

an unbiased manner to both the regulatory and regulated communities. The organization has offices in every

state and conducts studies in close collaboration with

local, state, and other federal agencies as well as universities. In choosing field laboratories, USGS selects

sites that are known to be contaminated with what the

agency considers to be the most important contaminants in the environment. USGS’s approach in performing field investigation is to integrate disciplines to

conduct research that produces thorough site analyses

and reports. 

In addition to field-based investigations, USGS research

includes a significant laboratory component. Laboratory

studies help to verify field observations and to understand controlling processes. The ultimate goal of USGS

research is to understand processes so that information

transfer can occur between sites about the fate and

transport of contaminants.

Research conducted at the site of a petroleum spill in

Galloway Township, New Jersey, illustrates the three

steps that USGS stresses regarding the use of bioremediation as a treatment technology:

• Performing site characterization

• Presenting multiple lines of evidence

• Gathering as much information as possible

Understanding site, system, and process helps to evaluate cleanup alternatives. At the Galloway site, a gasoline

tank on a fuel farm had leaked, but the extent and

duration of the leak were unknown. USGS assisted in

performing a site characterization that enabled the

agency to gain approval from New Jersey regulatory

authorities to conduct ongoing research at the site. 

Multiple lines of evidence, such as information based on

both field and laboratory observations, data on hydrogeologic parameters, and information on accumulation

of by-products, further strengthen the case for bioremediation. Periodic reappearances of petroleum by-products combined with information on hydraulic properties

of the unsaturated zone soil and other hydrogeologic

data helped USGS and its collaborators to infer the

overall degradation rate of hydrocarbons at the Galloway site. Comparing rates inferred from field studies

with those from laboratory experiments gave researchers confidence that the remedial approach was

effective. 

Having information such as degradation rates gives researchers a better understanding of the natural system

so that they can arrive at optimal engineered solutions.

For instance, degradation rates can be used in a passive

remediation system to predict how long hydrocarbons

are likely to persist in the unsaturated zone, or they can

be used in an engineered solution to design models that

optimize hydrocarbon removal and vapor extraction

(modeling currently being developed at Drexel University). Thus, the more that is known about a natural

system, the better the engineered solution will be. 

U.S. EPA Perspective

Fran V. Kremer, U.S. EPA, Office of Research and

Development, Cincinnati, OH

The primary mission of the U.S. EPA’s Office of Research and Development (ORD) consists of providing

technical assistance to regional offices and the states

while carrying out a longer term mission to conduct

research to fulfill needs in the environmental area. ORD

also supports activities for the regulatory offices to supply them with necessary data to issue regulations.

ORD has a staff of approximately 1,200 employees and

had a budget of about $535 million for the 1993 fiscal

16
year. The Agency had an overall increase in its budget

for 1994 and is currently undergoing evaluations for the

1995 budget. ORD’s projected total budget of $564

million includes $9.2 million tentatively reserved for

bioremediation activities and $11.7 million dedicated to

ground-water research.

ORD consists of offices and laboratories across the

country as well as 8 offices at headquarters in Washington, DC, 12 other laboratories, and 5 additional field

facilities. ORD is currently undergoing a major reorganization around four “mega” laboratories. Although the

consolidation will not force any labs into closing, at this

time, management will be realigned to reflect the

changes in office structure. The new Risk Management

Laboratory will encompass most of the engineering operations including the air engineering research being

conducted out of Research Triangle Park, North Carolina. The Environmental Criteria and Assessment Office

will move under the Risk Assessment Laboratory, and

the Monitoring Support Laboratory as well as the Ada

and Athens labs will become a part of the newly

structured Exposure Research Laboratory. Finally, the

Health Effects Research Laboratory and the Gulf

Breeze Laboratory will move under the Health and Environmental Effects Research Laboratory.

Some projects on natural attenuation include a collaborative effort with the U.S. Air Force Center for Environmental Excellence in developing a protocol for

evaluating the fate of dissolved-phase hydrocarbons in

ground water. Another project with the Public Service

Company in Colorado involves the development of a

model to establish a methodology to evaluate possible

risks associated with BTEX in soils after remediation.

Research also has been conducted toward distinguishing between attenuation due to destruction from biological treatment and attenuation due to dilution or sorption.

Evaluations also have been conducted on the geochemistry of ground water to confirm the existence of biological activity through examinations of product and

consumed compounds. Through these investigations,

EPA hopes to acquire a better understanding of bioremediation and its potential for environmental cleanup.

17
Overview of Process and Procedures

19
Site Characterization: What Should We Measure, Where (When?), and Why?

Michael J. Barcelona

Department of Civil and Environmental Engineering, University of Michigan, Ann Arbor, MI

Abstract

Site characterization represents the initial phase of the

active monitoring process that occurs as part of intrinsic

organic contaminant bioremediation efforts. Initial characterization work sets the stage for evaluating the progress of the natural transformation of contaminants. The

following have frequently been observed: parent compound disappearance, active microbial populations with

biotransformation capabilities, and the appearance or

disappearance of organic and inorganic constituents

that provide evidence of bioremediation at contaminated

sites. Quantitative evidence is lacking, however, for net

removal of toxic compounds from complex mixtures

solely by biological processes. This is due largely to the

reliance on monitoring well samples for evidence of

biological activity, rather than on identifying the mass of

contaminants (and total reactive organic carbon) and

estimating the net removal/transformation of reactive

compounds over time.

A dynamic approach to quantitative site characterization

is needed that recognizes intrinsic bioremediation as an

active cleanup approach. Careful attention must be paid

to the identification of the three-dimensional distribution

of contaminant mass. Then the correspondence between contaminant distribution and favorable physical,

geochemical, and microbial conditions in the subsurface

over time provides a basis for net contaminant-removal

estimates. Mere adaptations of detective ground-water

monitoring networks are insufficient for quantitative

evaluation of intrinsic bioremediation technologies.

Introduction

The practice of site characterization for remediation of

subsurface organic contaminants has evolved slowly in

the past decade. Early guidelines (1-3) for minimal

ground-water contamination detection monitoring (i.e.,

monitoring wells upgradient and downgradient) have

been applied to many sites of potential concern from

detection through remedial action selection phases.

This minimal approach has been applied widely, regardless of the physicochemical characteristics of contaminant mixtures or the complexity of hydrogeologic

settings. With soluble inorganic constituents, this approach may be adequate for detection purposes, but

assessment efforts require substantially more comprehensive approaches. For organic contaminant assessment efforts (i.e., determinations of the nature and

extent of contamination), wells alone have been found

to be inadequate monitoring tools. Recognition of the

value of subsurface soil vapor surveys for volatile

organic components of fuel and solvent mixtures has

generated a flurry of modified site characterization

approaches based on monitoring wells (4). These

approaches to site characterization and monitoring network design suffer also from a failure to identify the total

mass of contaminant in the subsurface.

This failure occurs for three main reasons. First, although volatile organic compounds (VOCs) are mobile

in ground water and are frequently early indicators of

plume movement (5), their detection in vapor or well

samples and their apparent aqueous concentration distribution do not identify the total mass distribution of

organic contaminant (6). Second, efforts to correlate

observed soil vapor or ground-water VOC concentrations with those in subsurface solid cores have often

been unsuccessful, because current bulk jar collection/refrigeration at 4o

C guidelines for solid core samples for VOC analyses lead to gross negative errors (7).

Third, “snapshots” (i.e., one-time surveys) of background and disturbed ground-water chemistry conditions have been interpreted as “constant,” ignoring

temporal variability in subsurface geochemistry.

The unhappy result of the slow improvement in site

characterization and monitoring practices has often

been the very low probability of detecting the source of

mobile organic contaminants. This outcome may be

followed by the misapplication of risk assessment or

remediation models and fiscal resources. Nonetheless,

good reasons exist for a more optimistic view of the

20
future reliability of site-characterization and monitoring

efforts.

The shortcomings of previous contaminant detection

and assessment efforts have been recognized. New

guidelines and recommendations for network design

and operations will lead to more comprehensive, costeffective site characterization (7,8) in general. Also, excellent reviews of characterization and long-term

monitoring needs and approaches in support of in situ

remediation efforts should guide us in this regard (9,10).

Site characterization efforts provide a basis for longterm monitoring design and actually continue throughout

the life of a remediation project.

Advanced Site Characterization and

Monitoring

How do we estimate the potential for subsurface intrinsic

bioremediation success and track its performance into

the future? Clearly, we should seek to design technically

defensible characterization and monitoring networks

that will provide reasonable estimates of in-place contaminant distributions over time. Therefore, a dynamic,

ongoing site-characterization effort includes the following objectives:

• Identify the spatial distribution of contaminants, particularly their relative fractionation in subsurface solids,

water, and vapor, along potential exposure pathways,

recognizing that the mass of contaminants frequently

resides in the solids.

• Determine the corresponding spatial distribution of total

reactive organic matter (e.g., degradable normal, aliphatic, and aromatic hydrocarbon compounds), because overall microbial activity and disruptions in

subsurface geochemical conditions (and bioremediation indicators) are due to the total mass of reactive

organic carbon.

• Estimate the temporal stability of hydrogeologic and

geochemical conditions that may favor microbial

transformations in background, source, and downgradient zones during the first year of characterization and monitoring.

• Derive initial estimates of net microbial transformations of contaminant-related organic matter over time

that may be built into an efficient long-term monitoring

network design.

The first three objectives establish the environment of

major contamination and the conditions under which

bioremediation may occur. The latter two objectives are

vitally important, because evaluating the progress of

intrinsic bioremediation processes depends on distinguishing compound “losses” due to dilution, sorption,

and chemical reactions from microbial transformations.

This approach has been suggested emphatically by

Wilson (9) and was recently developed into a draft technical U.S. Air Force (USAF) protocol by Wiedemeier et

al. (10).

The latter reference focuses directly on the implementation of intrinsic remediation for dissolved fuel contamination in ground water. The general approach is shown

in Figure 1, which has been modified from the original

work. The draft USAF protocol (10) has as its goals the

collection of data necessary to support:

• Documented loss of contaminants at the field scale.

• The use of chemical analytical data in mass balance

calculations.

• Laboratory microcosm studies using aquifer samples

collected from the site.

These data, if collected in three dimensions for an extended period, should be sufficient to implement intrinsic

remediation successfully (11). The data collected in the

initial site characterization effort (Figure 1) support the

development of a site-specific conceptual model. This

model is a three-dimensional representation of the

ground-water flow and transport fields based on geologic, hydrologic, climatologic, and geochemical data for

a site. The conceptual model, in turn, can be tested,

refined, and used to determine the suitability of intrinsic

remediation as a risk-management strategy. The validity

of the conceptual model as a decision tool depends on

the complexity of the actual hydrogeologic setting and

contaminant distributions relative to the completeness

of the characterization database. The draft USAF protocol is quite comprehensive in identifying important parameters, inputs, and procedures for data collection and

analysis. The major categories of necessary data are

listed in Table 1 from the draft protocol (10). Ongoing

work on the protocol has revised some of the detailed

guidance it provides on sampling and analytical protocols for these critical parameters; thus, recent drafts of

the protocol should be even more useful to practitioners.

Typical detective monitoring data sets available prior

to in-depth site characterization are more likely to

contain contaminant-related information rather than

the three-dimensional aquifer property, hydrogeologic, or geochemical data needed to formulate a

conceptual model. A recognition of the variability

inherent in these parameter distributions is critical to

site characterization efforts.

Sampling in Space

The initial site characterization phase should be designed to provide spatially dense coverage of critical

data over volumes corresponding to 10-yr to 100-yr

travel times along ground-water flow paths. If the flow

path intersects a discharge zone in less than 100 yr,

then the volume should be scaled accordingly. For

21
Evaluate Use of


Other Remedial


Options in


Conjunction With


Intrinsic Remediation

Free Product


Recovery Bioventing

Pump


and


Treat

Air


Sparging

Bioslurping Reactive


Barrier

Enhanced


Bioremediation

Assess Potential for


Intrinsic Remediation


With Remediation


System Installed

Refine Conceptual Model and


Complete Premodeling


Calculations

Model Intrinsic Remediation


Combined With Remedial


Option Selected Above


Using Numerical Models

Use Results of Modeling and


Site-Specific Information in


an Exposure Assessment

Is There


Unacceptable


Risk to Potential


Receptors


?

Yes Yes

Site Point-of-Compliance


Monitoring Wells and


Prepare Long-Term


Monitoring Plan

Present Findings


and Long-Term


Monitoring Plan to


Regulatory Agencies


and Reach Agreement


on Monitoring Strategy

No No

Examine Temporal/


Spatial Distributions

Prepare Refined


Monitoring Plan

Is There


Unacceptable


Risk to Potential


Receptors


?

Document Occurrence of


Intrinsic Remediation and


Model Intrinsic Remediation


Using Numerical Models

Use Results of Modeling and


Site-Specific Information in


an Exposure Assessment

Perform Site Characterization


in Support of Intrinsic Remediation

Refine Conceptual Model and


Complete Premodeling


Calculations

Make Preliminary


Assessment of Potential


for Intrinsic Remediation


Based on Existing Site


Characterization Data

- Contaminant Type


 and Distribution


- Hydrogeology


- Location of Receptors

Develop Preliminary


Conceptual Model

Review Available


Site Data

Figure 1. Intrinsic remediation flow chart.

22
example, if the flow path discharges after 10 yr, the

critical volume would be 1 yr of travel time. The “volumeaveraged” values of the contaminants, hydrogeological

and geochemical parameters within zones along the

flow path(s), should be derived from data sets that are

large enough to permit estimation of statistical properties (e.g., mean, median, correlation distance, and variance). In general, this means that the data sets for

derived mass loadings of contaminants, aquifer properties, and geochemical constituents (Table 2) derived

from spatial averages of data points must include

approximately 30 or more data points (12-14). Indeed,

this minimum data-set size strictly applies to points in

a plane.

Table 1. Site-Specific Parameters To Be Determined During Site Characterization (10)

Fractionation and Spatial Extent

of Contamination

Extent and type of soil and ground-water contamination

• Location and extent of contaminant source area(s) (i.e., areas containing free- or 

 residual-phase product)

• Potential for a continuing source due to leaking tanks or pipelines

Hydrogeologic and Geochemical

Framework

Ground-water geochemical parameter distributions (Table 2)

• Regional hydrogeology, including:

 - Drinking water aquifers

 - Regional confining units

• Local and site-specific hydrogeology, including:

 - Local drinking water aquifers

 - Location of industrial, agricultural, and domestic water wells

 - Patterns of aquifer use

 - Lithology

 - Site stratigraphy, including identification of transmissive and nontransmissive units

 - Grain-size distribution (sand versus silt versus clay)

 - Aquifer hydraulic conductivity determination and estimates from grain-size distributions

 - Ground-water hydraulic information

 - Preferential flow paths

 - Location and type of surface water bodies

 - Areas of local ground-water recharge and discharge

• Definition of potential exposure pathways and receptors

Table 2. Target Constituents for Site Characterization in Support of Intrinsic Bioremediation

Contamination

Area

Apparent/Geochemical Redox Zone Contaminant Mixture Inorganic Constituents Intrinsic Constituents

Source Reducing Fuels

Chlorinated solvents

O2, CO2, H2S; pH, Fe2+,

HS-

/S=, NO2-, NH3,

alkalinity

Organic carbons, CH4,

organic acids, phenols

As above and:

chlorinated metabolites,

ethylene, ethane

Downgradient Transitional/Suboxic Fuels

Chlorinated solvents

O2, CO2, H2S; pH, Fe2+,

alkalinity, NO2-, NO3-, 

NH3, HS-

/S=

Organic carbon, CH4,

organic acids, phenols

As above and:

chlorinated metabolites,

ethylene, ethane

Upgradient/Far-field

downgradient

Oxic Fuels

Chlorinated solvents

O2, CO2, H2S; 

alkalinity, Fe2+, NO3-,

NO2-, NH3

Organic carbon, CH4,

organic acids, phenols

As above and:

chlorinated metabolites,

ethylene, ethane

23
Two major decisions must be made with regard to how

spatially averaged masses of contaminants, electron

donors (e.g., organic carbon, Fe2+, S=, and NH3), and

electron acceptors (e.g., O2, NO3-, NO2-, Fe and Mn

oxides, and SO4=) are to be estimated.

The first question deals with identification of the media

in which the bulk of the constituent’s mass resides. For

aquifer properties (e.g., grain size and laboratory estimates of hydraulic conductivity), the answer is simple.

In this case, the solids are clearly the media of interest.

For constituents, particularly VOCs, which are sparingly

water soluble, the bulk of the mass may in fact reside in

the solids, though both solids and water samples must

be collected carefully.

The second question pertains to the depth interval over

which “planar” data points might be averaged. With

fuel-related aromatic contaminants, the depth interval

above and below the capillary fringe/water table interface typically exhibits order-of-magnitude differences in

solid-associated concentrations. In this situation, averaging data points over depths greater than 0.5 m could

easily lead to order-of-magnitude errors in estimated

masses for a site. Continuous coring of subsurface solids and close interval (i.e., <1 m) sampling of water

should be considered in many VOC investigations. To

approach this level of depth detail in sampling, “push”

technologies and/or multilevel sampling devices present

very useful tools for site characterization. Push technologies rely on hydraulic or hammer-driven, narrow

diameter (i.e., <2 in.) probes for solid or water sampling.

These technologies have the potential to provide greater

spatial coverage of the subsurface at less cost than

drilling techniques.

The approach to site characterization for chlorinated

hydrocarbons is significantly more difficult. Very few

models of site characterization for these contaminants

have estimated mass loadings in specific media. Many

of the previously referenced methods may work satisfactorily. Free-phase detection, assessment, and quantitation, however, may be more a matter of luck and

exhaustive sampling than intuition based on experience.

Sampling Over Time

VOC compounds (e.g., aromatic hydrocarbons and

chlorinated solvents) are among the target contaminants that are considered constituents of concern in

remedial investigations. Their aqueous solubility and

demonstrated association with aquifer solids require

sampling of these media during the site characterization

phase. This suggestion also applies to organic metabolites of complex organic mixtures (e.g., ethylene, vinyl

chloride, aromatic acids, and phenols). Aqueous plumes

that develop subsequent to the release of these organic

mixtures and byproduct compounds have received the

most attention in the past. The fact that the mass of

these contaminants frequently resides in the solids

strongly suggests that the solids should receive the

most attention in the initial site characterization effort.

This should also be the case for the physical, geochemical, and microbial determinations.

Initially, conventional nested monitoring wells with

screened lengths of 1 m or more will be useful for

estimating the spatial extent of the dissolved plume, for

delineating apparent geochemical zones, and for providing data on water level and aquifer property (e.g., slugand pump-test derived hydraulic conductivity estimates). Semiannual or annual sampling of wells,

particularly multilevels appropriately designed and completed, should be quite useful over the course of the

long-term monitoring program. Sampling should track

the downgradient progress of risk-associated target

compounds and permit testing predictions of intrinsic

bioremediation effects on risk reduction.

Proof of the effects of the net removal of specific solidassociated contaminants due to intrinsic bioremediation,

however, will depend on solid sampling and analysis at

annual or greater intervals, because solid-associated

concentrations may be expected to change slowly. Unless biotransformation can be shown to be a major loss

mechanism for contaminants mainly in solids over extended periods, it will remain an area of research rather

than practice.

Because very few contamination situations have been

monitored intensively for periods exceeding several

years, it is difficult to define specific sampling frequencies for the range of hydrogeologic and contaminant

combinations that may be encountered. The adoption

and future refinement of recently developed, technically

defensible protocols will improve intrinsic remediation

approaches to risk management in subsurface contamination situations.

Acknowledgments

The author would like to express his gratitude to the

following individuals who aided in the preparation of the

manuscript: Dr. Gary Robbins, Mr. Todd H. Wiedemeier,

Dr. John T. Wilson, Dr. Fran Kramer, and Ms. Rebecca

Mullin.

References

1. Scalf, M.R., J.F. McNabb, W.J. Dunlop, R.L. Cosby,

and J.S. Fryberger. 1981. Manual of ground-water

sampling procedures. National Water Well Association. 

2. Barcelona, J.J., J.P. Gibb, J.A. Helfrich, and E.E.

Garske. 1985. Practical guide for ground-water

sampling. Illinois State Water Survey, SWS Contract Report 374 U.S. Environmental Protection

Agency, Ada, OK. 

24
3. U.S. EPA. 1986. RCRA technical enforcement guidance document, OSWER-9950.1. Washington, DC.

4. Eklund, B. 1985. Detection of hydrocarbons in

ground water by analysis of shallow soil gas/vapor.

API Publication No. 4394. Washington, DC.

5. Plumb, R.H. 1987. A comparison of ground-water

monitoring data from CERCLA and RCRA sites.

Ground Water Monitor. Rev. 7:94-100.

6. Robbins, G.A. 1989. Influence of using purged and

partially penetrating wells on contaminant detection, mapping, and modeling. Ground Water 2:155-

162.

7. U.S. EPA. 1992. RCRA ground-water monitoring:

Draft technical guidance document. EPA/530/R93/001. Washington, DC.

8. U.S. EPA. 1994. Proceedings of the Ground Water

Sampling Workshop, Dallas, TX, December 8-10,

1993. U.S. Environmental Protection Agency, Ada,

OK. 

9. Wilson, J.T. 1993. Testing bioremediation in the field

In: National Research Council. In situ bioremediation—

when does it work? Washington, DC: National

Academy Press. pp. 160-184. 

10. Wiedemeier T.H., D.C. Downey, J.T. Wilson, D.H.

Kampbell, R.N. Miller, and J.E. Hansen. 1994. Draft

technical protocol for implementing the intrinsic remediation (natural attenuation) with long-term monitoring option for dissolved-phase fuel contamination

in ground water. Air Force Center for Environmental

Excellence, Brooks AFB, San Antonio, TX (March).

11. National Research Council. 1993. In situ bioremediation—when does it work? Washington, DC: National Academy Press. 

12. Journel, A.G. 1986. Geostatistics: Models and tools

for the earth sciences. Math. Geol. 18:119-140.

13. Hoeksema, R.J., and P.K. Kitanidis. 1985. Analysis

of the spatial structure of properties of selected

aquifers. Water Resour. Res. 21:563-572.

14. Gilbert, R.O., and J.C. Simpson. 1985. Kriging for

estimating spatial patterns of contaminants: Potential and problems. Environ. Monitor. Assess. 5:113-

135.

25
Processes Controlling the Distribution of Oil, Air, and Water

John L. Wilson

Department of Geoscience, New Mexico Institute of Mining and Technology,

Socorro, NM

Abstract

Oils and other nonaqueous-phase liquids (NAPLs) are

a major source of dissolved contamination in aquifer

systems. Three major forces control the movement and

distribution of NAPLs, as well as air and water, in both

porous and fractured media. These are capillary, viscous, and buoyancy forces. These forces interact with

the complex pattern of geological features, making fluid

behavior difficult to predict and fluid distribution difficult

to characterize. Intrinsic bioremediation within a NAPLcontaminated zone is probably of limited effectiveness

because of mass transfer limitations, toxicity concerns,

and the limited availability of other nutrients. Thus, intrinsic bioremediation is aimed at the contamination in

the downgradient aqueous phase of the dissolved

plume. Evaluating the effectiveness of bioremediation is

impossible within the plume without knowing the distribution of the NAPL—the plume’s source—or determining whether that source is still moving.

Introduction

Oil or another NAPL, such as gasoline or trichloroethylene (TCE), may be released at or near the

ground surface. These liquids are primary sources for

dissolved contaminant plumes in ground-water systems.

Even if the NAPL has ceased to move, trapped by

capillary forces as we describe below, it remains a longterm source of dissolved contamination. The limited

aqueous solubility of the chemicals composing these

oily liquids implies that even in small volumes they can

lead to ground-water plumes of enormous dimensions.

Intrinsic bioremediation is aimed at the contamination in

the aqueous-phase plume, for reasons that will become

clear. The plume location and concentrations depend on

understanding the spatial pattern of its source, the

NAPL. In the vadose zone, the problem is complicated

by the presence of a third phase, air, and the propensity

of the NAPL to spread at the air-water interface and to

volatilize. 

In this review, we first describe the three major forces

controlling the movement and distribution of fluids in the

subsurface, using natural processes such as infiltration

to illustrate. We then add NAPLs, to relate the discussion to the contamination issue, and aquifer heterogeneity, to relate the discussion to natural hydrogeological

systems. Finally, we discuss the implications for intrinsic

bioremediation. Many of these issues are illustrated with

photomicrographs taken of appropriate processes in situ.

Three Major Forces That Control

Processes

Ground-water systems are composed of porous and/or

fractured aquifer material containing water in the spaces

between the solids (see Figure 1a). Above the water

table, in the vadose zone (see Figure 2), air is also

present within this pore space (see Figure 1c). Three

major forces control both the movement and distribution

of each of the fluid phases: capillary forces, viscous

forces, and gravity or buoyancy forces (1).

Capillarity is the result of the cohesive forces within each

fluid phase and the adhesive forces between the solid

phase and each of the fluids. A capillary force is proportional to the interfacial tension at the fluid-fluid interface

Water Air Solid

Key:

(a) Water Saturated (b) Residual Gas (c) Continuous Gas

Figure 1. Diagram of fluid saturation in a porous media (4).

26
and the strength of fluid wetting to the solid surface, and

inversely proportional to the pore size. In the vadose

zone, with air and water present, interfacial tension is

the same as air-water surface tension. In the saturated

zone, beneath the water table (see Figure 2), the only

fluid is water, and capillary forces are absent (see Figure

1a). The exception occurs beneath a fluctuating water

table, where gas bubbles may become entrapped by

capillary forces (see Figure 1b and Figure 3; Figure 3 is

a photomicrograph from a visualization experiment [see

also the appendix to this paper]). For most aquifer materials, water is the wetting fluid; that is, the solid aquifer

material has a greater affinity for water than for air, and

the air or gas is almost always the nonwetting phase

(see Figures 1b and 1c). Water occurs as an interconnected film or layer of wetting liquid covering and connecting the solids in the vadose zone. The nonwetting

gas phase occupies the larger pores. 

Viscous or flowing forces within a fluid phase require an

expenditure of potential energy. For example, in the

saturated zone the ground-water flow rate is proportional to the hydraulic gradient. The flow rate is also a

function of the aquifer material and the structure of its

pore space, as represented by the effective permeability. If more than one fluid phase is present, as in Figures

1b and 1c, the interconnected paths in each phase are

more tortuous. The permeability of each phase is reduced, with a relative permeability closer to zero than to

one, the value that applies when only a single fluid

occupies the pore space.

Buoyancy is a gravitational force proportional to the

density difference between two fluids. The gas phase

has a much lower density than water, so that gravity

(buoyancy) forces play a significant role in the vadose

zone. Water infiltrates downward, toward the water table, primarily under the influence of gravity. As it moves

downward, water easily displaces the less dense and

less viscous air phase. In the saturated zone, with only

water present, gravity usually plays no direct role

(although, of course, it ultimately drives the hydraulic

gradient). In the saturated zone, if the chemical concentration varies enough to influence water density, gravity

can again play a direct role, causing the water and its

chemicals to move. 

Oil and Other NAPLs

A third fluid is often present at many hazardous waste

sites and most leaking underground storage tanks.

NAPLs often share the pore space with both gas and

Hazardous Waste Site Ground


Surface

Vadose


Zone

Leaking Tank

Vapor


 Phase


 Organic

Capillary


Fringe

Water Table

Residual


NAPL


Saturation

Dense NAPL

Dissolved


Organic

Rock

Ambient


Flow Ambient


Flow

Saturated


Zone

Floating NAPL

Figure 2. Diagram of the saturated and vadose zones, showing the migration pattern for a NAPL more dense than water (left), and

less dense than water (right) (3).

27
water, in the vadose zone, or water alone, in the saturated zone. The NAPL may be more dense than water

(a dense NAPL, or DNAPL), such as TCE, or it may be

lighter than water, such as gasoline. In either case, a

NAPL is heavier than air, so it easily moves downward

through the vadose zone. Once it reaches the vicinity of

the water table, it may continue to move deeper if it is

denser than water (DNAPL), or it may remain in the

vicinity of the water table if it is lighter, as illustrated in

Figure 2 (2,3); in the latter case, the term “water table”

begins to lose its meaning. In either case, the NAPL

leaves behind a trail of residual as it moves. 

Most NAPLs appear to be of intermediate wettability in

typical aquifer materials; that is, they are nonwetting

relative to water but wetting relative to gas. Thus, depending on whether it encounters gas or water, a NAPL

can then display either wetting or nonwetting behavior

or both. Many NAPLs, such as gasoline, also have low

internal cohesion and will spread at a gas-water interface, presumably forming a film between the water

phase, which because of capillarity preferentially occupies the smallest pores, and the gas phase, which preferentially fills the largest pores. In the vadose zone, this

film interconnects the pockets of NAPL, which even at

residual saturation should be largely continuous, as

shown in Figure 4. Some NAPLs, such as TCE, have

more internal cohesion and do not spread. In this case,

the sum of the interfacial tensions between the NAPL

and the water and air, σow + σao, exceed the surface

tension between the gas and water, σaw (see Figure 5,

theta>0). These NAPLs will not spread as films. On a

flat water surface (the gas-water interface), nonspreading liquids tend to coalesce into lenses that float on the

surface (much as depicted in Figure 5), even though

many of these liquids are denser than water. In porous

media, this leads to complications, as shown in Figure 6.

In the saturated zone, the NAPL fills the larger pores,

while water occupies the smaller pores and lies as a film

between the NAPL and the solid surfaces. After the

NAPL moves on, it leaves behind a residual saturation

trapped by capillary forces, which occupies the larger

pores. NAPL blobs occupying one or two pore bodies

are shown in Figure 7. A similar process occurs in the

vadose zone, when a rising water table entraps air

bubbles, as shown in Figures 2b and 3. Because of

lower interfacial tension, NAPL blobs are often much

more complex than air bubbles, forming interconnected

groups of larger pore bodies and pore throats, as shown

in Figure 8.

So far we have looked at water-wet porous media.

Wetting, however, may change for a variety of reasons,

including the adsorption at hazardous waste sites of

(polar) organic compounds onto the solid. Figure 9

shows the saturated zone residual in a pore space with

Figure 3. Photograph of a micromodel with entrapped air bubbles. The bubbles are trapped in the pore bodies,

which are roughly 200 μm in diameter, while the pore

throats are 50 μm in diameter (4). 

Figure 4. A micromodel after a spreading NAPL (Soltrol 130) is

drained with air. Air fills the pore bodies and is interconnected through some of the pore throats. Water

fills the other pore throats and some of the pore

wedges and is a wetting film everywhere else. The

NAPL forms a thick film filling some pore wedges and

surrounding the air everywhere. This ubiquitous and

interconnected film is particularly thick near the

water-filled pore throats (3).

NAPL


Lens

Gas (air)

Solid

Water

Interface

σaw

σao

σow

θ

Figure 5. Cross-sectional diagram of spreading potential for a

drop of NAPL floating on the air (gas)-water interface,

after Adamson (10) and others. The water is wetting,

the gas is nonwetting, and the NAPL is intermediate

wetting.

28
altered wetting and significant wetting hysteresis. The

residual NAPL is not trapped in the pore bodies, the

characteristic location for a nonwetting fluid. Some of it

is in the pore throats, some in the pore bodies, some

has an end in both. Because of the large wetting hysteresis of this surface, water wet conditions prevailed during the invasion of the NAPL, while intermediate wet

conditions prevailed when the water reentered to trap

the NAPL. The NAPL-water interfaces in the figure

clearly show evidence of both histories. In some locations, clearly water wet, we can infer that the interface

was not disturbed as the water reentered. In other locations, the contact angle shows the recent displacement

of the NAPL by water.

Heterogeneity

All of this is complicated by the geology of the groundwater system. The material composing an aquifer is

always heterogeneous. For example, fluvial-deposited

materials contain sand, gravel, and clay in a complex

geometric pattern of geologic facies. Many formations,

even those containing clay, are fractured and sometimes

faulted. Heterogeneities provide preferential paths for

fluid and chemical migration in an interplay of the geology with the forces of capillarity, viscosity, and gravity.

For example, a DNAPL reaching the saturated zone will

tend to seek out the heterogeneities with larger pore

spaces such as the coarser sands or, more insidiously,

fractures where its movement will be hard to trace.

Heavier than water, it tends to move farther downward

into the aquifer, fingering out the bottom of a sand lens

or moving snakelike down the inside of a fracture.

Figure 6. A micromodel after a nonspreading DNAPL (mineral

oil) is drained with air. The DNAPL is not connected.

As it leaves the photo at the bottom, the air is surrounded by a film of oil, similar to that seen with the

spreading oil. As the air leaves to the left, the film no

longer exists; it truncates somewhere in between. In

this area, the air-water interface is dimpled, with small

lenses of mineral oil. These lenses are not interconnected. The various pockets of oil, such as on the

right side of the figure, are also generally discontinuous and not connected (11).

Figure 7. A micromodel with a singlet and doublet NAPL blob

(dark gray) in a water-filled pore space (lighter gray)

of an irregular pore network. The pore bodies are up

to 1,000 μm across, and the model was 10 cm x 15 cm

in area (8).

Figure 8. A micromodel with a large, branched, nonaqueous,

nonwetting phase blob (dark gray) in an irregular pore

network (3).

Figure 9. Residual nonaqueous-phase saturation in a micromodel treated with an alkoxysilane of proprietary

composition, GlassClad 18 (Huls Americal Petrarch,

Bristol, PA). This treated surface has significant contact angle hysteresis. The water is light gray, and the

NAPL is darker gray in this image (9).

29
Figure 10 shows a pore system with embedded zones

of larger pores surrounded by a network of smaller

pores. This simulates lenses of coarse porous materials

(e.g., coarse sand) embedded in a matrix of finer material (e.g., fine sand). The NAPL invaded from the side,

mimicking the lateral migration of a NAPL plume along

the bottom of an aquifer. Afterwards, the NAPL was

displaced by water, leaving behind a residual saturation.

The photograph in Figure 10 was taken at this time. The

residual is dominated by the coarse lenses. In the fine

material, we see the typical single- and multiple-pore body

blobs, but the coarse lenses have been completely bypassed. Capillary forces are too strong, and the viscous

forces in the moving water too weak, to displace the NAPL

from these isolated coarser zones. Note that the zones

with higher saturated permeability trap the NAPL.

This experiment was repeated at a much higher water

flow rate. A photograph taken while this was occurring

is shown in Figure 11. A significant amount of NAPL has

been swept from the coarse lenses, even though the

initial condition was essentially the same. In the swept

area, on the right side, there is still some bypassed

NAPL, but it is now located on the downstream end of

the lenses. Sufficient viscous forces were generated to

overcome the capillary forces that held NAPL in the

coarse lenses. The displacement is incomplete because, as the wetting front reaches the end of a lens, it

closes together and surrounds the lens. At that point, the

nonwetting NAPL in the lens is no longer connected with

the downstream NAPL. These two photographs suggest

that in heterogeneous systems the residual nonwettingphase saturation is a function of the structure of the

heterogeneities and the fluid history. Because the residual is the source of the dissolved plume, we would like

to be able to predict its location and magnitude from

knowledge of the geology and the fluid flow history. We

use mathematical models to assist with this kind of

prediction. There is a problem, however. The current

generation of multiphase flow models assumes that

there is a single value of residual and, worse, that we

know what it is. These models offer little assistance with

the important issue of predicting the amount and distribution of residual.

These photographs were taken from experiments that

simulated the movement of DNAPL parallel to stratigraphic bedding planes—for instance, as it migrates

along the bottom of an aquifer. What happens as it

moves perpendicular to the bedding, for example, during its movement downward from the water table toward

the bottom of the aquifer? Figure 12 shows the same

kind of heterogeneous lenses, except for this change in

geometry. The flow is downward, across the lenses, with

both gravity and viscous components. As the DNAPL

encounters its first coarse lens, it tends to spread horizontally because of capillary forces. As soon as it fills a

lens, it begins to spill downward, but because of the

gravity instabilities—DNAPL is heavier than water—this

movement occurs as fingers on the scale of a few pores.

When one of these fingers encounters another, deeper

coarse lens, the downward migration of the DNAPL is

again arrested, until the lens is filled. In essence, the

lenses promote horizontal spreading, here confined by

the lateral boundaries of the model, and vertical fingering. The process leads to a very inefficient displacement, with lots of water bypassed by the DNAPL, and

illustrates some of the reasons why it is difficult to directly measure DNAPLs in the field, predict their behavior, or design effective remediation measures.

Influence on Intrinsic Bioremediation

The distribution of fluid and solid phases controls the

mass transfer and transport of chemicals and their

Figure 10. Water has swept through a heterogeneous micromodel containing a NAPL as a nonwetting fluid and

water as a residual wetting fluid. The nonwetting

NAPL in the isolated coarse-grained heterogeneities

has been bypassed (3).

Figure 11. Water is displacing the NAPL at a higher flow rate

through the same heterogeneous micromodel. The

flow is from right to left. With larger viscous forces,

the displacement is more efficient, and less NAPL is

bypassed. This photograph was taken while the displacement was still under way. The bypassing is

complete on the right side of the model and has not

yet occurred at the left side (3).

30
availability for biotransformation by the microbial population. For all practical purposes, biotransformation

takes place in the aqueous phase or in contact with the

aqueous phase. Bacteria colonize surfaces. In watersaturated media, the only surfaces available are the

solids, but in the vadose zone and in NAPL-contaminated areas there are also fluid-fluid (e.g., air-water,

oil-water) interfaces that some bacteria will colonize.

These colonies develop in part due to the same interfacial forces that attract abiotic colloids to solid-liquid or

fluid-fluid interfaces. Figure 13 shows fluorescing colloids that are repulsed from the solid surface by electrostatic forces but that are attracted to an air-water

interface by hydrophobic and capillary forces. Similar

observations have been made for microorganisms (4,

5). Other important processes also lead to bacterial

attachment and colonization at interfaces that are

unique to living bacteria, with their complex surface

chemistry and ability to produce extracellular polymers.

Where NAPLs are concerned, there may also be toxicity

effects repelling the bacteria, or, if the NAPL is an important nutrient source, it may be an attractant. Figure

14 shows a bacteria colony in the vicinity of a blob of

iso-octane. The bacteria have colonized the solid surface

and the surface of the blob, and a few were even observed

to be living in the iso-octane. In a similar experiment with

toluene and a toluene-degrading bacteria, the bacteria

found high concentrations of toluene toxic and set up

housekeeping at a distance that allowed ambient flow and

mixing to dilute the concentrations many times.

The geometry of the fluid distributions have other indirect influences on intrinsic bioremediation. Consider

mass transfer of chemical components between a multicomponent residual NAPL and the water phase in the

saturated zone. Many spilled NAPLs—from gasoline

and other fossil fuels to transformer oils containing polychlorinated biphenyls (PCBs), as well as the mixed bag

of organics that are sometimes found in industrial waste

pits—are mixtures having many (sometimes hundreds

of) organic chemical components. The more soluble

components of these mixtures can dissolve comparatively quickly, leaving behind less soluble components

to leach out more slowly.

There are also cosolvency effects to consider. Interaction between components can either enhance the solubility of a given component (cosolvency) or reduce the

solubility of that component (by a kind of salting-out

process). Capillary-trapped residual blob size and shape

influence the partitioning of NAPL components to the

aqueous phase. Mass transfer coefficients used in the

mathematical models of this partitioning often employ

Figure 12. DNAPL is moving downward through a watersaturated heterogeneous micromodel (9).

Figure 13. Negatively charged hydrophilic polystyrene particles attached to the surface of gas bubbles entrapped in a micromodel. There is almost no

sorption on the glass surface at this ionic strength

(4).

Figure 14. A colony of bacteria (Arthrobacter sp. ZAL001) in

the vicinity of an iso-octane blob after 48 hr of

growth. The blob is trapped in a pore body. The

bacteria are light gray (5). 

31
the analogy of an equivalent spherical blob. Certainly

singlet blobs can be represented by a spherical model,

but it is less clear that this model works for the tortuous

multiple pore-body elongated or branched blobs shown

in Figure 8, or in the presence of heterogeneity shown

in Figure 10. Large complex blobs hold the majority of

the NAPL volume, and mass transfer from them to the

aqueous phase is clearly rate limited (6-8); it can take

decades under natural conditions. Intrinsic bioremediation can do little to speed up this mass-transfer process,

except to possibly increase concentration gradients between the NAPL and the aqueous phase, or perhaps

increase the solubility of chemical components through

the production of biosurfactants. 

Even in the absence of toxicity effects and mass-transfer

limitations, there may be little intrinsic bioremediation

within the NAPL zone. The aqueous-phase flow field is

complicated by the presence of the NAPL. Among other

things, the effective permeability to water is lower. Consequently, it is difficult to supply the nutrients necessary

for growth (3,8). 

For these reasons, intrinsic bioremediation of the NAPL

itself may be of limited value. If instead we consider the

downgradient plume of dissolved contamination, the basic concern with the NAPL is as a source of that contamination. 

To understand the plume, we must know something

about the spacial distribution of the source and determine whether or not it is still moving. This both locates

the source and controls the effective mass-transfer rate

between the source and the ground water. If the source

is a NAPL, it is obvious from some of the earlier photographs that even a small amount of liquid may be spread

throughout a relatively large volume of aquifer material,

and in a most complex pattern (see Figures 11 and 12).

Characterization of this distribution and movement is

difficult and may be beyond the current state of the art.

Summary

Three major forces control the movement and distribution

of fluids in the vadose and saturated zones: capillary,

viscous, and buoyancy forces. These forces interact

with the complex pattern of geological features, making

fluid behavior difficult to predict and fluid distribution

difficult to characterize. Wetting fluids have an affinity for

the solid surfaces of aquifer materials. In general,

NAPLs are wetting relative to air, but nonwetting relative

to water. Consequently, they tend to fill the larger pores

of the saturated zone, leading to entrapment in what are

usually regarded as high-permeability materials: the

coarse sand but not the fine silt. This tendency, together

with the propensity for gravity fingering through the finer

materials, causes a complex pattern of migration and

distribution in the saturated zone. In the vadose zone,

behavior is further complicated by the presence of air

and the question of whether or not the NAPL tends to

spread at the air-water interface. Wettability can be

altered by sorption processes, changing the relative

roles of water and the NAPLs and leading to new pathways and fluid distributions. 

Intrinsic bioremediation within a NAPL-contaminated

zone is probably of limited effectiveness. Mass-transfer

limitations constrain the rate at which dissolved components reach the aqueous phase. There are toxicity concerns with the high concentrations that are found nearer

the NAPL blobs or films. Some other nutrients are only

limited in availability because of the more complex flow

patterns that water must take in this region. 

Consequently, intrinsic bioremediation should probably

be aimed at the contamination in the downgradient

aqueous phase of the dissolved plume. It is impossible

to evaluate the effectiveness of bioremediation within

the plume without knowing the answer to two questions:

Where is the source? and, How much contamination is

entering the flowing ground water from the source? The

answer to both questions requires that we know the

distribution of the NAPL—the plume’s source—and determine whether that source is still moving. It is a direct

answer to the “where” question. The “how much” question depends on the effective mass-transfer rate between the source and the passing ground water. The

trouble is that, even in a geologically simple aquifer, a

small amount of nonaqueous contamination may be

spread throughout a relatively large volume of aquifer

material, and in the most complex pattern. Characterization of this distribution and movement is difficult

and may be beyond the current state of the art.

Appendix: Visualization of Processes in

Micromodels

Micromodels are transparent physical models of a pore

space network, created by etching a pattern onto two

glass plates which are then fused together (3,8,9). The

resulting pores have complex three-dimensional structures, although the network is only two dimensional. The

micromodel in Figure 3 shows pore bodies connected

together by narrower pore throats. The fused glass located in between these connected channels represents

the solid material in this model of a porous media. Glass

is also at the top of the channels, toward the viewer, and

below the channels. When the pore body contains a

nonwetting fluid, a thin water film is attached to the top

and bottom of the pore body. The film is thin enough to

partially exclude the bacteria from colonizing the top

surface of the iso-octane blob in Figure 14. All of the

photographs in this paper can be found collected together and described in more detail in Wilson (9).

Greater detail is given in the original references.

32
Acknowledgments

The photographs presented in this paper were taken

from work sponsored by the Subsurface Science Program of the Department of Energy, or from work sponsored by the Robert S. Kerr Environmental Research

Laboratory of the U.S. Environmental Protection

Agency. M. Flinsch of New Mexico Institute of Mining

and Technology assisted with the preparation of this

manuscript.

References

1. Bear, J. 1972. Fluid flow in porous media. New York:

Elsevier.

2. Mercer, J.W., and R.M. Cohen. 1990. A review of

immiscible fluids in the subsurface: Properties,

models, characterization and remediation. J. Contam. Hydrol. 6:107-163.

3. Wilson, J.L., S.H. Conrad, E. Hagan, W.R. Mason,

W. Peplinski, and E. Hagan. 1990. Laboratory investigation of residual liquid organics. Report CR813571. U.S. Environmental Protection Agency,

Ada, OK.

4. Wan, J., and J.L. Wilson. 1994. Visualization of the

role of the gas-water interface on the fate and transport of colloids in porous media. Water Resour. Res.

30:11-23.

5. Wan, J., J.L. Wilson, and T.L. Kieft. 1994. The gaswater interface as an influence on the transport of

microorganisms through unsaturated porous media. Appl. Environ. Microbiol. 60:509-516.

6. Miller, C.T., M.M. Poirier-McNeill, and A.S. Mayer.

1990. Dissolution of trapped nonaqueous phase

liquids: Mass transfer characteristics. Water Resour. Res. 26:2783-2796.

7. Powers, S.E., C.O. Louriero, L.M. Abriola, and W.J.

Weber. 1991. Theoretical study of the significance

of nonequilibrium dissolution of NAPLs in subsurface systems. Water Resour. Res. 27:463-478.

8. Conrad, S.H., J.L. Wilson, W.R. Mason, and W.

Peplinski. 1992. Visualization of residual organic

liquid trapped in aquifers. Water Resour. Res.

28:467-478.

9. Wilson, J.L. 1994. Visualization of flow and transport at the pore level. In: Dracos, T.H., and F. Stauffer, eds. Transport and reactive processes in

aquifers. Rotterdam: Balkema. pp. 19-36.

10. Adamson, A.W. 1982. Physical chemistry of surfaces, 4th ed. New York: Wiley and Sons.

11. Wilson, J.L. 1992. Pore scale behavior of spreading

and nonspreading organic liquids in the vadose

zone. In: Weye, U., ed. Subsurface contamination

by immiscible fluids. Rotterdam: Balkema. pp. 107-

114.

33
New Tools To Locate and Characterize Oil Spills in Aquifers

Bruce J. Nielsen

Environics Directorate, Armstrong Laboratory, Tyndall AFB, FL

Abstract

The vision of the Tri-Services (Air Force, Army, and

Navy) scientists is about to become reality, as a partnership between the Department of Defense (DOD), academia, and private industry evolves into a combined

technology that can save millions of dollars in long-term

hazardous waste site cleanup costs.

DOD has about 20,000 contaminated sites, many of

which require characterization. These sites also may

require monitoring for more than 30 years, at a cost of

millions of dollars per year. The cost of site characterization and monitoring can be 25 percent to 30 percent of the total remediation costs.

A quick fix is needed—one that would take a “snapshot”

of the type and amount of contamination, help determine

the best method of remediation, monitor the effectiveness of treatment, and let technicians and engineers

know when the site is clean. The vision was that of a

cost-effective method that would replace the current

time- and labor-intensive processes of site characterization and monitoring. 

To solve the problem, the Tri-Services integrated a laser

spectrometer with fiberoptics with a cone penetrometer

to provide the required “snapshot” of subsurface conditions. To date, this technology has been field-tested at

numerous Air Force, Army, and Navy installations. Ongoing studies will extend its sensitivities; expand its

capability to other contaminants such as solvents, metals, and explosives; and make it more user-friendly for

operating technicians. The end result is a technology

that can significantly reduce the cost of site characterization and monitoring. 

Introduction

The Air Force’s Armstrong Laboratory is executing a

program to demonstrate, test, and evaluate the application of cone penetrometers in support of intrinsic bioremediation (natural attenuation) demonstrations. The Air

Force and its contractors, in cooperation with the U.S.

Environmental Protection Agency’s (EPA’s) Robert S.

Kerr Environmental Research Laboratory, have conducted investigations at Tinker Air Force Base (AFB),

Oklahoma; Plattsburgh AFB, New York; Patrick AFB,

Florida; and Dover AFB, Delaware. One objective of the

investigations was to assess the subsurface conditions,

obtaining parameters for input into BIOPLUME® II, a

numerical model useful for determining the potential for

intrinsic remediation. A second objective was to validate

the cone penetrometer/laser spectrometer system by

soil sampling and analysis. The demonstrations proved

the cone penetrometer’s capabilities to quickly locate

and define the areal and vertical extent of the liquid

oily-phase plume using laser-induced fluorescence

(LIF), and to rapidly install monitoring points and collect

soil samples to provide additional data necessary to

define the dissolved-phase plume.

This paper briefly describes the Tri-Services program for

developing this cone penetrometer technology and recent results of the Air Force program.

Methods

Cone Penetrometry

The Site Characterization and Analysis Penetrometer

System (SCAPS), developed jointly by the Tri-Services,

has proven to be an effective technology for characterizing contaminated DOD sites. The Army has provided leadership on developing SCAPS, including the

concept of using a sapphire window in the cone rod and

of using fiberoptics and spectroscopy to analyze the soil

for contamination. The Army has patented a “Device for

Measuring Reflectance and Fluorescence of In Situ Soil”

and is now licensing it.

The typical cone penetrometer is mounted on a 20-ton

truck and driven to the site requiring characterization,

where a conical rod is hydraulically pushed into the

ground to be characterized. The rod is equipped with a

34
variety of sensors or soil and ground-water sampling

tools. The cone penetrometer can characterize several

aspects of the subsurface, depending on the types of

sensors integrated into the penetrometer. Strain gauges

measure the forces against the tip and sleeve of the

cone tool, allowing determination of soil type (e.g., sand,

silt, clay) and stratification. Other sensors provide electrical resistivity, pore pressure, spectral characteristics,

and other properties of the soil and contamination. The

sensors provide information on hydrogeology and contamination; the samplers verify it. The real-time ability to

receive and assess monitoring data on site without laboratory analysis is critical, facilitating decision-making

during site investigation projects while ensuring accurate and efficient completion of site investigations and

optimization of remedial activities.

Laser Spectrometer System

Each of the services has significant programs for developing and demonstrating laser spectroscopy and other

sensor systems. One of the key components of the cone

penetrometer is a neodymium:yttrium aluminum garnet

(Nd:YAG) laser, which pumps a dye laser system to

induce fluorescence of fuel products as the cone penetrometer probe is advanced into soils. LIF has been

shown to be useful in identifying petroleum, oil, and

lubricant (POL) contamination, such as gasoline and

JP-4 jet fuel. The Armstrong Laboratory’s Environics

Directorate, working with North Dakota State University

(NDSU), has developed a tunable laser/fiberoptic spectrometer system, which uses laser-generated ultraviolet

light, optical fibers, and spectroscopy for hazardous

waste site monitoring. The basic detection approach

takes advantage of the fact that certain substances

fluoresce when a particular wavelength of light shines

on them. The spectral emission, including fluorescent

lifetime, is somewhat like a fingerprint and therefore is

useful in identifying the contaminant. The fluorescent

intensity indicates concentration of the contaminant.

The transportable laser system is unique because its

output may be tuned to the optimum frequency for detecting the pollutants of interest.

Optical fibers are used to transmit ultraviolet light to

subsurface monitoring points and return resulting light

for the spectroscopic analysis. The system can identify

aromatic hydrocarbons such as benzene, toluene, and

xylene (BTX) and naphthalene by their fluorescent spectra. Jet fuel, which contains naphthalene and BTX, is the

most common contaminant at Air Force sites.

This technology can provide semiqualitative and

semiquantitative information, on site, in minutes. The

LIF response can be correlated to the total petroleum

hydrocarbon (TPH) concentration within the soil. The

system has been tested in the field with detection limits

as low as parts-per-million levels on soil when used with

a cone penetrometer and in the laboratory at parts-perbillion levels for BTX in water using fiberoptic probes.

Laser spectroscopy technology could also be used to

monitor the progress of site remediation and to provide

baseline data for intrinsic bioremediation modeling studies.

Technology Transition

Armstrong Laboratory and Unisys Corp. signed a Cooperative Research and Development Agreement (CRADA)

to commercialize the Air Force-developed laser spectrometer system. The laser spectrometer was initially

developed for making ground-water measurements in

monitoring wells or in implanted monitoring points. A

consortia consisting of the CRADA partners, Dakota

Technologies Inc., and NDSU submitted a proposal to

the Advanced Research Projects Agency (ARPA), Technology Reinvestment Project (TRP). ARPA selected the

proposal to receive a 2-year, $1.6 million grant with two

1-year options for follow-on technology development.

The industry partners provide in-kind contributions and

matching funds.

The Rapid Optical Screening Tool (ROST) is the proposed product from the commercialization of the laser

spectrometer developed for the Air Force by NDSU.

ROST will build upon the previous Environics Directorate research by automating the collection and mapping

of data, making equipment components smaller and

more rugged, and developing a more user-friendly interface to allow use by environmental technicians involved

in site characterization and cleanup. ROST also has

potential for process monitoring and for medical diagnostics. Initial commercial use will be with cone penetrometers for soil characterization.

Use of the proposed ROST technology should result in

substantial savings in costs associated with characterization, monitoring, and remediation of hazardous

waste sites. The participants are committed to commercializing the resulting instrumentation for worldwide

sales by U.S. firms or companies. In short, DOD will

benefit from the technology and knowledge gained; the

private sector will receive a highly transferable and profitable technology; the U.S. economy will be helped; and

all will benefit from a cleaner environment.

Combined Technologies

The combined cone penetrometer and transportable laser spectrometer was demonstrated at Air Force installations having fuel-contaminated areas. To date, the

laser spectroscopy system’s primary function with the

cone penetrometer has been to define the oily-phase

35
plume. At Tinker AFB, the tunable laser system was

configured to optimize the system for jet fuel and heating

oil, the known petroleum contaminants. Laboratory

fluorescence spectra from these fuels suggest that

naphthalene produces the maximum fluorescence; consequently, a laser excitation wavelength appropriate for

the known fuels was utilized during the field program.

The system is designed to collect data in two different

modes: “push” or “static.” In the push mode, laser excitation frequency is fixed, and the LIF signal is monitored

as the cone penetrometer probe is advanced. Operation

in the static mode, or with the probe stopped, allows

collection of LIF multidimensional data sets, typically

fluorescence emission wavelength, intensity, and time of

decay matrices (WTM). WTMs have proven to be very

useful in identifying various fuels.

Results

Studies to demonstrate site amenability towards intrinsic

demonstrations were conducted at Plattsburgh, Patrick,

and Dover AFBs. The sites selected for the demonstrations included closed fire-fighter training areas, gasoline

service stations, and aircraft hydrant fueling systems.

The old fire-fighter training areas were composed of

unlined pits used to ignite fuels such as JP-4, waste oils,

and other flammable substances. The service stations

or hydrant fueling systems had leaking underground

storage tanks or piping. Substances from these operations have percolated through the vadose zone into the

unconfined aquifer. The penetrometer determined the

areal extent and volume of oily-phase contamination,

and obtained ground-water and soil core samples. To

determine site amenability, acquired data were then fed

into BIOPLUME® II, a computer model for in situ contaminant biodegradation. The cone penetrometer system rapidly located and defined the leading edge of the

oily-phase petroleum plume. The technology proved

that it can be used to provide timely and accurate data

for intrinsic bioremediation modeling.

The Tri-Services conducted a series of laboratory tests,

and some of the preliminary results are calibration

curves with different fuels on various soil matrices. The

calibration curve obtained in the laboratory for diesel fuel

marine (DFM) on a sand matrix indicates a detection

limit that is lower than 30 mg/kg (ppm) (Figure 1). The

collection of LIF multidimensional data sets (fluorescence emission wavelength, intensity, and lifetime) or

WTMs for diesel #2, JP-4, unleaded gasoline, and diesel

fuel marine show how each one has a characteristic

pattern. These patterns make possible reliable fuel-type

identification without the need for bringing samples to

the surface (Figure 2). In the field, the LIF count measurements can be correlated with collected samples and

analytical results. To assist in the correlation, several

WTMs were conducted at various depths. Color-coded

WTMs from the North Tank Area (NTA) and Fuel Purge

Area (FPA) at Tinker AFB indicate different fuel types.

The shapes of these spectra identify the contaminants

as fuel oil at the NTA and JP-4 at the FPA (Figure 3).

The fluorescence versus depth profile from push location 84L at Plattsburgh AFB indicates narrow bands of

contamination are in the “oily phase,” which rests just

above the water table. Note that discrete sampling at 5-ft

intervals (25, 30, 35 ft, etc.) could easily skip over the

contamination (Figure 4). A series of fluorescence versus depth profiles taken across a north-south transect

at the Plattsburgh AFB fire-fighter training area show the

extent of contamination (Figures 5 and 6). Location 84A

is upgradient, 84D is in the center of the burn pit, and

the remainder are downgradient. The contamination

traveled directly down from the burn pit and then along

the water table.

Conclusions

Currently, these technologies are being further developed and demonstrated within numerous DOD, Department of Energy, and EPA programs. Ongoing research

will develop techniques to monitor contaminants such as

chlorinated solvents, metals, and explosives that do not

naturally fluoresce. Refining this technology is at the

heart of site remediation because it provides cost-effective characterization before, during, and after remediation. We can use it to determine if remediation is

needed, what remediation technology we should apply,

whether the remediation is working, and whether the

cleanup effort has been successful—all with a minimum

of risk, time, labor, and cost.

100,000

10,000

1,000

100

10

1

1 10 100 1,000 10,000

Signal (pVs)

Concentration DFM (ppm)

Figure 1. Calibration curve for diesel fuel marine on fisher

sand.

36
References

1. Bratton, W.L., J.D. Shinn, S.M. Timian, G. Gillispie,

and R. St. Germain. 1993. The Air Force Site Characterization and Analysis Penetrometer System

(AFSCAPS), Vols. I-V. AL/EQ-1993-0009.

2. Gildea, M.L., W.L. Bratton, J.D. Shinn, G. Gillispie,

and R. St. Germain. 1994. Demonstration of the Air

Force Site Characterization and Analysis Penetrometer System (AFSCAPS) in support of the intrinsic bioremediation (natural attenuation) option

(interim report). 

3. Schroeder, J.D., and S.R. Booth. 1991. Cost effectiveness analysis of the Site Characterization and

Analysis Penetrometer System (SCAPS). Los

Alamos National Laboratory, Los Alamos, NM.

Diesel #2

JP-4

Diesel Fuel Marine

Unleaded Gasoline

80

0

40

60

20

10

30

50

70

80

0

40

60

20

10

30

50

70

Time (ns)

300 350 400 500 450

Wavelength (nm)

300 350 400 500 450

Wavelength (nm)

300 350 400 500 450

Wavelength (nm)

300 350 400 500 450

Wavelength (nm)

80

0

40

60

20

10

30

50

70

Time (ns)

120

0

40

80

20

60

100

Figure 2. Wavelength time matrices of various fuels.

37
300 400 325 375 350 425 475 450 500

Wavelength (nm)

FPA

NTA

300 400 325 375 350 425 475 450 500

Wavelength (nm)

120

0

100

80

60

40

20

Time (ns)

120

0

100

80

60

40

20

Time (ns)

Figure 3. Wavelength time matrices at two different sites:

North Tank Area versus Fuel Purge Area.

0 5 10 15 20 25 30

0


5


10


15


20


25


30


35


40

Depth (ft)

Equivalent (% of JP-4 standard)

Figure 4. Fluorescent versus depth profile from fire training

facility.

38
84A-LIF

BM LS 4906.2

84B-LIF

84D-LIF

0 200 400

ft 84H-LIF

CPT


Monitoring Well


Survey Point

84P-SS

84G-LIF

A'

B'

84L-LIF

A

B

84F-LIF

84K-LIF

84J-LIF

84I-LIF

Figure 5. Base map of the fire training area showing location of CPT soundings and wells.

Section A - A'

84D-LIF

Fuel Pit

Leading Edge of Oily Phase

84L-LIF 84F-LIF 84K-LIF

Smear Zone

260

240

220

200

180

Elevation

>=150.0

A = A' LIF Intensity

max=462.8


min=0.1600

0 0 100 200

Vertical Exaggeration: 6

Gr Sand


Sand


Sand Mix

Sample: Water Level (PE)

ft

Figure 6. Cross section showing contaminated zone and water table along section A-A′ in Figure 5.

39
Microbiological and Geochemical Degradation Processes

E. Michael Godsy

U.S. Geological Survey, Menlo Park, CA

Introduction

Ground-water contamination is perhaps the most dangerous and intractable environmental impact of toxic

chemical spills and leaks. The prevention of groundwater contamination is of foremost importance. In terms

of treatment processes, the present technology of excavation and relocation of contaminated soils to “secure

landfills” (which are seldom secure) and “pump and

treatment” of contaminated ground water has proven to

be totally inadequate; these processes just transfer contaminants from one environmental phase to another.

Bioremediation, however, achieves contaminant decomposition or immobilization by exploiting the existing

metabolic potential of microorganisms with novel catabolic functions derived through selection. In response to

the introduction of a toxic contaminant, an indigenous

bacterial population arises that is unique from the standpoint of physiological capabilities and species diversity.

Conditions that restrict life or inactivate microbial enzymes are incompatible with intrinsic bioremediation efforts. Although the physical and chemical characteristics

of the contaminants and the metabolic potential of microorganisms determine the feasibility of biotransformation reactions, actually achieving biotransformation also

depends on the prevailing geochemical conditions. The

ecological constraints to bioremediation can be classified as microbial, chemical, or environmental, and recognition of these constraints, as required, assesses the

feasibility of intrinsic bioremediation. This article focuses

on the microbiological and geochemical constraints influencing intrinsic bioremediation processes.

Bacterial Metabolic Diversity

Bacteria comprise a large and diverse group of microorganisms that obtain their energy from a variety of

sources, including 1) light (photosynthetic bacteria), 2)

the oxidation or decomposition of reduced organic matter (heterotrophic bacteria), or 3) the oxidation of reduced inorganic compounds (autotrophic bacteria).

Some bacteria derive energy from more than one

source, such as combinations of light and reduced inorganic or organic compounds; however, heterotrophic

bacteria are the major group responsible for the biodegradation of organic compounds.

All living organisms must generate reducing power for

the purpose of replenishing enzymatic systems and

maintaining the oxidation-reduction power cycle. This

involves the reduction of oxidized compounds by the

addition of electrons released from compounds oxidized

during energy production. The electron acceptor can be

either an organic or an inorganic compound. For many

bacteria, most fungi, and all higher organisms, the final

electron acceptor is O2 in the process termed aerobic

respiration. The final reduced substrate in aerobic respiration is H2O, and the final oxidized compound respired from energy production is CO2.

In the absence of O2, certain bacterial populations respire other less oxidized inorganic compounds (inorganic respiration) or use only organic compounds

(fermentation). Denitrification is a process promoted by

bacteria that can thrive under either aerobic or anaerobic conditions (facultative anaerobes) by using O2 as the

terminal electron acceptor, when available, or, in the

absence of O2, by using NO3- as the terminal electron

acceptor (inorganic respiration). When oxidation-reduction potentials within soils are even lower, other inorganic compounds are used by specific groups of

bacteria as terminal electron acceptors. Because O2 is

toxic to these bacteria, they are called obligate anaerobes. Several common alternative electron acceptors

and associated bacterial groups include ferric iron

(iron reducers), SO4

2- (sulfate reducers), and CO2

(methanogens).

Incorporating Contaminants Into Bacterial

Ecosystems

The ecosystem as a whole can be thought of as a series

of integrated oxidation-reduction (redox) reactions

driven ultimately by the radiant energy of the sun. Microorganisms catalyze many of these reactions and play an

40
essential role in maintaining the electron balance of

complex ecosystems. For a contaminant to be incorporated into these redox reactions, it must be able to serve

as either an electron donor or an electron acceptor.

Moreover, its tendency to either donate or accept electrons—and thus be oxidized or reduced—depends on

the chemistry of the compound. For example, many

halogenated organic compounds are highly oxidized

relative to their nonhalogenated counterparts, and thus

tend to accept electrons and to be reductively dehalogenated. The halogenated compounds must compete

with other physiological electron acceptors in order to be

incorporated into microbial energy cycles. Thus, the

effectiveness of reductive dehalogenations is often influenced by the presence of other electron acceptors, such

as NO3- or SO42-. This effect would explain why reductive dehalogenations are more frequently observed under methanogenic conditions, where generally a paucity

of these anions exist.

Bacteria-Contaminant Interactions in

Aquifer Material

A minimum of three conditions must be met before a

contaminant can be degraded or transformed by bacteria: 1) the bacteria must be in the immediate vicinity of

the contaminant; 2) the contaminant must be available

to the bacteria; and 3) the bacteria must have the capacity to participate in some part of the degradation or

transformation process. Specific bacterial populations

prefer or require particular environmental conditions. If

these conditions do not exist, these populations tend to

become quiescent until more ideal conditions return or

develop, or in some cases they may even die off. The

nature of limiting environmental factors often dictates

which bacterial populations exist. While the subsurface

environment can be modified, modification is often

accomplished with great difficulty and always at great

expense.

Bacterial Nutrition

Bacteria are composed of combinations of elements that

are the components of their genetic material, structural

molecules, enzymes, and intracellular plasma. Because

of the great diversity among bacteria, the proportion of

nutrient elements required for growth varies widely;

however, the major required elements that make up

bacteria are carbon, hydrogen, sulfur, nitrogen, and

phosphorus. In soils and aquifer materials contaminated

with most organic compounds, carbon and hydrogen are

not typically limiting because they are the major components of organic compounds, and sulfur is generally

available in sufficient quantities for growth. Thus, the

major limiting elements for growth are nitrogen and

phosphorus. The carbon/nitrogen/phosphorus ratio

(C/N/P) usually considered ideal is 300 to 100:10:1 to

0.05 (1); however, this ratio can vary depending on the

nature of the contaminant(s).

pH and Redox Potential

Near-neutral aquifer pH values are usually optimum for

the biodegradation of contaminant organic material. The

hydrogen ion concentration of the ground water is governed by the types of compounds produced by bacterial

activity, and is controlled especially by CO32-:HCO3-

:CO2 equilibrium rates (2). Because hydrogen ion transfer is commonly involved in electron transport, pH and

redox potential are interdependent (3).

The redox potential, termed Eh, is extremely important

in the biotransformation schemes of both organic and

inorganic contaminants. Usually, a heavily contaminated

site is anoxic because ongoing bacterial respiration has

depleted all available O2. The resulting anoxic conditions tend to favor different electron acceptors, with the

most oxidized compounds (higher Eh) being used first.

The resultant scheme is NO3- (denitrification) utilization

after O2 depletion, Fe3+ (iron reduction) utilization after

NO3- depletion, SO42- (sulfate reduction) utilization after

Fe3+ depletion, and finally CO2 reduction to CH4 after

depletion of the available SO42-. As a result, bacterial

populations having different degradative potentials can

be operative at different times at the same contamination site as the redox potential varies.

Although Eh measurements can provide valuable clues

about the functioning of geochemical systems in aquifer

material, they alone will not give definite information

about the chemical species present. These measurements indicate that a potential exists for certain redox

reactions, such as those involving bacterial respiration.

Various organic and inorganic redox reactions cannot be

predicted because of their complexity and different interdependent reaction rates (2).

Temperature

Ground-water temperature is often one of the most important factors controlling microbial activity and rates of

organic matter decomposition. Generally, rates of enzymatic degradation and bacterial metabolism double for

every 10o

C increase in temperature until close to inhibitory temperatures, which are usually around 40o

C to

50o

C for most bacteria. Except for subfreezing temperatures, however, bacteria are generally capable of degradation at most ambient temperatures (4). Perhaps more

importantly, temperature can influence biodegradation

of a compound or contaminant mixture by changing its

physical properties, bioavailability, or toxicity to bacteria.

For example, an increase in temperature usually increases the equilibrium vapor concentration, resulting in

an increased volatilization rate, but it can at times also

increase sorption to aquifer particles (5).

41
Physical Deterrents to Biodegradation

Physical or physicochemical factors can also affect the

biodegradation of contaminants. Some molecules are

recalcitrant to degradation because they are too large to

enter bacterial cells, which is usually required for complete degradation by membrane-bound enzymes. Some

substances are difficult to biodegrade because the number, length, or location of functional groups impede enzyme attack. Strong sorption on aquifer material can

greatly hinder the ability of bacteria to attach to, absorb,

or enzymatically attack the molecule (6).

Sorption and solubility of organic contaminants are

complex interdependent phenomena that vary with the

composition of the aquifer material and complex contaminant mixtures. For example, aromatic hydrocarbon

concentrations in water extracts of 31 gasoline samples

varied over an order of magnitude (7). Although gasoline

variability could account for this, the solubility of each

component of mixtures can vary from ideal conditions,

with each compound acting as a cosolvent to increase

hydrophobic hydrocarbon solubilities (8). Organic solvents can also affect the sorption of organics on soils in

general (9).

Contaminant Metabolism: Aerobic Versus

Anaerobic

A common misconception is that all organic contaminants are biodegraded most rapidly and thoroughly under aerobic conditions. Although this is commonly the

case, anaerobic conditions promote some very important degradative processes. For example, compounds

that are highly oxidized, such as polychlorinated

biphenyls or chlorinated solvents, are more susceptible

to reductive processes than to oxidative processes (10)

during the initial stages of mineralization.

Most organic compounds found in crude oil, refined oils,

and fuels are known to degrade under aerobic conditions (4); however, current research efforts have shown

that the biodegradation of many monoaromatic compounds common to most fuels also occurred in the

laboratory under anaerobic conditions. This biodegradation readily occurs not only with NO3- serving as the

terminal electron acceptor (11) but also under SO42-

reducing conditions (12), Fe3+ reducing conditions (13),

and methanogenic conditions (14,15).

Benzene has been especially recalcitrant to anaerobic

biodegradation in laboratory studies under denitrifying

and sulfate-reducing conditions (10,12). Yet some laboratory (16) and field studies (17) have shown the depletion of all common monoaromatic hydrocarbons found

in gasoline under denitrifying conditions. In 1986, Vogel

and GrbiŒ-GaliŒ (18) demonstrated that benzene and

toluene were degradable to CH4 and CO2 in laboratory

microcosms; however, the confirmation of the degradation of benzene in field situations under methanogenic

conditions has been rather elusive. One of the few

well-documented instances of the methanogenic degradation of benzene in the field is a crude oil spill in

Bemidji, Minnesota (19), where many of the watersoluble monoaromatic hydrocarbons present in crude

are undergoing intrinsic bioremediation.

Conclusion

Although bioremediation in general has gained considerable attention, obstacles remain before bacteria can

be used effectively for detoxifying wastes affecting

ground water. A lack of knowledge or misunderstanding

concerning what can and cannot be done with bioremediation has resulted in unrealistic expectations, leading

in turn to disappointments and ultimate failures. Continued research into and application of sound bioremediation schemes will undoubtedly prove the viability of

intrinsic bioremediation in the overall remediation efforts

of contaminated ground water.

References

1. Torpy, M.F., H.F. Stroo, and G. Bruebaker. 1989.

Biological treatment of hazardous wastes. Poll.

Eng. 21:80-86.

2. Stumm, W., and J.J. Morgan. 1981. Aquatic chemistry. New York, NY: John Wiley & Sons.

3. Grundi, T. 1994. A review of the current understanding of redox capacity in natural, disequilibrium

systems. Chemosphere 28:613-626.

4. Leahy, J.G., and R.R. Colwell. 1990. Microbial degradation of hydrocarbons in the environment. Microbiol. Rev. 54:305-315.

5. Lyman, W.J., W.F. Reehl, and D.H. Rosenblatt.

1982. Handbook of chemical property estimation

methods: Environmental behavior of organic chemicals. New York, NY: McGraw-Hill. 

6. Cheng, H.H., K. Haider, and S.S. Harper. 1983.

Catechol and chlorocatechols in soil: Degradation

and extractability. Soil Biol. Biochem. 15:311-317.

7. Cline, P.V., J.F. Delfino, and P.S.C. Rao. 1991. Partitioning of aromatic constituents into water from

gasoline and other complex solvent mixtures. Environ. Sci. Technol. 25:914-920.

8. Groves, F.R. 1988. Effect of cosolvents on the solubility of hydrocarbons in water. Environ. Sci. Technol. 22:282-286.

9. Fu, J.-K., and R.G. Luthy. 1986. Effect of organic

solvent on sorption of aromatic solutes onto soils.

J. Environ. Eng. 112:346-366.

42
10. Kuhn, E.P., and J.M. Sulflita. 1989. Dehalogenation

of pesticides by anaerobic microorganisms in soils

and ground water: A review. In: Sawhney, B.L., and

K. Brown, eds. Reactions and movement of organic

chemicals in soils. Special Publication 22. Madison,

WI: Soil Science Society of America. pp. 111-180.

11. Evans, P.J., D.T. Mang, and L.Y. Young. 1991. Degradation of toluene and m-xylene and transformation of o-xylene by denitrifying enrichment cultures.

Appl. Environ. Microb. 57:450-454.

12. Edwards, E.A., L.E. Wills, D. GrbiŒ-G aliŒ, and M.

R e inhard. 1 991 . Anae robic de gradation of tolue ne

and xyle ne : E vide nce for sulfate as the te rm inal

e le ctron acce ptor. In: H inche e, R . E . , and R .F. O lfe nbutte l, e ds. In situ bioreclamation: Applications and

investigations for hydrocarbon and contaminated

site remediation. Boston, MA: Butterworth-Heinemann. pp. 463-471.

13. Lovely, D.R., and D.J. Lonergan. 1990. Anaerobic

oxidation of toluene, phenol, and p-cresol by the

dissimilatory iron-reducing organism, GS-15. Appl.

Environ. Microb. 56:1,858-1,864.

14. GrbiŒ-G aliŒ, D. , and T. M. Voge l. 1 987. Transform ation of tolue ne and be nze ne by m ixe d m e thanoge nic culture s. Appl. E nviron. Microb. 53: 254-260.

15. Wilson, B.H., G.B. Smith, and J.F. Rees. 1986.

Biotransformations of selected alkylbenzenes and

halogenated aliphatic hydrocarbons in methanogenic aquifer material: A microcosms study. Environ. Sci. Technol. 20:997-1,002.

16. Major, D.W., C.I. Mayfield, and J.F. Barker. 1988.

Biotransformation of benzene by denitrification in

aquifer sand. Ground Water 26:8-14.

17. Berry-Spark, K.L., J.F. Barker, D. Major, and C.I.

Mayfield. 1986. Remediation of gasoline-contaminated ground-waters: A controlled experiment. In:

Proceedings of petroleum hydrocarbons and organic chemicals in ground water: Prevention, detection, and restoration, NWWA/API. Dublin, OH:

Water Well Journal Publishing.

18. Vogel, T.M., and D. GrbiŒ-G aliŒ. 1 986. Incorporation

of oxyge n from w ate r into tolue ne and b e nze ne

during anae robic fe rm e ntative transform ation. Appl.

E nv iron. Microb. 52:200-202.

19. Cozzarelli, I.M., R.P. Eaganhouse, and M.J.

Baedecker. 1990. Transformation of monoaromatic

hydrocarbons to organic acids in anoxic groundwater environment. Environ. Geol. Water Sci.

16:135-141.

43
Field and Laboratory Results: Getting the Whole Picture

Mary Jo Baedecker

U.S. Geological Survey, Reston, VA

Abstract

Concern over contamination of ground water in the last

decade has led to an increased awareness of the need

to understand the transport and fate of organic contaminants and the geochemical processes that result from

their presence in the subsurface. A large number of

contaminated sites contain petroleum-derived hydrocarbons. Many of the hydrocarbons are biodegraded in

ground-water environments, and the extent of their removal by natural processes has been evaluated in field

and laboratory investigations. Both types of investigations demonstrate that natural biodegradation can be an

important component in remediation strategies for some

contaminated sites. 

Results and Discussion

The processes that control the attenuation of organic

compounds in the subsurface are complex, and many

investigations have been undertaken in the field and in

laboratories to understand better the factors that control

degradation reactions. Parts of contaminant plumes

often become anoxic, and the fate of contaminants at

anoxic field sites has been reported in several studies

(1-5). One of the most widespread types of contaminants is petroleum-derived hydrocarbons from pipeline

breaks, leaking storage tanks, spills, and disposal of

wastes. Many of these sites, such as those contaminated by leaking small underground storage tanks, are

easier to remediate than sites with contaminants such

as chlorinated compounds. The number of such sites is

large, however, and much effort has been spent trying

to understand processes and to develop effective remediation strategies to deal with petroleum-derived hydrocarbons. Several research efforts have been undertaken

in the field and laboratory to determine processes that

affect the fate and transport of individual hydrocarbons.

In field investigations, degradation of soluble aromatic

hydrocarbons has been shown to occur downgradient

from source areas (6-10). Hydrocarbon-degrading bacteria were found and quantified in soil and ground water

at a fuel-oil contaminated site (11). Degradation of petroleum-derived hydrocarbons is generally considered

to occur more rapidly in aerobic or suboxic environments, where oxygen or nitrate is available as an electron acceptor (7,12-13), but anaerobic biodegradation

also may remove significant amounts of hydrocarbons

from ground water (14-17).

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. An underground pipeline carrying crude oil ruptured and sprayed oil over land surface.

Part of the oil was removed during remediation, but part

of it infiltrated through the unsaturated zone and accumulated as an oil body floating on the ground water. A

more detailed description of the processes that occurred

at the site is available in Baedecker et al. (5), Bennett et

al. (18), Eganhouse et al. (9), Baedecker and Cozzarelli

et al. (19), Cozzarelli et al. (20), and Eganhouse et al. (21).

The concentrations of benzene, toluene, ethylbenzene

and o-, m-, and p-xylene (BTEX) in the upper 1.5-m

thickness of the aquifer downgradient from the oil body

are shown in Figure 1. The concentrations of BTEX

decreased with increasing distance from the oil body

and were attenuated under anoxic and oxic conditions.

Where oxygen was encountered, concentrations decreased by several orders of magnitude (56 m to 137 m).

The mass of BTEX lost near the oil body (26 m to 56 m),

however, was as high as that lost farther downgradient.

The decrease in concentrations of BTEX near the oil

body (26 m to 56 m) was in the anoxic part of the plume,

where no oxygen was detected in the ground water over

an 8-yr period. Another indicator that this part of the

aquifer was anoxic is that ferrous iron concentrations

were 1.8 mg/L at the water table and 71 mg/L at only 0.3

m deeper in the aquifer (22). Ferrous iron precipitates

where it encounters trace quantities of oxygen. Thus,

the high concentrations of ferrous iron near the oil

body indicate that dissolved oxygen was not transported to the water table. Degradation of hydrocarbon

44
vapors in the unsaturated zone most likely consumed

the oxygen by aerobic respiration (23). The downgradient movement of the BTEX plume averaged about

8 m/yr from 1987 to 1992, but movement has not been

at a steady rate. Near the oil body, microbial degradation

is the primary process of hydrocarbon attenuation. As

the hydrocarbons are transported farther from the oil

body, additional processes such as dispersion, mixing

(24), and sorption become more important.

Laboratory experiments demonstrated that these hydrocarbons can degrade under controlled laboratory conditions. To verify that the hydrocarbons were degrading

under anoxic conditions, microcosm experiments were

undertaken with sediment and water from the anoxic

part of the plume (5,20). In two separate experiments,

benzene and a mixture of toluene and naphthalene were

added to microcosms under anaerobic conditions. These

compounds were also added to microcosms that were

poisoned and sterilized for controls. In the microbially

active microcosms, benzene decreased in concentration by 98 percent in 125 d, and toluene decreased by

99 percent in 45 d (5). No loss of naphthalene was

observed during the same period. For the controls, no

loss of benzene, toluene, or naphthalene was observed.

These laboratory experiments and field results provide

strong evidence that hydrocarbons are degrading in an

anoxic environment. By comparing the results from the

microbially active microcosms with results from controls,

sorption and abiological chemical oxidation were eliminated as possible explanations for the loss of benzene

and toluene.

Field and laboratory studies of anaerobic biotransformation or aromatic hydrocarbons were reviewed by Barker

and Wilson (25) who found evidence at five methanogenic sites for biodegradation of benzene, toluene,

ethylbenzene, and the xylenes. The estimated half-lives

were 0.5 yr and 3.8 yr. The longest half-lives were for

benzene. Laboratory experiments also have indicated

biodegradation by several pathways (26); however, for

benzene the results have been contradictory. Large concentrations of benzene are biodegraded in the subsurface (27), yet in some ground-water environments

benzene is the most persistent hydrocarbon among the

monoaromatics. 

Natural processes can remove significant concentrations of hydrocarbons and prevent the spreading of a

plume. At sites where the rates of solubilization, volatilization, and biodegradation of hydrocarbons are such

that the plume is either contained or spreading at a slow

rate, these natural processes can be considered in the

design of a site remediation program. Even at sites

where plumes are spreading at a fast rate, knowledge

of the natural processes that are attenuating contaminants provides information that can be used to accelerate biodegradation processes. 

Note

The data and interpretation in this report were previously

published in Baedecker and Cozzarelli (19) and Baedecker

et al. (5).

References

1. Baedecker, M.J., and W. Back. 1979. Hydrogeological processes and chemical reactions at a landfill.

Ground Water 17:429-437.

2. Nicholson, R.V., J.A. Cherry, and E.J. Reardon.

1983. Migration of contaminants in ground water at

a landfill: A case study 6. Hydrogeochemistry. J.

Hydrol 63:131-176.

3. Lesage, S., R.E. Jackson, M.W. Priddle, and P.G.

Riemann. 1990. Occurrence and fate of organic

solvent residues in anoxic ground water at the

Gloucester landfill, Canada. Environ. Sci. Technol.

24:559-566.

4. Godsy, E.M., D.F. Goerlitz, and D. GrbiŒ-G aliŒ.

1 992. Me thanoge nic biode gradation of cre osote

contam inants in natural and sim ulate d groundw ate r e cosyste m s. G round W ate r 30: 232-242.

5. Baedecker, M.J., I.M. Cozzarelli, R.P. Eganhouse,

D.I. Siegel, and P.C. Bennett. 1993. Crude oil in a

shallow sand and gravel aquifer, III. Biogeochemical reactions and mass balance modeling in anoxic

ground water. App. Geochem. 8:569-586.

6. Barker, J.F., J.S. Tessman, P.E. Plotz, and M. Reinhard. 1986. The organic geochemistry of a sanitary landfill leachate plume. Contam. Hydrol.

1:171-189.

100

10

1.0

0.1

0.01

0.001

0.0001

BTEX

1987

1990

1992

0 100 200 50 150

Meters Downgradient from Oil Body

Concentration, Milligram per Liter

Figure 1. Concentrations of benzene, toluene, ethylbenzene,

and o-, m-, and p-xylene (BTEX) in the upper 1.5-m

thickness of an aquifer contaminated with crude oil

(19).

45
7. Major, D.W., C.I. Mayfield, and J.F. Barker. 1988.

Biotransformation of benzene by denitrification in

aquifer sand. Ground Water 26:8-14.

8. Cozzarelli, I.M., R.P. Eganhouse, and M.J.

Baedecker. 1990. Transformation of monoaromatic

hydrocarbons to organic acids in anoxic groundwater environment. Environ. Geol. Water Sci.

16:135-141.

9. Eganhouse, R.P., M.J. Baedecker, I.M. Cozzarelli,

G.R. Aiken, K.A. Thorn, and T.F. Dorsey. 1993.

Crude oil in a shallow sand and gravel aquifer, II.

Organic geochemistry. Appl. Geochem. 8:551-567.

10. Davis, J.W., N.J. Klier, and C.L. Carpenter. 1994.

Natural biological attenuation of benzene in ground

water beneath a manufacturing facility. Ground

Water 32(2):215-226.

11. Kämpfer, P., M. Steiof, and W. Dott. 1991. Microbiological characterization of a fuel oil-contaminated site

including numerical identification of heterotrophic

water and soil bacteria. Microb. Ecol. 21:227-251.

12. Kuhn, E.P., J. Zeyer, P. Eicher, and R.P. Schwarzenback. 1988. Anaerobic degradation of alkylated

benzene in denitrifying laboratory aquifer columns.

Appl. Environ. Microb. 54:490-496.

13. Hutchins, S.R., G.W. Sewell, D.A. Kovacs, and G.A.

Smith. 1991. Biodegradation of aromatic hydrocarbons by aquifer microorganisms under denitrifying

conditions. Environ. Sci. Technol. 25:68-76.

14. Wilson, B.H., G.B. Smith, and J.F. Rees. 1986.

Biotransformations of selected alkylbenzenes and

halogenated aliphatic hydrocarbons in methanogenic aquifer material: A microcosm study. Environ.

Sci. Technol. 20:997-1,002.

15. GrbiŒ-G aliŒ, D. , and T. M. Voge l. 1 987. Transform ation of tolue ne and be nze ne by m ixe d m e thanoge nic culture s. Appl. E nviron. Microb iol. 53: 254-260.

16. Lovley, D.R., M.J. Baedecker, D.J. Lonergan, I.M.

Cozzarelli, E.P. Phillips, and D.I. Siegel. 1989. Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339:297-299.

17. Haag, F.M., M. Reinhard, and P.L. McCarty. 1991.

Degradation of toluene and p-xylene in anaerobic

microcosms: Evidence for sulfate as a terminal electron acceptor. Environ. Toxicol. Chem. 10:1,379-

1,390.

18. Bennett, P.C., D.I. Siegel, M.J. Baedecker, and M.F.

Hult. 1993. Crude oil in a shallow sand and gravel

aquifer, I. Hydrology and inorganic geochemistry.

Appl. Geochem. 8:529-549.

19. Baedecker, M.J., and I.M. Cozzarelli. 1994. Biogeochemical processes and migration of aqueous constituents in ground water contaminated with crude

oil. In: Toxic substances and the hydrologic sciences. American Institute of Hydrology. pp. 69-79.

20. Cozzarelli, I.M., M.J. Baedecker, R.P. Eganhouse,

and D.F. Goerlitz. 1994. The geochemical evolution

of low-molecular-weight organic acids derived from

the degradation of petroleum contaminants in ground

water. Geochim. Cosmochim. Acta 58(2):863-877.

21. Eganhouse, R.P., M.J. Baedecker, and I.M. Cozzarelli.

1994. Biogeochemical processes in an aquifer contaminated by crude oil: An overview of studies at

the Bemidji, Minnesota, research site. Presented at

the U.S. EPA Symposium on Intrinsic Bioremediation of Ground Water, Denver, CO (August 30 to

September 1).

22. Cozzarelli, I.M., M.J. Baedecker, G. Aiken, and C.

Phinney. 1994. Small-scale chemical heterogeneities

in a crude-oil-contaminated aquifer, Bemidji, Minnesota. In: Morganwalp, D.W., and D.A. Aronson, eds.

Proceedings of the Technical Meeting of the U.S.

Geological Survey Toxic Substances Hydrology Program, Colorado Springs, CO (September 20-24,

1993). Water Res. Invest. Rep. 94-4014. In press.

23. Hult, M.F., and R.R. Grabbe. 1988. Distribution of

gases and hydrocarbon vapors in the unsaturated

zone. In: Ragone, S.E., ed. U.S. Geological Survey

toxic waste/ground water contamination study. Proceedings of the Technical Meeting of the U.S. Geological Survey, Cape Cod, MA (October 21-25,

1985). Open-File Report 86-481. pp. C21-C26.

24. Essaid, H.I., M.J. Baedecker, and I.M. Cozzarelli.

1994. Use of simulation to study field-scale solute

transport and biodegradation at the Bemidji, Minnesota, crude-oil spill site. In: Morganwalp, D.W., and

D.A. Aronson, eds. U.S. Geological Survey toxic substances hydrology program. Proceedings of the Technical Meeting of the U.S. Geological Survey Water

Resources, Colorado Springs, CO (September 20-24,

1993). Water Res. Invest. Rep. 94-4014. In press.

25. Barker, J.F., J.T. Wilson. 1992. Natural biological

attenuation of aromatic hydrocarbons under anaerobic conditions. Proceedings of the Subsurface

Restoration Conference, Dallas, TX. pp. 57-58.

26. GrbiŒ-G aliŒ, D. 1 990. Anae robic m icrobial transform ation of nonoxyge nate d arom atic and alicyclic

com pounds in soil, subsurface, and fre shw ate r

se dim e nts. In: Bollag, J.M. , and G . Stotzky, e ds. Soil

Bioche m istry, Vol. 6. N ew York, N Y: Marce l D e kke r,

Inc. pp. 1 1 7-1 89.

27. Hadley, P.W., and R. Armstrong. 1991. Where’s the

benzene? Examining California ground-water quality surveys: Ground Water 29(1):35-40.

46
Case Studies: Public Service’s Site

47
In Situ Bioremediation at the Seventh Avenue Site in Denver: 

Remediation of Soils and Ground Water

Christopher Nelson

Groundwater Technology, Inc., Englewood, CO

In situ bioremediation is a cost-effective method for the

remediation of soils and ground waters contaminated

with petroleum hydrocarbons. This paper presents a

case study of the successful application of in situ bioremediation at a public utility site in Denver, Colorado.

The site was used as a truck maintenance facility for

almost 30 years. During that time, nonvolatile petroleum

hydrocarbons from used motor oil, diesel, gasoline, and

other automotive fluids were released to a used oil sump

at the site. The local geology of the site includes

claystone and sandstone bedrock, covered with recent

deposits of interbedded alluvial sands and gravels,

as well as silty sands and clays. The principal aquifer at

the site lies within the interbedded alluvial sands and

gravels.

Soil and water sampling confirmed the presence of petroleum hydrocarbons in both the vadose and saturated

zones. Laboratory studies showed that the chemical,

microbiological, and hydrogeological characteristics of

the site were conducive to bioremediation.

Site Assessment/Treatability Study

An extensive site assessment and treatability study was

conducted at the site to determine physical, chemical,

microbiological, and hydrogeological characteristics

controlling the biodegradation of contaminants and the

mass transport of nutrients and oxygen. Nine monitoring

wells were installed at depths of about 7.6 m, using a

3-m screen interval. The site assessment indicated that

the primary contaminant at the site was waste oil located

in the saturated and unsaturated sediments beneath the

former used oil sump. Samples showed high levels of

benzene, toluene, ethylbenzene, and the xylenes

(BTEX); total petroleum hydrocarbons (TPHs); and total

organic carbon (TOC) in ground water, with localized but

detectable levels of chlorinated organics. A relatively

large population of bacteria was found within the

contamination zone; however, its growth appeared to be

restricted by nutrient and oxygen conditions. 

Feasibility studies were performed on soil and sediment

samples to determine the biodegradability of contaminants under various nutrient loads and aerobic conditions. Aerobic testing was performed to simulate optimal

conditions for the bioremediation of hydrocarbons. Column studies also were conducted to determine how

nutrient and hydrogen peroxide loading would affect the

hydraulic conductivity of sediments in the subsurface

above the ground-water table. 

The results showed that the loading of nutrients and

hydrogen peroxide would be critical to the success of in

situ bioremediation at this site and that the loading

should be minimized in the silty sand zone because of

high reactivity. A nutrient adsorption test indicated that

ammonia and phosphate loading was feasible in the soil

at the site. 

System Design and Installation

After reviewing several remedial operations, in situ

bioremediation was selected. The conceptual design

included stimulating indigenous bacterial populations

through the introduction of oxygen and inorganic nutrients. The primary functions of the remediation system

included ground-water recovery, treatment, and reinjection; vapor extraction and discharge; stimulation of

in situ bioremediation by subsurface inorganic nutrient

and oxygen additions; and phase-separated hydrocarbon

recovery.

Laboratory tests showed that hydrogen peroxide and

nutrient loading worked best in sediments from the

coarse sand interval. Nutrient injections and the addition

of atmospheric oxygen from the vapor extraction system

stimulated bioremediation in the unsaturated zone and

enhanced the desorption of adsorbed hydrocarbons for

recovery in the monitoring wells.

48
Operation, Monitoring, and Results

Once the bioremediation system was installed, it was

inspected weekly to adjust and maintain the water-table

depression pump, the hydrogen peroxide and nutrient

injection equipment, and the soil vapor extraction system. Crews conducted field analyte tests, sampled

ground water, and gauged monitoring wells. The bioremediation system began operating in July 1989 and

continued through March 1992. 

Approximately 36,000 lb of hydrocarbons have been

removed. Nearly 94 percent of the contaminant mass

was degraded biologically, as evidenced by the low

concentrations of dissolved oxygen and the relatively

high concentrations of background heterotrophs and

hydrocarbon-utilizing bacteria. Approximately 9 million

gal of ground water were recovered, amended with nutrients, and reinfiltrated. The site is currently undergoing

closure with the Colorado Department of Health.

Reference

For more information about the Denver site and supporting data, refer to:

Nelson, C.H., R.J. Hicks, and S.D. Andrews. 1994. An

integrated system approach for in situ bioremediation of

petroleum hydrocarbon contaminated soil and ground

water. In: Flathman, P.E., D.E. Jerger, and J.H. Exner,

eds. Bioremediation: Field experience.

49
The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through

In Situ Bioremediation: The Role of Mathematical Models

Tissa H. Illangasekare and David C. Szlag

Department of Civil, Environmental, and Architectural Engineering, University of Colorado, Boulder, CO

John T. Wilson

Robert S. Kerr Environmental Research Laboratory, U.S. Environmental Protection Agency, Ada, OK

Abstract

This paper discusses the important processes in developing mathematical models to be used as tools for the

design and evaluation of ground-water remediation

schemes in aquifers contaminated with nonaqueous

phase organic chemicals. The paper also presents the

application of a mathematical model that considers

these processes in developing a methodology for retrospective evaluation of bioremediated sites.

Introduction

Mathematical models of water flow and chemical transport have been extensively used in ground-water quantity and quality management. Some of the applications

of these models involve bioremediation. The basic processes of flow of water and transport of dissolved substances are fairly well understood, and numerical

models for the solution of the governing equations have

been developed. Field applications of models as prediction and design tools, however, have not been very

successful for many reasons, including the complexities

associated with natural heterogeneities and the inadequacy of available field techniques for physical and

chemical characterization; this is especially true in

ground-water contamination situations involving chemicals in the form of separate phase organics. The models

that have been developed to simulate the transport and

entrapment behavior of nonaqueous phase chemicals

and waste products have not been adequately validated

due to the scarcity of laboratory and field data. These

models sometimes fail to simulate flow and entrapment

behavior under heterogeneous soil conditions that are

commonly encountered in the field. Accurate calibration

and prediction become difficult due to the limitations of

field and laboratory techniques that are used to obtain

model parameters. Some of the scaling issues related

to multiphase flow model parameters are not very well

understood. The assumptions that are made in modeling

mass transfer from entrapped chemicals to the aqueous

phase become questionable under some conditions of

ganglia formation and macroscale entrapment. 

The movement of nonaqueous phase liquids (NAPLs) in

the subsurface is a complicated process leading to large

amounts of NAPLs becoming trapped in the soil. Most

of the common organic wastes in the form of NAPLs are

only sparingly soluble in water and thus act as long-term

sources of ground-water contamination. In a case study

presented in this paper, we will show that after bioremediation, pockets of NAPLs remained in the soil. These

entrapped fluids have the potential to contribute to

ground-water pollution after active remediation has

been discontinued. 

In our research, we have identified two modes of entrapment: microscale and macroscale. Microscale entrapment, which occurs at residual levels, is primarily

governed by fluid properties and pore characteristics

(1). Macroscale entrapment is defined as entrapment at

saturations higher than residual (or irreducible) due to

heterogeneities in the soil (2). 

Many models employed in remediation design make use

of the local equilibrium assumption (LEA); in other

words, any water exiting a zone of entrapped NAPL will

be completely saturated with the contaminant, regardless of the system parameters. The LEA is conservative

in predicting the maximum concentration observed in

ground water but may lead to gross underprediction of

the contaminant source lifetime. It is our hypothesis that

entrapment itself will change the system parameters

controlling mass transfer into the flowing water, and the

LEA assumption is not a physically realistic way to quantify mass transfer. 

50
Laboratory experiments were conducted to obtain a fundamental understanding of the processes that govern

the transport and distribution of organic chemicals in

soils and to generate data for validation of models that

will be used as tools for the design of remediation

schemes and monitoring systems. A detailed investigation of these processes under controlled conditions was

done in small soil cells, columns, and large flumes. 

In our ongoing research, new models and modeling

approaches have been developed. To improve these

models, we have focused on issues related to entrapment, mobilization, and mass transfer associated with

organic waste chemicals in heterogeneous aquifers.

The effectiveness of models as tools to design and

evaluate treatment and remediation technologies depends on their ability to accurately represent the above

processes. A case study conducted in Colorado identifies some of these basic processes that are of importance in remediating and monitoring sites contaminated

with organic fluids, and demonstrates the use of a

mathematical model.

Study Objectives

Conventional methods for determining the extent of

cleanup at a bioremediation site can often be misleading. Monitoring wells may show very low or zero levels

of contaminants after active bioremediation, but levels

may increase over time. In most cases, regulatory

authorities require a direct measure of the residual

NAPLs after bioremediation in addition to monitoring

well data. Often the relative composition of the oily

phase is assumed to remain constant during bioremediation. This is a conservative assumption and generally

leads to target levels of total petroleum hydrocarbon

(TPH) concentrations on the order of 10 to 100 mg/kg

aquifer material. Many bioremediation schemes, however, may preferentially degrade the compounds of

regulatory concern, leaving relatively high TPH levels in

the soil that pose a minimal risk. This modeling study

focuses on developing a methodology to evaluate the

possible risk, if any, associated with benzene, toluene,

ethylbenzene, and xylene (BTEX) sources left in soils

after the implementation of a bioremediation scheme.

This developed methodology will assist in providing answers to the following questions: 1) Will BTEX reappear

in ground water? 2) How long will it take the plume to

reappear? 3) What concentration level may be expected? The results of the case study will also assist in

providing a technical basis for implementing monitoring

schedules, locating compliance wells, and constructing

rational criteria for site closure. 

Problem Description

A temporary holding tank under a garage in an industrial

area in Denver, Colorado, leaked used crank case oil,

diesel fuel, gasoline, and other material into a shallow

water table aquifer. 

Remediation involved removal of separate oily phases,

in situ bioremediation with hydrogen peroxide and mineral nutrients, and bioventing. An estimated 2,147 lb of

hydrocarbons have been removed by pumping hydrocarbon emulsion from monitoring wells or by volatilization through soil aeration. In the treatment system, the

ground water pumped from a downgradient recovery

well was treated and amended with organic nutrients

and hydrogen peroxide. The solution was then injected

into the subsurface system upgradient of the contaminated site. The system operated from October 1989 to

March 1992. Table 1 compares the reduction in concentration of benzene and total BTEX compounds in ground

water achieved by this remediation scheme (3,4).

Water from the monitoring wells and the recirculation

well contained low concentrations of contaminants by

March 1992. Active remediation was terminated, and the

site entered a period of postremediation monitoring. In

June 1992, core samples were taken from the aquifer to

determine the extent of hydrocarbon remaining and to

determine whether a plume of contamination could return once active remediation ceased. The site was cored

along a transect downgradient of the release. The cores

were extracted and analyzed for total petroleum hydrocarbons and for the concentration of individual BTEX

compounds. Data from one of the boreholes with the

highest concentrations of hydrocarbons are given in

Table 2. At the time of sampling, the elevation of the

Table 1. Reduction in Concentration (μg/L) of Hydrocarbon Contaminants in Ground Water Achieved by In Situ Bioremediation

Benzene Total BTEX

Well Before During After Before During After

MW-1 220 <1 <1 2,030 164 <6

MW-8 180 130 16 1,800 331 34

MW-2A — 11 0.8 — 1,200 13

MW-3 11 5 2 1,200 820 46

RW-1 <1 2 <1 <1 2 <1

51
water table was 5,280.5 ft above mean sea level

(AMSL).

To hydraulically characterize the aquifer, selected soil

cores that had previously been used to measure hydrocarbon saturations were reconstructed in a load cell, and

hydraulic conductivity was determined with a constant

flux apparatus. The hydraulic conductivity varied two

orders of magnitude across the site, with some highly

permeable channels evident.

The results of this field investigation suggest that hydrocarbon in the form of nonaqueous fluids moved into

preferential flow channels created by the local heterogeneities. After direct recovery, the fluids remained entrapped at saturations that may be higher than residual.

During remediation, the treating agents did not reach

some of the locations where the chemicals were

entrapped. 

A modeling study was conducted at the site to make a

retrospective evaluation of the effectiveness of the remediation scheme. The results of this study are reported

in Szlag et al. (3). 

Model Selection

The following observations at the site and in the laboratory indicated the need for three-dimensional simulation:

1) visual inspection of aquifer material indicated the

presence of coarse gravel lenses, clayey sands, and

sands of varying gradation; 2) light nonaqueous phase

liquid (LNAPL) plumes are inherently three-dimensional,

forming thin, pancake-like plumes in the capillary fringe

and just beneath the water table; 3) LNAPL can become

entrapped in coarse lenses that act as preferential flow

channels well beneath the water table; 4) solute plumes

are not vertically homogeneous, and biological activity

will not be uniformly distributed vertically. Bypassing due

to the lowered hydraulic conductivity of the central part

of the LNAPL plume by nutrients and electron acceptors

resulted in high TPH and BTEX levels in some cores.

Coupled with the clear need for a three-dimensional

model are other criteria such as availability, ease of use,

reliability, and cost. We have selected MODFLOW, a

three-dimensional ground-water flow model developed

by the U.S. Geological Survey (USGS), to simulate

ground-water flow. Solute transport is simulated with a

three-dimensional random walk called RAND3D. 

Modeling Approach

The problem domain was modeled as a rectangular area

300 ft long and 200 ft wide. Two wells outside of the

modeled domain were used as reference head locations

for general head boundaries. The NAPL contaminant

zone covers approximately 1,600 ft2

 of area and a soil

depth of approximately 1.7 ft. To accurately assess the

mass flux from the LNAPL contaminant source, three

layers were chosen in the model. The upper two layers

are 1 ft thick, and the bottom layer is 18 ft thick. The

LNAPL organic contaminant is confined to the upper two

layers. 

Data for the period June 8, 1989, to April 1, 1990, were

used for calibration. The goodness of fit between the

model and measured well data was characterized by a

mean residual and standard deviation of the mean residual. The best fit model was obtained by assigning the

pump test average hydraulic conductivity to the bottom

layer, which carried the majority of water.

The main focus of this modeling effort is to determine

how much contaminant mass will be transported from

the remaining residual and whether it will generate a

plume of regulatory concern. A monitoring well screened

in only the upper two layers would see the highest

contaminant concentrations. A pumping well screened

over the entire aquifer thickness is also being considered; in this case, however, dilution will play a major role

in reducing the maximum concentrations. Two significant assumptions are used in the solute transport

modeling: 1) the concentration of BTEX in the source

zone remains constant, and 2) water flowing from the

contaminated cells is in equilibrium with the residual

NAPL. Using the heads generated by MODFLOW,

Table 2. Vertical Extent of Total BTEX and TPH (mg/kg) at a Borehole

Elevation (Feet AMSL) TPH BTEX Benzene Color and Texture

281.14-5,280.31 <44 <1 <0.02 Brown sand

5,280.31-5,279.97 227 5.1 <0.2 Brown sand

5,279.97-5,279.56 860 101 <0.2 Brown sand

5,279.56-5,279.14 1,176 206 4.3 Brown sand

5,279.14-5,278.97 294 27 0.68 Brown sand

5,278.97-5,278.64 273 7.4 0.26 Brown sand

5,278.64-5,278.22 <34 <1 <0.2 Brown sand

5,278.22-5,277.14 <24 <1 <0.2 Brown/yellow sand

52
ground-water velocity through each source cell was calculated. The known NAPL BTEX concentration was then

used to calculate the equilibrium concentration and,

consequently, the mass flux. Estimated benzene mass

fluxes were converted into particle inputs for each layer.

The particle tracking model was used to simulate solute

transport using the velocity field generated from

MODFLOW. 

Results and Discussion

Any simulation of solute transport requires specification

of the contaminant source. For the case of a NAPL spill,

the source function will consist of a continuous mass flux

of solute from the residual NAPL phase to the aqueous

phase. Many researchers have shown that equilibrium

is quickly reached in spill scenarios if ground-water

velocity is low and the “residence time” of the water in

contact with NAPL is “long enough.” From a regulatory

standpoint, the assumption of equilibrium is conservative, as greater mass fluxes cannot be achieved. The

water flux can vary significantly in the source zone,

which often gives misleading indications that the contaminant transport is rate-limited. This primary problem

has been the focus of our work. Preferential flow paths

often develop within the source zone in areas with low

BTEX and TPH concentrations, allowing water flow to

bypass the more highly contaminated areas. Laboratory

determination of the hydraulic conductivity in the

samples containing high amounts of TPH confirms this

observation. 

A key result of the modeling study is that the solute

plume emanating from a NAPL source is not homogeneous. In general, the solute plume will consist of subplumes at different depth intervals and widely different

concentrations, and moving at different velocities. A

regulatory question posed earlier in this paper is, “How

long should the compliance wells be monitored?” The

answer is when all the subplumes have reached steadystate. The plumes in the middle and bottom layer have

reached or are close to equilibrium by 330 days. The

plume in the top layer, which has the highest benzene

concentration, has not reached equilibrium at 420 days.

The design of the compliance wells will have tremendous impact on the actual sampled concentration. If the

wells are bailed or pumped so that the well volume is

completely mixed, significant dilution will occur. The existing monitoring wells at the site are screened over the

top 5 ft of the aquifer. The maximum concentration

achieved in the well screened over a 5-ft interval

reaches a steady-state concentration of 26 ppb. If that

well is screened over the entire saturated thickness, a

concentration of 15 ppb is achieved. Even greater dilution will occur if the well is pumped. 

Several operational considerations for risk assessment

and compliance well monitoring can be made from the

modeling study: 1) a benzene plume will reestablish

itself at the site, but it will be three orders of magnitude

lower than the federal maximum contaminant level—

new standards may be set, however, and the risk from

this plume may be deemed significant; 2) local hydraulic

conductivity plays a significant role in determining the

contaminant mass flux and in creating subplumes of

different concentration and velocity; 3) compliance well

monitoring will have to be continued past August 1993

so that solute plumes in all levels will reach steady-state;

4) retardation coefficients and effective porosity data

would significantly improve the time of arrival estimate

of the solute plume; 5) compliance well design should

be carefully considered when sampling a three-dimensional plume because well design can lead to significant

contaminant dilution. 

Conclusions

A modeling methodology for the retrospective evaluation

of bioremediated aquifers contaminated with organic

chemicals was developed. The primary hypothesis on

which the methodology was based is that during a spill,

NAPL contaminant becomes entrapped preferentially in

coarse formations in the saturated zone and fine formations in the unsaturated zone. This hypothesis is supported by laboratory experimental (2) and field data.

Flow channels created by naturally occurring aquifer soil

heterogeneities as well as macroscale entrapment of

the NAPL will also produce preferential paths for the

treating agents. The proposed methodology requires

that these local heterogeneities in the contaminant zone

of the spill be captured. Standard pump tests, which

provide regional values for transmissivity, will not have

adequate resolution to capture these spill-site-scale heterogeneities. Even though hydraulic conductivity values

determined in the laboratory on disturbed soil samples

were used in this study, a more appropriate characterization method would be well-designed bail tests (or

slug tests) that capture the local layered heterogeneities

more accurately. These local hydraulic conductivity values

allow us to obtain the velocity field in the contaminant zone

and to subsequently determine the contaminant mass flux.

Solute breakthrough curves determined by this method

can then be used to conduct risk analysis and to provide

a rational basis for postremediation well monitoring. 

Acknowledgments

The support of the U.S. Environmental Protection

Agency through the Hazardous Substance Research

Center at Kansas State University (agreement R815709) is gratefully acknowledged. We would also like

to thank Ms. Lisa Weers of the Colorado Department of

Health for her assistance.

53
References

1. Szlag, D., and T.H. Illangasekare. 1994. Quantification of residual entrapment of nonaqueous phase

organic fluids in soils. Ground Water. In review.

2. Illangasekare, T.H., D. Szlag, J. Campbell, J. Ramsey, M. Al-Sherida, and D.D. Reible. 1991. Effect of

heterogeneities and preferential flow on distribution

and recovery of oily wastes in aquifers. Proceedings

of the Conference on Hazardous Waste Research,

Manhattan, KS. Manhattan, KS: Kansas State University.

3. Szlag, D.C., T.H. Illangasekare, and J.T. Wilson.

1993. Use of a three-dimensional ground-water

model for retrospective evaluation of a bioremediated aquifer contaminated with organic chemicals.

Proceedings of the Ground-Water Modeling Conference, Golden, CO (June 10).

4. Wilson, J. 1993. Retrospective performance evaluation on in situ bioremediation: Site characterization.

In: U.S. EPA. Symposium on bioremediation of hazardous wastes: Research, development, and field

evaluations (abstracts). EPA/600/R-93/054. Washington, DC (May).

54
The Importance of Knowledge About Intrinsic Bioremediation for Cost-Effective

Site Closure: The Client’s Perspective

Harry E. Moseley

Public Service Company of Colorado, Denver, CO

Public Service Company of Colorado (PSCo) is Colorado’s largest electric and gas utility. During an underground storage tank removal and replacement project in

1987 at PSCo’s Seventh Avenue Service Center, a facility used for automotive maintenance, an oil sump was

discovered. Because the sump did not have a concrete

bottom or other type of platform, the soil beneath the

sump was saturated with used oil.

After an extensive study to determine the nature and

extent of the contamination beneath the facility, PSCo

officials decided that in situ bioremediation of the site

would be more cost effective than removing the building

that housed the center. Construction of a new facility

would cost at least $1 million, excluding demolition of

the existing facility and removal or remediation of the

contaminated soil and ground water. After discussions

with the Colorado Department of Health and 18 months

of ground-water monitoring, the site received approval

for final closure in March 1994, 7 years after discovery

of the contamination. 

PSCo, a business without extensive expertise in environmental restoration, undertook the project with a fairly

direct goal: to remediate the site as efficiently and costeffectively as possible. Intrinsic bioremediation was

found to be a key element of the project’s success. In

1988, little was known about intrinsic bioremediation,

and some opportunities may have been lost; indeed,

additional knowledge of these fields might have reduced

costs.

For example, in investigating whether bioremediation

was feasible, emphasis was placed on enhancing the

activities of the bacteria that would perform the remediation, rather than on determining how long the enhancement could sustain bacterial growth, which

might have revealed a more effective strategy for

delivering nutrients. Once the nutrients penetrated

across the site, batch feeding of the nutrients may have

been more effective for the sustained growth of the

bacteria than the continuous recirculation of the enriched ground water. If this fact had been considered

before the start of the project, operation and maintenance costs incurred during the project’s life cycle could

have been reduced.

The risk assessment and ground-water data suggested

that the chance of any contaminants appearing in any

ground water being consumed or used by humans was

minimal. The natural ground-water flow across the site

was very slow, and the site might have remediated itself

with the help of a small amount of oxygen and other

nutrients.

To evaluate the progress of the project and to adhere to

Underground Injection Control regulations, many tests

were required that monitored the concentration of the

injected chemicals, or “nutrients”; these tests necessitated costly site visits and report submissions. The

chemicals were not injected at harmful levels, however,

nor was the risk of human exposure significant due to

the location of the injections and the velocity of the

ground water. Three years into the project, the reporting

requirements were reduced to quarterly, but a better

understanding of the bioremediation process would

have reduced these requirements earlier.

The bioremediation process came in on budget and

successfully removed all threat to human health and the

environment (see Table 1).

Information from past projects, sound scientific judgment, and risk analyses will provide an understanding

of intrinsic bioremediation that could increase the number of cleanups, speed up the closure process, and

reduce testing and reporting requirements, thereby substantially reducing costs to industry.

55
Table 1. Resources Used at the Seventh Avenue

Bioremediation Project

1990 Labor hours 2,024 hours total

168 hours per month average

Activities O&M costs

Quarterly and monthly reporting

Testing cost $12,778

Materials cost $16,473

1991 Labor hours 1,746 hours total

145 hours per month average

Activities Onsite inspections (once or twice a

week)

Risk assessment formulation

O&M costs

Quarterly and monthly reporting

Testing cost $13,451

Materials cost $12,000

1992 Labor hours 551 hours total

46 hours per month average

Activities Quarterly reports

Monthly site visits

Closure report submittal

Testing cost $11,539

1993 Labor hours 94.25 hours total

7.85 hours per month average

Activities Closure monitoring

Testing

(quarterly)

$8,420

56
A Regulator’s Perspective of a Bioremediation Site

Lisa C. Weers

Colorado Department of Health, Denver, CO

The regulator’s review and perception of a bioremediation

site differs from those of the person preparing the site

assessment plan/corrective action report and the site

representatives. This paper describes the regulator’s role

at the Public Service Facility at 2701 W. 7th Avenue in

Denver, Colorado. At the time of corrective action plan

approval, the Colorado Department of Health (CDH)

reviewed the technology to be used. After 2 years of

postremediation monitoring, Public Service provided information indicating that the points of compliance were not

affected, which is a consideration for closure. 

The Public Service site is located within 1,000 feet of the

South Platte River, which supplies public water. On the

site is a garage and an office building, while the perimeter of the site is secured with a 6-foot fence. The location

of the building and the fence put some constraints on

the placement of the monitoring wells and the bioremediation system. 

The information CDH requested for this site is roughly

the same information found in any underground storage

tank site assessment; however, this site used a new

technology, which required that additional information

be provided. The site was first reviewed in 1990, when

there were no other active bioremediation sites in the

Underground Storage Tank Program (Storage Tank Remedial Section). Also, no written guidance documents

were available for performing site assessments. Because the technique at that time was relatively unknown,

finding a coworker with prior experience was impossible.

In addition, each staff member was assigned a large

number of sites because only three people were reviewing sites for the state at the time. As a result, going to

the library during work time was out of the question.

When a new technology is reviewed, the person assigned to the site looks for information on the safety of

this remediation in regard to what the effects may be on

human health and the environment, how the technology

affects the sampling and analysis program, and whether

the proposed remediation has special considerations

that need to be incorporated in the approved corrective

action plan. Another question we often ask is, are other

appropriate permits obtained? GTI, the consulting firm

working on the project, was challenged on several occasions by some very tough questions, such as, are the

fatty acids produced during bioremediation interfering

with your laboratory tests? Are any intermediate compounds produced? 

For a corrective action plan to be approved, the site

should receive a complete assessment. While reviewing

a site assessment, a technical reviewer examines the

report for completeness. The site assessment should

reveal the following information:

• Is there contamination in the subsurface?

• If so, how much and what type?

• Is the site the source?

• Where is it going?

• Has contamination reached ground water?

• Is ground water used?

• Is the contamination migrating off site?

• What are the potential receptors?

• What is the qualitative risk to potential receptors?

• How will subsurface conditions affect migration?

• Is remediation necessary?

• What are remediation options to consider?

Some of the main points we look for in approving a corrective action plan are determination of the full extent of soil

and ground-water contamination; definition of the hydrology of the site, the location of utility corridors, and the

geology of the site; a nearby surface and ground-water

assessment; a risk assessment; and a remediation plan.

After all the information is reviewed, a decision is made to

either approve or deny the corrective action plan.

When the Public Service site corrective action plan was

approved, EPA also evaluated the site for its Superfund

Innovative Technology Evaluation (SITE) program. We

were very excited that the site would be the subject of

research. The additional information that was provided

was useful in helping to determine a monitoring schedule.

57
The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through

In Situ Bioremediation: The Regulator’s Perspective

Mark E. Walker and Lisa C. Weers

Colorado Department of Health, Denver, CO

Because staff are assigned such a large number of

sites, reviewing all of the hard data for each and every

site is not conceptually possible. Therefore, we do not

want to receive all of the technical information and correspondence generated at every site; instead, we prefer

the technical interpretation. Our approach to site evaluation is to review this interpretation, reserving the right

to request any and all hard data generated during the

course of the investigation.

Although the technical staff at the Colorado Department

of Health (CDH) review an abridged version of the total

information generated at a particular site, our version of

the file on the Public Service Company of Colorado

(PSCo) occupied 6 feet of shelf space. Even a review of

the abridged version of this site presented a daunting

task. 

I needed to get up to speed quickly on the file to consider

this site for closure. My first task was to verify if an

approved corrective action plan was in place and to

determine what conditions had been placed on this

approval, if any. In any case where ground-water contamination has been documented, we check to see if the

point of compliance (POC) wells have been affected in

excess of state maximum contaminant levels (MCL).

The second task was to conduct an evaluation of the

existing monitoring program to determine if the program

was adequate for detecting contamination emanating

from the release. Based on the calculated ground-water

velocity, was the monitoring period long enough to detect contamination from the source? With respect to

monitoring well construction and location, were these

wells placed downgradient? (Has direction changed with

the seasons?)

After the monitoring plan was completed and the closure

borings had been drilled and analyzed, the next topic for

examination prior to closure was the extent of the remaining contamination and its potential for adversely

affecting the environment. This information is considered with the identified receptors in the immediate vicinity and the potential for these receptors to be affected

by the remaining contamination.

After closure is complete, one might ask, “Well, where

does this site stand now, from a regulatory point of

view?” The following items should be considered: 1)

after closure is complete, CDH does not release the

owner/operator from any liability regarding the contamination remaining in the subsurface; 2) the closure decision is based solely on the information submitted; and

3) similar to the owner’s liability issue, CDH wants to be

informed of any developments that could increase the

potential for this site to affect human health and the

environment adversely, at which point we assess the

need for additional work/remediation.

In conclusion, CDH is willing to consider innovative

remedial technologies. Those who are proposing such

technologies should be prepared to demonstrate that

these technologies can be implemented in a manner

that does not adversely affect human health and the

environment. Proposal of an innovative technology

should be approached from the viewpoint of a regulator:

consider receptors and aspects of the technology that

might be detrimental to the environment, and address

these concerns in a responsible, forthright manner.

58
Case Studies: 

Sites Where Natural Attenuation Has Been Documented

59
Intrinsic Bioremediation of JP-4 Jet Fuel 

John T. Wilson, Fredrick M. Pfeffer, James W. Weaver, and Don H. Kampbell

U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK

Todd H. Wiedemeier

Engineering Science, Denver, CO

Jerry E. Hansen and Ross N. Miller

Air Force Center for Environmental Excellence, Brooks AFB, TX

Introduction

Intrinsic bioremediation is a risk management option

that relies on natural biological processes to contain the

spread of contamination from spills. The option is most

appropriate when the concentration of contaminants is

reduced to regulatory limits before ground water discharges to surface water or is collected by a pumped well.

In the past, remedial action plans have proposed the

intrinsic remediation option based solely on the apparent attenuation of contamination in water from monitoring wells that are distant from the spill. These plans were

often criticized because it was impossible to distinguish

attenuation due to contaminant destruction from attenuation due to simple dilution in the aquifer or in the

monitoring well. Convincing regulators that the wells

with low concentrations of contaminants actually sample

the plume of contaminated ground water has been difficult. This lack of credibility has led to the “one-more-well”

syndrome, with excessive investment in a monitoring

approach that focuses on the compounds of regulatory

concern but fails to earn the confidence of the regulatory

community. 

During characterization of JP-4 jet fuel spills at Eglin Air

Force Base (AFB), Florida, and Hill AFB, Utah, three

approaches were used to distinguish contaminant attenuation due to destruction from attenuation due to

dilution or sorption. 

To distinguish attenuation due to biological destruction

of the contaminants from attenuation due to dilution, the

attenuation of the compounds of regulatory concern—

benzene, toluene, ethylbenzene, and the xylenes

(BTEX)—was compared with the attenuation of other

components of the fuel that were relatively recalcitrant.

Tracers have been used successfully to correct for dilution and sorption of hydrocarbons in ground water. Cozzarelli et al. (1) used 1,2,3,4-tetramethylbenzene to

normalize the concentrations of other alkylbenzenes

and their anaerobic degradation products in ground

water that had been contaminated by a spill of light

crude oil. Wilson et al. (2) used 2,3-dimethylpentane to

normalize the concentrations of BTEX compounds in

ground water contaminated with gasoline from an underground storage tank. In both cases, the tracer was a

component of the spilled fuel.

To distinguish attenuation due to biological destruction

from attenuation due to sorption, core samples were

analyzed for the total quantity of individual BTEX compounds and for total petroleum hydrocarbons. Partitioning theory was used to predict the concentration of

individual hydrocarbons in ground water in contact with

the core material. The predicted concentrations were

compared with concentrations in water from monitoring

wells to determine if the ground water in the plume was

in sorptive equilibrium with the spilled fuel.

To prove that the attenuation was due to biological activity, the geochemistry of the ground water was examined. Microbial metabolism of petroleum hydrocarbons

has predictable geochemical consequences. The hydrocarbons can be respired, resulting in the consumption

of oxygen, nitrate, sulfate, or iron II minerals in the

aquifer matrix and in the production of water, dinitrogen,

sulfide, or iron II. Microbiologists often refer to the substrates for microbial respiration as electron acceptors.

Alkylbenzenes can be fermented, resulting in the production of methane. Simple stoichiometry can be used to

predict the quantity of electron acceptors consumed or the

quantity of methane produced during biotransformation

60
of a given concentration of petroleum-derived hydrocarbons. 

Intrinsic Bioremediation of Ground Water

at Eglin AFB, Florida

Leaking distribution pipes from an underground storage

tank released JP-4 jet fuel to the water table aquifer

under the petroleum, oil, and lubricants storage depot

(POL) at Eglin AFB. The POL is located over sands and

silty peats characteristic of a barrier island complex. At

the time of the study, a plume of contaminated ground

water moved away from the residual JP-4 and discharged to a small creek approximately 300 ft downgradient (Figure 1). The elevation of the water table is

approximately 8.4 ft above sea level in the area with

residual JP-4. The water table in the creek is 1.4 ft above

mean sea level. Hydraulic conductivity determined by

pumping tests in monitoring wells varied from 48 ft to

102 ft per day. Based on these data, and assuming an

effective porosity of 30 percent, the residence time along

the flow path from the spill to the creek is on the order

of 10 wk. Water samples were acquired with a geoprobe,

using an 18-in. screen. Seasonally, the temperature of

ground water at the site varies from 19o

C to 28oC, and

the pH varies from 5.6 to 6.7. Samples producing the

data in Table 1 were taken in August 1993, when groundwater temperatures varied from 24o

C to 28oC.

Correcting for Dilution at Eglin AFB

Table 1 presents the changes in concentrations of BTEX

compounds and the three trimethylbenzenes (TMB)

along the flow path from the spill to the creek (Figures

1 and 2). Sample 80H-3 is from a location just outside

PL-2

Extent of


JP-4 Spill

80-H 8

2

3

7

4

6

5

A

G

D

C

B

1-P1

Ground-Water


Flow Direction

Geoprobe Location


Soil Boring


2" Monitoring Well

Legend

83-Z Creek

A' Creek

83-U

Scale 5

Line of Equal


Water Table Elevation


(Feet above Mean


Sea Level) 0' 50'

A A' Line of Hydrogeologic Section

Figure 1. Flow path of ground water from a JP-4 spill at Eglin

AFB to the point of discharge to surface water: plan

view.

Table 1. Bioattenuation in Methanogenic Ground Water: Changes in Ground-Water Chemistry Along a Flow Path From a JP-4

Spill to the Point of Discharge to Surface Water 

Location Along Flow Path to Surface Water

Compound 80H-3 83Z-2 83U-3 83U-2 83U-1

(μg/L)

Benzene 100 153 198 6.9 <1

Toluene 5,150 18.3 1.1 <1 <1

Ethylbenzene 1,700 227 1.4 1.4 <1

p-Xylene 3,120 594 13.8 25.1 <1

m-Xylene 6,750 1,270 13.5 39.8 <1

o-Xylene 5,480 <1 <1 <1 <1

1,3,5-TMB 327 114 70.9 119 2.3

1,2,4-TMB 1,090 420 172 299 <1

1,2,3-TMB 406 182 115 187 <1

(mg/L)

BTEX and TMB 24.1 2.97 0.59 0.68 0.002

Oxygen 0.4 0.2 0.3 0.6 3.8

Nitrate + Nitrite - N 0.14 0.12 <0.05 <0.05 <0.05

Sulfate 1.6 0.62 5.6 1.76 <0.5

Methane 3.7 16.8 12.5 14.2 0.7

Iron II 2.3 7.8 3.3 2.8 <0.5

61
the JP-4 spill and appears to be in chemical equilibrium

with the weathered residual fuel. Samples 83H-1 and

83Z-2 are from locations approximately 150 ft and 300 ft

downgradient from the spill. Samples 83U-2 and 83U-3

were taken 0.5 ft and 4.1 ft below the sediments of the

creek receiving discharge from the plume. Sample 83U-1

is water from the creek at the sediment boundary, taken

when the tide was going out and the plume was actively

discharging to the creek. 

When trimethylbenzene concentrations in samples 80H-3

and 83U-3 were compared, the reduction in the concentration of these compounds was found to be remarkably

uniform. Concentrations under the creek were 36, 27,

and 46 percent of the concentrations near the spill, while

the concentrations of toluene, ethylbenzene, p-xylene,

m-xylene, and o-xylene were 0.02, 0.08, 0.44, 0.20,

and less than 0.02 percent of the initial concentrations,

respectively.

Apparently, concentrations were reduced from one-half

to one-fourth of the initial concentration due to dilution,

with further reductions due to biological activity. Benzene was not degraded in the anaerobic portion of the

flow path.

As the plume moved up into the sediments of the creek,

the concentration of benzene was reduced more than

20-fold (compare 83U-2 with 83U-3 in Table 1). If we

assume that the trimethylbenzenes are recalcitrant, the

dilution of 1,2,4-trimethylbenzene can be used to correct

for the dilution of benzene and determine the true removal due to biodegradation. The corrected concentration of a biologically transformed compound in a

downgradient well would be the measured concentration in the downgradient well, multiplied by the measured concentration of 1,2,4-trimethylbenzene in the

upgradient well, and divided by the measured concentration of 1,2,4-trimethylbenzene in the downgradient

well.

The concentration of benzene measured in sample 83U-2

was 6.9 μg/L. Compared with sample 83Z-2, the concentration corrected for dilution would be 9.7 μg/L. Compared with 153 μg/L in 83Z-2, there was a 15-fold

attenuation in benzene concentration between 83Z-2

and 83U-2. The benzene attenuation at 83U-2 compared with 83U-3 would be 50-fold.

The concentration of oxygen in the sample 83U-2, taken

0.5 ft below the sediment surface, was higher than the

concentration in sample 83U-3, taken 4.5 ft below the

sediment surface. Tidal action may reverse the hydraulic

gradient in the area proximate to the creek. This would

produce a reciprocating flow of oxygenated creek water

into the sediments and mix oxygen into the contaminated ground water. Benzene may have been degraded

aerobically. In any case, benzene and the other BTEX

compounds did not discharge to the stream at detectable concentrations (see sample 83U-1).

Kinetics of Bioattenuation in Ground

Water at Eglin AFB

First-order rate constants were calculated by correcting

the downgradient concentration for dilution. The rate

constants were calculated as:

Rate = In(corrected conc. downgradient /conc. upgradient )

residence time

Based on this relationship, rates were calculated for flow

path segments from samples 80H-3 to 83Z-2 and 83Z-2

to 83U-3. The residence time in each segment was

assumed to be 5 wk.

The rates of anaerobic bioattenuation (Table 2) were

some of the fastest that have ever been encountered by

staff of the Robert S. Kerr Environmental Research

Laboratory, probably due to the high water temperatures. There seemed to be preferential removal of toluene and o-xylene in the segment close to the spill.

Rates of removal of ethylbenzene and m+p-xylene increased after toluene and o-xylene were depleted.

A


North

A'


South

15

10

0

-10

-20

Elevation (Feet Mean Sea Level)

80-H

80 H-3

Sand

Sand

80 Z-2

Ground


Water Flow

83-U

Creek

15

10

0

-10

-20

Elevation (Feet Mean Sea Level)

U-1

U-2

U-3

Legend

Sample Location


Water Table

5 Feet


0

0 50

Feet

Scale

83 Z

Peat

Figure 2. Flow path of ground water from a JP-4 spill at Eglin

AFB to the point of discharge to surface water: cross

section.

Table 2. First-Order Rate Constants for Bioattenuation of

BTEX Compounds in a Plume of Ground Water

Contaminated by JP-4

Compound 80H-3 to 83Z-2 83Z-2 to 83U-3

(per week)

Benzene None None

Toluene -0.94 -0.38

Ethylbenzene -0.21 -0.38

p-Xylene -0.14 -0.57

m-Xylene -0.14 -0.73

o-Xylene -1.5 Cannot calculate

62
Stoichiometry of Bioattenuation at 

Eglin AFB

There was very little oxygen, nitrate, or sulfate available

for respiration of the BTEX compounds in the plume of

contaminated ground water (see Table 1). There was an

increase in the concentration of methane and iron and

a corresponding decrease in aromatic fuel hydrocarbons along the flow path, however. Assuming the following stoichiometry for methanogenesis from BTEX

compounds,

2 CH + 3/2 H2O → 5/4 CH4 + 3/4 CO2,

approximately 1.0 mg of methane is produced for each

1.3 mg of BTEX destroyed. If 80H-3 and 83Z-2 are

compared, the concentration of methane at 83Z-2, corrected for dilution, would be 43.5 mg/L. The increase in

methane from 80H-3 to 83Z-2 would be 39.8 mg/L.

Corrected for dilution, the concentration of BTEX plus

TMB in 83Z-2 is 7.69 mg/L, a decrease of 16.4 mg/L

compared with 80H-3. This decrease in aromatic petroleum-derived hydrocarbons would be expected to produce 12.6 mg/L of methane. Some of the methane

sampled at 83Z-2 may have come from the natural

degradation of the peat. In any case, the accumulation

of methane is sufficient to rationalize the destruction of

the aromatic fuel hydrocarbons along the flow path.

Intrinsic Bioremediation of Ground Water

at Hill AFB, Utah

Hill AFB is situated on a bird’s foot delta formed by the

Weber River in Pleistocene Lake Bonneville. Leaking

distribution pipes from an underground storage tank

released JP-4 jet fuel to the water table aquifer under

the POL. An area with oily phase JP-4 extends approximately 1,000 ft downgradient of the spill (Figure 3). The

oily phase hydrocarbons and the plume of contaminated

ground water are confined to the channel sands of the

delta deposits (Figure 4). The elevation of the water

table drops 56 ft across the length of the spill. Slug

testing of monitoring wells indicates a hydraulic conductivity near 8.5 ft per day, corresponding to an interstitial

seepage velocity of 1.6 ft per day.

Continuous cores were taken near the source of the spill

(82I in Table 3), near the midpoint (82D), at the lower

edge of the spill (82C), and just beyond the edge of the

spill (82B). The continuous cores started in clean material above the spill and extended through the spill into

clean material underneath the interval containing hydrocarbon. The continuous cores were subdivided into core

samples representing 0.3 vertical ft of the subsurface.

Near the spill, JP-4 appeared in the capillary fringe

(Table 3). The concentration maximum of 14,800 mg/kg

was located 0.5 ft above the water table. A second

interval that contained JP-4 extended from 3.8 to 5.2 ft

below the water table, with TPH concentrations that

ranged from 1,290 to 3,830 mg/L. At the midpoint of the

spill (82D), only one core sample contained significant

concentrations of hydrocarbons. That core sample

came from 1.0 ft below the water table. At the lower

edge, two core samples representing 0.6 vertical ft had

significant concentrations of hydrocarbons. Again these

core samples were at or just below the water table. At

the midpoint and lower edge of the spill, hydrocarbons were

not detected in core material collected above or below the

samples reported in Table 3 (detection limit 10 mg/kg).

Concentrations of hydrocarbons in the midpoint and lower

edge of the spill were less than 700 mg/kg.

Pol

EPA-82-1

EPA-82-D

Extent of JP-4 Spill

MW-10

Ground Water

Flow Direction

MW-11

EPA-82-B

EPA-82-H

EPA-82-F EPA-82-C

EPA-82-A

EPA-82-E

B'

B

Scale

MW-10 Monitoring Well Location

B-B' Line of Hydrogeologic Section

Legend

0 300 600

Feet

Storm


Drain

Figure 3. Flow path of ground water through a JP-4 spill at Hill

AFB to the point of discharge in a storm drain: plan

view.

SW B B' NE

4690


4680


4670


4660


4650


4640


4630


4620


4610


4600


4590


4580


4570

EPA-82-D

MW-11

EPA-82-C

EPA-82-E

EPA-82-F

EPA-82-A

Elevation Above Mean Sea Level (msl) (feet)

Legend

Poorly Sorted Sands


Silty or Clayey Sands


Ground Water Level

Silt or Silty Clay


Storm Sewer

0 150 300

Feet

EPA-82-I

Figure 4. Flow path of ground water from a JP-4 spill at Eglin

AFB to the point of discharge in a storm drain: cross

section.

63
Monitoring wells were installed in the boreholes used to

acquire the core samples. Additional monitoring wells

were installed downgradient of the lower edge of the

spill.

Correcting for Dilution at Hill AFB

The TMBs were remarkably persistent in ground water

from the area containing oily phase hydrocarbons (Table 4). Concentrations of the TMBs in water from the

lower edge of the spill varied from 50 percent to 147

percent of the concentration at the source, while the

concentrations of benzene, toluene, ethylbenzene, pxylene, m-xylene, and o-xylene were reduced to 0.02,

0.9, 5.6, 5.5, 3.4, and 0.23 percent of the initial concentration, respectively. The saturated thickness of

the channel sands was at most a few feet, and, therefore, there was little opportunity for dispersive dilution

of the contaminated ground water into clean ground

water underneath the plume. Once the ground water

moved past the spill, it was remediated rapidly. Concentrations of all the aromatic hydrocarbons are low in

well 82B just past the lower edge of the spill.

Stoichiometry of Bioattenuation at 

Hill AFB

Similar to the plume at Eglin AFB, neither oxygen nor

nitrate is available for respiration of the BTEX compounds within the spill area (Table 4). Oxygen concentrations as high as 8.0 mg/L occur outside the spill.

Unlike the case at Eglin, there was little accumulation of

iron II and practically no methane production in the spill.

Sulfate concentrations were high throughout the spill.

Assuming the following stoichiometry,

8 CH + 5 SO4-2 → 5 S-2 + 8 CO2 + 4 H2O,

4.6 mg of sulfate would be required to degrade 1.0 mg

of BTEX and TMB. The 7.7 mg/L of BTEX and TMB

present at the source would have a total theoretical

demand of 35 mg/L of sulfate. Concentrations in excess

Table 3. Vertical Distribution of Oily Phase Hydrocarbons at Hill AFB

Core Elevation TPH Benzene Toluene

Ethylbenzene1,2,4-

TMB

(ft) (mg/kg)

82I, near the spill area, water table elevation 4,664.53 ft

82I-5 4,665.7 4,330 0.0326 0.0266 14.5 49.9

82I-4 4,665.4 3,770 0.517 0.235 4.8 42.7

82I-3 4,665.0 14,800 4.55 2.73 47.7 167

82I-2 4,664.6 5,870 0.401 12.6 17.5 69.8

82I-1 4,664.3 398 <0.01 0.142 0.556 4.43

82I-27 to 82I-19 4,663.9 to 4,661.0 <300 — — — —

82I-17 4,660.7 1,290 0.653 0.591 3.39 2.34

82I-37 4,659.3 1,370 0.712 0.182 2.72 5.38

82I-32 4,659.0 3,830 0.0136 0.032 1.24 8.37

82D, downgradient of spill, water table elevation 4,631.7 ft 

82D-24 4,631.0 77.1 — — — —

82D-23 4,630.7 572.0 0.271 <0.01 1.48 3.11

82C, lower end of spill, water table elevation 4,603.4 ft

82C-20 4,603.5 593.0 <0.01 0.0176 0.00618 1.03

82C-19 4,603.2 638.0 0.0062 <0.01 0.0180 1.16

82B, below the spill, water table elevation 4,608.3 ft

82B-20 4,608.0 0.7 <0.01 <0.01 <0.01 <0.01

82B-19 4,607.7 1.0 <0.01 <0.01 <0.01 <0.01

64
of the theoretical demand remain in the water after it has

moved away from the spill. There is an adequate supply

of sulfate to remediate the plume through sulfate reduction alone.

Concentrations of sulfate were higher in ground water near

the source. It is possible that the lower concentrations of

sulfate near the lower edge of the spill represent the sulfate

demand exerted. It is just as possible that they represent

natural variations in sulfate concentrations.

Hydrogen sulfide did not accumulate in ground water

undergoing intrinsic remediation at Hill AFB. Unimpacted sediments at the site are tan or brown colored,

while contaminated samples are black. Iron minerals in

the aquifer matrix must act to precipitate sulfide. Sulfide

very likely precipitated as iron II sulfide.

The concentrations of electron acceptors were compared in ground water at various distances from the

lower edge of the spill. Well 82B is just outside the lower

edge; wells 82F, 82H, and 82E are 300, 400, and 500 ft

downgradient of the spill. These wells are screened in

the water-table aquifer (Figure 4); well 82A is adjacent

to 82F but is screened in the first confined aquifer underneath the water table aquifer. 

None of the wells outside the spill had significant concentrations of aromatic hydrocarbons (Table 5). Water adjacent to the spill (82B) was depleted in oxygen and nitrate.

Intrinsic Bioremediation of Oily Phase

JP-4 Jet Fuel at Hill AFB

Two samples of floating oil from monitoring well MW-10,

near the midpoint of the spill, were analyzed by gas

chromatography/mass spectrometry (GC/MS) to determine the number-average mean molecular weight of the

weathered fuel. Values of 156 and 160 daltons were

determined on duplicate samples.

The concentration of individual hydrocarbons in ground

water in contact with oily phase JP-4 was estimated

using Raoult’s Law. With data from Table 3, the concentration of an individual petroleum hydrocarbon (mg/kg)

was divided by its molecular weight to express its concentration in moles/kg core material. The concentration

of total petroleum hydrocarbons (TPHs) was divided by

160 to express TPH in moles/kg core material. The mole

fraction of the individual hydrocarbon was calculated by

dividing moles of individual hydrocarbon per kilogram by

moles of TPH per kilogram. Then the mole fraction was

multiplied by the water solubility of the individual petroleum hydrocarbons to estimate the equilibrium concentration in ground water.

Concentrations of individual petroleum hydrocarbons in

water from monitoring well 82I, near the source of the

spill, were in reasonable agreement with the concentrations predicted from the cores that had the highest TPH

Table 4. Changes in Ground-Water Chemistry Along a Flow Path Through a JP-4 Spill Undergoing Sulfate Reduction

Location Along Flow Path to Surface Water

Compound 82I MW-11 82D 82C 82B

(μg/L)

Benzene 2,740 336 96 4.9 <1

Toluene 327 90 10 3.1 4.3

Ethylbenzene 486 139 147 27 <1

p-Xylene 784 230 149 43 <1

m-Xylene 1,370 635 383 47 <1

o-Xylene 1,140 204 103 2.6 <1

1,3,5-TMB 162 71 129 238 1.1

1,2,4-TMB 495 165 183 324 1.4

1,2,3-TMB 240 69 89 120 <1

(mg/L)

BTEX and TMB 7.7 2.1 1.3 2.1 0.001

Oxygen — 0.1 1.3 0.5 1.2

Nitrate + Nitrite - N — 0.4 0.5 0.1 0.4

Sulfate — 98 193 50 64

Methane 0.68 0.022 <0.001 0.002 <0.001

Iron II — 0.05 1.7 0.84 0.11

65
values (see cores 82I-3, 82I-17, and 82I-37 in Table 6

and Table 3). The JP-4 in the capillary fringe appeared

to be aerobically weathered. There was extensive removal of benzene and toluene in the more shallow cores

82I-5 and 82I-4, while ethylbenzene and 1,2,4-TMB

were not removed.

Concentrations in water from the well at the midpoint of

the spill (82D) were considerably lower than concentrations predicted from the core material. Oily-phase material

at this location could not remove individual hydrocarbons from ground water through partitioning processes.

Core material at the lower edge of the oil spill (82C) was

very highly weathered with respect to benzene, toluene,

and ethylbenzene, while weathering of 1,2,4-TMB was

less extensive. When core 82I-37 near the source and

82C-19 at the lower edge were compared, the mole

fraction of benzene was reduced by a factor of 53, while

the mole fraction of 1,2,4-TMB was reduced by a factor

of 2.2. When concentrations in monitoring well 82I near

the source and 82C near the lower edge were compared, the concentration of benzene was reduced by a

factor of 560 while the concentration of 1,2,4-TMB was

reduced by a factor of 1.5.

Within an order of magnitude, the concentrations in

ground water were in equilibrium with the mole fraction

in the oily phase. Concentrations predicted for ground

water were higher than measured concentrations, indicating that sorption or partitioning could not be responsible for removal of the individual hydrocarbons from

ground water.

Kinetics of Bioattenuation at Hill AFB

There is no straightforward approach to calculate the

kinetics of bioattenuation in ground water in contact with

oily-phase material. For the ground-water plume at Eglin

AFB, the distance between wells along a flow path and

the estimated interstitial seepage velocity of ground

water was used to estimate residence time of the contaminants. At Hill AFB, the contaminated ground water

is in contact with oily-phase material. The major fraction

of individual contaminant hydrocarbons in the aquifer is

partitioned to the oily phase, which is moving slowly if it

is moving at all. Travel time of ground water is not related

to residence time of contaminants, making it impossible

to determine the kinetics of attenuation from data at

different locations collected at the same time. Kinetics

must be inferred from long-term monitoring data, which

are not available at the present time.

If we assume that the spill started 30 years ago, that the

JP-4 at the lower edge was the first oil spilled, and that

the mole fraction of benzene in JP-4 used at Hill AFB

has not changed appreciably over time, a comparison

of the mole fraction of benzene in 83I-3 near the source

and 83C-19 near the lower edge of the spill indicates a

reduction in the mole fraction of benzene to 3 percent of

the original concentration. If kinetics are first order on

time, the rate would be -0.11/year.

References

1. Cozzarelli, I.M., R.P. Eganhouse, and M.J.

Baedecker. 1990. Transformation of monoaromatic

hydrocarbons to organic acids in anoxic groundwater environment. Environ. Geol. Water Sci.

16(2):135-141.

2. Wilson, J.T., D.H. Kampbell, and J. Armstrong. 1994.

Natural bioreclamation of alkylbenzenes (BTEX)

from a gasoline spill in methanogenic ground water.

In: R.E. Hinchee, B.C. Alleman, R.E. Hoeppel, and

R.N. Miller, eds. Hydrocarbon bioremediation. Ann

Arbor, MI: Lewis Publishers.

Table 5. Changes in Ground-Water Chemistry Along a Flow Path Downgradient of a JP-4 Spill

Location Along Flow Path to Surface Water

Compound 82B 82F 82E 82H 82A

(μg/L)

BTEX <1 <1 <1 <1 <1

TMB <1 <1 <1 0.9 2.0

(mg/L)

Oxygen 1.2 1.1 2.7 5.4 0.3

Nitrate + Nitrite - N 0.4 5.1 5.6 2.0 0.1

Sulfate 72.2 55.5 65.8 55.7 60.5

Methane <0.001 0.006 <0.001 <0.001 0.001

Iron II 0.11 0.04 0.02 0.19 0.1

66
Table 6. Comparison of the Concentration of Benzene, Toluene, and Ethylbenzene Measured in Ground Water to the

Concentration Predicted Using Partitioning Theory From the Mole Fraction of Those Compounds in Extracts 

of Core Material

Core Elevation Benzene Toluene Ethylbenzene 1,2,4-TMB

(ft) Predicted Concentration (μg/L)

82I, near the spill area, water table elevation 4,664.53 ft

11/93 — 2,740* 372* 486* 495*

82I-5 4,665.7 22.8 5.5 769 900

82I-4 4,665.4 513 55.6 289 884

82I-3 4,665.0 1,150 165 740 881

82I-2 4,664.6 254 1,923 684 929

82I-1 4,664.3 103 320 321 869

82I-27 to 82I-19 4,663.9 to 4,661.0 <300 — — —

82I-17 4,660.7 1,870 410 603 135

82I-37 4,659.3 1,940 119 456 307

82I-32 4,659.0 13.2 7.5 74.3 171

82D, downgradient of spill, water table elevation 4,631.7 ft 

7/93 4,631.7 96* 10* 147* 183*

11/93 4,631.7 174* 4.6* 30.8 119*

82D-23 4,630.7 1,769 15.6 594 460

82C, lower end of spill, water table elevation 4,603.4 ft

7/93 4,603.4 4.9* 3.2* 27* 324*

11/93 4,603.0 <1* 8.4* 6.8* 68.9*

82C-20 4,603.5 <63 26.6 2.4 135

82C-19 4,603.2 36.3 13.9 6.5 142

*Measured Concentration (μg/L)

67
A Natural Gradient Tracer Experiment in a Heterogeneous Aquifer With Measured

In Situ Biodegradation Rates: A Case for Natural Attenuation

Thomas B. Stauffer and Christopher P. Antworth

Armstrong Laboratory, Tyndall AFB, FL

J. Mark Boggs

Engineering Laboratory, Tennessee Valley Laboratory, Norris, TN

William G. MacIntyre

School of Marine Science, College of William and Mary, Gloucester Point, VA

Biodegradation rates of organic compounds have been

measured in a heterogeneous unconfined aquifer at

Columbus Air Force Base (AFB), Columbus, Mississippi, during a pulse release experiment. Reaction rate

calculations were based on a kinetic model that includes

the hydrologic characteristics of the aquifer. Degradation kinetics were approximately first order with the following rate constants: benzene, 0.0066 d-1; p-xylene,

0.0141 d-1; naphthalene, 0.0063 d-1; o-dichlorobenzene,

0.0059 d-1.

Introduction

Biodegradation rates of organic contaminants in aquifers are needed for use in fate and transport models.

Field experiments for determination of in situ biodegradation rates are desirable because laboratory measurements may not relate to conditions in an aquifer. Madsen

(1) notes in a recent review “that determination of microbial activity in disturbed, displaced environmental

samples incubated in the laboratory is likely to be quantitatively, even qualitatively, different from the same determination in situ.” Reliable estimation of an in situ

biodegradation rate requires introduction of a known

mass of contaminant at a defined time zero, and observation of contaminant concentration variation in space

and time. Confirmation of biodegradation requires maintenance of mass balances and determination of organic

compound degradation products. To our knowledge, no

experiments to accurately measure biodegradation

rates and confirm biodegradation in an aquifer are found

in the literature.

The objective of this research was to measure in situ

biodegradation rates of organic compounds. Accordingly,

an experimental pulse injection of tritiated water and

organic solutes has been conducted at Columbus AFB

at the macrodispersion experiment (MADE) site. The

injection was into the saturated zone of the heterogeneous unconfined aquifer formed by fluvial sedimentation.

Kinetics of in situ biodegradation of benzene, p-xylene,

naphthalene, and o-dichlorobenzene in the Columbus

aquifer are reported here and related to the structural

and hydrologic properties of the aquifer. The Columbus

aquifer material has a wide range of particle sizes and

large spatial variation in horizontal hydraulic conductivity

(Kh), from <10-4 to 1 cm/s. The heterogeneity is in contrast to the generally lower (10-4 to 10-3 cm/s) values and

less variable Kh field at the Borden Canadian Forces

Base, Ontario (2).

A summary of hydrogeologic properties of the Columbus

aquifer and a description of the MADE site are given by

Boggs et al. (3). Rehfeldt et al. (4) measured the spatial

distribution of hydraulic conductivity at the MADE site

using borehole flow meters and other techniques. Figure

1 shows the Kh distribution over a vertical section directed along the test plume axis, and indicates largescale heterogeneity and structures that Rehfeldt et al.

(4) refer to as channels. Close to the injection wells Kh

is relatively low, with values near 10-3 cm/s from the

phreatic surface to the lower confining layer and extending about 40 m downgradient from the injection location.

Immediately beyond this region, Kh increases to values

near 10-1 cm/s in the upper 3 m of the aquifer, which are

maintained out to 200 m, while Kh in the lower portion

of the aquifer remains low. Thus, a near surface channel

crosses the intermediate and far field at the site. These

observations suggest that solute transport and distribution

68
might be analogous to the behavior of a hypothetical

leaking reactor located at the injection wells, with the

near field region of the aquifer representing the reactor

vessel and the upper portion of the far field serving as

the leak. A leaky reactor kinetic model including radioactive decay and biodegradation was developed for use

with field data to obtain in situ biodegradation rates for

organic compounds in the Columbus aquifer. Radiocarbon measurements using 14C-labeled p-xylene were

done to confirm biodegradation of p-xylene to its degradation products.

Experimental Methods

The experiment (MADE2) is generally similar in design

to the Borden site test described by Mackay et al. (5). A

2-day pulse injection of water containing 3

H2O and organic solutes was begun on June 26, 1990. Injection

wells were closely spaced on a line normal to the flow

direction at locations given in Boggs et al. (3). The

intersection line formed by the well plane and a vertical

plane along the flow path is shown in Figure 1. Locations

of injection and sampling wells are shown in plan view

in Figure 2.

Experimental parameters for the MADE2 test were injection volume, 9,600 L; injection time, 47.5 hr; injection

wells screened over a 0.6-m interval at 4 m below the

phreatic surface in the saturated zone of the aquifer; and

five injection wells in a line normal to the hydraulic

gradient and spaced at 1-m intervals with equal flow to each

well. Injection concentrations were: tritium, 55.6 nCi/mL;

benzene, 68.1 mg/L; naphthalene, 7.23 mg/L; p-xylene,

51.5 mg/L containing ring-labeled 14C p-xylene,

2.77 nCi/mL; o-dichlorobenzene, 32.8 mg/L. Tritium

and 14C were analyzed by liquid scintillation counting of

water samples. Concentrations of the organic compounds in water were determined by solvent extraction

into pentane containing toluene as an internal standard,

and analysis of the extract was made by gas chromatography (GC) with flame ionization detection. Most

sampling wells contained multilevel samplers, but BARCAD water samplers were used in a few wells. Water

was collected from the sampling wells and analyzed to

provide three-dimensional snapshots of solute concentrations at 27, 132, 224, 328, and 440 d after injection.

Statistical moments for each snapshot were calculated

by establishing a triangular grid followed by vertical

integration of concentrations at each well, and spatial

integration was calculated over the grid by the method

of Garabedian (6), which was previously applied at the

MADE site by Adams and Gelhar (7). The agreement

between the calculated zeroth moments and the mass

of 3

H2O injected confirmed mass balance for 3H2O. Selected wells from high and low solute concentration

regions of the plume were analyzed for p-xylene degradation by 14C counting methods 421 d after injection.

Each water sample was subsampled in the field. A 2-mL

subsample was mixed with cocktail and counted to obtain total 14C in the water as either p-xylene or its

degradation products. A 20-mL subsample was extracted with 2 mL of unlabeled p-xylene, and 1 mL of the

xylene layer was transferred to scintillation cocktail and

counted for 14C p-xylene, which remained in the organic

layer. Finally, a 10-mL subsample was made basic by

adding 1 mL 3N NaOH and 1 mL 3N Ca(NO3)2, and

centrifuged to isolate a precipitate containing Ca14CO3.

The washed precipitate and 2 mL of supernatant water

were placed in separate scintillation vials and counted.

These analyses determined the amount of 14C carbonate produced by complete degradation of 14C p-xylene

and the amount of 14C in water-soluble organic intermediate products, respectively. 

Leaky Reactor Model

Organic solutes remaining in the low Kh region near the

injection wells may be regarded as semiconfined in a

65


60


55


50


45

0 25 50 75 100 125 150 175 200 225 250 275

Distance (m)

Flow 10 (cm/s) 0 10-1 10-2 10-3 10-4

Elevation (m)

Figure 1. Distribution of hydraulic conductivity over a vertical

section containing the center line of the plume.

-100 -50 0 50 -100 150 200

x (m)

Injection


Point

-15 -10 -5 0 5 10 15

20

15

10

5

0

Inset

Barcad

Multilevel Sampler

Injection Well

y (m)

0

50

-50

100

150

200

250

300

Figure 2. Plan view of the injection and sampling wells at the

MADE2 site.

69
reactor zone that is stirred by advection and dispersion.

This reactor is considered to leak by advection through

a near surface channel with relatively high Kh, with

replacement water supplied by ground-water flow

across the upstream boundary of the reactor zone. The

conceptual model of this situation is shown in Figure 3.

This model compares well with the Kh distribution shown

in Figure 1. Reactions occurring in the reactor zone are

tritiated water (Tr) decay, with a rate constant (kd) of

1.548 x 10-3 d-1, and biodegradation of organic compounds, assumed to be first order with rate constants kb,

kx, kn, and kc for benzene (B), xylene (X), naphthalene

(N), and o-dichlorobenzene (C), respectively. The leak

rate for each solute from a well-mixed reactor should be

first order, with the same constant, kl

, for all solutes.

Sorption-desorption processes at aquifer material surfaces are assumed to be very fast relative to the other

processes considered here, and to be rate-limited by

physical transport of organic solute molecules to and

from the solid-solution interface. Biodegradation and

leakage occur simultaneously in the reactor zone by the

scheme given in Figure 3. The overall kinetics can be

expressed by the following set of linear differential equations with initial conditions: at t = t0 = 0, [Tr]0 = 0.539 Ci,

[B]0 = 660 g, [X]0 = 402 g, [N]0 = 70 g, [C]0 = 318 g, where

square brackets represent solute mass and t is time

following injection.

Analytic solutions of equations 1 to 5 with these conditions are the integrated rate laws (equations 6 to 10)

describing solute concentrations as functions of time.

The leak rate constant kl

 is calculated by solving equation 6, using the known kd of tritium. Constants kb, kx, kn,

and kc are then determined by substitution of kl

 and

solving equations 7 through 10.

− d [Tr]

dt = kd [Tr] + kl

 [Tr] (1)

− d [B]

dt = kb [B] + kl

 [B] (2)

− d [X]

dt = kx [X] + kl

 [X] (3)

− d [N]

dt = kn [N] + kl

 [N] (4)

− d [C]

dt = kc [C] + kl

 [C] (5)

kl =

ln 





[Tr]0

[Tr]







t − kd (6)

kb =

ln 





[B]0

[Tr]0







t +

ln 





[Tr]

[B]







t + kd (7)

kx =

ln 





[X]0

[Tr]0







t +

ln 





[Tr]

[X]







t + kd (8)

kn =

ln 





[N]0

[Tr]0







t +

ln 





[Tr]

[N]







t + kd (9)

kc =

ln 





[N]0

[Tr]0







t +

ln 





[Tr]

[C]







t + kd (10)

Organic solutes were retarded slightly relative to 3

H2O

by weak sorption on the aquifer material (8), but this

effect was negligible when compared with concentration

changes due to biodegradation. Laboratory batch sorption coefficients on a composite sample of Columbus aquifer material for naphthalene, o-dichlorobenzene, p-xylene,

and benzene were 0.085, 0.065, 0.048, and 0.059 L/kg,

respectively. Accordingly, this model does not incorporate

terms for sorption of solutes in the reactor zone.

Results

The distribution of 3H2O in sampling wells on the plume

axis 224 d after injection is presented in Figure 4a. It is

apparent that a large fraction of the 3

H2O mass remained near the injection wells after 224 d. Biodegradation of organic solutes occurred primarily in a reactor

zone approximately delineated by the p-xylene distribution at 224 d after injection. Figure 4b shows that

p-xylene was approximately confined within 20 m of the

injection wells. The same distribution pattern also held

for the other organic solutes. Confinement of organic

solutes to the reactor zone occurred because their

biodegradation was rapid relative to the leak rate from

the reactor zone.

Figure 3. Diagram of the leaky reactor model.

70
The spatial distributions of Kh in Figure 1 and of 3

H2O in

Figure 4a are similar, which is expected because hydraulic conductivity is the controlling parameter for

transport of an unretained solute. High 3

H2O concentrations remained in the neighborhood of the injection wells

224 d after injection, with lower concentrations in the

upper portion of the aquifer from about 30 m to 250 m

downgradient. There was little 3

H2O downgradient in the

lower portion of the aquifer between 30 m and 175 m

depth. This distribution indicates that solutes were transported downgradient from the low Kh near the injection

wells through an upper channel whose size, location, and

Kh govern the solute loss rate from the source region. Very

little transport occurred in the lower portion of the aquifer.

The observed transport of 3

H2O during this experiment

was consistent with the leaky reactor model.

Leakage of solutes was determined by the loss of tritium

from the reactor zone, defined as the portion of the

plume volume within 10 m downstream from the injection wells. The amount of tritium in the reactor zone at

a given time was calculated by spatial integration. Total

tritium in this zone decreased exponentially with time,

as shown in Figure 5. The tritium values have been

corrected for radioactive decay. This plot indicates that

solute leakage from the reactor is first order, with leakage rate constant, kl

 = 5.45 x 10-3 d-1. A mass balance

for tritium calculated using spatial integration over the

entire plume showed that the injected tritium mass was

accounted for at the times given in Figure 5. This implies

that the tritium decrease in the reactor zone was indeed

due to a process analogous to leakage.

p-Xylene

65

60

55

50

45

0 25 50 75 100 125 150 175 200 225 250 275

100

10-1 10-2 10-3 10-4

Elevation (m)

Tritium

Elevation (m)

65

60

55

50

45 0 25 50 75 100 125 150 175 200 225 250 275

100

10-1 10-2 10-3 10-4

(a)

(b)

C/Co

C/Co

Figure 4. Distribution of (a) normalized tritium and (b) normalized p-xylene concentration zone over a vertical section containing

the center line of plume motion.

1

0

-1

-2

-3

0 100 300 500 200 400

Days After Injection

In [Tr/Tr ] 0

Figure 5. Tritiated water content in the reactor organic as a

function of time.

71
Biodegradation rates within the reactor zone are approximately first order, as shown in Figure 6. Departure

of the curves from linearity is attributed to microbial

processes. There is an apparent lag period for microbial

activity soon after the injection, during which degradation rates are low. This lag period is followed by degradation at a maximum rate, which is characteristic of the

microbial metabolism. Finally, the rate decreases late in

the reaction, as the solutes (substrates) are depleted. 

Biodegradation rate constants for the organic solutes in

the reactor zone (kmax(reactor)), determined from the

maximum slopes in Figure 6, are presented in Table 1.

The maximum rate constants (kmax(corr)(reactor)) in the

second column have been corrected for organic solute

leakage from the reactor zone. Maximum biodegradation rate constants for each organic solute were also

calculated by spatial integration over the entire well field,

and are included in Table 1 as kmax(whole field). These

values are independent of the leaky reactor model. 

Degradation rate calculations for p-xylene are based on

analyses of p-xylene by GC. Ring-labeled 14C p-xylene

was included in the injection solution to demonstrate

that reductions in p-xylene concentration were a result

of biodegradation. Microorganisms mineralize 14C pxylene to water-soluble labeled intermediates and 14CO2

(predominantly as H14CO3- at ground-water pH). 14C

counting of whole water samples does not distinguish

between degraded and intact p-xylene, and thus does

not measure degradation of 14C-labeled organic compounds. Detection of these products provides a strong

indication that p-xylene has biodegraded in the aquifer.

Results of measurement of 14C p-xylene degradation

are given in Table 2. Total 14C in the water sample and

the amount of 14C in the water after extraction were used

to calculate the fraction of p-xylene converted to all

products. Total 14C in the water sample and the 14C in

the carbonate precipitate were used to calculate the

fraction of p-xylene converted to CO2. 

These p-xylene conversion figures compare well with

mass-balance-based conversions calculated from the

total p-xylene remaining in the plume at day 421, which

were interpolated from snapshot data. Mass-balancebased conversions are based on GC analysis and on

the known mass of p-xylene injected. They are included

in the last column of Table 2.

The values in the second column are on the high, but

normal, side for 14CO2 release. This implies that most of

the p-xylene went to energy production and was not

converted to biomass. The difference between degradation in the high and low concentration regions of the

aquifer is not significant. The difference of the means for

p-xylene converted to all products and p-xylene converted to CO2 may imply that some intermediate products are present, but this difference may also be

attributable to analytical anomalies (i.e., loss of Ca14CO3

during precipitate transfer prior to counting). Agreement

between the first and third columns of Table 2 indicates

the consistency of p-xylene degradation measurements

by GC and 14C counting methods. The small difference

between these numbers is due to incorporation of 14C

into biomass and insoluble carbonates. Most p-xylene

degradation products apparently remained in the local

ground water.

0

-1

-2

-3

-4

-5

Benzene


Naphthalene


p-Xylene


o-Dichlorobenzene

In [M/M ] 0

Days After Injection

0 100 200 300 400 500

Figure 6. Degradation curves for the compounds in the reactor

zone.

Table 1. Maximum Biodegradation Rate Constants From the

MADE2 Site

kmax(reactor)

(d-1)

kmax(corr)

(reactor)

(d-1)

kmax(whole

field)(d-1)

Benzene 0.0120 0.0066 0.0104

p-Xylene 0.0196 0.0141 0.0187

Naphthalene 0.0118 0.0063 0.0104

o-Dichlorobenzene0.0114 0.0059 0.0100

Table 2. Degradation of p-Xylene in Water Samples Taken

After 421 Days, Expressed as Weight Percent

Converted

% 14C

p-Xylene

Converted

to All

Products

%14C

p-Xylene

Converted

to CO2

MassBalanceBased %

p-Xylene

Converted

[xylene]>1 ppm

(n=8)

85.1 (±6.3) 73.3 (±11.1) 98

[xylene]>1 ppm

(n=10)

82.6 (±6.1) 74.2 (±6.2) 98

72
Discussion

Biodegradation rates given in Table 1 are based on field

observations of solute behavior and are an essential

input in modeling organic contaminant fate and transport

in this aquifer material. The relative rates are as expected, with p-xylene most rapidly biodegraded. Figure

6 indicates that biodegradation in the Columbus aquifer

is a first-order process. Simkins and Alexander (9) have

indicated that biodegradation can be expected to follow

Monod kinetics for a microbial population not limited by

nutrients and with a sufficient substrate concentration,

but first-order kinetics are observed at low substrate

concentrations. Organic solute concentrations in the

MADE2 test were quite low, so this conclusion is consistent with experimental results. Larson (10) notes that

the rate of decrease in concentration as a function of

substrate concentration can often be expressed as a

first-order equation, and that first-order kinetics are generally expected for biodegradation at low organic substrate concentrations. Applicability of the leaking reactor

model is site specific, and its success with the MADE2

test data is a fortuitous circumstance dependent on the

arbitrarily selected positions of the injection wells in the

Columbus aquifer. It is encouraging that the two methods of calculating biodegradation rates gave similar results, thus providing confidence that these rates are

useful for predictive purposes.

Organic solutes in MADE2 biodegraded under aerobic

conditions in the aquifer. The organic solute concentrations injected were chosen to be too low to significantly

deplete dissolved oxygen in the reactor zone of the

aquifer. Maintenance of oxic conditions was confirmed

by monitoring dissolved oxygen and redox potential of

water samples during the experiment. Thus the biodegradation rates reported here were not affected by oxygen limitation. For very large releases of similar organic

compounds, the conditions near the source might rapidly become reducing due to the biological oxygen demand. Kinetic models to determine biodegradation rates

in this situation would be complex due to the need to

include oxygen transport terms and the potential for

degradation by anaerobic bacteria. It is therefore wise

to use small injection amounts in field experiments to

determine biodegradation rates. 

In MADE2, the organic solutes degraded quickly. These

results suggest that, for similar solutes, aquifer remediation activities should be restricted to the source region,

which might include pumping to remove nonaqueousphase liquids or excavation of contaminated aquifer material. The source would be reduced but not eliminated

by this treatment. In an aquifer with approximately

steady flow, the plume of organic solutes from the reduced source would reach a steady state, with the

boundary determined by the hydrology of the site,

sorption, in situ biodegradation, and oxygen and nutrient

supply. 

Conclusions

Controlled-release experiments similar to the MADE2

test are needed to determine accurate biodegradation

rates for use in ground-water contaminant fate and

transport models of aquifer situations. The MADE2

study has demonstrated the practicality of these experiments and obtained in situ degradation rates for four

organic contaminants in the Columbus aquifer. These

rates will be used in the design and modeling stages of

a new field test at the MADE site, which is now in

preparation. 

References

1. Madsen, E.L. 1991. Determining in situ biodegradation: Facts and challenges. Environ. Sci. Technol.

25(10):1,663-1,673. 

2. Robin, M.J.L., E.A. Sudicky, R.W. Gillham, and R.G.

Kachanoski. 1991. Spatial variability of strontium

distribution coefficients and their correlation with

hydraulic conductivity in the Canadian Forces Base

Borden aquifer. Water Resour. Res. 27(10):2,619-

2,632.

3. Boggs, J.M., S.C. Young, and L.M. Beard. 1992.

Field study of dispersion in a heterogeneous aquifer, 1. Overview and site description. Water Resour.

Res. In press.

4. Rehfeldt, K.R., J.M. Boggs, and L.W. Gelhar. 1992.

Field study of dispersion in a heterogeneous aquifer, 3. Geostatistical analysis of hydraulic conductivity. Water Resour. Res. In press.

5. Mackay, D.M., D.L. Freyberg, and P.V. Roberts.

1986. A natural gradient experiment on solute

transport in a sand aquifer, 1. Approach and overview of plume movement. Water Resour. Res.

22(13):2,017-2,029. 

6. Garabedian, S.P., D. LeBlanc, L.W. Gelhar, and

M.A. Celia. 1991. Large-scale natural gradient

tracer test in sand and gravel, Cape Cod, Massachusetts, 2. Analysis of spatial moments for a nonreactive tracer. Water Resour. Res. 27(5):911-924.

7. Adams, E.E., and L.W. Gelhar. 1992. Field study of

dispersion in a heterogeneous aquifer, 2. Spatial

moments analysis. Water Resour. Res. In press.

8. MacIntyre, W.G., T.B. Stauffer, and C.P. Antworth.

1991. A comparison of sorption coefficients determined by batch, column, and box methods on a low

carbon aquifer material. Ground Water 29(6):908-

913. 

73
9. Simkins, S., and M. Alexander. 1984. Models for

mineralization kinetics with the variables of substrate concentration and population density. Appl.

Environ. Microbiol. 47(6):1,299-1,306.

10. Larson, R.J. 1979. Role of biodegradation kinetics

in predicting environmental fate. In: Maki, A.W., et

al., eds. Biotransformation and fate of chemicals in

the aquatic environment. Washington DC: American Society for Microbiology.

74
Traverse City: Distribution of the Avgas Spill

David W. Ostendorf

Civil and Environmental Engineering Department, University of Massachusetts, Amherst, MA

Abstract

The capillary tension/liquid saturation equations of

Parker and Lenhard (1) and Lenhard and Parker (2)

provide a reasonably accurate description of the vertical

distribution of the separate phase aviation gasoline

(avgas) in solid core samples taken from the U.S. Coast

Guard Air Station in Traverse City, Michigan. 

Introduction

The depth of a continuous liquid below the ground surface is directly related to its capillary tension in the

absence of vertical velocity, since the sum of the potential and pressure heads forms a vertically uniform hydraulic head under static conditions. Thus, the theory (1)

describing capillary tension/saturation relations for immiscible, continuous fluids also finds expression as vertical profiles of water and avgas in the absence of

infiltration. Lenhard and Parker (2) recognize this

equivalence and propose profiles for total liquid and

avgas saturation that are reasonably borne out by Ostendorf et al. (3) in their analysis of solid core data from

five stations in the avgas plume at Traverse City (Figure

1). We summarize the theory here and discuss some

implications of the distribution.

Vertical Distribution of Free Avgas

The vertical distribution of water and free avgas in the

contaminated soil is idealized in Figure 2. The water is

more strongly attracted to the solid grains than the

avgas and fills the smaller pores in the soil, forming an

interface with free avgas. The free avgas in turn faces

the soil gas in the large pores, which control the overall

saturation of the soil as a consequence. Residual avgas

may be trapped as discontinuous bubbles within the

water phase due to hysteresis and a fluctuating water

table.

The volumetric water θW, residual avgas θLR, and irreducible water θWR contents combine to define an effective apparent water saturation SW and comprise the total

saturation S when they are added to the free avgas

content θLF (1).

Sw = θw + θLR − θWR

n − θWR

(1a)

S = θW + θLF + θLR − θWR

n − θWR

(1b)

with porosity n. The total saturation is controlled by the

tension across the avgas/air interface and varies with

depth b below the ground surface in the absence of

dynamic effects associated with infiltration

S =SW (bM > b) (2a)

S = $1 + [βL(bL − b)]

α%(1/α)−1 (bL >b >bM) (2b)

S = 1 (b > bL) (2c)

Figure 1. Site plan, U.S. Coast Guard Air Station.

75
Lenhard and Parker (2) derive equation 2 as an application of van Genuchten’s (4) pore size distribution to

separate phase contamination in soil. Figure 2 displays

the minimum depth bM of free avgas occurrence, along

with the avgas table bL and water table bW existing in a

hypothetical monitoring well. Note that the avgas does

not occupy all the pores at a given elevation and extends

above the avgas table in the soil due to capillary tension.

Thus, the monitoring well levels, though necessary to

determine liquid tensions, do not explicitly determine the

vertical distribution of avgas in the soil. The pore size

distribution of the soil is also needed to specify the

profile.

These latter data are characterized by the pore size

uniformity exponent α and the scaling factor βL appearing in equation 2b. The latter parameter may be expressed in terms of the mean pore radius r

_

 by noting

that the avgas/air surface tension σLA relates fluid tension to interfacial radius

βL =

ρLgr

_

2σLA

(3)

with gravitational acceleration g, avgas density ρL, and

an assumed zero contact angle. Table 1 lists parameter

values calibrating the data from five stations at Traverse

City (3). Observations (circles) and theory (curve) for a

typical total saturation profile are sketched in Figure 3.

A mean pore radius of 5.4 x 10-5 m is implied by equation

3. A simple extension of a classical grain size based

model of r

_

 checks this estimate. We note that r

_

/2 is the

hydraulic radius of the mean pore size flowing full, and

equate this to a modified Fair and Hatch (5) estimate of

the quantity

r

_

2 = VVOIDS

ASOLIDS

(4a)

r

_

= d

_

(n − θWR)

3(1 − n + θWR) (4b)

The observed mean grain size of 3.8 x 10-4 m at Traverse City leads to a mean pore size value of 5.7 x 10-5 m,

in excellent agreement with the equation 3 value.

The water/avgas interface controls the water saturation

in the presence of the free avgas

SW = $1 + [βW (bW − b)]

α%(1/α)−1

 (bW >b >bM) (5a)

SW = 1 (b >bW) (5b)

βW = βL

[(ρ⁄ρL) − 1]σLA

σWL

(5c)

Figure 2. Vertical distribution of liquids in soil and a monitoring

well.

Table 1. Total and Avgas Saturation Profile Parameters at

Traverse City

Symbol Parameter Value

n Porosity 0.367

θWR Irreducible moisture content 0.059

ρL Avgas density 707 kg/m3

α Pore size uniformity 3.00

βL Avgas scaling factor 8.20 m-1

βW Water scaling factor 1.53 m-1

∆ Water table amplitude 0.35 m

Υ Trapping factor 40

Figure 3. Total saturation, Core 50BT.

76
with water density ρ and scaling factor βW predicted on

the water/avgas surface tension σWL. The upper extent

bM of free avgas may be estimated by equating the total

saturation (equation 2b) and the water saturation (equation 5a) at this elevation. A water/air interface with a

surface tension σWA governs the water saturation above

this depth

bM = βLbL − βW bW

βL − βw

(6a)

SW = (1 + $

βLσLA

σWA

[bL − bW + ρ

ρL

(bW − b]%

α)(1/α)−1

 (bM > b)

(6b)

The free avgas saturation SLF, in view of equation 1, is

simply given by

SLF = S − SW (7)

Figure 4 shows a typical free avgas profile at the site,

based upon the scaling factors of Table 1.

Vertical Distribution of Residual Avgas 

Residual avgas may be attributed to hysteretical trapping of product as it rises and falls through the water wet

soil over a fluctuating water table. Free liquids at a given

depth b experience progressively stronger capillary tensions as the water table falls due to their higher position

above this reference level. After the table attains a maximum depth and begins to rise, the water/avgas interface

becomes steadily larger due to a decreasing capillary

tension. The encroaching water occludes some of the

avgas, giving rise to a discontinuous residual fraction

within the water phase.

Historical maximum and minimum effective water saturations SWMAX, SWMIN are established by corresponding

minimum (bWMIN) and maximum (bWMAX) water table

depths, respectively

SWMIN = $1 + [βW (bWMAX−b)]

α%(1/α)−1 (8a)

SWMAX = $1 + [βW (bWMIN−b)]

α%(1/α)−1 (8b)

bWMIN = bWMAX − ∆ (8c)

Ostendorf et al. (3) infer the historical water table excursion amplitude ∆ cited in Table 1 from water level variations in nearby Lake Michigan. Parker and Lenhard (1)

suggest that the extreme saturations induced by extreme water positions trap residual avgas saturation SLR

that is given by

SLR = 1 −SWMIN

1 + Υ(1 −SWMIN)

 (bWMAX > b >bWMIN)

with empirical trapping factor Υ. Ostendorf et al. (3)

calibrate residual saturation profiles from their five stations with the trapping factor cited in Table 1, with the

typical results sketched in Figure 5.

Discussion

The vertical distribution of free and residual avgas has

important implications for the estimation of separate

phase contamination from monitoring well observations.

The depth integrated mass MLF of free avgas may be

estimated formally by integrating equations 2b, 5a, 6a,

Figure 4. Free avgas, Core 50CE. Figure 5. Residual avgas, Core 50CL.

77
and 7 from the water table to the upper limit of free avgas

occurrence, with the result

MLF = ρL∫

bM

bW

SLF db (10a)

MLF = ρL

βW

$βW (bW −bL) − (1 − (

BW

BL

) Ι [

βW (bW − bL)

1 − βW

βL

] %

(10b)

Ι(y) = ∫

0

y

(1 +y ′

α)(1/α)−1

dy ′ (10c)

The integral function Ι(y) varies with the uniformity exponent of the pore sizes, while its argument is determined by the scaling factors and the product thickness

in the monitoring well. Figure 6 displays the equilibrium

variation of the depth integrated mass predicted at Traverse City as a function of the monitoring well avgas

thickness. We note the exclusion of the residual mass

from this estimate, since the trapped avgas is not continuously connected to product in the monitoring well.

For a point of reference, Figure 5 implies a depth integrated residual avgas mass of about 5.3 kg/m2

; if this

mass were free avgas, we would observe a 0.3-m product thickness in a monitoring well.

The methylene chloride extract chromatograms are

composed primarily of known compounds (6), so that

the composition of the avgas can be examined at a

given depth. We note distillation of avgas as a function

of depth in the solid core samples, as suggested by a

profile from station 50BS (Figure 7). The open circles

correspond to relatively volatile avgas compounds, with

pure phase vapor densities of 0.27 to 1.00 kg/m3

 (at

12o

C). This fraction becomes more important in the

lower elevations due presumedly to decreased volatilization losses in the wetter region of the soil. The midrange volatiles (closed circles), with vapor densities of

0.14 to 0.27 kg/m3

, are quite uniformly distributed, while

the heavy compounds (triangles), with vapor densities

less than 0.14 kg/m3

 in magnitude, tend to dominate the

blend of hydrocarbons in higher, drier soil due to stripping away of the lighter fraction.

The partitioning of avgas into free and residual fractions

has important implications on its downgradient fate and

transport as well. Figure 2 suggests that the free avgas

occupies a relatively high-permeability, saturated soil

region below the avgas table and a lower-permeability,

unsaturated zone above the table. Since both these

regions share a common horizontal gradient due to the

slope of the avgas table, we anticipate slower and faster

zones of horizontal separate phase transport. This vertical profile of horizontal specific avgas discharge is very

nonuniform for sands with high values like that at the

site. The phenomenon can be approximated as a

pseudosorptive process, with a linear balance between

a mobile fraction and a reversible, immobile fraction

“sorbed” by capillary tension (7). Pursuing this analogy

further, the residual avgas can be thought of as an

irreversibly sorbed partition, lost from the mobile fraction

by hysteretical trapping. The net effect of these mechanisms is that the avgas travels with the underlying

ground water, but at a retarded velocity.

The free avgas is relatively easy to strip out of the soil

due to its direct interface with air. The residual avgas is

much more difficult to remediate, since it is surrounded

by essentially immiscible water. We accordingly expect

an initial period of relatively rapid remediation in response to soil venting or air sparging, followed by an

Figure 6. Depth integrated free avgas mass.

Figure 7. Avgas composition, 50BS.

78
asymptotically lower removal rate exacerbated by avgas

distillation. 

References 

1. Parker, J.C., and R.J. Lenhard. 1987. A model for

hysteretic constitutive relations governing multiphase flow, 1. Saturation pressure relations. Water

Resour. Res. 23:2,187-2,196. 

2. Lenhard, R.J., and J.C. Parker. 1990. Estimation of

free hydrocarbon volume from fluid levels in monitoring wells. Ground Water 28:57-67. 

3. Ostendorf, D.W., R.J. Richards, and F.P. Beck. 1993.

LNAPL retention in sandy soil. Ground Water 31:285-

292. 

4. van Genuchten, M.T. 1980. A closed form equation

for predicting the hydraulic conductivity of unsaturated soils. Soil Sci. Soc. Am. J. 44:892-898.

5. Fair, G.M., and L.P. Hatch. 1933. Fundamental factors governing the streamline flow of water through

sand. J. Am. Waterworks Assoc. 25:1,551-1,565. 

6. Ostendorf, D.W., L.E. Leach, E.S. Hinlein, and Y.F.

Xie. 1991. Field sampling of residual aviation gasoline in sandy soil. Ground Water Monitor. Rev.

11:107-120. 

7. Ostendorf, D.W. 1990. Long-term fate and transport

of immiscible aviation gasoline in the subsurface

environment. Water Sci. Tech. 22:37-44. 

79
Traverse City: Geochemistry and Intrinsic Bioremediation of BTX Compounds

Barbara H. Wilson, John T. Wilson, Don H. Kampbell, and Bert E. Bledsoe

U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK

John M. Armstrong

The Traverse Group, Inc., Ann Arbor, MI

Introduction

Loss of petroleum products from underground storage

tanks, pipelines, and accidental spills is a major source

of contamination of unsaturated soils, aquifer solids, and

ground water. Volatile aromatics such as benzene, toluene, ethylbenzene, and the xylenes (BTEX) are more

soluble in water than the aliphatic and higher molecularweight aromatic constituents of petroleum products (1).

Once released to the subsurface, petroleum compounds

are subject to aerobic microbial processes. The lowmolecular-weight alkanes and aromatics are readily

biodegraded in oxygenated ground water, depleting the

ground water of available oxygen (2,3). Reoxygenation

of the ground water may occur through reaeration from

soil gases, ground-water recharge, and usually inefficient mixing with surrounding oxygenated ground waters (4,5). Although ground waters near the perimeter of

the contaminant plume may be reoxygenated, the interior of the plume will remain anoxic for a distance downgradient. Anaerobic biological processes can account

for most of the removal of BTEX from the plume (6,7).

The biogeochemical mechanisms that contribute to anaerobic processes in the subsurface, however, are not

well understood.

The impact of anaerobic microbial processes on the fate

of monoaromatics, substituted aromatics, and chlorinated hydrocarbons in anoxic subsurface environments

has been studied in laboratory and field situations (8-

12). Field evidence of biotransformation of o-, m-, and

p-xylene was observed in methanogenic landfill

leachate with their preferential removals compared with

other alkylbenzenes present (13). Methanogenesis has

been observed at two sites with ground water contaminated with creosote (14,15). Denitrifying, iron-reducing,

sulfate-reducing, and methanogenic activities found in

ground water at the first site were highly correlated with

the biodegradation of creosote; methane was only

detected in the ground water that had been contaminated with creosote. Intermediate products of methane

fermentation, formate and acetate, were found in the

ground water at the second site.

In 1969 the unsaturated soil and ground water underlying

the U.S. Coast Guard Air Station at Traverse City, Michigan, were contaminated with an estimated 25,000 gal of

aviation gasoline when a flange in an underground storage tank failed. Dissolution of the aromatics in the

ground water resulted in concentrations of 36 mg to

40 mg of total alkylbenzenes per liter near the center of

the plume. The subsurface contamination existed as

residual-phase hydrocarbon, dissolved-phase aromatics, and gaseous hydrocarbons resulting from volatilization (16-18). The plume of dissolved-phase aromatics

extended from the air station into the East Arm of Grand

Traverse Bay and affected numerous drinking water

wells. The area near the leaking underground storage

tank was used to store degreasing solvents and to

conduct degreasing operations during aircraft maintenance. A small plume of chlorinated solvents lies adjacent to the gasoline plume.

Geochemical Characterization

Geochemical analyses of the water samples collected

at Traverse City revealed waters of four distinct geochemistries: 1) the heart of the plume, 2) an anaerobic

zone of treatment, 3) an aerobic zone of treatment, and

4) a pristine or renovated zone. The water from the heart

of the plume contained high concentrations of methane

and BTEX, with no detectable oxygen. These waters

were surrounded by an anaerobic zone of treatment with

greatly reduced concentrations of dissolved aromatics,

no oxygen, and substantial concentrations of methane.

Surrounding the anaerobic zone of treatment was an

aerobic zone of treatment with measurable oxygen, small

quantities of methane, and very low concentrations of

80
the alkylbenzenes. The perimeter of the plume was

surrounded by a renovated or pristine zone with high

concentrations of oxygen, no detectable alkylbenzenes,

and no methane.

Gas chromatography/mass spectrometry (GC/MS) analyses of the waters confirm the presence of BTEX at Traverse City. Also found were the chlorinated compounds

1,2-dichloroethane (1,2-DCA), tetrachloroethylene (PCE),

and 1,1,1-trichloroethane (TCA). The GC/MS analyses of

waters from wells R, S, and Q (Figure 1a) identified phenols and aromatic acids indicative of anaerobic microbial

action on the soluble aromatic constituents of petroleum

products (7,19,20). These compounds are found in portions of the plume with substantial concentrations of methane and no detectable oxygen, and are probably

precursors of the methane. Complete information on the

geochemical characterization of the ground waters at Traverse City, collection of aquifer material, analytical methods

used, and microcosm construction may be found in Wilson

et al. (21).

Laboratory Studies

To confirm field evidence of intrinsic bioremediation,

laboratory microcosm studies were conducted on aquifer material from the U.S. Coast Guard Air Station.

Material for aerobic and anaerobic fate studies was

collected from three locations in the plume (Figure 1a).

Aquifer material from site A (11.7 m to 12.3 m below land

surface), the zone of anaerobic treatment, was used to

construct the microcosms for the anaerobic fate study.

Aquifer material from site B (9.6 m to 10.2 m below land

surface), the aerobic zone of active biological treatment,

and from site C (6.6 m to 7.2 m below land surface), the

pristine or renovated zone, was used to prepare the

aerobic fate studies. Autoclaved controls were prepared

from the site C material. The compounds added to the

microcosms were benzene, toluene, p-xylene, o-xylene,

TCA, TCE, and chlorobenzene.

The initial compound concentrations and results of the

fate studies at various incubation times are shown in

Table 1. Benzene, toluene, p-xylene, and o-xylene were

biodegraded in both aerobic and anaerobic aquifer material. The removals were quite rapid for all compounds

in each of the three geochemical zones studied, whether

in the anaerobic zone of treatment, the aerobic zone of

treatment, or the renovated (pristine) material. By the

end of 8 wk of incubation of the compounds in the anaerobic aquifer material, the concentrations of benzene, toluene, p-xylene, and o-xylene had been reduced one order

of magnitude. The biotransformation of the four compounds in the aerobic zone of biological treatment occurred even more rapidly. By the end of 2 wk of incubation,

the concentrations for all four compounds were decreased

by two orders of magnitude. Similar losses were seen in

material from the renovated/pristine zone.

At the end of 4 wk of incubation, the respective concentrations for benzene, toluene, p-xylene, and o-xylene in

the autoclaved samples were 57 percent, 61 percent,

46 percent, and 49 percent of the original concentrations. Due to the removals of BTX in the controls, a

second set of autoclaved samples was prepared. The

concentrations after 16 wk of incubation in the duplicated controls were benzene, 83 percent; toluene, 66

percent; m- + p-xylene, 51 percent; o-xylene, 50 percent; TCA, 86 percent; TCE, 73 percent; and chlorobenzene, 58 percent. The cause for the removal of organics

in the controls has not been determined; however, sorption to aquifer solids probably occurred. Chlorobenzene

was also biodegraded in each of the three geochemical

zones studied. Decreases of one order of magnitude

were observed for chlorobenzene in both the aerobic

and anaerobic zones of biological treatment after 4 wk

U.S. Coast Guard


Air Station Boundary

Purge Well Field

Hangar 150 Meters

C

B

S

D & Q

P

R

A

Figure 1a. Locations of wells in an aviation gasoline plume at

the U.S. Coast Guard Air Station at Traverse City.

Well A is located in the plume below the purge field;

well B is located in the aerobic zone of treatment;

well C is located in the pristine region surrounding

the plume; and wells D, P, Q, R, and S are located

in the plume above the purge field.

μg/L

30,000


10,000


3,000


1,000


300


100


30


10

Toluene Benzene Xylenes

Figure 1b. Concentration of BTX in monitoring wells along a

flow path down the central axis of the plume in the

first quarter of 1986.

81
of incubation. Similar removal was seen in the renovated/pristine material after 14 wk of incubation. In the

autoclaved samples, 44 percent of the chlorobenzene

remained after 4 wk of incubation.

No significant biotransformation of TCA or TCE, compared with the controls, was observed at the end of

8 wk of incubation. Evidence of reductive dechlorination

of both compounds, however, was indicated by GC/MS

analyses of anaerobic microcosms at 95 wk of incubation by the identification of 1,1-dichloroethylene (1,1-

DCE) and 1,1-dichloroethane (1,1-DCA) (22,23).

Headspace concentrations of methane were measured

immediately before sampling to determine the maintenance of methanogenic conditions in the microcosms.

Methane was found in all the anaerobic samples, with

concentrations ranging from 50 ppm to 100 ppm; no

methane was found in the headspace of the aerobic or

autoclaved samples. 

Correspondence Between Laboratory and

Field Data

As part of a settlement with the State of Michigan, the

U.S. Coast Guard monitors alkylbenzene concentrations in selected monitoring wells quarterly. Three of the

monitoring wells (M30 near site S, M31 near site Q, and

M2 near site A in Figure 1) lie along a flow path. The

time required for water to move from one well to the next

can be estimated by dividing the distance between the

wells by the flow velocity (approximately 1.5 m per day).

This value was determined directly from tracer tests, and

is confirmed by calculations based on the hydraulic

conductivity of the aquifer and its hydraulic gradient (24).

Water takes 10 wk to flow from S to Q, and 24 wk to flow

from S to A. The first-order rate of biodegradation along

a segment of aquifer between the monitoring wells can

be estimated by dividing the concentration in the well

distal to the spill by the concentration in the proximate

well, taking the natural logarithm, then dividing by the

time required for water to flow between the wells.

Table 2 portrays the depletion of total BTX between S

and Q and between S and A for the years 1984 through

1987. The rate constants are surprisingly consistent. A

purge field was installed and put on line in mid-1985 to

prevent further migration of the plume from Coast Guard

property. As soon as the purge field was put into operation, the water behind the field near site A (Figure 1a)

became stagnant, and concentrations of BTX began to

drop. Solution concentrations of BTX dropped to low

values by late 1985 (data not shown). Because the

water was not moving, the decline in concentration over

time could be used to estimate the first-order rate constant for anaerobic BTX biotransformation in the part of

the aquifer near site A (Table 3).

Sediment samples for the microcosm study were acquired from site A in late 1985, and this study was

conducted in the first quarter of 1986. Figure 1b depicts

the concentration of BTX in monitoring wells along a flow

path down the central axis of the plume at that time.

Table 3 compares the depletion of BTX along the aquifer

Table 1. Behavior of Benzene, Alkylbenzenes, TCAa, TCEb, and Chlorobenzene in Aquifer Material From an Aviation Gasoline

Plume

Compound (μg/L Pore Water)

Subsurface

Material Week Benzene Toluene

m- + pXylene o-Xylene TCA TCE Chlorobenzene

A, anaerobic

treatment, 11.7 m

to 12.3 m below

land surface

0

4

8

95c

450

12

6

ndd

420

56

40

nd

440

78

17

nd

410

41

6

nd

570

420

580

73

540

260

340

54

500

66

34

3

B, active aerobic

treatment, 9.6 m

to 10.2 m below

land surface

0

2

4

14

450

2

5

2

420

2

5

2

390

1

2

1

390

2

1

nd

600

440

540

430

650

440

480

260

500

53

50

30

C, aerobic

renovated, 6.6 m

to 7.2 m below

land surface

0

2

4

14

420

4

1

—

380

3

3

—

370

3

1

—

370

3

1

—

600

580

620

650

540

540

600

500

500

100

106

67

C, autoclaved,

6.6 m to 7.2 m

below land

surface

0

2

4

420

380

240

380

290

230

370

200

170

370

190

180

600

590

560

650

540

430

500

300

210

a

1,1,1-Trichloroethane

b

Trichloroethylene

c

Concentrations at this time interval were determined by GC/MS; concentrations at other time intervals were determined by GC. All values are

means of triplicate analyses. d

Not detected, detection limit of 0.1 μg/L

82
segment S to Q during the first quarter of 1986 and the

depletion in the static water at site A in the last half of

1985 with the rates of anaerobic BTX biotransformation

in the microcosm study.

The rate of disappearance of BTX compounds as measured in aerobic microcosms does not compare well with

actual rates of aerobic BTX degradation in the field. The

rate of degradation in the field is controlled by masstransport limitations for oxygen (4,5), while laboratory

studies are limited by reaction rates. The anaerobic fate

study compared quite favorably with those rates measured by field data. Methanogenesis and other anaerobic

processes are not limited by the availability of oxygen in

either microcosms or subsurface materials. Mass-transport considerations are therefore not as critical to the

comparison of microcosm and field data. Anaerobic microcosms might prove to be a valuable tool to evaluate

intrinsic biorestoration of aquifers contaminated with petroleum products.

Conclusions

The results of the laboratory study confirm field evidence

of both aerobic and anaerobic transformation of alkylbenzenes and suggest that intrinsic aerobic and anaerobic in situ biorestoration of ground water contaminated

with petroleum products can occur. The anaerobic transformations seen at this site and confirmed by the laboratory study provide an attractive alternative to aerobic

restoration. The removals of the alkylbenzenes in the

anaerobic material were quite rapid and compared favorably with removals seen in the aerobic zone of treatment. Comparison of first-order rates of disappearance

in anaerobic microcosms with those calculated from

field data show acceptable agreement. Anaerobic processes in the subsurface are probably limited by in situ

reaction rates rather than by mass-transport limitations

for nutrients. Potentially, anaerobic microcosm studies

could be useful in the evaluation of intrinsic bioremediation of petroleum-contaminated subsurface materials.

The aerobic degradation of alkylbenzenes in subsurface

environments has been well documented (25,26) and is

currently the state-of-the-art for restoration of petroleum

contamination. Anaerobic biotransformation, however,

can enhance in situ biorestoration in oxygen-depleted

regions of a plume where heavily contaminated ground

water has excessive oxygen demand. Naturally occurring anaerobic biological processes can potentially remediate ground water contaminated with petroleum

products and significantly increase the reliability of existing remediation technologies.

References

1. Coleman, W.E., J.W. Munch, R.P. Streicher, H.P.

Ringhand, and F.C. Kopfler. 1984. The identification

and measurement of components in gasoline, kerosene, and no. 2 fuel oil that partition into the aqueous phase after mixing. Arch. Environ. Contam.

Toxicol. 13:171-178.

2. Atlas, R.M. 1981. Microbial degradation of petroleum hydrocarbons: An environmental perspective.

Microbiol. Rev. 45(1):180-209.

3. Gibson, D.T., and V. Subramanian. 1984. Microbial

degradation of aromatic hydrocarbons. In: D.T. Gibson, ed. Microbial degradation of organic compounds. New York: Marcel Dekker. pp. 181-252.

4. Borden, R.C., and P.B. Bedient. 1986. Transport of

dissolved hydrocarbons influenced by oxygenlimited biodegradation, 1. Theoretical development.

Water Resour. Res. 22(13):1,973-1,982.

Table 2. First-Order Rates of Anaerobic Biotransformation (per Week) of Total BTX Along Segments in the Aquifer

S to Q, Quarter of Year S to A, Quarter of Year

Year First Third First Third

1984 0.11 0.17 0.10 0.09

1985 0.02 0.27 0.07 cnca

1986 0.43 0.14 0.20 cnc

1987 0.34 0.10 0.33 cnc

a

Cannot calculate: the plume was intercepted by a purge well field and did not reach site A.

Table 3. Comparison of the First-Order Rates (per Week) of

Anaerobic Biotransformation of BTX

Compound Microcosmsa

Aquifer

Segmentb 

S to Q

Change Over

Time at Ac

,

6/85 to 10/85

Benzene 0.5 0.05 0.17

Toluene 0.3 1.3 0.47

m- +

p-Xylene

0.4 — —

o-Xylene 0.5 — —

All xylenes — 0.03 0.10

a

Laboratory microcosm studies

b

Along flow path segments in the aquifer

c

At a site with stagnant ground water behind a purge well field

83
5. Borden, R.C., P.B. Bedient, M.D. Lee, C.H. Ward,

and J.T. Wilson. 1986. Transport of dissolved hydrocarbons influenced by oxygen-limited biodegradation, 2. Field application. Water Resour. Res.

22(13):1,983-1,990.

6. Baedecker, M.J., D.I. Siegel, P. Bennett, and I.M.

Cozzarelli. 1988. The fate and effects of crude oil

in a shallow aquifer, 1. The distribution of chemical

species and geochemical facies. In: Mallard, G.E.,

and S.E. Ragone, eds. Proceedings of the Technical Meeting of the U.S. Geological Survey Toxic

Substances Hydrology Program, Phoenix, AZ (September 26-30). Water Res. Invest. Rep. 88-4220.

pp. 13-20.

7. Cozzarelli, I.M., R.P. Eganhouse, and M.J.

Baedecker. 1988. The fate and effects of crude oil

in a shallow aquifer, II. Evidence of anaerobic degradation of monoaromatic hydrocarbons. In: Mallard, G.E., and S.E. Ragone, eds. Proceedings of

the Technical Meeting of the U.S. Geological Survey Toxic Substances Hydrology Program, Phoenix, AZ (September 26-30, 1988). Water Res.

Invest. Rep. 88-4220. pp. 21-33.

8. Kuhn, E.P., P.J. Colberg, J.L. Schnoor, O. Wanner,

A.J.B. Zehnder, and R.P. Schwarzenbach. 1985.

Microbial transformations of substituted benzenes

during infiltration of river water to ground water:

Laboratory column studies. Environ. Sci. Technol.

19(10):961-968. 

9. Lovley, D.R., M.J. Baedecker, D.J. Lonergan, I.M.

Cozzarelli, E.J.P. Phillips, and D.I. Siegel. 1989.

Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339:297-300.

10. Lovley, D.R. and D.J. Lonergan. 1990. Anaerobic

oxidation of toluene, phenol, and p-cresol by the

dissimilatory iron-reducing organism, GS-15. Appl.

Environ. Microbiol. 56(6):1,858-1,864.

11. Suflita, J.M., S.A. Gibson, and R.E. Beeman. 1988.

Anaerobic biotransformations of pollutant chemicals in aquifers. J. Ind. Microbiol. 3:179-194.

12. Wilson, B.H., G.B. Smith, and J.F. Rees. 1986.

Biotransformations of selected alkylbenzenes and

halogenated aliphatic hydrocarbons in methanogenic aquifer material: A microcosm study. Environ.

Sci. Technol. 20(10):997-1,002.

13. Reinhard, M., N.L. Goodman, and J.F. Barker.

1984. Occurrence and distribution of organic chemicals in two landfill leachate plumes. Environ. Sci.

Technol. 18(12):953-961.

14. Ehrlich, G.G., D.F. Goerlitz, E.M. Godsy, and M.F.

Hult. 1982. Degradation of phenolic contaminants

in ground water by anaerobic bacteria: St. Louis

Park, Minnesota. Ground Water 20(6):703-710.

15. Goerlitz, D.F., D.E. Troutman, E.M. Godsy, and B.J.

Franks. 1985. Migration of wood-preserving chemicals in contaminated ground water from a sand

aquifer in Pensacola, Florida. Environ. Sci. Technol.

19(10):955-961.

16. Kampbell, D.H., J.T. Wilson, and D.W. Ostendorf.

1990. Simplified soil gas sensing techniques for

plume mapping and remediation monitoring. In:

Kostecki, P.T., and E.J. Calabrese, eds. Petroleum

Contaminated Soils. Chelsea, MI: Lewis Publishers. pp. 125-139. 

17. Ostendorf, D.W. 1990. Long-term fate and transport

of immiscible aviation gasoline in the subsurface

environment. Water Sci. Technol. 22:37-44.

18. Rifai, H.S., P.B. Bedient, J.T. Wilson, K.M. Miller,

and J.M. Armstrong. 1988. Biodegradation modeling at aviation fuel spill site. J. Environ. Eng.

114(5):1,007-1,029.

19. GrbiŒ-GaliŒ, D. 1989. Microbial degradation of homocyclic and heterocyclic aromatic hydrocarbons

under anaerobic conditions. Develop. Ind. Microbiol. 30:237-253.

20. GrbiŒ-GaliŒ, D., and T.M. Vogel. 1987. Transformation of toluene and benzene by mixed methanogenic

cultures. Appl. Environ. Microbiol. 53(2):254-260.

21. Wilson, B.H., J.T. Wilson, D.H. Kampbell, B.E. Bledsoe, and J.M. Armstrong. 1990. Biotransformation

of monoaromatic and chlorinated hydrocarbons at

an aviation gasoline spill site. Geomicrobiol. J.

8:225-240.

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Lorenzo. 1986. Sequential dehalogenation of chlorinated ethenes. Environ. Sci. Technol. 20(1):96-99.

23. Klecka, G.M., S.J. Gonsior, and D.A. Markham.

1990. Biological transformations of 1,1,1-trichloroethane in subsurface soils and ground water. Environ. Toxicol. Chem. 9:1,437-1,451. 

24. Sammons, J. 1994. Personal communication between John Sammons, The Traverse Group, Inc.,

and the authors.

25. Lee, M.D., J.M. Thomas, R.C. Borden, P.B. Bedient,

C.H. Ward, and J.T. Wilson. 1988. Biorestoration of

aquifers contaminated with organic compounds.

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contaminated waters. U.S. Patent Office 3,846,290.

84
Mathematical Modeling of Intrinsic Bioremediation at Field Sites

Hanadi S. Rifai

Energy and Environmental Systems Institute, Rice University, Houston, TX

Introduction

Intrinsic bioremediation is an important attenuation mechanism at contaminated field sites because it limits pollutant

migration and reduces contaminant mass in the subsurface. Quantifying the impact of intrinsic bioremediation at

a field site involves conducting a comprehensive field

sampling program in conjunction with extensive modeling

of contaminant transport and fate. This paper reviews the

basic data requirements for modeling intrinsic bioremediation at field sites and discusses some of the available

models that can be used for that purpose. The limitations

of existing models for simulating biodegradation are presented, as well as the difficulties in validating and verifying

these models. Finally, a few case studies of modeling

intrinsic bioremediation are reviewed.

Intrinsic Bioremediation Processes

Intrinsic bioremediation refers to the reduction of contaminant mass at a field site due to biodegradation. This reduction can occur under aerobic or anaerobic conditions.

When oxygen is utilized as the electron acceptor, the

process is referred to as aerobic respiration. When oxygen

is not present (anoxic conditions), microorganisms can use

organic chemicals or inorganic anions as alternate

electron acceptors. Anaerobic biodegradation refers to

fermentative, denitrifying, iron-reducing, sulfate-reducing,

or methanogenic processes. To quantify the impact of

intrinsic bioremediation on contaminant concentrations at

a field site, one needs to develop an accurate picture of

the distribution of electron acceptors both in pristine and

contaminated areas. The concentrations of the electron

acceptor in pristine areas provide an indication of the

biodegradation potential at the site. The disappearance or

decline of electron acceptor concentrations in contaminated areas provides an indication that biodegradation

may be occurring.

Existing Biodegradation Models

The biodegradation of contaminants in ground water is

mainly controlled by the rate of the reaction and the

availability of the electron acceptor. A mathematical expression that represents the chemical reaction can be

written to account for the effect that the rate of the

reaction has on biodegradation. This mathematical expression can then be combined with the transport equation to account for the electron acceptor limitation effect

on the biodegradation process in the subsurface.

Many biodegradation models have been developed in

recent years, most of which utilize some kinetic expression for biodegradation (see Table 1). The models listed

in Table 1 simulate a number of aerobic and anaerobic

biodegradation processes subject to specified conditions and assumptions. The difficulties involved in applying these models include 1) the data required as input

to the model is lacking, such as the kinetic rate parameters or estimates of the hydraulic conductivity; 2) the

majority of these models are proprietary and very few

are public domain models; and 3) modeling in general

is complicated and time-consuming, and requires a certain level of expertise.

Modeling of Intrinsic Bioremediation at

Field Sites

Many case studies of simulating biodegradation at field

sites exist in the general literature. In this paper, four

case studies are reviewed that combine field and laboratory investigation programs with modeling studies.

Conroe Superfund Site, Texas

Borden et al. (2) applied the first version of the

BIOPLUME model to simulate biodegradation at the

Conroe Superfund site in Texas. The United Creosoting

Company (UCC) site was operated as a wood preserving facility from 1946 to 1972. Wastes disposed in two

unlined ponds were composed of predominantly polycyclic aromatic hydrocarbons and pentachlorophenol

(PCP). Monitoring of the site has shown elevated levels

of organic contaminants in the soil and ground water, as

well as elevated levels of chloride in the ground water.

85
The ground-water velocity at the UCC site is approximately 5 m/yr.

Oxygen exchange with the unsaturated zone was simulated by Borden et al. (2) as a first-order decay in

hydrocarbon concentration. The loss of hydrocarbon

due to horizontal mixing with oxygenated ground water

and resulting biodegradation was simulated by generating oxygen and hydrocarbon distributions independently

and then combining them by superposition. Simulated

oxygen and hydrocarbon concentrations closely

matched the observed values. The Conroe Superfund

site was one of the first sites to be modeled using a

biodegradation expression in a transport model.

Traverse City Site, Michigan

The Traverse City field site is a U.S. Coast Guard Air

Station located in Grand Traverse County in the northwestern portion of the lower peninsula of Michigan. The

ground water at the site is contaminated with organic

chemicals from a leaking underground storage tank. The

main contaminants at the site are benzene, toluene, and

xylenes (BTX). The contaminant plume ranges from

150 ft to 400 ft wide and is about 4,000 ft long. A

pumping wellfield system was installed at the downgradient end of the dissolved plume to control offsite

migration.

A modeling effort of natural attenuation at the site was

completed by Rifai et al. (8) with the BIOPLUME II

model. Modeling was performed for the period before

the pumping wells were installed and also for the period

after the wells were turned on. The data in Figure 1 show

the results of the model simulation along the center line

of the plume for the period before the wellfield was

turned on. The model predictions by Rifai et al. (8)

matched the observed concentrations at the monitoring

wells reasonably well except in the vicinity of well M31.

Rifai et al. (8) indicated that this was because the simulation did not account for anaerobic biodegradation,

which was occurring in the interior of the plume.

Gas Plant Facility in Michigan

Soluble hydrocarbon and dissolved oxygen (DO) were

characterized in a shallow aquifer beneath a gas plant

facility in Michigan by Chiang et al. (14). The distributions

Table 1. Biodegradation Models

Name Description Author(s)

 – 1-D, aerobic, microcolony, Monod Molz et al. (1)

BIOPLUME 1-D, Monod Borden et al. (2)

 – 1-D, analytical first-order Domenico (3)

BIO1D 1-D, aerobic, anaerobic, Monod Srinivasan and Mercer (4)

 – 1-D, co-metabolic, Monod Semprini and McCarty (5)

 – 1-D, aerobic, anaerobic, nutrient limitations, microcolony,

Monod

Widdowson et al. (6)

 – 1-D, aerobic, co-metabolic, multiple substrates,

fermentative, Monod

Celia et al. (7)

BIOPLUME II 2-D, aerobic, instantaneous Rifai et al. (8)

 – 2-D, Monod MacQuarrie et al. (9)

BIOPLUS 2-D, aerobic, Monod Wheeler et al. (10)

ULTRA 2-D, first-order Tucker et al. (11)

 – 2-D, denitrification Kinzelbach et al. (12)

 – 2-D, Monod, biofilm Odencrantz et al. (13)

Observed Data


BIOPLUME II Model

80


70


60


50


40


30


20


10


0

M28 M26 TP4 TP3 M30 M31 IN2 M2


IN4 M4

Total BTX Concentration (mg/L)

Figure 1. BIOPLUME II model predictions for the Traverse City

field site (8).

86
of BTX in the aquifer had been monitored in 42 wells for

a period of 3 years. The site geology is characterized as

a medium to coarse sand with interbeds of small gravel

and cobbles. The general direction of ground-water flow

is northwesterly. The depth to water table ranges from

10 ft to 25 ft below land surface, and the slope of the

water table was estimated as 0.006. Based on groundwater and soil sampling data, Chiang et al. (14) concluded that the flare pit was the major source of the

hydrocarbons found in the aquifer, while the slope oil

tank was a secondary source.

Chiang et al. (14) evaluated a first-order decay biodegradation approach and the BIOPLUME II model for

simulating biodegradation at the gas plant facility. Using

the model and assuming first-order decay, several simulations were made to match the observed benzene concentration distribution of 1/22/85 by setting the observed

concentration distribution of 11/1/84 as the initial condition. The variables involved included the distribution of

the leakage/spill rates between the flare pit and the

slope oil tanks and macrodispersivities of the aquifer. 

The BIOPLUME II model was used to simulate the July

1987 data by setting the observed concentration distribution of February 1987 as an initial condition. The data

in Figure 2 show the comparison between the measured

and the simulated soluble BTX concentrations of July

1987. As can be seen from Figure 2, the model predictions for BTX were reasonable. The model predictions

for oxygen concentrations, however, were not as similar.

The authors attributed the differences to the fact that the

BIOPLUME II model assumes a requirement of 3 ppm

of oxygen for 1 ppm of benzene, whereas the actual

requirement is in the range of 1 ppm to 3 ppm.

Cliffs-Dow Superfund Site

Ground water at the Cliffs-Dow site is contaminated with

low levels of phenolic and polycyclic compounds. The

aquifer sediments at the site consist of mostly coarse

sands and gravels. The hydraulic conductivity ranges

between 3.5 x 10-3 to 4.6 x 10-2 cm/s. The principal

contaminants found at the site near the source area

include phenol, several methyl-substituted phenols, and

naphthalene at concentrations ranging from 220 μg/L to

860 μg/L. Based on the analysis of samples obtained

from monitoring wells, Klecka et al. (15) found that the

levels of organic contaminants are reduced to near or

below the detection limit within a distance of 100 m

downgradient from the source. Further analyses of the

ground-water chemistry were used to verify that biodegradation was occurring at the site and causing the

disappearance of the contaminants.

The migration of organic constituents in the aquifer was

simulated using the BIO1D model and assuming a firstorder decay expression. Half-lives for the contaminants

at the site were estimated from the results of soil

microcosm experiments based on the time required for

50-percent disappearances of the parent compound.

The velocity was varied over a range from 0.2 to 0.46 m/d,

which is representative of the range of ground-water

flow rates at the site.

Figure 3 illustrates the impact of biodegradation on contaminant concentrations at the site. Model simulations

performed using a half-life of 2 d indicated that levels of

the phenolic components were reduced by greater than

99 percent within a distance of 30 m downgradient of

the source. When the half-life was increased by a factor

of 10, the concentrations were reduced to a similar

extent within 75 m. Because of the dominance of biodegradation, increases in ground-water velocity from 0.2

to 0.46 m/d had minor effects on the level of attenuation

predicted with the model.

Conclusions

A number of biodegradation models have been developed over the last 10 years. These models are generally

4.0


3.0


2.0


1.0


0.0

4.0


3.0


2.0


1.0


0.0

0 50 100 150 200 250

Retardation factor = 5


Ground-water velocity = 0.2 m/d

0 50 100 150 200 250

Retardation factor = 5


Ground-water velocity = 0.46 m/d

t = 2 days 1/2

t = 10 days 1/2

t = 20 days 1/2

Distance (m)


(a)

t = 2 days 1/2

t = 10 days 1/2

t = 20 days 1/2

Distance (m)


(b)

Concentration (mg/L) Concentration (mg/L)

Figure 2. BIO1D model predictions for the Cliffs-Dow Superfund site (15).

87
similar in that they simulate the transport and biodegradation of a number of components in the ground water.

The models differ in the mathematical biodegradation

expressions that they use and in the numerical procedures used to solve the complicated system of equations. Application of these models to field sites has

proven to be complicated due to the lack of biodegradation parameters that can be measured in the field for

model input. As a result, most modeling applications at

the field scale have resorted to first-order decay or

instantaneous representation of the biodegradation

process.

References

1. Molz, F.J., M.A. Widdowson, and L.D. Benefield.

1986. Simulation of microbial growth dynamics coupled to nutrient and oxygen transport in porous

media. Water Resour. Res. 22(8):1,207-1,216.

2. Borden, R.C., P.B. Bedient, M.D. Lee, C.H. Ward,

and J.T. Wilson. 1986. Transport of dissolved hydrocarbons influenced by oxygen-limited biodegradation, 2. Field application. Water Resour. Res.

13:1,983-1,990.

3. Domenico, P.A. 1987. An analytical model for multidimensional transport of a decaying contaminant

species. J. Hydrol. 91:49-58.

4. Srinivasan, P., and J.W. Mercer. 1988. Simulation

of biodegradation and sorption processes in ground

water. Ground Water 26(4):475-487.

5. Semprini, L., and P.L. McCarty. 1991. Comparison

between model simulations and field results for in situ

biorestoration of chlorinated aliphatics, 1. Biostimulation of methanotrophic bacteria. Ground Water

29(3):365-374.

6. Widdowson, M.A., F.J. Molz, and L.D. Benefield.

1988. A numerical transport model for oxygen- and

nitrate-based respiration linked to substrate and nutrient availability in porous media. Water Resour.

Res. 24(9):1,553-1,565.

7. Celia, M.A., J.S. Kindred, and I. Herrera. 1989.

Contaminant transport and biodegradation, 1. A numerical model for reactive transport in porous media. Water Resour. Res. 25(6):1,141-1,148.

8. Rifai, H.S., P.B. Bedient, J.T. Wilson, K.M. Miller,

and J.M. Armstrong. 1988. Biodegradation modeling at aviation fuel spill site. J. Environ. Eng.

114(5):1,007-1,029.

9. MacQuarrie, K.T.B., E.A. Sudicky, and E.O. Frind.

1990. Simulation of biodegradable organic contaminants in ground water, 1. Numerical formulation in

principal directions. Water Resour. Res. 26(2):207-

222.

10. Wheeler, M.F., C.N. Dawson, P.B. Bedient, C.Y.

Chiang, R.C. Borden, and H.S. Rifai. 1987. Numerical simulation of microbial biodegradation of hydrocarbons in ground water. Proceedings of the

Solving Ground Water Problems With Models Conference, Denver, CO (February 10-12). Dublin, OH:

National Water Well Association (NWWA). 

11. Tucker, W.A., C.T. Huang, J.M. Bral, and R.E. Dickinson. 1986. Development and validation of the underground leak transport assessment model

(ULTRA). Proceedings of Petroleum Hydrocarbons

and Organic Chemicals in Ground Water: Prevention, Detection, and Restoration, Houston, TX (October-November). Dublin, OH: National Water Well

Association (NWWA). pp. 53-75.

12. Kinzelbach, W., W. Schäfer, and J. Herzer. 1991.

Numerical modeling of natural and enhanced denitrification processes in aquifers. Water Resour. Res.

27(6):1,123-1,135.

13. Odencrantz, J.E., A.J. Valocchi, and B.E. Rittman.

1990. Modeling two-dimensional solute transport

with different biodegradation kinetics. Proceedings

of Petroleum Hydrocarbons and Organic Chemicals

in Ground Water: Prevention, Detection, and Restoration, Houston, TX (October-November). Dublin,

OH: National Water Well Association (NWWA).

14. Chiang, C.Y., J.P. Salanitro, E.Y. Chai, J.D. Colthart,

and C.L. Klein. 1989. Aerobic biodegradation of

benzene, toluene, and xylene in a sandy aquifer:

Data analysis and computer modeling. Ground

Water 6:823-834.

15. Klecka, G.M., J.W. Davis, D.R. Gray, and S.S. Madsen. 1990. Natural bioremediation of organic contaminants in ground water: Cliffs-Dow Superfund

site. Ground Water 4:534-543.

Figure 3. BIOPLUME II model predictions for the gas plant facility in Michigan (14)—top number: observed data;

bottom number: simulated data.

88
Biogeochemical Processes in an Aquifer Contaminated by Crude Oil: 

An Overview of Studies at the Bemidji, Minnesota, Research Site

Robert P. Eganhouse, Mary Jo Baedecker, and Isabelle M. Cozzarelli

U.S. Geological Survey, Reston, VA

Abstract

Crude oil inadvertently released from a pipeline in a

remote area of north-central Minnesota has altered the

geochemistry of a shallow aquifer. Part of the oil was

sprayed over a large area to the west of the pipeline,

and another portion accumulated in an oil body that now

floats on the water table to the east of the point of

discharge. Dissolution of oil components into the ground

water and microbial degradation of the oil have resulted

in the formation of distinct geochemical zones in which

a variety of natural biogeochemical reactions can be

observed. Upgradient from the oil body in the “spray

zone,” concentrations of total dissolved organic carbon

(TDOC), Ca, Mg, and HCO3- are greater and pH is lower

than measurements observed in native ground water.

These differences reflect the transport of oil constituents

to the water table by recharge, oxidation under aerobic

conditions, and dissolution of carbonates. Beneath and

just downgradient from the oil body, oxygen is depleted,

and anaerobic degradation reactions (Fe and Mn reduction, methanogenesis) dominate. This is evidenced by

increased concentrations of Fe2+, Mn2+, CH4, and TDOC

in ground water. Volatile hydrocarbons, mainly benzene

and its alkylated derivatives, represent 26 percent of the

TDOC in this zone. Microbially mediated removal of

these compounds within the anoxic zone is indicated by

the presence of structurally related oxygenated intermediates (for example, alkylated benzenecarboxylic acids)

and differences in the removal rates of isomeric alkylbenzenes. Downgradient from the anoxic zone, mixing

of oxygenated water with plume constituents leads to

removal of iron (via precipitation) and virtually all of the

oil-derived organic constituents. Within a distance of

200 m downgradient from the oil body, the geochemistry

of the ground water is virtually indistinguishable from

that of native ground water. Data collected over an 8-yr

period demonstrate that while the contaminant plume has

become increasingly reducing in character, its size has not

changed significantly. This attests to the efficiency of

natural processes in removing/attenuating the oil-derived

contaminants of primary concern at this site.

Introduction

Contamination of ground water by intentional or inadvertent releases of crude oil or refined petroleum products

is a widespread problem, and a great deal of effort (and

money) is presently being devoted to remediation efforts. The efficacy of current engineering approaches is

subject to considerable debate. One thing is clear, however. For future remediation efforts to be effective and

useful, we must improve our understanding of how nature responds to such impacts. In principle, this should

permit the manipulation of natural processes for purposes of contaminant removal. Investigations that have

been carried out at the Bemidji, Minnesota, crude oil spill

site by the U.S. Geological Survey since 1984 are aimed

at developing such an understanding. The present paper

discusses some of the major findings of those studies.

Methods

Sampling and Analysis

Wells were installed at the study site using a hollow stem

auger without lubricants or grease on the equipment (1).

Augers for the drill rig and stainless steel well screens

were steam cleaned prior to use. The polyvinyl chloride

casings were 5 cm in diameter. Two types of wells were

installed. Water table wells were emplaced so that stainless steel screens (1.5 m long) intersected the water

table at the approximate midpoint of the screen; locations are shown in Figure 1. Deeper wells were installed

below the water table with screen lengths of 0.15 m or

0.61 m (Figure 2).

Samples of water, sediment, and oil were collected and

subjected to a variety of chemical analyses. Water samples to be used for determination of inorganic constituents and dissolved organic carbon were collected with

89
submersible pumps, whereas a Teflon bailer was used

to collect water for determination of organic constituents. The analyses included 1) water—dissolved oxygen (DO), volatile dissolved organic carbon (VDOC),

nonvolatile dissolved organic carbon (NVDOC), methane, pH, Eh, major cations (Ca, Mg, Na, K), alkalinity,

NH4+, NO3-, Cl-

, SO42-, sulfide, iron (total δ Fe2+), Mn2+,

Si, Ba, Al, Sr, δ13Cmethane, δ13CTIC (TIC=total dissolved

inorganic carbon), volatile hydrocarbons (VHCs), extractable hydrocarbons, XAD resin isolates, and lowmolecular-weight organic acids; 2) oil—δ13C, elemental

analysis (C,H,N,S), hydrocarbons, 13C-NMR, and heavy

metals; and 3) sediment—δ13C, hydrocarbons, and elemental analysis (C,H,N,S). Details of the methods of

sample collection and analysis are given by Eganhouse

et al. (2,3) for hydrocarbons; by Leenheer and Huffman

(4), Huffman and Stuber (5), and Thorn and Aiken (6) for

XAD resin isolates; by Baedecker and Cozzarelli (7),

Baedecker et al. (8), and Bennett et al. (9) for inorganic

constituents, methane, and elemental analyses; and by

Cozzarelli et al. (10,11) for low-molecular-weight organic

acids. 

Site Description 

The study site is in north central Minnesota near the

town of Bemidji (Figure 1). The aquifer is a pitted and

dissected glacial outwash underlain by a poorly permeable till at about 24 m below land surface. The outwash

sediments are heterogeneous and composed of moderately calcareous (6 percent carbonates), moderately to

poorly sorted sands consisting primarily of quartz and

feldspar of fine-to-medium grain size (1,12). Coring

studies have revealed that the sands are variably interbedded with gravel deposits and clay lenses. The water

table is 6 m to 10 m below land surface, and groundwater flow is to the east-northeast (Figure 1), discharging

into an unnamed lake approximately 300 m downgradient

of the spill site. Estimates of the flow velocities near the

water table range from 0.05 meters/day (m/d) to 0.5 m/d

(9) for fine-grained and coarse-grained sediment, respectively.

A pipeline rupture occurred in August 1979, spilling

1,670 m3

 of crude oil. A portion of the 410 m3 of crude

oil unaccounted for after the cleanup effort is present in

a body of oil floating on the water table. The oil body

described here is irregularly distributed over a 7-m to

8-m vertical interval of unconsolidated sediment above

the water table; by 1990, it had spread to a length of 70 m

to 80 m in the direction of ground-water flow (13). An

area upgradient of this oil body received oil spray during

the pipeline rupture, and crude oil coated only the surface sediment. This area, extending 140 m to 180 m to

the west-southwest of the pipeline and encompassing

approximately 6,500 m2

, is hereafter referred to as the

“spray area” (see Figure 1).

Minnesota

Study Site

Direction of Ground-Water Flow

318

518

532

534

317

522

301

421 319

315

604

Oil Body

Bemidji

Pipeline

Sampling Sites

Ground Water


Sediment


Oil


Oil and Sediment


Oil and Ground Water

N

0 50 100

310 Meters

603

 Oil


 Spray


Area

707

Figure 1. Map of the field site near Bemidji showing location

of the ruptured pipeline, approximate location of the

oil body, area over which oil was sprayed, and sampling locations (2).

-200 -150 -100 -50 0 50 100 150 200

Direction of


Ground-Water Flow

Zone IV

Oil


Body

522

532

518

310

603

707

604

534

435

430

425

420

415

Distance From Center of Oil Body (Meters)

Altitude Above Sea Level (Meters)

Land


Surface

Zone II

Zone I Zone V

Zone III

Water Table

Screened


interval of 


sampling sites

Figure 2. Cross section of the aquifer near Bemidji along main

sampling transect. Locations of water table and deep

wells indicated as filled bars (2). For a detailed description of zonation, see Baedecker et al. (8).

90
Results and Discussion

Effects of the Crude Oil on Ground-Water

Geochemistry

Formation of Geochemical Zones. The inadvertent

introduction of crude oil to this aquifer has resulted in

marked alteration of geochemical conditions in the

ground water. The driving force for these changes has

been the microbial degradation of metabolizable organic

matter, in this case represented by soluble constituents

of the oil. After several years of research during the mid

to late 1980s, it became evident that the saturated zone

of this aquifer could be characterized by geochemically

distinct zones. These zones, depicted in a cross section

centered on the main contaminant plume in the direction

of ground-water flow (Figure 2), correspond to I—uncontaminated native ground water, II—ground water upgradient of the oil body but within the “spray area,”

III—anoxic ground water immediately beneath and

downgradient of the oil body, IV—suboxic transitional

zone where anoxic ground water from zone III mixes

with more oxygenated ground water further downgradient, and V—oxygenated ground water that increasingly

approaches conditions of the native ground water. The

geochemistry of the ground water in each zone reflects

that of the water upgradient from it and processes (e.g.,

sorption, dilution, degradation, dissolution, gas exchange)

occurring within that zone.

The native ground water (zone I) is a dilute Ca-MgHCO3

-

 water with total dissolved solutes <400 mg/L,

a Ca:Mg ratio of 2.2:1, a median DO concentration of

7.68 mg/L, and a pH of 7.6 to 7.8. TDOC concentrations

are approximately 2 mg/L to 3 mg/L. The chemical composition is controlled by carbonate equilibria and, to a

lesser extent, by dissolution of quartz, feldspar, and clay

minerals and by the degradation of naturally occurring

organic material (8).

Ground water within zone II is influenced by the aerobic

degradation of soluble oil constituents originally deposited at land surface by the spraying of oil from the

ruptured high pressure pipeline. In this area, oil is known

to have penetrated the upper 4 cm to 6 cm of the soil.

The quantities were insufficient, however, for formation

of a discrete oil phase. Ground water in zone II exhibits

elevated concentrations of TDOC, Ca, Mg and HCO3-;

lower pH (about 0.5 to 1 pH unit); and a slightly reduced

DO concentration. Carbon dioxide partial pressures

(PCO2

) calculated by Bennett et al. (9) are as high as 10-1.5

atm (as compared with ambient atmospheric values of

10-3.5 atm). The TDOC concentrations are nearly an

order of magnitude greater than those found in the

native ground water. While the additional TDOC in zone

II is undoubtedly derived from the oil, virtually all of it

represents partially oxidized transformation products,

not hydrocarbons (2). No VHCs were detected. The few

petroleum hydrocarbons that are present have

compositions that indicate extensive biodegradation.

Presumably, some of the degradation/oxidation occurred in the unsaturated zone followed by transport of

soluble transformation products to the water table during

recharge. The slight depression of oxygen concentration

in ground water in this zone indicates that some degradation is likely to be occurring within the ground water

as well. If removal of the NVDOC is occurring via aerobic

respiration, however, the rate of supply of NVDOC must

exceed its rate of removal because NVDOC concentrations tend to increase with approach to zone III. Bennett

and others (9) have suggested that the large increases

in Ca, Mg, and HCO3- with increasing PCO2

, as well as

the similarity of the [Ca+Mg]/[HCO3-] mole ratio of

ground water in this zone to that of the native ground

water, indicate that the principal reaction responsible for

generating the plumes of alkali-earth solutes is dissolution of carbonates by reaction with carbonic acid. Because these processes are occurring under oxic (as

opposed to anoxic) conditions, the primary chemical

signatures of the contaminant plumes reflect the products of organic remineralization (HCO3-) and transformation (TDOC) and those resulting from acidification

(pH, Ca, Mg), not redox reactions.

Zone III encompasses ground water near, downgradient

from (about 75 m), and in immediate contact with the oil

to a depth of about 3 m below the water table. Dramatic

changes in ground-water geochemistry result from the

dissolution of soluble oil constituents and their metabolism, leading to complete consumption of oxygen and

the dominance of anoxic degradation reactions, including iron and manganese reduction and methanogenesis. Sulfate reduction and denitrification are not

important in this system because of the very low concentrations of SO42- and NO3- found in the native

ground water. In this zone, TDOC concentrations rise to

a maximum of 48 mg/L, and a significant fraction of this

is VDOC (42 percent). The volatile compounds are

dominated (approximately 63 percent) by a mixture of

saturated, aromatic, and alicyclic hydrocarbons derived

from the oil, the most important constituents of which are

benzene and a complex assemblage of alkylbenzenes.

Methane is found in the ground water, and field measurements of Eh indicate a strongly reducing environment. In 1987 the average stable carbon isotope ratios

for TDOC (δ13CTIC) and methane (δ13Cmethane) were -8.23

and -55.45 per mil, respectively. The average δ13C value

for TIC in native ground water is -12.55 per mil. The

heavier ratio for TIC in zone III, thus, reflects fractionation resulting from methanogenesis. The most dramatic

changes in inorganic chemistry are seen as large increases in the concentrations of dissolved iron and manganese that result from mobilization of these redox

species in response to the microbially mediated

oxidation of hydrocarbons. Silica concentrations also

91
increase due to enhanced dissolution of silicate minerals. Other compounds present in the ground water of

zone III, but not found in the oil or (except in trace

amounts) ground water in zones I and II, are a complex

mixture of oxygenated products of hydrocarbon degradation, including low-molecular-weight organic acids

(10,11). The acids are structurally related to coexisting

monoaromatic hydrocarbons from the oil. Those organic

acids that have been identified correspond to aromatic

hydrocarbons whose concentrations decrease most

rapidly within the anoxic zone, whereas no potential acid

intermediates were found for aromatics which appear to

be more stable within zone III. These results signal the

partial oxidation of hydrocarbons to more soluble metabolites.

Zone IV is a transition zone characterized by small but

detectable quantities of oxygen. Dissolved iron and silica decrease to below detection limits at the boundary

of this zone due to precipitation reactions. Low-molecular-weight organic acids are at or below detection limits,

and the concentrations of all oil-derived hydrocarbons

are much lower than in the ground water of zone III. Ca,

Mg, and Sr also decrease but rather gradually. Bennett

and others (9) have hypothesized that this is the result

of dispersive mixing and that transport of these constituents is conservative.

Evolution of the Contaminant Plume. Figure 3 depicts

the concentrations of dissolved Fe2+, Mn2+, methane,

and δ13CTIC from ground water taken from zone III near

the downgradient edge of the oil body for the years 1984

to 1992 (8,14). At this site, concentrations of methane

and Fe2+ increased by factors of 100 and 25, respectively, during the first 5 years. Thereafter, the concentrations of methane and Fe2+ have virtually leveled off.

Over the same period, manganese concentrations first

increased and then declined, whereas δ13CTIC has increased continuously. These variations in the contaminant

plume chemistry reflect evolutionary changes in the biology and geochemistry of this perturbed system with

continued supply of hydrocarbons and depletion of terminal electron acceptors.

In particular, it would appear that manganese reduction

was an important degradative process at the earliest

stages of plume evolution, but as time has progressed

iron reduction and methanogenesis have become dominant, presumably due to depletion of the supply of reducible manganese. As of 1992, the dominant process

at this location appears to be methanogenesis. This is

supported by the steady increase in δ13CTIC. During this

same period, the distribution and size of the contaminant

plume, as measured by the concentrations of monoaromatic hydrocarbons (2), have not changed appreciably.

While our understanding of the factors controlling plume

evolution remains limited and the size and shape of the

contaminant plume itself has not changed, it is clear that

the biogeochemistry within the plume is dynamic.

Transport and Fate of Oil-Derived

Contaminants

The following discussion considers the transport and

fate of the monoaromatic hydrocarbons within zones III

and IV. Benzene is the dominant individual VHC in

anoxic ground water near the leading edge of the oil

body (70 percent), reaching concentrations in excess of

9,300 μg/L. The C1-4 alkylated benzenes are second in

abundance, representing approximately 14 percent of

the total VHCs. Alkanes, consisting of a complex mixture

of normal and branched hydrocarbons with four to seven

carbon atoms, account for 12 percent of the VHCs.

Cyclic hydrocarbons (cyclopentanes and cyclohexanes),

cycloaromatics (e.g., alkylated indans), and heteroatomic species (e.g., tetrahydrothiophenes) are relatively

minor constituents (total less than 4 percent). Thirty-one

meters farther downgradient from this site (but still within

zone III), the VHC composition is markedly different. The

relative abundance of alkanes (and toluene; see Figure

4) is reduced, and benzene plus monoaromatic hydrocarbons with two to four alkyl carbon atom substituents

dominate (96 percent). Benzene represents 90 percent

of the total VHCs. These compositional differences are

also found among individual C1-4 (alkylated) benzenes.

There are four isomeric C2-benzenes, eight isomeric

C3-benzenes, and 22 isomeric C4-benzenes. All of these

compounds, with the exception of t-butylbenzene, which

was not detected in the crude oil, are present in the

contaminated ground water within zone III. The composition of the monoaromatic hydrocarbons changes systematically with distance downgradient from the oil body

and with increasing depth in the saturated zone (15,16).

Most importantly, isomeric alkylbenzenes show dramatically different apparent removal rates downgradient

from the oil body. Because these isomers have similar

physical properties, the attenuation of VHCs (and

10

1

0.1

0.01

0.001

1984

C

1988 1990 1992

13

δ

CH4

Mn2+

Fe2+

1986

-25

-20

-15

-10

-5

0

C13

δ TIC (per mil)

Figure 3. Concentrations of dissolved ferrous iron, manganese, methane (in millimoles), and δ13C in per mil of

TDOC for the years 1984 to 1992 (14).

92
therefore VDOC) is attributable to biological, rather than

physical, processes. 

This hypothesis is reinforced by data illustrated in Figure

4. Here concentrations of benzene, toluene, ethylbenzene, and a conservative tracer (see Baedecker et al.

[8] for discussion) are shown as a function of distance

downgradient from the center of the oil body. Two features of these data are readily apparent. First, the concentration changes observed for these aromatic

hydrocarbons greatly exceed what would be predicted

on the basis of conservative transport. Second, because

benzene, toluene, and ethylbenzene form an homologous series with increasing KOWs (octanol-water partition coefficients), one would expect that a systematic

pattern of removal rates would be found if sorption was

the dominant process. This obviously is not the case, as

toluene is very rapidly removed by comparison with

either benzene or ethylbenzene. Clearly, biodegradation

is the most important process limiting the transport of

these hydrocarbons.

Conclusions

At the Bemidji field site, the introduction of crude oil to

the subsurface has resulted in dramatic changes in the

geochemistry of the ground water. Even so, natural

processes, with biodegradation being most important,

have effectively limited the transport of oil-derived contaminants to a distance of 200 m from the source.

Continued monitoring of the contaminant plume has

revealed the dynamic nature of the coupled biological

and geochemical processes operative at this site. An

understanding of these processes and the factors that

affect plume evolution will be essential if we are to

develop environmentally responsible remediation methods in the future.

References

1. Hult, M.F. 1984. Ground-water contamination by

crude oil at the Bemidji, Minnesota, research site:

An introduction. In: Proceedings of the U.S. Geological Survey toxic-waste technical meeting, Tucson, AZ (March 20-22). Water Res. Invest. Rep.

94-4188. pp. 1-15.

2. Eganhouse, R.P., M.J. Baedecker, I.M. Cozzarelli,

G.R. Aiken, K.A. Thorn, and T.F. Dorsey. 1993.

Crude oil in a shallow sand and gravel aquifer, II.

Organic geochemistry. Appl. Geochem. 8:551-567.

3. Eganhouse, R.P., T. Dorsey, C. Phinney, and S.

Westcott. 1993. Determination of C6-C10 aromatic

hydrocarbons in water by purge-and-trap capillary

gas chromatography. J. Chromatog. 628:81-92.

4. Leenheer, J., and E.W.D. Huffman. 1979. Analytical

method for dissolved organic carbon fractionation.

Water Res. Invest. Rep. No. 79-4.

5. Huffman, E.W.D., and H.A. Stuber. 1985. Analytical

methodology for elemental analysis of humic substances. In: Aiken, G.R., D.M. McKnight, R.L. Wershaw, and P. MacCarthy, eds. Humic substances in

soil, sediment and water. New York, NY: John Wiley

and Sons. pp. 433-456.

6. Thorn, K.A., and G.R. Aiken. 1989. Characterization of

nonvolatile organic acids resulting from the biodegradation of crude oil by nuclear magnetic resonance

spectrometry. In: Mallard, G.E., and S.E. Ragone, eds.

Proceedings of the technical meeting of the U.S. Geological Survey Toxic Substances Hydrology Program,

Phoenix, AZ (September 26-30, 1988). pp. 41-52.

7. Baedecker, M.J., and I.M. Cozzarelli. 1992. The

determination and fate of unstable constituents in

contaminated ground water. In: Lesage, S., and

R.E. Jackson, eds. Ground-water quality and analysis at hazardous waste sites. New York, NY: Marcel

Dekker. pp. 425-461.

8. Baedecker, M.J., I.M. Cozzarelli, D.I. Siegel, P.C.

Bennett, and R.P. Eganhouse. 1993. Crude oil in a

shallow sand and gravel aquifer, III. Biogeochemical reactions and mass balance modeling in anoxic

ground water. Appl. Geochem. 8:569-586.

9. Bennett, P.C. D.I. Siegel, M.J. Baedecker, I.M. Cozzarelli, and M.F. Hult. 1993. Crude oil in a shallow

sand and gravel aquifer, I. Hydrogeology and inorganic geochemistry. Appl. Geochem. 8:529-549.

10. Cozzarelli, I.M., M.J. Baedecker, R.P. Eganhouse,

and D.F. Goerlitz. 1994. The geochemical evolution

of low-molecular-weight organic acids derived from

105

104

103

102

101

10-2

10-1

100

20 40 60 80 100 120 140 160

Distance From Center of Oil Body (meters)

Concentration (μg/L)

Conservative Solute

Benzene

Ethylbenzene

Toluene

Figure 4. Concentrations of benzene, toluene, and ethylbenzene (μg/L) along the flowpath of the contaminant

plume. Also shown is the concentration of a conservative solute with an assumed starting concentration

of 10,000 μg/L. Model assumptions discussed in

Baedecker et al. (8); figure from Mallard and

Baedecker (17).

93
the degradation of petroleum contaminants in ground

water. Geochim. Cosmochim. Acta 58:863-877.

11. Cozzarelli, I.M., R.P. Eganhouse, and M.J. Baedecker.

1990. Transformation of monoaromatic hydrocarbons

to organic acids in anoxic ground-water environment.

Environ. Geol. Water Sci. 16:135-141.

12. Franzi, D.A. 1985. The surficial and subsurface distribution of aquifer sediments at the Bemidji research site, Bemidji, Minnesota. Presented at the

Second Toxic-Waste Technical Meeting, Cape Cod,

MA (October 21-25). 

13. Essaid, H.I., W.N. Herkelrath, and K.M. Hess. 1991.

Air, oil, and water distributions at a crude-oil spill

site, Bemidji, Minnesota. In: Mallard, G.E., and D.A.

Aronson, eds. Proceedings of the technical meeting

of the U.S. Geological Survey Toxic Substances

Hydrology Program, Monterey, CA (March 11-15).

pp. 614-620.

14. Baedecker, M.J., and I.M. Cozzarelli. 1994. Biogeochemical processes and migration of aqueous constituents in ground water contaminated with crude

oil. In: Dutton, A.R., ed. Toxic substances and the

hydrologic sciences. Minneapolis, MN: American

Institute of Hydrology. pp. 69-70.

15. Cozzarelli, I.M., R.P. Eganhouse, and M.J.

Baedecker. 1989. The fate and effects of crude oil

in a shallow aquifer, II. Evidence of anaerobic degradation of monoaromatic hydrocarbons. In: Mallard, G.E., and S.E. Ragone, eds. In: Proceedings

of the technical meeting of the U.S. Geological Survey Toxic Substances Hydrology Program, Phoenix, AZ (September 26-30, 1988). Water Res.

Invest. Rep. 88-4220. pp. 21-34.

16. Eganhouse, R.P., T.F. Dorsey, C.S. Phinney, M.J.

Baedecker, and I.M. Cozzarelli. 1987. Fate of

monoaromatic hydrocarbons in an oil-contaminated

aquifer: Evidence for the importance of microbial

activity. In: Proceeding of the Geological Society of

America 1987 annual meeting, Phoenix, AZ.

19:652.

17. Mallard, G.E., and M.J. Baedecker. 1993. Hydrocarbon transport and degradation in ground water:

U.S. Geological Survey investigations. In: Pare,

K.M., ed. Proceedings of the Air Combat Command

environmental quality 1993 symposium, Langley

AFB, VA (March 1-5) pp. 102-108.

94
Simulation of Flow and Transport Processes at the Bemidji, Minnesota,

Crude-Oil Spill Site

Hedeff I. Essaid

U.S. Geological Survey, Menlo Park, CA

Abstract

The Bemidji, Minnesota, field site has provided an opportunity to study in detail the natural processes that

occur following a crude-oil spill. Detailed field studies

have characterized the subsurface oil distribution and

the characteristics of the contaminated ground-water

plume. Numerical models that simulate flow and transport processes are useful tools for integrating information collected in the field, for testing hypotheses, and for

studying the relative importance of simultaneously occurring processes in complex, real-world systems. Numerical modeling of multiphase flow at the Bemidji site

has illustrated the importance of spatial variability on the

movement and distribution of oil in the subsurface. Solute-transport modeling that includes aerobic and anaerobic degradation processes is being used as a tool

to study the field-scale, naturally occurring solute-transport and degradation processes occurring at the site. 

Introduction

On August 20, 1979, a buried oil pipeline near Bemidji

broke, spilling about 1.7 x 106

 L (11,000 barrels) of crude

oil (Figure 1). The site is located in a pitted and dissected

glacial outwash plain. Depth to the water table ranges

from 0 m to 8 m below land surface, and the flow through

the aquifer is generally horizontal and northeastward

towards an unnamed lake 300 m downgradient from the

point of pipeline rupture. An estimated 1.2 x 106

 L (7,800

barrels) of the spilled oil was removed by pumping from

surface pools, trenching, burning, and excavating soil

(1). The petroleum in the pipeline was under pressure,

causing oil to be sprayed over approximately 6,500 m2

when the pipeline broke. The oil collected in topographic

depressions and trenched areas where large volumes

of oil infiltrated into the subsurface, forming two main

bodies of oil floating on the water table. The subsurface

oil bodies provide a long-term, continuous source of

hydrocarbon components that dissolve in and are transported with the flowing ground water. 

Numerical models that simulate flow and transport processes are useful tools for integrating information collected in the field, for testing hypotheses, and for studying

the relative importance of simultaneously occurring processes in complex, real-world systems. Many researchers

working at the Bemidji site have focused considerable

effort on characterizing the subsurface distribution of the

oil and the nature of the resulting contaminated groundwater plume. This paper reviews this work, then briefly

summarizes the results of numerical simulations of multiphase flow and of ground-water transport and biodegradation at the site.

The Observed Subsurface Oil Distribution

The Bemidji field site has provided an opportunity to

study in detail the subsurface oil distribution following a

Direction of


Ground- Water Flow

B'

Lake

95 5'40" 95 5'4"

47 34'32"

Area of


Pipeline


Break

B

Trace of North Pool


Simulated Section

A

A'

Pipeline

Railroad


Track

Scale

0 400 Feet

0 50 100 150 Meters

Minnesota

Bemidji

47 34'14"

Figure 1. The north oil pool at the Bemidji crude-oil spill site.

Plus symbols are locations of boreholes where

samples were collected for saturation analyses, and

circles are wells sampled for ground-water concentrations. The line A-A′ is the trace of the multiphase

flow simulation section, and B-B′ is the trace of the

transport simulation section.

95
spill. Field cores were collected for the purpose of determining the oil-saturation distribution (the fraction of

the pore space that is occupied by oil) in the subsurface.

Determination of the subsurface fluid-saturation distributions required the implementation of a sampling technique that could recover relatively undisturbed core

samples from the unsaturated and saturated zones

while maintaining the in situ pore-fluid distribution (2). To

improve field sample collection, a freezing-tip core barrel was developed and used for sample collection (3).

To allow visual inspection of the cores in the field, clear

polycarbonate liners, 47 mm in diameter and 1.5 m long,

were used within the core barrel.

Following retrieval, the cores were frozen and cut into

78-mm long subsamples using a circular saw fitted with

a masonry blade. The oil saturation of each core was

determined in the laboratory using a porous polyethylene (PPE) technique (2,4). In this process, strips of

hydrophobic PPE are placed into a slurry created by

adding water to a core sample. The PPE absorbs the oil

from the sample but does not take up water. The amount

of oil present in the core is calculated from the change

in weight of the oily PPE strips. The sample is then dried

in an oven, and water saturation is determined gravimetrically. Air saturation can then be calculated by subtracting the sum of oil and water saturation from unity.

Following the saturation analysis, each sample was

sieved to obtain the particle-size distribution.

To date, samples have been collected and analyzed

from 10 boreholes (550 subsamples) at the site of the

north oil pool (Figure 1). The three-dimensional distribution of oil saturation at the north pool obtained by kriging

the observed data is shown in Figure 2 (5,6). The oil

distribution at this site is complex. A considerable

amount of oil remains in the unsaturated zone at locations where oil infiltrated from a trench excavated following the spill. The maximum oil saturation measured at

this site (0.74) was located downgradient from the zone

of oil infiltration. The body of oil floating on the water

table is not lens-shaped. The oil-saturation contours

follow a sinuous path that is roughly parallel to the

direction of ground-water flow. Thin silt lenses at the

north pool site appear to have considerable influence on

the observed oil-saturation distribution. These layers

result in high residual oil saturations in the unsaturated

zone and cause the shape of the oil body floating on the

water table to be complex and irregular rather than

lens-shaped. 

Geostatistical and multiphase flow simulation have been

used to assess the effect of spatial variability of hydraulic properties on the oil-saturation distribution at the site.

Because of the difficulty of measuring sediment hydraulic properties for oil-contaminated samples, the particlesize distribution data were used to estimate the

permeabilities (k) of the field samples and the retention

curves (7). The mean and variance of the log(k) distribution

(k in m2

) were -11.2 and 0.25, respectively, at the north

pool. Figure 3 shows a cumulative probability plot of

log(k) at the site. A linear relation indicates a lognormal

distribution of k, and the slope of the line is related to

the variance of k. The north pool plot of log k appears

to consist of two linear segments, a distribution that

suggests that the permeability distribution at this site

consists of two lognormal populations: a coarse fraction

(log(k) >-11.64) and a fine fraction (log(k) <-11.64).

To obtain a regular grid of k values needed for the

multiphase flow simulations, geostatistical simulation

techniques (8) were used to generate k distributions for

the north pool site that were conditioned on the values

estimated from the core samples. A permeability realization was obtained that reproduced the geometry of the

fine and coarse fractions and also reproduced the variability structure within these fractions. The details of this

process are explained by Dillard (9) and Dillard et al. (10).

0 Oil Saturation .75

Kriged North Pool Oil Saturations

422

430

Elevation


(m)

8

-3

20 Distance (m) 90

Figure 2. Three-dimensional oil saturation distribution at the

north oil pool site. The water table is at an elevation

of about 423.5 m above sea level.

-15. -14. -13. -12. -11. -10. -9.0

Log (k)

0.01

0.1

0.2

1

2

5

10

20

99.99

99.9

99.8

99

98

95

90

80

70

6050

40

30

Cumulative Probability

Figure 3. Cumulative probability plot of log (permeability) at the

north pool site.

96
Multiphase Flow Modeling

A two-dimensional numerical model was developed and

used to simulate multiphase flow along a longitudinal

vertical transect parallel to the direction of flow at the

north oil pool. The model solves a mass balance equation for the oil and water phases, assuming that the air

phase is maintained at atmospheric pressure. An important feature of the model is that it incorporates hysteretic

relations between capillary pressures and fluid saturations. The details of the model and the approaches used

are given by Essaid et al. (7).

The model was used to simulate subsurface flow from

the time of the spill in August 1979 until the samples

were collected in June 1990. The simulated section was

120 m long and 10 m deep. The initial condition for the

simulation was a hydrostatic water pressure distribution

corresponding to the measured water-table elevation at

the time of sampling. The lateral and bottom boundaries

were assumed to be hydrostatic water-pressure boundaries. Oil was assumed to infiltrate through five constant

oil pressure nodes at the top boundary representing the

trench that was excavated following the spill. Oil infiltration was stopped when the cumulative oil infiltration was

equal to the estimated oil mass in the observed transect.

More details of the north pool simulation are given by

Dillard (9) and Dillard et al. (10).

Data from the first sample transect (Figure 4a) were compared with the two-dimensional simulated oil-saturation

distribution. In the first simulation, a uniform mean k of

5 x 10-12 m2 and uniform mean capillary pressure/saturation function parameters were used. A uniform k resulted in a symmetric, lens-shaped oil body (Figure 4b)

that did not reflect the features of the observed oil

distribution. In the second simulation, spatial variability

of hydraulic properties was introduced, resulting in variability in oil saturations within the lens (Figure 4c). Very

little oil was trapped in the unsaturated zone. The shape

of the oil body floating on the water table became complex and irregular, with zones of low oil saturation corresponding to the low permeability silt layers. In the third

simulation, hysteresis was introduced, causing more oil

to become entrapped in the unsaturated zone (Figure

4d). The modeling results suggest that the silt layers and

spatial variability exert a strong control on the oil distribution in the subsurface.

The Observed Ground-Water Plume

Numerous researchers working at the Bemidji crude-oil

spill site have documented evidence for microbial activity and degradation of petroleum hydrocarbons in the

field (11-18). As a result of these studies, five geochemical zones in ground water (Figure 5) have been

identified at the site (11,12,19). 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 downgradient from

the separate-phase oil body, consists of an anoxic

plume of ground water 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.

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 body has remained

relatively stable with time. In the anoxic zone (Zone 3),

0 Oil Saturation 1.0

0 Distance (m) 120

Elevation


(m)

420

d 430

0 Distance (m) 120

Elevation


(m)

420

430 c

0 Distance (m) 120

Elevation


(m)

420

430 b

15 Distance (m) 100

Elevation


(m)

420

430

a

Figure 4. Oil saturation distributions along the sample transect: a) observed oil saturations, b) simulated oil

saturations with uniform mean properties, c) simulated oil saturation with spatially variable properties,

d) simulated oil saturation with spatially variable

properties and hysteresis.

97
concentrations of reduced manganese (Mn) and iron

(Fe) and of methane have increased with time, indicating a sequence of Mn reduction followed by Fe reduction

and methanogenesis.

Ground-Water Transport and

Biodegradation Modeling

A two-dimensional, multispecies solute-transport model

that incorporates biodegradation is being developed

and applied to the ground-water system at the Bemidji

site. The model is being used to quantify the field-scale

degradation processes and to identify the important factors affecting the distribution of solute species in the

field. The model simulates the aerobic and anaerobic

degradation processes that have been observed in the

contaminated ground-water plume at the spill site. The

U.S. Geological Survey’s Method of Characteristics

transport model (20,21) was expanded to handle multiple solutes and to include biodegradation terms. The

approach of Kindred and Celia (22) was used to represent the biodegradation terms in the transport equation.

Details of the model are given by Essaid et al. (23).

A vertical cross section of unit width that is approximately parallel to the direction of ground-water flow was

simulated using the transport model for the period from

the time of the spill in 1979 until September 1990 (Figure

5). Ground-water samples from numerous wells along

this section have been analyzed over time (11,12,15).

Steady-state flow, no sorption, and isothermal conditions were assumed. Four solutes and two microbial

populations were modeled. TDOC was split into two

fractions: degradable dissolved organic carbon (DDOC)

and refractory dissolved organic carbon (RDOC). The

remaining two solutes modeled were DO and methane.

To represent the aerobic and anaerobic degradation

processes, aerobic and methanogenic populations of

bacteria were included in the simulations. 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. Iron and manganese

reductions were not included because of the complexity

of incorporating the rock-water interactions of dissolution and precipitation into 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, it was assumed that DDOC

and RDOC dissolved and entered the aquifer with recharge water. The estimated values of initial concentrations and recharge water concentrations for each solute

are given in Table 1. The oil present in the pore space

within the oil body reduces water 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 recharge rate

in the zone of the oil body were reduced to 25 percent

of the aquifer values. No measurements or estimates

are available for many of the transport and biodegradation parameters under natural field conditions. Therefore, reasonable estimates of these values were used in

the simulations. The details of the simulation parameters

and boundary conditions are given by Essaid et al. (23).

Transport Simulation Results

Observed and simulated profiles of TDOC, DO, and

methane near the water table are plotted in Figure 6.

The observed and simulated concentration profile of

TDOC is shown in Figure 6a. 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.

Recharge Zones

A DE B C

Oil Body

Zone 2 Zone 3

Zone 1

Explanation

Midpoint of Well


Screen


Datum Is Sea Level

0 100 200

420

416

424

Altitude, in Meters

Distance, in Meters

Direction of


Ground-Water Flow

Zone 4

Zone 5

300 400

Figure 5. The simulated cross section showing the five geochemical zones in ground water.

98
Table 1. Initial and Recharge Water Concentration (mg/L)*

Recharge Zone

Solute

Initial

Concentration A B C D E

DDOC 0.0 0.0 10.0 100.0 0.0 0.0

RDOC 2.0 2.0 20.0 30.0 2.0 2.0

TDOC 2.0 2.0 30.0 130.0 2.0 2.0

DO 9.0 9.0 3.0 0.0 0.0 9.0

Methane 0.0 0.0 0.0 0.0 0.0 0.0

*Recharge zones A through E are shown in Figure 5.

a c TDOC Methane

80

60

40

20

0

30

20

0

0 100 200 300 400

Concentration, in Milligrams per LiterConcentration, in Milligrams per Liter

0 100 200 300 400

b

10

6

4

2

0

0 100 200 300 400

Concentration, in Milligrams per Liter

DO

10

Distance, in Meters Distance, in Meters

Distance, in Meters

8

Explanation

1986 Simulated


1990 Simulated


1986 Observed


1987 Observed


1988 Observed


1990 Observed


Figure 6. Graphs of simulated and observed concentrations at the water table: a) TDOC, b) DO, c) methane.

99
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. Downgradient from the oil body, the TDOC concentration

decreases gradually to the background concentration.

This decrease is a result of microbial and physical processes. There is anaerobic degradation of DDOC within

the anoxic zone near the oil body and aerobic degradation of DDOC at the margins of the plume, where oxygenated recharge and ground water are encountered.

Also, there is dilution of TDOC downgradient from the

oil body as a result of the physical processes of displacement and mixing of flowing ground water with recharge water.

The simulated DO concentration (Figure 6b) decreases

in the spray zone because of the assumed decrease in

DO concentration in recharge water, caused by degradation of hydrocarbons in the unsaturated zone and by

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. Farther downgradient, DO begins to increase as oxygenated recharge

water enters the system.

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 predicted decline in methane concentration at a

distance of 230 m (Figure 6c) is a result of the upwelling

of oxygenated water caused by the upward bending of

flow lines around the oil body. The simulated profiles

show a marked increase in methane production from

1986 to 1990 as the population of methanogens increases. This increase in methanogenesis results in a

slight decrease in TDOC concentrations from 1986 to

1990 (Figure 6a). The rate of increase in methane production was quite sensitive to the biodegradation parameters used in the simulation.

To examine the effect of degradation on DDOC in the

aquifer, two-dimensional distributions of DDOC for three

different simulations are shown in Figure 7. In the first

simulation, there is no degradation (Figure 7a); in the

second simulation, degradation occurs (Figure 7b); and

in the third simulation, degradation occurs and the hydraulic conductivity distribution is heterogeneous (Figure 7c). The distribution of DDOC for the case with no

degradation (Figure 7a) reflects the physical processes

of dispersion, diversion of flow around the oil body, 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 (Figure 7b). 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 (7,9). To make the

simulation more realistic, a heterogeneous hydraulic

conductivity distribution was created using the methods

of Dillard et al. (10) and was used in the transport model.

Because of the complex flow field, an irregularly shaped

plume develops (Figure 7c). The variability in flow paths

and flow velocities results in increased mixing and dispersion of ground water. This, in turn, results in increased biodegradation. In this simulation, 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 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 caused by heterogeneity of the hydraulic

conductivity.

Summary

The Bemidji crude-oil spill site has provided an opportunity to study in detail the processes that occur following

a spill of an organic immiscible fluid that is slightly

soluble in water. Detailed field studies have characterized the subsurface oil distribution and the characteristics of the contaminated ground-water plume.

Numerical modeling of multiphase flow has illustrated

the importance of spatial variability on the movement

and distribution of oil in the subsurface. Solute-transport

modeling that includes aerobic and anaerobic degradation processes is being used as a tool to study the

field-scale solute-transport and degradation processes.

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. 

References

1. Hult, M.F. 1984. Ground-water contamination by

crude oil at the Bemidji, Minnesota, research site:

100
An introduction. In: Hult, M.F., ed. Ground-water

contamination by crude oil at the Bemidji, Minnesota, research site. Water Res. Invest. Rep. 84-

4188. pp. 1-15.

2. Hess, K.M., W.N. Herkelrath, and H.I. Essaid. 1992.

Determination of subsurface fluid contents at a

crude-oil spill site. J. Contam. Hydrol. 10:75-96.

3. Murphy, F., and W.N. Herkelrath. 1994. A samplefreezing drive shoe for a wireline-piston core sampler. In: Morganwalp, D.W., and D.A. Aronson, eds.

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Program, Colorado Springs, CO (September 20-24,

1993). Water Res. Invest. Rep. 94:4014.

4. Cary, J.W., J.F. McBride, and C.S. Simmons. 1991.

Assay of organic liquid contents in predominantly

water-wet unconsolidated porous media. J. Contam. Hydrol. 8:135-142.

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

1994. Field and modeling studies of multiphase fluid

flow at the Bemidji, Minnesota, crude-oil spill site.

In: Dutton, A.R., ed. Toxic Substances and the Hydrologic Sciences. Minneapolis, MN: American Institute of Hydrology. pp. 52-68. 

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

1994. Multiphase fluid distributions in a shallow

water-table aquifer contaminated by crude oil. In:

Dracos, T.H., and F. Stauffer, eds. 

7. Essaid, H.I., W.N. Herkelrath, and K.M. Hess. 1993.

Simulation of fluid distributions observed at a crudeoil spill site incorporating hysteresis, oil entrapment,

and spatial variability of hydraulic properties. Water

Resour. Res. 29(6):1,753-1,770.

8. Deutsch, C.V., and A.G. Journel. 1992. GSLIB:

Geostatistical Software Library and user’s guide.

New York: Oxford University Press.

Recharge Zones

A B C D E

50

10

1

10

1

1

420

424

416

Altitude, in Meters

420

424

416

Altitude, in Meters

420

424

416

Altitude, in Meters

c

b

a

50

10 1

10 1

1

50

10 1

Direction of


Ground-Water Flow

1

1

Distance, in Meters

Explanation

0 100 200 300 400

Midpoint of Well Screen

Zone Boundary

Isoconcentration Line

Datum Is Sea Level

1

Figure 7. Simulated two-dimensional distributions of DDOC: a) with no degradation, b) with degradation, c) with degradation and

spatial variability.

101
9. Dillard, L.A. 1993. Multiphase flow modeling of a

crude-oil spill site using geostatistical simulation of

soil hydraulic properties. M.S. thesis. Stanford, CA:

Stanford University.

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

1994. Multiphase flow modeling at the Bemidji, Minnesota, crude-oil spill site using geostatistical simulation of soil hydraulic properties. In: Morganwalp,

D.W., and D.A. Aronson, eds. Proceedings of the

Technical Meeting of the U.S. Geological Survey

Toxic Substances Hydrology Program, Colorado

Springs, CO (September 20-24, 1993). Water Res.

Invest. Rep. 94-4014. In press.

11. Baedecker, M.J., I.M. Cozzarelli, D.I. Siegel, P.C.

Bennett, and R.P. Eganhouse. 1993. Crude oil in a

shallow sand and gravel aquifer, 3. Biogeochemical

reactions and mass balance modeling in anoxic

ground water. Appl. Geochem. 8:569-586.

12. Bennett, P.C., D.I. Siegel, M.J. Baedecker, and M.F.

Hult. 1993. Crude oil in a shallow sand and gravel

aquifer, 1. Hydrogeology and inorganic geochemistry. Appl. Geochem. 8:529-549.

13. Chang, F.H., H. Wang, B. Denzin, and J. Buller.

1991. Kinetics of crude-oil biodegradation by bacteria indigenous to sediment and ground water. In:

Mallard, G.E., and D.A. Aronson, eds. Proceedings

of the Technical Meeting of the U.S. Geological

Survey Toxic Substances Hydrology Program, Monterey, CA (March 11-15, 1991). Water Res. Invest.

Rep. 91-4034. pp. 646-649.

14. Cozzarelli, I.M., R.P. Eganhouse, and M.J.

Baedecker. 1990. Transformation of monoaromatic

hydrocarbons to organic acids in anoxic groundwater environment. Environ. Geol. Water Sci.

16:135-141.

15. Eganhouse, R.P., M.J. Baedecker, I.M. Cozzarelli,

G.R. Aiken, K.A. Thorn, and T.F. Dorsey. 1993.

Crude oil in a shallow sand and gravel aquifer, 2.

Organic geochemistry. App. Geochem. 8:551-567.

16. Hiebert, F.K., and P.C. Bennett. 1992. Microbial

control of silicate weathering in organic rich ground

water. Science 258:278-281.

17. Hult, M.F., M.K. Landon, and H.O. Pfannkuch.

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 D.A. Aronson, eds.

Proceedings of the Technical Meeting of the U.S.

Geological Survey Toxic Substances Hydrology

Program, Monterey, CA (March 11-15, 1991). Water

Res. Invest. Rep. 91-4034. pp. 621-626.

18. Lovley, D.R., M.J. Baedecker, D.J. Lonergan, I.M.

Cozzarelli, E.J.P. Phillips, and D.I. Siegel. 1989.

Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339:297-299.

19. Baedecker, M.J., D.I. Siegel, P. Bennett, and I.M.

Cozzarelli. 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 S.E. Ragone, eds. Proceedings of the Technical Meeting of the U.S. Geological Survey Toxic

Substances Hydrology Program, Phoenix, AZ (September 26-30, 1988). Water Res. Invest. Rep. 88-

4220. pp. 13-20.

20. Goode, D.J., and L.F. Konikow. 1989. Modification

of a method-of-characteristics solute-transport

model to incorporate decay and equilibrium-controlled sorption or ion exchange. Water Res. Invest.

Rep. 89-4030. p. 65.

21. Konikow, L.F., and J.D. Bredehoeft. 1978. Computer model of two-dimensional solute transport

and dispersion in ground water. U.S. Geological

Survey Techniques of Water Resources Investigations. Book 7. 

22. Kindred, J.S., and M.A. Celia. 1989. Contaminant

transport and biodegradation, 2. Conceptual model

and test simulations. Water Resour. Res. 26(6):1,149-

1,160.

23. Essaid, H.I., M.J. Baedecker, and I.M. Cozzarelli.

1994. Use of simulation to study field-scale solute

transport and biodegradation at the Bemidji, Minnesota, crude-oil spill site. In: Morganwalp, D.W., and

D.A. Aronson, eds. Proceedings of the Technical

Meeting of the U.S. Geological Survey Toxic Substances Hydrology Program, Colorado Springs, CO

(September 20-24, 1993). Water Res. Invest. Rep.

94-4014. In press.

102
Natural Attenuation of Trichloroethylene and 

Similar Compounds:

Case Studies

103
An Overview of Anaerobic Transformation of Chlorinated Solvents

Perry L. McCarty

Department of Civil Engineering, Stanford University, Stanford, CA

Abstract

Intrinsic cometabolic transformation of chlorinated solvents commonly occurs at sites where co-contaminants

are present as primary substrates to support the energy

needs and metabolic activities of transforming bacteria.

The extent of transformation that occurs depends upon

the relative concentration of primary substrates and the

microorganisms and environmental conditions present.

Reduction of tetrachloroethene (PCE) and trichloroethene (TCE) to ethene has occurred at many sites,

although transformations are often not complete.

Evidence for intrinsic biotransformation of chlorinated

aliphatic hydrocarbons (CAHs) is provided by the presence of CAH transformation products and indicators of

anaerobic biogical activity, such as the disappearance

of dissolved oxygen, nitrates, and sulfates, and the production of methane and soluble iron (II). 

Introduction

Chlorinated solvents and their natural transformation

products represent the most prevalent organic groundwater contaminants in the country. These solvents, consisting primarily of CAHs, have been used widely for

degreasing of aircraft engines, automobile parts, electronic components, and clothing. Only during the past

15 years has it become recognized that CAHs can be

transformed biologically (1). Such transformations sometimes occur under the environmental conditions present in

an aquifer in the absence of planned human intervention,

a process called intrinsic biotransformation (2). Conditions under which this is likely to occur with CAHs and

the end products that can be expected are discussed in

this paper. 

The major chlorinated solvents are carbon tetrachloride

(CT), PCE, TCE, and 1,1,1-trichloroethane (TCA).

These compounds can be transformed by chemical and

biological processes in soils to form a variety of other

CAHs, including chloroform (CF), methylene chloride

(MC), cis- and trans-1,2-dichloroethene (c-DCE, t-DCE),

1,1-dichloroethene (1,1-DCE), vinyl chloride (VC),

1,1-dichloroethane (DCA), and chloroethane (CA). In

CAH transformation, the microorganisms responsible

cannot obtain energy for growth from the transformations. The transformations are brought about through

co-metabolism or through interactions of the CAHs with

enzymes or cofactors produced by the microorganisms

for other purposes. In co-metabolism, other organic

chemicals must be present to serve as primary substrates to satisfy the energy needs of the microorganisms. Chemical transformations of some CAHs can also

occur within the timeframe of interest in ground water.

Transformations that are likely, and the environmental

conditions required, are discussed below.

Chemical Transformation

TCA is the only major chlorinated solvent that can be

transformed chemically in ground water under all conditions likely to be found and within the one- to two-decade

time span of general interest, although chemical transformation of CT through reductive processes is a possibility. TCA chemical transformation occurs by two

different pathways, leading to the formation of 1,1-DCE

and acetic acid (HAc):

kDCE

 CH2 = CCI2 + H+ + Cl− (elimination)

1,1−DCE (1)

CH3CCl3 TCA

 kHAc +H2O

 CH3COO + 3H++ 3Cl− (hydrolysis)

HAc (2)

The rate of each chemical transformation is given by the

first-order reaction: 

C = Coe-kt (3)

104
where C is the concentration of TCA at any time t, Co

represents the initial concentration at t = 0, and k is a

transformation rate constant. The overall rate constant

for TCA transformation (kTCA) is equal to the sum of the

individual rate constants (kDCE + kHAc). The transformation rate constants are functions of temperature:

k = Ae-E/0.008314K (4)

where A and E are constants and K is the temperature

in degrees Kelvin. Table 1 provides a listing of values

reported for A and E for TCA abiotic transformation by

various investigators, and calculated values for the TCA

transformation rate constant for 10o

C, 15oC, and 20oC

using equation 4. Also given is the average calculated

TCA half-life based upon t1⁄2

 = 0.69/k. The temperature

effect on TCA half-life is quite significant.

Cline and Delfino (4) found that kDCE equaled about

21 percent of KTCA, and Haag and Mill (3) found it to be

22 percent. This means that almost 80 percent of the

TCA is transformed into acetic acid. The 20-plus percent

that is converted to 1,1-DCE, however, is of great significance because 1,1-DCE is considered more toxic

than TCA, with a maximum contaminant level of 7 μg/L

compared with 200 μg/L for TCA. Whenever TCA is

present as a contaminant, 1,1-DCE can also be expected. In general, TCA is probably the main source of

1,1-DCE contamination found in aquifers. 

Chloroethane, formed through biological transformation

of TCA, can also be chemically transformed with a halflife on the order of months by hydrolysis to ethanol,

which can then be biologically converted to acetic acid

and nonharmful products (6).

Biological Transformations

CAHs can be oxidized or reduced, generally through

co-metabolism, as noted in Table 2. In ground waters,

intrinsic reductive transformations are most often noted,

perhaps because the presence of intermediate products

that are formed provide strong evidence that reductive

transformations are taking place. Intrinsic aerobic transformation of TCE is also possible, although if it did occur,

the intermediate products are unstable and more difficult, analytically, to measure. Thus, convincing evidence

for the latter is difficult to obtain. Also, aerobic co-metabolism of TCE would only occur if sufficient dissolved

oxygen and a suitable electron donor, such as methane,

ammonia, or phenol, were present. Since circumstances

under which the proper environmental conditions for

significant aerobic co-metabolism are unlikely to occur

often, intrinsic aerobic co-metabolism of TCE is probably of little significance. Evidence is ample, however,

that anaerobic reductive transformations of CAHs occur

frequently, and this process is important to the transformation of all chlorinated solvents and their transformation products. The major environmental requirement is

the presence of sufficient concentrations of other organics that can serve as electron donors for energy

metabolism, which often is the case in aquifers. Indeed,

the extent to which reductive dehalogenation occurs

may be limited by the amount of such co-contaminants

present. Theoretically, only a 0.4-g chemical oxygen

demand (COD) equivalent of primary substrate would

be required to convert 1 g of PCE to ethene (7), but

much more is actually required because of the cometabolic nature of the transformation. 

Figure 1 illustrates the potential chemical and biological

transformation pathways for the four major chlorinated

solvents under anaerobic environmental conditions (6).

Freedman and Gossett (8) provided the first evidence

for conversion of PCE and TCE to ethene, and de Bruin

et al. (9) reported completed reduction to ethane. Table

3 indicates that while some transformations, such as CT

to chloroform and carbon dioxide, may take place under

mild reducing conditions such as those associated with

denitrification, complete reductive transformation to inorganic end products and of PCE and TCE to ethene

generally requires conditions suitable for methane fermentation. Extensive reduction, although perhaps not

complete, can also occur under sulfate-reducing conditions. For methane fermentation to occur in an aquifer,

the presence of sufficient organic co-contaminant is

required to reduce the oxygen, nitrate, nitrite, and

sulfate present. Some organics will be required to reduce the CAHs, and perhaps Fe(II) as well, if present

Table 1. Reported First-Order TCA Abiotic Transformation Rates (kTCA)

kTCA (yr-1)

A

yr-1

E

kJ

10

o

C

15

o

C

20

o

C References

3.47 (10)20 118.0 0.058 0.137 0.32 Haag and Mill (3)

6.31 (10)20 119.3 0.060 0.145 0.34 Cline and Delfino (4)

1.56 (10)20 116.1 0.058 0.137 0.31 Jeffers et al. (5)

Average half-life (yr) 12 4.9 0.95

105
in significant amounts. If the potential for intrinsic transformation of CAHs is to be evaluated, then the concentrations of nitrate, nitrite, sulfate, Fe(II), and methane,

and of organics (as indicated by COD or total organic

carbon [TOC]) should be determined. Unfortunately,

such analyses are not considered essential in remedial

investigations, but it is evident that they should be.

Case Studies

Major et al. (10) reported field evidence for intrinsic

bioremediation of PCE to ethene and ethane at a chemical transfer facility in North Toronto. PCE was stored at

the site 10 years prior to the study and contaminated the

ground water below with both free and dissolved PCE.

In addition to high concentrations of PCE (4.4 mg/L),

high concentrations of methanol (810 mg/L) and acetate

(430 mg/L) were found in the contaminated ground

water; methanol and acetate are co-contaminants that

served as the primary substrates for the transforming

organisms. Where high PCE was found, TCE (1.7 mg/L),

cis-DCE (5.8 mg/L), and VC (0.22 mg/L) were also

found, but little ethene (0.01 mg/L) was found. At one

downgradient well, however, no PCE or TCE was found,

but cis-DCE (76 mg/L), VC (9.7 mg/L), and ethene

(0.42 mg/L) were present, suggesting that significant

dehalogenation had occurred. Other dichloroethylenes

(1,1-DCE and trans-DCE) were not significant in concentration, indicating that cis-DCE was the major transformation intermediate. Microcosm studies also supported

that biotransformation was occurring at the site, with complete disappearance of PCE, TCE, and cis-DCE and

production of both VC and ethene. The conversions

were accompanied by significant methane production,

indicating that suitable redox conditions were present for

the transformation.

Fiorenza et al. (11) reported on PCE, TCE, TCA, and

dichloromethane (DCM) contamination of ground water

at two separate locations at a carpet-backing manufacturing plant in Hawkesbury, Ontario. The waste lagoon

was the major contaminated area, with ground water

containing 492 mg/L of volatile fatty acids and 4.2 mg/L

of methanol, organics that appeared to provide the cocontaminants that served as primary sources of energy

for the dehalogenation reactions. Here, the sulfate concentration was nondetected, but the concentration in

native ground water was about 15 to 18 mg/L. Total

dissolved iron was quite high (19.5 mg/L) and well above

the upgradient concentration of 2.1 mg/L. Methane was

present, although quite low in concentration (0.06 mg/L).

These parameters are all supportive of conditions suitable for intrinsic biodegradation of the chlorinated solvents. While some chemical transformation of TCA was

indicated (0.4 mg/L), biotransformation was quite extensive, as indicated by a 1,1-DCA concentration of 7.2 mg/L

compared with the TCA concentration of 5.5 mg/L.

Some CA was also present (0.19 mg/L). Transformation

was also indicated for PCE and TCE, which remained

at concentrations of only 0.016 mg/L and 1.5 mg/L,

respectively, while the cis-DCE, VC, and ethene

concentrations were 56, 4.2, and 0.076 mg/L, respectively. Only traces of ethane were found. Trans-DCE

concentration was only 0.57 mg/L, again providing evidence that cis-DCE is the most common transformation

intermediate from TCE and PCE. Downgradient from the

Table 2. Biodegradability of Chlorinated Solvents Under Aerobic or Anaerobic Conditions and Through Use as a Primary

Substrate for Energy and Growth or Through Co-metabolism

Carbon

Tetrachloride (CT)

Tetrachloroethylene

(PCE)

Trichloroethylene

(TCE)

1,1,1-

Trichloroethane

(TCA)

Aerobic Biotransformation

Primary substrate No No No No

Co-metabolism No No Yes Perhaps

Anaerobic Biotransformation

Primary substrate No Perhaps Perhaps No

Co-metabolism Yes Yes Yes Yes

Hazardous intermediates Yes Yes Yes Yes

Chemical Transformation Perhaps No No Yes

CCI =CCI 2 2

CHCI=CCI2

PCE

TCE

CHCI = CHCI

TCA

CT

a

a

CH =CH 2 2

CH =CCI 2 2

CH =CHCI 2

CH =CH 3 3 CO + H O + CI 2 2

CH COOH 3

CH CH CI 3 2

CH CHCI 3 2

CH CCI 3 3 CHCI3

CH CI 2 2

CH CI 3

CCI4

a

-

Figure 1. Anaerobic chemical and biological transformation

pathways for chlorinated solvents.

106
lagoon, the dominant products were cis-DCE (4.5 mg/L),

VC (5.2 mg/L), and 1,1-DCA (2.1 mg/L). While good

evidence for intrinsic biotransformation is provided for

this site, the ethene and ethane concentrations appear

very low compared with VC concentration, suggesting

that biotransformation was not eliminating the chlorinated solvent hazard at the site, although it was producing compounds that may be more susceptible to aerobic

co-metabolism.

Evidence for intrinsic biotransformation of chlorinated

solvents has also been provided from analyses of gas

from municipal refuse landfills where active methane

fermentation exists. A summary by McCarty and Reinhard (12) of data from Charnley et al. (13) indicated

average gaseous concentrations in parts per million by

volume from eight refuse landfills to be PCE, 7.15; TCE,

5.09; cis-DCE, not measured; trans-DCE, 0.02; and VC,

5.6. While these averages indicate that, in general,

transformation was not complete, the presence of high

VC indicates the transformation was significant. For

TCA, gaseous concentrations were TCA, 0.17; 1,1-

DCE, 0.10; 1,1-DCA, 2.5; and CA, 0.37. These data

indicate that TCA biotransformation was quite extensive,

with the transformation intermediate, 1,1-DCA, present

at quite significant levels, as is frequently found in

ground water.

Perhaps the most extensively studied and reported intrinsic chlorinated solvent biodegradation is that at the

St. Joseph, Michigan, Superfund site (7,14-17). Groundwater concentrations of TCE as high as 100 mg/L were

found present, with extensive transformation to cisDCE, VC, and ethene. A high but undefined COD (400

mg/L) in ground water, resulting from waste leaching

from a disposal lagoon, provided the energy source for

the co-metabolic reduction of TCE. Nearly complete

conversion of the COD to methane provided evidence

of the ideal conditions for intrinsic bioremediation (7).

Extensive analysis near the source of contamination

indicated that 8 percent to 25 percent of the TCE had

been converted to ethene, and that up to 15 percent of

the reduction in COD in this zone was associated with

reductive dehalogenation (15). Through more extensive

analysis of ground water farther downgradient from the

contaminating source, Wilson et al. (17) found a 24-fold

reduction in CAHs across the site. A review of the data

at individual sampling points indicated that conversion

of TCE to ethene was most complete where methane

production was highest and removal of nitrate and sulfate by reduction was most complete. 

References

1. McCarty, P.L., and L. Semprini. 1994. Ground-water

treatment for chlorinated solvents. In: Norris, R.E.,

ed. Handbook of bioremediation. Boca Raton, FL:

Lewis Publishers, Inc. pp. 87-116.

2. Council, N.R. 1993. In situ bioremediation: When

does it work? Washington, DC: National Academy

Press. 

3. Haag, W.R., and T. Mill. 1988. Effect of subsurface

sediment on hydrolysis of haloalkanes and epoxides. Environ. Sci. Technol. 22:658-663.

4. Cline, P.V., and J.J. Delfino. 1989. Transformation

kinetics of 1,1,1-trichloroethane to the stable product 1,1-dichloroethene. In: R.A. Larson, ed. Biohazards of drinking water treatment. Chelsea, MI:

Lewis Publishers, Inc. pp. 47-56.

5. Jeffers, P., L. Ward, L. Woytowitch, and L. Wolfe.

1989. Homogeneous hydrolysis rate constants for

selected chlorinated methanes, ethanes, ethenes,

and propanes. Environ. Sci. Technol. 23(8): 965-

969.

6. Vogel, T.M., C.S. Criddle, and P.L. McCarty. 1987.

Transformations of halogenated aliphatic compounds. Environ. Sci. Technol. 21:722-736.

7. McCarty, P.L., and J.T. Wilson. 1992. Natural anaerobic treatment of a TCE plume, St. Joseph,

Michigan, NPL site. In: U.S. EPA. Bioremediation of

hazardous wastes. EPA/600/R-92/126. Cincinnati,

OH. pp. 47-50.

8. Freedman, D.L., and J.M. Gossett. 1989. Biological

reductive dechlorination of tetrachloroethylene and

trichloroethylene to ethylene under methanogenic

conditions. Appl. Environ. Microbiol. 55(9):2,144-

2,151.

9. de Bruin, W.P., et al. 1992. Complete biological

reductive transformation of tetrachloroethene to

Table 3. Environmental Conditions Generally Associated With Reductive Transformations of Chlorinated Solvents

Redox Environment

Chlorinated Solvent All Denitrification Sulfate Reduction Methanogenesis

Carbon tetrachloride — CT → CF CT → CO2+Cl- —

1,1,1-Trichloroethane TCA → 1,1-DCE + CH3COOH — TCA → 1,1-DCA TCA → CO2 + ClTetrachloroethylene — — PCE → 1,2-DCE PCE → ethene

Trichloroethylene — — TCE → 1,2-DCE TCE → ethene

107
ethane. Appl. Environ. Microbiol. 58(6):1,996-

2,000.

10. Major, D.W., W.W. Hodgins, and B.J. Butler. 1991.

Field and laboratory evidence of in situ biotransformation of tetrachloroethene to ethene and ethane

at a chemical transfer facility in North Toronto. In:

Hinchee, R.E., and R.F. Olfenbuttel, eds. Onsite

bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 147-171.

11. Fiorenza, S., et al. 1994. Natural anaerobic degradation of chlorinated solvents at a Canadian manufacturing plant. In: Hinchee, R.E., A. Leeson, L.

Semprini, and S.K. Ong, eds. Bioremediation of

chlorinated and polycyclic aromatic hydrocarbon

compounds. Boca Raton, FL: Lewis Publishers, Inc.

pp. 277-286.

12. McCarty, P.L., and M. Reinhard. 1993. Biological

and chemical transformations of halogenated aliphatic compounds in aquatic and terrestrial environments. In: Oremland, R.S., ed. The biogeochemistry

of global change: Radiative trace gases. New York,

NY: Chapman & Hall, Inc.

13. Charnley, G., E.A.C. Crouch, L.C. Green, and T.L.

Lash. 1988. Municipal solid waste landfilling: A

review of environmental effects. No. Meta Systems,

Inc.

14. Haston, Z.C., P.K. Sharma, J.N. Black, and P.L.

McCarty. 1994. Enhanced reductive dechlorination

of chlorinated ethenes. In: U.S. EPA. Bioremediation of hazardous wastes. San Francisco, CA.

15. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T.

Wilson. 1993. Natural anaerobic bioremediation of

TCE at the St. Joseph, Michigan, Superfund site.

In: U.S. EPA. Symposium on bioremediation of hazardous wastes: Research, development, and field

evaluations (abstracts). EPA/600/R-93/054. Washington, DC (May). Cincinnati, OH. pp. 47-50.

16. McCarty, P.L., et al. 1991. In situ methanotrophic

bioremediation for contaminated ground water at

St. Joseph, Michigan. In: Hinchee, R.E., and R.G.

Olfenbuttel, eds. Onsite bioreclamation processes

for xenobiotic and hydrocarbon treatment. Boston,

MA: Butterworth-Heinemann. pp. 16-40.

17. Wilson, J.T., J.W. Weaver, and D.H. Kampbell.

1994. Intrinsic bioremediation of TCE in ground

water at an NPL site in St. Joseph, Michigan. Presented at the U.S. EPA Symposium on Intrinsic

Bioremediation of Ground Water, Denver, CO (August 30 to September 1).

108
Contamination of Ground Water With Trichloroethylene at the Building 24 Site at

Picatinny Arsenal, New Jersey

Mary Martin and Thomas E. Imbrigiotta

U.S. Geological Survey, West Trenton, NJ

Abstract

Ground water at the Building 24 site at Picatinny Arsenal

in Morris County, New Jersey, is contaminated with

trichloroethylene (TCE). Estimated average linear

ground-water flow velocities are 0.3 to 1.0 m/d, and

estimated travel time from the source area to Green

Pond Brook is 2 y to 5 y. The total mass of dissolved

TCE within the 130,000-m2

 plume area is estimated to

be 970 kg. About 65 percent of the mass is in areas

where TCE concentrations exceed 10,000 μg/L,

whereas about 30 percent is in areas where TCE concentrations are 1,000 to 10,000 μg/L.

The average flux of TCE discharged to Green Pond

Brook from the plume area, estimated from measured

TCE concentrations in samples of water from the unconfined aquifer and measured base-flow discharge in

Green Pond Brook, is 1 to 2 mg/s. Biotransformation is

the most important mechanism by which dissolved TCE

leaves the ground-water system.

Introduction

Picatinny Arsenal, a U.S. Army armament research and

development center, is located in a glaciated valley in

north-central New Jersey (Figure 1). In 1960, Building

24 was remodeled, and a new metal-plating facility and

industrial wastewater treatment plant were installed.

From 1960 to 1981, the wastewater treatment system

discharged tens of thousands of liters of wastewater

daily into two 2.5-m deep, sand-bottomed settling lagoons behind the building (1). The metal-plating wastewater contained trace metals, such as cadmium,

chromium, copper, lead, nickel, tin, vanadium, and zinc,

and other ions used in plating solutions, such as sodium, potassium, sulfate, chloride, and cyanide (2).

From 1973 to 1985, an improperly installed relief system

of the degreasing unit allowed pure chlorinated solvents

to condense in an overflow pipe and discharge to a 1-m

deep dry well in front of Building 24. The condensate

from the overflow system contained TCE and, after

1983, 1,1,1-trichloroethane (2).

The infiltration of wastewater from the lagoons and of

chlorinated solvents from the dry well has created a

plume of contaminated ground water downgradient from

Building 24. Most of the contamination is limited to

unconfined sediments, where estimated ground-water

flow velocities (based on estimated horizontal hydraulic

conductivities, measured head gradients, estimated porosities, and results of calibrated solute-transport model

simulations) generally range from 0.3 to 1.0 m/d in the

plume area. On the basis of these estimated velocities,

the residence time within the unconfined aquifer of a

conservative solute entering the aquifer near Building 24

and discharging at Green Pond Brook is 2 y to 5 y.

Ground-water contamination in the unconfined aquifer

has been discussed previously by Sargent et al. (3),

Fusillo et al. (2), and Imbrigiotta et al. (4,5). Results of

water-quality sampling during 1986 to 1991 confirm that

TCE remains the dominant contaminant in the unconfined aquifer and that the extent of the plume of TCEcontaminated ground water has changed little since

September 1986. These water-quality measurements

also show that 1) Building 24 is the source of the TCE

plume; 2) the highest TCE concentrations near Building

24 are found at depths of less than 6 m and within 10 m

downgradient from the dry well; 3) the plume extends

500 m from Building 24 to Green Pond Brook and follows the ground-water head gradients as it flows downward through the unconfined aquifer, then upward

toward Green Pond Brook; 4) the plume disperses as it

moves downgradient, is about 350 m wide where it

enters Green Pond Brook, and has an areal extent of

about 130,000 m2

; 5) the highest concentrations of TCE

are found at the base of the unconfined aquifer midway

between Building 24 and Green Pond Brook; and 6) the

highest TCE concentrations at Green Pond Brook are

found at depths of less than 6 m (Figure 1).

109
Figure 1. (A) Location of Building 24 study area at Picatinny Aresenal, and areal extent of TCE plume and altitude of water table,

January 1993, and (B) vertical distribution of TCE concentrations, October to November 1991. (Location of section A-A′

is shown in Figure 1A.)

110
The U.S. Geological Survey is conducting an interdisciplinary research study of ground-water contamination by

the chlorinated solvents and other contaminants at the

Building 24 site at Picatinny Arsenal. The objectives of

the study are to 1) describe the chemical, physical, and

biological processes that affect the movement and fate

of these contaminants, particularly TCE, in the subsurface; 2) determine the relative importance of these processes; and 3) develop predictive models of contaminant

transport.

This paper describes a conceptual model of the processes that affect the fate and transport of chlorinated

solvents at Picatinny Arsenal research site. A preliminary solute mass balance of TCE in the unconfined

aquifer at the site is presented and compared with results of numerical solute-transport simulations.

Conceptual Model

A conceptual model of the physical, chemical, and biological processes that affect the transport and mass

balance of TCE within the plume was developed by

Imbrigiotta and Martin (6) (Figure 2). TCE is present at

the site in several forms: solute phase (dissolved in

water), vapor phase (soil gas), and matrix phase (sorbed

onto solid surfaces or associated with biota). TCE also

may be present as a dense nonaqueous-phase liquid

(DNAPL). Nonaqueous-phase flow of TCE from Building

24 to the base of the unconfined aquifer is hypothesized

to be the cause of high concentrations of TCE at the

base of the unconfined aquifer (5).

Results of chemical analyses of samples of ground

water, soil gas, and aquifer sediments from the plume

area indicate that the fate and transport of TCE are

affected by the physical, chemical, and biological processes at the site. Advection and dispersion in the saturated zone affect the movement of dissolved TCE and

cause it to be removed from the system in the discharge

to Green Pond Brook. TCE also is removed from the

system by anaerobic biotransformation (reductive dehalogenation), volatilization at the water table, and sorption

to saturated-zone sediments. Volatilization of TCE to the

unsaturated-zone soil gas was measured by Smith and

others (7). Transport in the unsaturated zone was

determined to be driven principally by molecular diffusion. The detection of the biotransformation products

Building 24 Water Table

Volatilization


(0.1 mg/s)

Estimated Top of Confining Unit

Anaerobic


Biotransformation


(1-30 mg/s)

Land


Surface

Advective


Transport to


Green Pond Brook


(1-2 mg/s)

Dissolution


(Not Estimated)

Long-Term


Desorption


(15-85 mg/s)

Short-Term


Desorption


(Not Estimated)

Infiltration


(<0.1 mg/s)

Short-Term


Adsorption


(Not Estimated)

Trichloroethylene Plume

Not to Scale

Trichloroethylene Mass Balance Components


(mg/s, milligrams per second; <, less than)

Gains


Long-Term Desorption 


Infiltration


Dissolution


Short-Term Desorption

15-85 mg/s


< 0.1 mg/s


Not Estimated


Not Estimated


Losses


Anaerobic Biotransformation 


Advective Transport to


Green Pond Brook


Volatilization


Short-Term Absorption

1-30 mg/s


1-2 mg/s


0.1 mg/s


Not Estimated


Figure 2. Conceptual model and preliminary mass-balance estimates of TCE fluxes resulting from processes that affect the fate

and transport of TCE in the ground-water system at Picatinny Arsenal.

111
cis-1-2-dichloroethylene (cis-DCE) and vinyl chloride

(VC) in water from 75 percent of the wells screened

within the plume is indicative of biologically mediated,

reductive dehalogenation of TCE. The occurrence of

methanogenesis is indicated by the detection of dissolved methane in water from 85 percent of the wells in

which cis-DCE was detected and in water from 94 percent of the wells in which VC was detected.

Desorption from contaminated sediments is a source of

TCE in the ground-water system. If TCE is present as a

DNAPL, dissolution will result in a source of dissolved

TCE in the ground water. Relatively constant TCE concentrations measured near Building 24 and at the base

of the unconfined aquifer from 1986 to 1991 indicate that

slow desorption of TCE from contaminated aquifer sediments, dissolution of DNAPL TCE in the aquifer, or both

are a continuing source of dissolved TCE in the groundwater system.

Estimated Mass Distribution of TCE

For a given sample set, the total mass of dissolved TCE

below the water table is estimated to be 970 kg or about

660 L of liquid TCE. This estimate was calculated by

using results of six sets of water-quality analyses made

from 1986 to 1991 (Figure 3). Each measured TCE concentration is assumed to represent the TCE concentration

of a volume of ground water extending half the horizontal and vertical distance to the adjacent sampling points.

These volumes are limited vertically by the extent of the

unconfined aquifer and horizontally by the 10-μg/L maximum TCE concentration line. Aquifer porosity is

assumed to be 30 percent. The total volume of dissolved

TCE within the ground water outside the 10-μg/L

maximum TCE concentration line is estimated to be

about 1 L.

The estimated total volume of TCE appears to be unrelated to the total number of samples collected. Four sets

of data plotted in Figure 3 (A, C, E, and F) show that

about 65 percent of the total volume of TCE is found in

ground water that contains TCE concentrations greater

than 10,000 μg/L, and about 30 percent is found in

areas where TCE concentrations range from 1,000 to

10,000 μg/L. The estimate of the volume of TCE in the

water appears to depend mostly on the number of samples that contain relatively high concentrations of TCE

and the volume of water each sample is assumed to

represent.

The relative amounts of TCE in the dissolved, vapor, and

matrix phases in a block of aquifer and unsaturated zone

immediately downgradient from Building 24 were estimated on the basis of measured TCE concentrations in

samples of all three phases (8) (Figure 4). Most of the

900


800


700


600


500


400


300


200


100


0

1,300


1,200


1,100


1,000


900


800


700


600


500


400


300


200


100


0

Volume of


Trichloroethylene, in Liters

Mass of


Trichloroethane, in Kilograms

Sample Set A B C D E F

Number of Samples 96 39 40 59 50 42

Volume of trichloroethylene


calculated using samples


with trichloroethylene


concentrations of:

Explanation


[μg/L, micrograms per liter; ≤, less than or equal to; >, greater than]

1 to ≤100 μg/L


>100 to ≤1,000 μg/L


>1,000 to ≤10,000 μg/L


>10,000 μg/L

Sample


Set


A


B


C


D


E


F

April and August-September 1986


August 1987


June 1989


November-December 1989


April-March 1990


October-November 1991

Date of Collection

1,319

Figure 3. Estimated volume of dissolved TCE in ground water at the Building 24 site at Picatinny Arsenal, 1986-91.

112
mass of TCE in the system near Building 24 is in the

matrix phase and is associated with the sediments in

both the unsaturated and saturated zones. The mass

distribution shown in Figure 4 is based on a representative sample of each TCE phase in the area of

Building 24. The ratio of the mass of TCE sorbed to the

soil to the mass of dissolved TCE in six sets of samples

collected near the water table throughout the site ranged

from 2:1 to 5:1. One sample, collected downgradient

from the wastewater lagoons, had an unusually high

ratio of about 20:1. The amount of DNAPL TCE at the

site has not been estimated.

Preliminary Solute Mass Balance

Desorption of TCE from soils that have undergone longterm adsorption (years) and dissolution of DNAPL TCE

probably are the processes by which most TCE enters

the ground-water system, because the direct release of

TCE from Building 24 stopped in 1985. Because the

amount of DNAPL TCE has not been estimated, the rate

of TCE dissolution cannot be estimated. Three firstorder rate constants of TCE desorption from shallow

aquifer sediments at the arsenal calculated by Koller et

al. (8) ranged from 0.003 to 0.015 per week. The rate

constants were measured in laboratory flowthrough columns using uncontaminated water as the influent fluid.

Desorption rates in the field, where ground water containing TCE is flowing past the desorbing sediments,

probably would be lower. By using an estimated mass

of TCE sorbed to the aquifer sediments of 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 15 to 85 mg/s. This flux estimate is made

by assuming that, over long periods (years), the shortterm desorption rate (weeks and months) is equal to the

short-term adsorption rate, and that soils previously

have undergone long-term adsorption, which is no

longer occurring.

Preliminary estimates of the flux of TCE into and out of

the ground-water system at the Building 24 site for each

of the mass-balance components are shown in Figure 2.

The estimated flux of TCE discharged to Green Pond

Brook from the plume area, calculated on the basis of

measured TCE concentrations in ground water and

measured base-flow discharge in 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 concentration gradients and estimates of

the physical characteristics of the unsaturated zone.

Biotransformation probably is the mechanism by which

most of the dissolved TCE leaves the ground-water

system. First-order rate constants for TCE transformation ranging from less than 0.001 to 0.02 per week were

estimated by Wilson et al. (9) on the basis of results of

laboratory microcosm studies of soil from five sites

within the plume area. By using these rate constants and

the estimated mass of dissolved TCE, the rate of TCE

loss from the plume through biotransformation is calculated to be about 1 to 30 mg/s. Analogous first-order rate

constants for TCE biotransformation calculated by Ehlke

et al. (10) from field-measured TCE concentrations and

time-of-travel data generally were higher than those

measured in the laboratory experiments. Thus, the actual flux of TCE lost through biotransformation may be

greater than that shown in Figure 2.

A reactive multispecies transport model of a two-dimensional vertical section along the central axis of the plume

is being used to analyze the laboratory and field estimates of the physical and chemical transport characteristics and the estimated TCE mass balance. The

model design, calibration, and sensitivity analysis have

been described by Martin (11). Transport, desorption,

volatilization, and microbial degradation of TCE are

simulated. The formation and transport of the degradation products cis-DCE and VC also are simulated.

Specified-flux boundary conditions are used to represent ground-water recharge and flows across the horizontal and vertical boundaries of the cross-sectional

area. Constant-concentration nodes and desorption

within the plume area are solute sources to the simulated

system. The constant-concentration solute sources are

assumed to represent high rates of desorption or

Unsaturated Zone

Saturated Zone

100 feet

100 feet

Building 24

Not to Scale

Mass distribution of trichloroethylene near Building 24


[kg, kilogram; <, less than]

Phase


Vapor


Matrix


Solute


Matrix

Mass (kg)


0.001


3


1


4

Percent


of Total


<0.1


99.9


20


80

Unsaturated


zone


Saturated


zone

10 feet


10 feet

Figure 4. Mass distribution of TCE in the saturated and unsaturated zones immediately downgradient from Building 24 at Picatinny

Arsenal.

113
dissolution of TCE near the settling lagoons, at the

overflow dry well, and near the base of the unconfined

aquifer 230 m downgradient from Building 24.

The simulations were designed to represent average

steady-state flow conditions and virtually steady-state

transport conditions after 1985. Simulated concentrations of TCE and cis-DCE from the calibrated model are

shown in Figure 5. Most concentrations were simulated

to within an order of magnitude of average concentrations measured in micrograms per liter in water samples

from each well. Because the degradation of VC was not

simulated, concentrations are higher than measured

concentrations and are not shown.

The calibrated model does not provide a unique estimate of the magnitudes of the various mass-balance

components of the plume of TCE-contaminated ground

water at the Building 24 site; however, sensitivity simulations were used to test hypotheses concerning the fate

of TCE in ground water. Results of a series of sensitivity

simulations discussed by Martin (11) support the general

conceptual model as defined by the estimated solute

mass balance presented above. 

Results of sensitivity simulations made with various

desorption and degradation rates typically showed that

use of the laboratory estimates resulted in reasonable

simulated concentrations. Although volatilization is not a

major mass-balance component, this process was

shown to be an important mechanism for removing solutes and thereby affecting solute concentrations near

the water table. The overall flux of TCE into and out of

the system was not simulated well because the total

simulated mass of TCE in the system was too low.

Increasing the simulated solute-source area near the

base of the unconfined aquifer might result in a reasonable simulated TCE mass balance.

Summary

Infiltration of wastewater from the lagoons and chlorinated solvents from the dry well at Building 24 at Picatinny Arsenal, New Jersey, has created a plume of

contaminated ground water downgradient from the

building. TCE is the predominant contaminant in the

130,000-m2

 plume, which extends 500 m to Green Pond

Brook. Ground-water velocities typically range from 0.3

to 1.0 m/d.

Results of water-quality sampling conducted from 1986

through 1991 show that the TCE contaminant plume has

changed little since September 1986, and that the highest TCE concentrations are found near the water table

near Building 24 and near the base of the unconfined

aquifer about midway between Building 24 and Green

Pond Brook. TCE is present at the 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 DNAPL. The total volume

of dissolved TCE below the water table is estimated to

be about 660 L, but most of the mass of TCE in the

system is associated with the sediments in the saturated

and unsaturated zones.

A conceptual model of the physical, chemical, and biological processes that affect the transport and mass

balance of TCE within the plume includes 1) transport

of TCE from the Building 24 source area to Green Pond

Brook by advection and dispersion; 2) loss of dissolved

TCE from the ground-water system by discharge to

Green Pond Brook, biotransformation, and volatilization;

and 3) gain of dissolved TCE by slow desorption from

contaminated aquifer sediments and possibly from

A A'

Distance From Building 24 (B-24), in Meters

A A'

Distance From Building 24 (B-24), in Meters

(A) Trichloroethylene

(B) cis-1-2-Dichloroethylene

Line of Equal Solute Concentration


Shows solute concentration, in micrograms per liter.

100

CAF-7

Location of Well Screen


Location of Well and Local Identifier

Explanation

Figure 5. Simulated concentrations from calibrated model with

simulated desorption, volatilization, and microbial

degradation: (A) TCE and (B) cis-DCE. (Location of

section A-A′ is shown in Figure 1A.)

114
dissolution of DNAPL TCE in the aquifer. Preliminary

estimates of the flux of dissolved TCE discharged to

Green Pond Brook from the plume is 1 to 2 mg/s. Most

dissolved TCE leaves the ground-water system by

means of biotransformation. The estimated flux of TCE

out of the system by this process may be about an order

of magnitude greater than the flux of TCE discharged to

Green Pond Brook. Although the estimated flux of TCE

out of the ground-water system by volatilization is estimated to be about an order of magnitude less than the

flux of TCE discharge to Green Pond Brook, volatilization is an important mechanism for removing solutes and

thereby affecting solute concentrations near the water

table.

References

1. Benioff, P.A., M.H. Bhattacharyya, C. Biang, S.Y.

Chiu, S. Miller, T. Patton, D. Pearl, A. Yonk, and C.R.

Yuen. 1990. Remedial investigation concept plan

for Picatinny Arsenal, Vol. 2. Descriptions of and

sampling plans for remedial investigation sites. Argonne, IL: Argonne National Laboratory, Environmental Assessment and Information Sciences

Division. pp. 22-1 to 22-24.

2. Fusillo, T.V., T.A. Ehlke, M. Martin, and B.P. Sargent. 1987. Movement and fate of chlorinated solvents in ground water: Preliminary results and

future research plans. In: Franks, B.J., ed. Proceedings of the U.S. Geological Survey Program on

Toxic Waste—Ground-Water Contamination, Pensacola, FL (March 23-27). U.S. Geological Survey

Open File Rep. 87-109. pp. D5-D12.

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Water Res. Invest. Rep. 88-4220. pp. 351-359.

5. Imbrigiotta, T.E., T.A. Ehlke, and M. Martin. 1991.

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6. Imbrigiotta, T.E., and M. Martin. 1991. Overview of

research activities on the movement and fate of

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Survey Toxic Substances Hydrology Program, Monterey, CA (March 11-15). Water Res. Invest. Rep.

91-4034. pp. 673-680.

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Smith. 1994. Desorption of trichloroethylene from

aquifer sediments at Picatinny Arsenal, New Jersey.

In: Morganwalp, D.W., and D.A. Aronson, eds. Proceedings of the U.S. Geological Survey Toxic Substances Hydrology Program, Colorado Springs, CO

(September 20-24, 1993). Water Res. Invest. Rep.

94-4014. In press.

9. Wilson, B.H., T.A. Ehlke, T.E. Imbrigiotta, and J.T.

Wilson. 1991. Reductive dechlorination of trichloroethylene in anoxic aquifer material from Picatinny

Arsenal, New Jersey. In: Mallard, G.E., and D.A.

Aronson, eds. Proceedings of the U.S. Geological

Survey Toxic Substances Hydrology Program, Monterey, CA (March 11-15). Water Res. Invest. Rep.

91-4034. pp. 704-707.

10. Ehlke, T.A., B.H. Wilson, J.T. Wilson, and T.E. Imbrigiotta. 1994. In situ biotransformation of trichloroethylene and cis-1,2-dichloroethylene at Picatinny

Arsenal, New Jersey. In: Morganwalp, D.W., and D.A.

Aronson, eds. Proceedings of the U.S. Geological

Survey Toxic Substances Hydrology Program, Colorado Springs, CO (September 20-24, 1993). Water

Res. Invest. Rep. 94-4014. In press.

11. Martin, M. 1994. Simulation of transport, desorption,

volatilization, and microbial degradation of trichloroethylene in ground water at Picatinny Arsenal,

New Jersey. In: Morganwalp, D.W., and D.A. Aronson, eds. Proceedings of the U.S. Geological

Survey Toxic Substances Hydrology Program,

Colorado Springs, CO (September 20-24, 1993).

Water Res. Invest. Rep. 94-4014. In press.

115
Intrinsic Bioremediation of TCE in Ground Water at an NPL Site in 

St. Joseph, Michigan

John T. Wilson, James W. Weaver, and Don H. Kampbell

U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK

Introduction

The ground water at the St. Joseph, Michigan, National

Priority List (NPL) site is contaminated with chlorinated

aliphatic compounds (CACs) at concentrations in the

range of 10 mg/L to 100 mg/L. The chemicals are

thought to have entered the shallow sandy aquifer either

through waste lagoons that were used from 1968 to

1976 or through disposal of trichloroethene (TCE) into

dry wells at the site (1). The contamination was determined to be divided into eastern and western plumes,

as the suspected sources were situated over a groundwater divide. Both plumes were found to contain TCE,

cis- and trans-1,2-dichoroethene (c-DCE and t-DCE),

1,1-dichloroethene (1,1-DCE), and vinyl chloride (VC). 

Previous investigation of the site indicated that natural

anaerobic degradation of the TCE was occurring, because transformation products and significant levels of

ethene and methane (2,3) were present. The purpose of

this presentation is to present the results of later sampling of the western plume near Lake Michigan, to estimate the contaminant mass flux, and to estimate

apparent degradation constants. The estimates are

based on visualization of the data that represent each

measured concentration by a zone of influence that is

based on the sample spacing. The presentation of the data

is free from artifacts of interpolation, and extrapolation of

the data beyond the measurement locations is controlled.

Data Summary

In 1991, three transects (1, 2, and 3 on Figure 1) were

completed near the source of the western plume (2).

The three transects consisted of 17 borings with a slotted auger. In 1992, two additional transects (4 and 5 on

Figure 1) were completed, consisting of 9 additional

slotted auger borings. In each boring, water samples

were taken at roughly 1.5 m (5 ft) depth intervals. Onsite

gas chromatography was performed to determine the

width of the plume and find the point of highest concentration. Three of the transects (2, 4, and 5) are roughly

perpendicular to the contaminant plume. Of the remaining transects, transect 1 crosses the plume at an angle

and transect 3 lies along the length of the plume. The

perpendicular transects form logical units for study of

TCE biotransformation.

The site data from the transects are visualized as sets

of blocks that are centered around the measurement

point. The blocks are defined so that the influence of a

particular measured concentration extends halfway to

the next measurement location both horizontally and

vertically. Thus, presentation of the data is simple and

direct. The visualization of the data is performed on a

Silicon Graphics Indigo workstation using a two-dimensional version of the fully three-dimensional field-data

analysis program called SITE-3D that is under development at the Robert S. Kerr Environmental Research

Laboratory. 

The mass of each chemical per unit thickness and the

advective mass flux of each chemical are calculated by

summing over the blocks. By following this procedure,

the measured chemical concentrations are not extrapolated into the clay layer under the site, nor are they

extrapolated beyond a short distance from the measurement locations (5 ft vertically and 50 ft to 100 ft horizontally). Other interpolation schemes such as inverse

distance weighting or kriging could also be used to

estimate the concentration field and perform the mass

estimates. Figures 2 and 3 show the distributions of the

VC and TCE at transect 5 using a logarithmic, blackand-white “color” scale. Notably, the maximum VC concentration at transect 4 was 1,660 μ/L and at transect 5

was 205 μg/L. The maximum TCE concentration at transect 4 was 8,720 μg/L and at transect 5 was 163 μg/L.

As noted previously for other portions of the site (2,4),

the contamination is found near the bottom of the aquifer. The highest concentrations of VC and TCE do not

appear to be co-located. In Table 1, mass estimates are

presented for the perpendicular transects ordered from

116
farthest upgradient (transect 2) to farthest downgradient

(transect 5). The data in Table 1 represent the mass in

a volume of aquifer that has an area equal to the crosssectional area of the transect and is 1.0 m thick in the

direction of ground-water flow.

Advective Mass Flux Estimates

Results from the calibrated MODFLOW model of Tiedeman and Gorelick (4) were used to estimate the groundwater flow velocity at each transect. The estimate is an

upper bound because the modeled vertical component

of flow was neglected in the present analysis. The head

drop from one location to the next was assumed to

generate horizontal flow only. Tiedeman and Gorelick (4)

also represented the aquifer by single values of hydraulic

conductivity and porosity. They gave, however, 95 percent

confidence limits for the hydraulic conductivity. Well

yields estimated for each sample location indicate declining hydraulic conductivity toward the west (i.e., towards

Lake Michigan and transects 4 and 5). Thus, using the

single parameter values from the MODFLOW simulations

may overestimate the flux of water into the lake. 

As would be expected, the advective mass fluxes decline toward the downgradient edge of the plume (Table

2). There the concentrations are lower, due to either

transient flow or degradation of the TCE. Notably the

mass fluxes using the average hydraulic conductivity

result in a total flux of 13 kg/y of TCE, c-DCE, t-DCE,

1,1-DCE, and VC at transect 5. This value contrasts with

the total flux of these CACs of 310 kg/y at transect 2

Transect 5

Transect 4

Transect


2

Transect


3

Lake


Michigan

Legend

Slotted Auger Boring


Ground-water Elevation


8/23/87 588

T5 - 1

200

200

0 200 FT

T5-1

T5-2

T5-3 T5-4

T5-5

587

588

589

590

591

592

593

594 599

595 596

597

598

599

600

T4-4

T4-1

T4-2

T4-3

Approx.


Extent of 


TCE Plume

Transect


1

Figure 1. Site plan, St. Joseph, Michigan, NPL site.

Concentration


μg/L

Ground Surface

10 Feet

Approx. N

100 Feet

St. Joseph, Michigan


Vinyl Chloride


Transect: 5


Mass: 0.4811E-01 kg/m

250,000.


25,000.


2,500.


250.0


25.00


2.500


0.2500


0.0250

t51 t52 t54 t53 t55

Figure 2. VC distribution at transect 5.

250,000.


25,000.


2,500.


250.0


25.00


2.500


0.2500


0.0250

Concentration


μg/L

t51 t52 t54 t53 t55

Ground Surface

10 Feet

Approx. N

100 Feet

St. Joseph, Michigan


Trichloroethene


Transect: 5


Mass: 0.2821E-01 kg/m

Figure 3. TCE distribution at transect 5.

117
near the source of contamination. Thus, there is a 24.4-

fold decrease in mass flux of CACs across the site.

Using the 95 percent confidence limits on the hydraulic

conductivity determined by Tiedeman and Gorelick (4),

the range total of mass flux of these five chemicals

ranges from 205 kg/y to 420 kg/y at transect 2 and from

8.4 kg/y to 17 kg/y at transect 5. The range of fluxes at

transect 5 is an upper bound on, and best estimate of,

the flux into Lake Michigan.

Apparent Degradation Constants

The mass per unit thickness of TCE at transects 2, 4,

and 5 was used to estimate apparent first-order degradation constants. The constants are estimated by applying the first-order rate equation

ln 





cj+1

cj







= λ ∆ t (1)

Table 1. Mass per Unit Thickness (kg/m) at St. Joseph, Michigan

Transect

Chemical 2 1 4 5

Vinyl chloride 0.457 0.569 0.146 0.0144

1,1-DCE 0.0713 0.0245 0.00435 0.00031

t-DCE 0.170 0.152 0.0109 0.00211

c-DCE 3.70 1.53 0.567 0.0850

TCE 3.20 1.67 0.419 0.00846

Methane 1.76 1.64 1.39 0.412

Ethene 0.205 0.268 0.0524 0.00147

Ethane No data No data 0.0626 0.00051

TOC No data No data 3.79 2.49

Chloride 39.0 44.6 63.9 46.9

Sulfate 11.1 10.3 28.7 19.9

NO3-Nitrogen 0.871 0.741 1.33 2.47

NH4-Nitrogen 0.551 0.768 0.137 0.0677

TKN-Nitrogen 0.896 1.15 0.191 0.109

Table 2. Mass Flux (kg/y) at St Joseph, Michigan

Transect

Chemical 2 1 4 5

Vinyl chloride 18.81 36.03 10.69 1.676

1,1-DCE 2.934 1.551 0.3185 0.03648

t-DCE 6.995 9.609 0.7963 0.2453

c-DCE 152.1 97.11 41.48 9.868

TCE 131.7 106.0 30.67 0.9829

Methane 72.29 104.1 101.4 47.86

Ethene 8.453 16.95 3.836 0.1708

Ethane No data No data 4.577 0.05885

TOC No data No data 277.2 289.7

Chloride 1,604 2,826 4,678 5,444

Sulfate 457.4 652.9 2,102 2,306

NO3-Nitrogen 35.85 46.93 97.05 287.4

NH4-Nitrogen 22.66 48.64 10.01 7.861

TKN-Nitrogen 36.88 72.85 13.95 12.70

118
to the site data, where cj

 is the average concentration in

the transect j, cj+1 is the average concentration in the

downgradient transect j+1, ∆t is the advective travel time

for TCE to move between the transects, and λ is the

apparent degradation constant. The mass per unit thickness data for TCE and the cross-sectional area were

used to determine the average concentrations cj

 and cj+1

in the up- and downgradient transects. The porosity,

bulk density, fraction organic carbon, organic carbon

partition coefficient (5), ground-water gradient, and distance between the transects were used to determine the

advective travel times. The values used in equation 1

are given in Table 3. From these quantities, the apparent

degradation constant for TCE was determined to be

-0.0076/wk from transect 2 to 4 and -0.024/wk from

transect 4 to 5.

References

1. Engineering Science, Inc. 1990. Remedial investigation and feasibility study, St. Joseph, MI, phase I

technical memorandum. Liverpool, NY.

2. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T.

Wilson. 1993. Natural anaerobic bioremediation of

TCE at the St. Joseph, Michigan, Superfund site. In:

U.S. EPA. Symposium on bioremediation of hazardous wastes: Research, development, and field evaluations. EPA/600/R-93/054. Washington, DC (May).

pp. 57-60.

3. McCarty, P.L., and J.T. Wilson. 1992. Natural anaerobic treatment of a TCE plume, St. Joseph, Michigan,

NPL site. In: U.S. EPA. Bioremediation of hazardous

wastes. EPA/600/R-92/126. pp. 47-50.

4. Tiedeman, C., and S. Gorelick. 1993. Analysis of

uncertainty in optimal ground-water contaminant

capture design. Water Resour. Res. 29(7):2,139-

2,153.

5. U.S. EPA. 1990. Subsurface remediation guidance

table 3. EPA/540/2-90/011b.

Table 3. Chemical and Hydraulic Values Used in Estimating Apparant Degradation Rates

Transect

Area With

Nonzero TCE 

Concentration 

(m2)

Mass Per

Unit 

Thickness

From 

SITE-3D

(kg/m)

Average TCE 

Concentration

in the 

Transect

(kg/m3)

cj and cj+1 in

Equation 1

Distance

Between

Transects

(m)

Gradient 

Estimated

From

Tiedeman

and

Gorelick

(1993)

Retarded

Seepage 

Velocity

for TCEa

(m/d)

Estimated

Travel

Time 

Between 

Transects

(weeks)

∆t in 

Equation 1

2 1,592 10.67 6.70e-3

260 7.3e-3 0.11 340

4 2,774 1.397 5.04e-4

160 1.1e-2 0.156 145

5 1,943 0.0282 1.44e-5

 aConstants used in seepage velocity calculation:

Hydraulic conductivity: 7.5 m/d 

Retardation factor for TCE: 1.78 = 1 + Kocfoc ρb/θ

Porosity, θ: 0.30

Bulk density ρb: 1.86 g/cm3

Koc: 126 mL/g, foc: 0.001

119
Posters

121
Technical Protocol for Implementing the Intrinsic Remediation With 

Long-Term Monitoring Option for Natural Attenuation of 

Fuel-Hydrocarbon Contamination in Ground Water

Todd H. Wiedemeier

Engineering-Science, Inc., Denver, CO

John T. Wilson and Donald H. Kampbell

U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK

Ross N. Miller and Jerry E. Hansen

U.S. Air Force Center for Environmental Excellence, Technology Transfer Division, Brooks AFB, TX

This paper presents a brief overview of the technical

protocol, currently under development by the U.S. Air

Force Center for Environmental Excellence (AFCEE),

Technology Transfer Division, for data collection,

ground-water modeling, and exposure assessment in

support of intrinsic remediation (natural attenuation) with

long-term monitoring for restoration of fuel-hydrocarbon

contaminated ground water (1). The material presented

herein was prepared through the joint effort of AFCEE,

the Bioremediation Research Team at the U.S. Environmental Protection Agency’s Robert S. Kerr Environmental Research Laboratory in Ada, Oklahoma, and

Engineering-Science, Inc., to facilitate implementation

of intrinsic remediation at fuel-hydrocarbon contaminated sites. Specifically, this protocol is designed to

evaluate the fate of dissolved-phase fuel hydrocarbons

having regulatory maximum contaminant levels (MCLs).

The intended audience for this document is U.S. Air

Force (USAF) personnel, scientists, consultants, regulatory personnel, and others charged with remediating

fuel-hydrocarbon contaminated ground water at USAF

facilities.

Intrinsic remediation is achieved when naturally occurring attenuation mechanisms, such as biodegradation

(aerobic and anaerobic), bring about a reduction in the

total mass of a contaminant dissolved in ground water.

During intrinsic remediation, contaminants are ultimately

transformed to innocuous byproducts (e.g., carbon dioxide and water), not just transferred to another phase or

location within the environment. Intrinsic remediation

results from the integration of several subsurface attenuation mechanisms that are classified as either destructive or nondestructive. Destructive processes include biodegradation, abiotic oxidation, and hydrolysis.

Nondestructive attenuation mechanisms include sorption, dilution (caused by dispersion and infiltration), and

volatilization.

In some cases, intrinsic remediation reduces dissolvedphase contaminant concentrations to below MCLs before the contaminant plume reaches potential receptors,

even if little or no source removal or reduction takes

place. In situations where intrinsic remediation will not

reduce contaminant concentrations to below regulatory

MCLs, in an acceptable time frame, less stringent

cleanup goals may be implemented. This is especially

likely if it can be demonstrated that intrinsic remediation

will result in a continual reduction in contaminant concentrations over time such that calculated risk values

are reduced.

Intrinsic remediation is gaining regulatory acceptance

and has been implemented at several sites over the past

few years (2-5). In addition to bringing about complete

mineralization of contaminants, intrinsic remediation is

nonintrusive and allows continuing use of infrastructure

during remediation. The main limitation is that intrinsic

remediation is subject to natural and institutionally induced changes in local hydrogeologic conditions. In

addition, aquifer heterogeneity may complicate site

characterization as it will with any remedial technology.

Evaluating the effectiveness of intrinsic remediation requires the quantification of ground-water flow and solute

transport and transformation processes, including rates

of natural attenuation. Quantification of contaminant

122
migration and attenuation rates, and successful implementation of the intrinsic remediation option, requires

completion of the following steps, each of which is discussed in the following sections and outlined in Figure 1:

1. Review existing site data.

2. Develop a preliminary conceptual model for the site,

and assess the potential significance of intrinsic remediation.

3. Perform site characterization in support of intrinsic

remediation.

4. Refine the conceptual model based on site characterization data, complete premodeling calculations,

and document indicators of intrinsic remediation.

5. Model intrinsic remediation using numerical fate and

transport models that allow incorporation of a biodegradation term (e.g., BIOPLUME II or BIOPLUME III).

6. Conduct an exposure assessment.

7. Prepare a long-term monitoring plan and site long-term

monitoring wells and point-of-compliance monitoring

wells.

8. Present findings to regulatory agencies, and obtain

approval for the intrinsic remediation with long-term

monitoring option.

Collection of an adequate database during the iterative

site characterization process is an important step in the

documentation of intrinsic remediation. At a minimum,

the site characterization phase should provide data on

the location and extent of contaminant sources; data on

the location, extent, and concentration of dissolvedphase contamination; ground-water geochemical data;

geologic data on the type and distribution of subsurface

materials; and hydrogeologic parameters such as hydraulic conductivity, hydraulic gradients, and potential

contaminant migration pathways to human or ecological

receptors. Contaminant sources include nonaqueousphase liquid (NAPL) hydrocarbons present as mobile

NAPL (NAPL occurring at sufficiently high saturations to

drain, under the influence of gravity, to a well) or residual

NAPL (NAPL occurring at immobile residual saturations

that are unable to drain to a well by gravity).

The following analytical protocol should be used for

analysis of soil and ground-water samples. This analytical protocol includes all of the parameters necessary to

document intrinsic remediation of fuel hydrocarbons,

including the effects of sorption and biodegradation

(aerobic and anaerobic) of fuel hydrocarbons. Soil samples should be analyzed for total volatile and extractable

hydrocarbons, aromatic hydrocarbons, and total organic

carbon. Ground-water samples should be analyzed for

dissolved oxygen, oxidation-reduction potential, pH,

temperature, conductivity, alkalinity, nitrate, sulfate, sulfide, ferrous iron, carbon dioxide, methane, chloride,

total petroleum hydrocarbons, and aromatic hydrocarbons. The extent and distribution (vertical and horizontal) of contamination and electron acceptor and

metabolic byproduct concentrations and distributions

are of paramount importance in documenting the occurrence of biodegradation of fuel hydrocarbons and in

numerical model implementation. Dissolved oxygen

concentrations below background in an area with fuelhydrocarbon contamination are indicative of aerobic hydrocarbon biodegradation. Similarly, nitrate and sulfate

concentrations below background in an area with fuelhydrocarbon contamination are indicative of anaerobic

biodegradation through denitrification and sulfanogenesis. Elevated concentrations of metabolic byproducts

such as ferrous iron and methane in areas with fuel

hydrocarbon contamination are indicative of hydrocarbon biodegradation by the processes of ferric hydroxide

reduction and methanogenesis. Contour maps can be

used to provide visible evidence of these relationships.

To support implementation of intrinsic remediation, the

property owner must scientifically demonstrate that degradation of site contaminants is occurring at rates sufficient to be protective of human health and the environment. Three lines of evidence can be used to support

intrinsic remediation: 1) documented loss of contaminants at the field scale, 2) the use of chemical analytical

data in mass balance calculations of microbial metabolism, and 3) laboratory microcosm studies using aquifer

samples collected from the site.

The first line of evidence involves using measured dissolved-phase concentrations of biologically recalcitrant

tracers found in fuels in conjunction with aquifer hydrogeologic parameters such as seepage velocity and dilution to show that a reduction in the total mass of

contaminants is occurring at the site. The second line of

evidence involves the use of chemical analytical data in

mass balance calculations to show that a decrease in

contaminant and electron acceptor concentrations can

be directly correlated to increases in metabolic byproduct concentrations. This evidence can be used to show

that electron acceptor concentrations are sufficient to

degrade dissolved-phase contaminants. Numerical

models can be used to aid mass-balance calculations

and to collate information on degradation. The third line

of evidence, the microcosm study, involves studying site

aquifer materials under controlled conditions in the laboratory to show that indigenous biota are capable of

degrading site contaminants and to confirm rates of

contaminant degradation measured at the field scale. 

The primary objective of the intrinsic remediation investigation is to determine if natural processes of degradation will reduce contaminant concentrations in ground

water to below regulatory standards before potential

exposure pathways are completed. This requires that a

projection of the potential extent and concentration of

123
Evaluate Use of


Other Remedial


Options in


Conjunction With


Intrinsic Remediation

Free Product


Recovery Bioventing

Pump


and


Treat

Air


Sparging

Bioslurping Reactive


Barrier

Enhanced


Bioremediation

Assess Potential for


Intrinsic Remediation


With Remediation


System Installed

Refine Conceptual Model and


Complete Premodeling


Calculations

Model Intrinsic Remediation


Combined With Remedial


Option Selected Above


Using Numerical Models

Use Results of Modeling and


Site-Specific Information in


an Exposure Assessment

Is There


Unacceptable


Risk to Potential


Receptors


?

Yes Yes

Site Point-of-Compliance


Monitoring Wells and


Prepare Long-Term


Monitoring Plan

Present Findings


and Long-Term


Monitoring Plan to


Regulatory Agencies


and Reach Agreement


on Monitoring Strategy

No No

Is There


Unacceptable


Risk to Potential


Receptors


?

Document Occurrence of


Intrinsic Remediation and


Model Intrinsic Remediation


Using Numerical Models

Use Results of Modeling and


Site-Specific Information in


an Exposure Assessment

Perform Site Characterization


in Support of Intrinsic Remediation

Refine Conceptual Model and


Complete Premodeling


Calculations

Make Preliminary


Assessment of Potential


for Intrinsic Remediation


Based on Existing Site


Characterization Data

- Contaminant Type


 and Distribution


- Hydrogeology


- Location of Receptors

Develop Preliminary


Conceptual Model

Review Available


Site Data

Figure 1. Intrinsic remediation flow chart.

124
the contaminant plume in time and space is made based

on governing physical, chemical, and biological processes. This projection should be based on historic variations in—and the current extent and concentration

of—the contaminant plume, as well as on the measured

rates of contaminant attenuation.

The data collected during site characterization can be

used to model the fate and transport of contaminants in

the subsurface. Such modeling allows an estimate of the

future extent and concentration of the dissolved-phase

plume to be made. Several models, including

BIOPLUME II (6), have been used successfully to model

dissolved-phase contaminant transport and attenuation.

Additionally, a new version of the BIOPLUME model,

BIOPLUME III, is under development by AFCEE. The

intrinsic remediation modeling effort has three primary

objectives: 1) to estimate the future extent and concentration of a dissolved-phase contaminant plume by modeling the combined effects of advection, dispersion,

sorption, and biodegradation; 2) to assess the possible

risk to potential downgradient receptors; and 3) to provide technical support for the natural attenuation remedial option at postmodeling regulatory negotiations. 

Microorganisms generally utilize dissolved oxygen and

nitrate in areas with dissolved-phase fuel-hydrocarbon

contamination at rates that are instantaneous relative to

the average advective transport velocity of ground

water. This results in the consumption of these compounds at a rate approximately equal to the rate at which

they are replenished by advective flow processes. For

this reason, the use of these compounds as electron

acceptors in the biodegradation of dissolved-phase fuel

hydrocarbons is a mass-transport-limited process (7,8).

The use of dissolved oxygen and nitrate in the biodegradation of dissolved-phase fuel hydrocarbons can be

modeled using BIOPLUME II. Microorganisms generally

utilize sulfate, iron III, and carbon dioxide (used during

methanogenesis) in areas with dissolved-phase fuel-hydrocarbon contamination at rates that are slow relative

to the advective transport velocity of ground water. This

results in the consumption of these compounds at a rate

slower than the rate at which they are replenished by

advective flow processes. Therefore, the use of these

compounds as electron acceptors in the biodegradation

of dissolved-phase fuel hydrocarbons is a reaction-limited process that can be approximated by first-order

kinetics. The BIOPLUME II model utilizes a first-order

rate constant to model such biodegradation. First-order

decay constants can be determined by simple calculations based on ground-water chemistry or through the

use of laboratory microcosm studies. In addition, the use

of radiolabeled materials in a microcosm study can be

used to provide evidence of the ultimate fate of the

contaminants. 

The results of the modeling effort are not in themselves

sufficient proof that intrinsic remediation is occurring at

a given site. The results of the modeling effort are only

as good as the original data input into the model and the

model itself. Because of the inherent uncertainty associated with such predictions, it is the responsibility of the

proponent to provide sufficient evidence to demonstrate

that the mechanisms of intrinsic remediation will reduce

contaminant concentrations to acceptable levels before

potential receptors are reached. This requires the use of

conservative input parameters and numerous sensitivity

analyses so that consideration is given to all plausible

contaminant migration scenarios. When possible, both

historical data and modeling should be used to provide

information that collectively and consistently supports

the natural reduction and removal of the dissolvedphase contaminant plume. In some cases, simple calculations of contaminant attenuation rates are all that are

required to successfully support intrinsic remediation. 

Upon completion of the fate and transport modeling

effort, model predictions can be used in an exposure

assessment. If intrinsic remediation is sufficiently active

to mitigate risks to potential receptors, the proponent of

intrinsic remediation has a reasonable basis for negotiating this option with regulators. The exposure assessment allows the proponent to show that potential

exposure pathways will not be completed. 

The long-term monitoring plan consists of locating groundwater monitoring wells and developing a ground-water

sampling and analysis strategy. This plan is used to monitor plume migration over time and to verify that intrinsic

remediation is occurring at rates sufficient to protect

potential downgradient receptors. The long-term monitoring plan should be developed based on the results of

a numerical model such as BIOPLUME II.

Point-of-compliance (POC) monitoring wells are wells

that are installed at locations downgradient of the contaminant plume and upgradient of potential receptors.

POC monitoring wells are generally installed along a

property boundary or at a location approximately 5 yr

downgradient of the current plume at the seepage velocity of the ground water of 1 yr to 2 yr upgradient of

the nearest downgradient receptor, whichever is more

protective. The final number and location of POC monitoring wells depends on regulatory considerations. 

Long-term monitoring wells are wells that are placed

upgradient of, within, and immediately downgradient of

the contaminant plume. These wells are used to monitor

the effectiveness of intrinsic remediation in reducing the

total mass of contaminant within the plume. The final

number and location of long-term monitoring wells depends on regulatory considerations. Figure 2 shows a

hypothetical long-term monitoring scenario. The results

of a numerical model such as BIOPLUME II can be used

125
to help site both the long-term and POC monitoring

wells.

References

1. Wiedemeier, T.H., D.C. Downey, J.T. Wilson, D.H.

Kampbell, R.N. Miller, and J.E. Hansen. 1994. Draft

technical protocol for implementing the intrinsic remediation with long-term monitoring option for natural attenuation of dissolved-phase fuel contamination

in ground water. U.S. Air Force Center for Environmental Excellence, Technology Transfer Division.

2. Klecka, G.M., J.W. Davis, D.R. Gray, and S.S. Madsen. 1990. Natural bioremediation of organic contaminants in ground water—Cliffs-Dow Superfund

site. Ground Water 28(4):534-543.

3. Downey, D.C., and M.J. Gier. 1991. Supporting the

no-action alternative at a hydrocarbon spill site. In:

Proceedings of the USAF Environmental Restoration

Technology Symposium, San Antonio, TX (May 7-8).

Section U. pp. 1-11.

4. Wiedemeier, T.H., P.R. Guest, R.L. Henry, and C.B.

Keith. 1993. The use of Bioplume to support regulatory negotiations at a fuel spill site near Denver,

Colorado. In: Proceedings of the Petroleum Hydrocarbons and Organic Chemicals in Ground Water:

Prevention, Detection, and Restoration Conference.

NWWA/API. pp. 445-459.

5. Wiedemeier, T.H., B. Blicker, and P.R. Guest. 1994.

Risk-based approach to bioremediation of fuel hydrocarbons at a major airport. In: Proceedings of the

1994 Federal Environmental Restoration III and

Waste Minimization Conference and Exhibition, New

Orleans, LA. Hazardous Materials Control Resources Institute. pp. 51-60.

6. Rifai, H.S., P.B. Bedient, J.T. Wilson, K.M. Miller, and

J.M. Armstrong. 1988. Biodegradation modeling at

aviation fuel spill site. J. Environ. Eng. 114(5):1,007-

1,029.

7. Borden, R.C., and P.B. Bedient. 1986. Transport of

dissolved hydrocarbons influenced by oxygen limited

biodegradation—theoretical development. Water

Resour. Res. 22(13):1,973-1,982.

8. Wilson, J.T., J.F. McNabb, J.W. Cochran, T.H. Want,

M.B. Tomson, and P.B. Bedient. 1985. Influence of

microbial adaptation on the fate of organic pollutants

in ground water. Environ. Toxicol. Chem. 4:721-726.

Legend

Point-of-Compliance Monitoring Well

Long-Term Monitoring Well

Not to Scale

Aerobic


Treatment


Zone

Extent of Dissolved-


Phase BTEX Plume

Anaerobic


Treatment


Zone

Direction of


Plume Migration

Figure 2. Hypothetical long-term monitoring strategy.

126
Wisconsin’s Guidance on Naturally Occurring Biodegradation as a Remedial

Action Option

Michael J. Barden

Wisconsin Department of Natural Resources, Emergency and Remedial Response Section, Madison, WI

In February 1993, the Wisconsin Department of Natural

Resources issued an interim guidance on naturally occurring biodegradation as a remedial action option for

contaminated sites. The focus of this guidance was

primarily on soil contamination by petroleum hydrocarbons and on the requirements for site characterization

and monitoring necessary to use this approach. Subsequent implementation of the interim guidance has

resulted in further refinement to make this an effective

approach for both soil and ground-water contamination.

The application of naturally occurring biodegradation as

a remedial action requires that the site be adequately

characterized and that an adequate monitoring program

be developed and implemented. This is a long-term

remediation option, likely requiring years or decades to

effect adequate cleanup. From a regulatory perspective,

the primary concerns are that the site conditions are

amenable to naturally occurring biodegradation and that

the process is effective in reducing contaminant concentrations to acceptable levels within a reasonable period

with respect to potential contaminant migration or impacts to receptors.

Adequate site characterization is required during the site

investigation so that naturally occurring biodegradation

can be evaluated with other possible remedial action

options. This also provides baseline information for the

potential application of enhanced bioremediation as

well, because the basic site characterization requirements are essentially the same. Characterization involves identification of 1) the contaminants present and

their concentrations and biodegradability, 2) physical

and chemical parameters affecting availability of oxygen

and alternative electron acceptors, 3) nutrients, and 4)

microbiological parameters indicating the presence and

viability of appropriate microbial populations. A sufficient

number of samples should be used to represent the

extent of contamination and site heterogeneity. The

guidance provides a framework for interpretation and

evaluation of the results.

If site conditions are favorable, a monitoring plan must

be developed and implemented. Monitoring indicates

that contaminant concentrations are decreasing over

time, ensures that no unexpected contaminant migration

is occurring, and provides information regarding the nature and rate of biodegradation at the site. A variety of

monitoring approaches and techniques are available for

soil and ground water. In general, monitoring changes

in contaminant concentrations and/or concentrations of

co-reactants are appropriate.

Experience with implementation indicates that many responsible parties are unlikely to select this option due to

the long time frame involved. This suggests that naturally occurring biodegradation is more viable as an option for stable entities where time is not an issue. The

availability of the guidance, however, has encouraged

consideration of bioremediation in general as a viable

remedy due to perceived regulatory acceptance of the

technology and because the required consideration of

biodegradation potential provides baseline site information that can be used in the evaluation and design of

enhanced bioremediation systems.

127
Assessing the Efficiency of Intrinsic Bioremediation

Francis H. Chapelle

U.S. Geological Survey, Columbia, SC

Abstract

The efficiency of intrinsic bioremediation to contain contaminant migration in ground-water systems can be

quantitatively assessed by comparing rates of contaminant transport with rates of biodegradation. If transport

rates are fast relative to rates of biodegradation, contaminants can migrate freely with ground-water flow and

possibly reach a point of contact with human or wildlife

populations. Conversely, if transport rates are slow relative to biodegradation rates, contaminant migration will

be more confined and less likely to reach a point of

contact. In either case, the efficiency of intrinsic bioremediation can be assessed by evaluating the presence

or absence of contaminant transport to predetermined

points of contact. Thus, this assessment includes hydrologic (rates of ground-water flow), microbiologic (rates

of biodegradation), and sociopolitical (points of contact)

considerations. 

The U.S. Geological Survey, in cooperation with the

Naval Facilities Engineering Command, has developed

a framework for assessing the efficiency of intrinsic

bioremediation that is based on these three considerations. In this framework, hydrologic and microbiologic

information is synthesized using a solute-transport code

(SUTRA) and used to estimate rates of contaminant

transport to predefined points of contact (adjacent water

supply wells or surface water bodies). This framework

is applied to two sites, in Beaufort and Hanahan, South

Carolina, contaminated with aviation fuel. At the

Beaufort site, rates of biodegradation are slow due to

anaerobic conditions (Kbio ~0.01 d-1), but because rates

of ground-water flow are low (~0.02 ft/d), soluble contaminants are effectively contained and are not transported to adjacent points of contact. At the Hanahan site,

biodegradation rates are similarly slow under the ambient anaerobic conditions (Kbio ~0.01 d-1), but because

rates of ground-water flow are relatively high (~1.0 ft/d),

contaminants are transported to multiple points of contact with humans. These examples illustrate the complex interplay that develops between hydrologic,

microbiologic, and sociopolitical considerations, and

show that the efficiency of intrinsic bioremediation can

only be assessed on a site-by-site basis.

128
A Practical Approach to Evaluating Natural Attenuation of Contaminants in

Ground Water

Paul M. McAllister and Chen Y. Chiang

Shell Development Company, Houston, TX

The extent of natural attenuation is an important consideration in determining the most appropriate corrective

action at sites where ground-water quality has been

affected by releases of petroleum hydrocarbons or other

chemicals. The objective of this presentation is to provide guidelines for evaluating natural attenuation based

on easily obtainable field and laboratory data. 

The primary indicators that can be used to evaluate

natural attenuation include dissolved oxygen (DO) levels in ground water and plume characteristics. Background DO levels greater than 1 to 2 mg/L and inverse

correlation between DO and soluble hydrocarbon concentration have been identified through laboratory and

field studies as key indicators of aerobic biodegradation.

Several unique plume characteristics include 1) plumes

migrate more slowly than expected; 2) plumes reach a

steady state; and 3) plumes decrease in extent and

concentration, which may indicate the effects of natural

attenuation. 

When DO is depleted in an aquifer, anaerobic conditions

prevail. For biodegradation to occur, an alternative electron

acceptor such as nitrate, carbonate, or iron III must be

available. Between aerobic and anaerobic conditions (i.e.,

0.1 ppm to 2 ppm), there is a region sometimes labeled the

hypoxic zone. Studies in the hypoxic zone have indicated

that biodegradation of benzene, toluene, ethylbenzene, and

the xylenes (BTEX) may occur at relatively low DO levels

provided a secondary electron acceptor is available.

Other secondary indicators (e.g., geochemical data) and

more intensive methods (e.g., contaminant mass balances, laboratory microcosm studies, and detailed

ground-water modeling) can be applied to demonstrate

natural attenuation as well. The recommended approach for evaluating natural attenuation is to design site

assessment activities so that required data such as DO

levels and historical plume flow path concentrations are

obtained. With the necessary data, the primary indicators should be applied to evaluate natural attenuation.

129
The Use of Low Level Activities To Assist Intrinsic Bioremediation

Robert D. Norris

Eckenfelder Inc., Nashville, TN

Jeffrey C. Dey

Resource Control Corporation, Rancogas, NJ

Daniel P. Shine

Sun Company, Inc., Aston, PA

Intrinsic bioremediation as discussed in the recent report of the National Academy of Sciences Committee on

Bioremediation (1) can reduce the concentrations of

some common contaminants to levels generally considered protective of human health. Since the observations

of the role of biodegradation in limiting the extent of the

ground-water plume at the Conroe, Texas, wood-preserving site (2), many other sites have been observed

to have undergone natural attenuation at sufficient rates

to limit the size of contaminant plumes; in several instances, contaminant levels decreased to below cleanup

levels.

The use of intrinsic bioremediation, while generally attractive from a cost perspective, may actually be desirable environmentally because secondary effects of

active remediation are avoided. Intrinsic bioremediation,

however, has several costs associated with its use.

These include some level of documentation that biological degradation of constituents of concern is taking

place and costs associated with monitoring and management of the site. Managing a small plume, such as

is frequently found at service stations, includes sampling

and reporting to the responsible environmental agency.

At one typical site (3), the annual environmental management costs have been approximately $11,900 per

year, consisting of site visits and sampling ($1,000),

chemical analysis ($7,200), reporting ($2,400), and consulting ($1,300). At other sites, costs have exceeded

these values by 50 percent or more. Documenting that

biodegradation is occurring adds substantially to these

costs.

Intrinsic bioremediation will be effective where the electron acceptor requirements are relatively small. While

oxygen may reach the affected zones at rates sufficient

to prevent and shrink contaminant plumes, and thus

eventually achieve remedial goals, the time frame may

be unacceptably long from the site owner’s perspective

because of long-term monitoring costs and management burdens.

Addition of appropriate electron acceptors would accelerate reduction in constituents such as monoaromatic

hydrocarbons. In some cases where intrinsic bioremediation is technically feasible, it may not be the most

cost-effective approach. To evaluate the concept of applying limited engineering solutions at sites where intrinsic bioremediation appears to be slowly reducing the

contaminant mass, we are testing the use of air sparging

wells (3). The air sparging wells are placed immediately

outside the plumes and operated intermittently at low

flow. The cost of installing three shallow air sparging

wells, routine maintenance, limited additional sampling,

and reporting was budgeted at $8,500 per site for each

of the three sites. If the time to reach closure is shortened by 1 yr or more, the cost of treatment will have

been less than the cost that would have been incurred

by only monitoring and managing the site.

Another approach to providing electron acceptors is to

add aqueous solutions of hydrogen peroxide or nitrate

to wells located within or immediately upgradient of the

plume; however, nitrate addition has not been shown to

be effective for degradation of benzene, and introduction

of hydrogen peroxide is more labor intensive and thus

more costly than air sparging. Alternatively, a slowrelease oxygen compound such as magnesium peroxide can be placed in wells and oxygen allowed to diffuse

into the formation.

Regulatory acceptance for natural attenuation may be

more easily attained if a migration barrier is created

130
along the downgradient edge of the plume. Migration

barriers can be created through a series of low-flow air

sparging wells. Alternatively, a row of wells containing a

slow-release oxygen compound can be placed perpendicular to the ground-water gradient near the downgradient edge of the plume. As demonstrated by Bob Borden

(4) of the University of North Carolina at a commercial

site in North Carolina and by Doug Mackay (5) of Stanford University in tests conducted at the Borden Landfill

in Canada, this approach can successfully prevent migration of monoaromatic hydrocarbons.

Intrinsic bioremediation should be viewed as one approach in a continuum of methods of utilizing biodegradation processes to remediate soil and ground water. It

should be used alone, in combination with other approaches, or as a polishing step based on evaluation of

the site conditions, regulatory issues, technical feasibility, implementability, and cost.

This poster session will present cost and analytical data

from three New Jersey service stations during the monitoring-only phase, site maps, diagrams of the air

sparging systems that have been installed, costs of

installation and operation, and analytical data available

at the time of the meeting. The cost of adding hydrogen

peroxide at two similar sites will be discussed. Additionally, the costs and relative advantages of the use of

slow-release oxygen compounds and air sparging will

be presented. 

References

1. National Research Council. 1993. In situ bioremediation: When does it work? Washington, DC: National

Academy Press.

2. Wilson, J.T., J.F. McNabb, J. Cochran, T.H. Wang,

M.B. Tomson, and P.B. Bedient. 1985. Influence of

microbial adaptation on the fate of organic pollutants

in ground water. Environ. Toxicol. Chem. 4:721-726.

3. Norris, R.D., J.C. Dey, and D.P. Shine. 1993. The

advantages of concerted bioremediation of lightly

contaminated sites compared to intrinsic bioremediation. Presented at the American Chemical Society

I&E Special Symposium, Atlanta, GA.

4. Kao, C.-M., and R.C. Borden. 1994. Enhanced aerobic bioremediation of a gasoline contaminated aquifer by oxygen-releasing barriers. In: Hydrocarbon

bioremediation. Boca Raton, FL: Lewis Publishers. 

5. Bianchi-Mosquera, G.C., R.M. Allen-King, and D.D.

Mackay. 1994. Enhanced degradation of dissolved

benzene and toluene using a solid oxygen-releasing

compound. Ground Water Monitor. Rev. pp. 120-128.

131
Natural Attenuation of Jet Fuel in Ground Water

Greg Doyle, Dwayne Graves, and Kandi Brown

International Technologies Corporation, San Bernardino, CA

Natural attenuation is a minimum action remedial strategy that permits the biodegradation of organic contaminants under natural, in situ conditions. Mechanisms that

act to affect contaminant biodegradation include aerobic

biodegradation at the plume boundary and various

anaerobic processes within the plume. Nitrate, iron,

manganese, and sulfate reduction and methanogenesis

are thought to support contaminant biodegradation.

Naturally occurring levels of nitrate, sulfate, oxidized

iron and manganese, and carbonates support anaerobic

biodegradation.

Natural attenuation was proposed as a feasible remedial

alternative for the dissolved jet fuel plume at George Air

Force Base in Victorville, California. The plume, covering approximately 1.1 million square feet, was located in

a poorly yielding perched aquifer, 120 ft below the

ground surface. Ground-water flow rate was approximately 20 ft/yr, and the edge of the plume was about

2,000 ft from the property line. The ground water in the

affected aquifer was not being used. Because of subsurface conditions, slow migration, poor water yield, and

lack of use, natural attenuation represented the most

efficient and cost-effective approach for remediating dissolved jet fuel at this site. 

A separate-phase layer of jet fuel floated on the water

table near the center of the plume. An aggressive skimming operation was employed to remove all recoverable

separate-phase jet fuel from the subsurface. Bioventing

will be used to further remediate contaminated vadose

zone soil in areas where jet fuel was detected and

removed.

Natural attenuation for aerobic and aerobic/anaerobic conditions was modeled using BIOPLUME II. The BIOPLUME

II model simulates the transport of dissolved hydrocarbons

under the influence of oxygen-limited biodegradation.

Using BIOPLUME II with site-specific parameters, various treatment scenarios were evaluated. Assuming that

60 percent of the separate-phase product was removed,

both anaerobic and aerobic biodegradation occurred,

and residual jet fuel diffused into the water based on

Fick’s Law of Diffusion, a remediation time of 42 yr was

determined. This treatment time was adequate to remediate the plume before it migrated off site. 

Based on this prediction and the cost savings realized

by applying natural attenuation, a 5-yr performance period was established prior to the issuance of a finalized

record of decision. Joint efforts by the Air Force, the U.S.

Environmental Protection Agency’s Robert S. Kerr Environmental Research Laboratory, and International Technologies Corporation will provide data verifying the

accuracy of the BIOPLUME II model predictions and

demonstrating the level of natural biological activity occurring in the ground water. These efforts are expected

to lead to regulatory acceptance of natural attenuation

for the full-scale remediation of the jet fuel plume on site.

132
Evaluation of Intrinsic Bioremediation at an Underground Storage Tank Site in

Northern Utah

R. Ryan Dupont, Darwin L. Sorensen, and Marion Kemblowski

Utah Water Research Laboratory, Utah State University, Logan, UT

A 2-yr field study was initiated by the Utah Water Research Laboratory for the U.S. Environmental Protection

Agency’s Office of Underground Storage Tanks at two

former underground storage tank (UST) sites in northern

Utah: a U.S. Air Force site at Hill Air Force Base (AFB)

near Ogden, Utah, and a private site in Layton, Utah.

The study sought to evaluate rapid site assessment

techniques and data collection and summary methods

that could be used to provide a comprehensive description of the potential for intrinsic bioremediation at UST

sites. This poster presentation focuses on the Hill AFB

site, where tank and line leaks from an 18,000 gal UST,

removed in 1989, were the probable sources of observed ground-water contamination. Total dissolved

petroleum hydrocarbon contaminant mass at the beginning of the study was estimated to be 950,000 mg, while

approximately 6,000 mg of benzene and 8,000 mg of

toluene were quantified in the plume underlying the site.

The site was covered with a permeable gravel surface

layer during the study, and dissolved oxygen concentrations at a number of locations in uncontaminated areas

surrounding the plume were over 2 mg/L.

Initial site assessment activities utilized cone penetrometry for rapid collection of subsurface soil stratigraphic data and for the placement of more than 60

small-diameter ground-water sampling points at the two

shallow field sites. Ambient temperature headspace

analyses were conducted using a polyethylene bag

method developed by the Utah Water Research Laboratory in addition to a commercial Lag-in-a-Bag apparatus to provide rapid field-determined measurements of

ground-water total hydrocarbons. Field total petroleum

hydrocarbon (TPH) measurements were collected on a

near real-time basis to guide the initial placement of

ground-water monitoring points along and perpendicular

to the axis of the contaminant plume at each site. These

field methods proved that the original conceptual model

of the nature and extent of contamination and of the

potential contaminant migration pathways at the Hill

AFB site, based on conventional site assessment techniques (soil gas survey, collection of limited soil core

samples, and placement of more limited numbers of

large-diameter ground-water monitoring wells), was

greatly in error. The model could be improved significantly with these cost-effective field techniques, which

are now widely available to the consulting community.

Seven field sampling events, beginning in April 1992,

were conducted over the 2-yr study period as part of a

proposed intrinsic bioremediation strategy developed in

the project. Ground-water quality data were collected from

the small-diameter sampling points and existing groundwater monitoring wells to assess the distribution and transport of contaminants, along with the predominant microbial

reactions taking place within the contaminant plume.

Ground-water quality data collected included field measurements of pH, dissolved oxygen, temperature, and

ambient headspace total hydrocarbon concentrations;

laboratory determinations included nitrate-N, sulfate,

dissolved iron and manganese, total hydrocarbons

(purge and trap and semivolatile constituents), specific

organic contaminants (C-6 to C-15 alkanes and benzene, toluene, ethylbenzene, p-xylene, naphthalene,

and methylnaphthalene), and contaminant boiling point

split concentrations.

Ground-water data were used to generate total and

specific compound dissolved contaminant and dissolved electron acceptor mass data for each sampling

period. The location of the center of the mass of contaminants and electron acceptors were evaluated, and

changes in these parameters over time, with respect to

bulk ground-water flow, were used to assess the rate

and magnitude of natural degradation processes taking

place at each site.

Over the 2-yr study period, the mass of TPH showed

exponential decay, while all specific constituents displayed zero-order decay rates. Dissolved TPH mass

declined to less than 1,000 mg (99.9 percent removal)

133
by the end of the 630-d monitoring period, while

dissolved benzene and toluene mass remaining in

the contaminant plume declined to less than 200 mg

(97 percent removal). TPH, benzene, and toluene mass

decay rates were found to be -0.013/d (p = 0.03,

r

2 = 0.933), -11.2 mg benzene/d (p = 0.01, r2

 = 0.976),

and -14.2 mg toluene/d (p = 0.01, r2

 = 0.998), respectively. Mass center data indicate that while ground-water

velocities at the site through the study period averaged

0.45 ft/d, the net movement of contaminants was attenuated significantly, with measured center of mass velocities of TPH, benzene, and toluene being 0.03 ft/d, 0.05

ft/d, and 0.07 ft/d, respectively. Corresponding utilization

of oxygen and other terminal electron acceptors occurred across the plume.

Intrinsic bioremediation of the dissolved plume at the Hill

AFB site successfully attenuated and removed the residual hydrocarbon mass existing at the site at the beginning of the study. In January 1994, only 1 of 34

monitoring well/piezometer ground-water samples contained a benzene concentration (20.7 μg/L) above regulatory concern, and the site is expected to be eligible

for closure at the next routine, semiannual sampling

event.

This poster session will detail the physical/chemical

characteristics of the field sites and the rapid site assessment techniques and typical results collected, as

well as highlight the data collection/reduction/interpretation methodology developed in this study. Finally, more

complete results demonstrating natural degradation of

hydrocarbon contaminants at this site will be presented,

along with a summary of a natural attenuation decision

support system to aid investigators in assessing the

viability of intrinsic bioremediation for the selection of a

“no action”/natural attenuation monitoring alternative at

their sites.

134
Case Studies of Field Sites To Demonstrate Natural Attenuation of BTEX

Compounds in Ground Water

Chen Y. Chiang and Paul M. McAllister

Shell Development Company, Houston, TX

The most definitive indicators of natural attenuation

such as plume characteristics and dissolved oxygen

(DO) concentrations are based on actual concentrations

obtained during periodic monitoring events. Based on

appropriate data from monitoring wells, the following

parameters can be used to indicate and demonstrate

that natural attenuation is occurring: 1) the mass of

benzene, toluene, ethylbenzene, and the xylenes

(BTEX) present and 2) the extent and rate of migration

and distribution of BTEX concentrations. Data collected

from two field sites will be used to demonstrate natural

attenuation mechanisms.

The first site is characterized by 42 monitoring wells to

show the relationship between soluble BTEX and DO

plumes. Results from 10 sampling periods over 3 years

show a significant reduction in total BTEX mass with

time in ground water. These reduction and leakage rates

from sources are determined from material balance and

nonlinear least-squares analyses. The natural attenuation rate is calculated to be 0.95 percent/d. Spatial

relationships between DO and total BTEX are shown to

be strongly correlated by statistical analyses and solute

transport modeling. In addition, laboratory microcosm

biodegradation experiments are performed to determine

possible threshold limits for aromatic hydrocarbon oxidation under varying levels of DO. The results are remarkably consistent with field data on the presence of

high or low levels of BTEX and DO in several monitoring

well-water samples.

The second site data will be used to demonstrate natural

attenuation from a cost-effectiveness perspective

through evaluation of plume characteristics over time.

The benzene concentrations along the primary flow path

at this site are observed to decrease from 2,600 ppb at

the source to 2.7 ppb at a distance 1,425 ft downgradient. The decrease in concentrations with distance from

the source is a direct indication that some degree of

natural attenuation is occurring. If no natural attenuation

was occurring, then concentrations would remain relatively constant out to the leading edge, where a sharp

front would be observed. It is emphasized that natural

attenuation also includes other mechanisms than biodegradation: dispersion, sorption, volatilization, and chemical

transformation.

135
Demonstrating Intrinsic Bioremediation of BTEX at a Natural Gas Plant

Keith Piontek, Tom Sale, and Jake Gallegos

CH2M Hill, St. Louis, MO, and Denver, CO

Steve de Albuquerque and John Cruze

Phillips Petroleum Company, Bellaire, TX, and Bartlesville, OK

Intrinsic bioremediation is being characterized at the site

of a former natural gas plant. Gas plant operations

resulted in the release of nonaqueous-phase liquid

(NAPL) to the eolian deposits beneath the site. Previous

investigation of the site had established that NAPL

(natural gas condensate) had been released to the subsurface, and had delineated to the extent of benzene,

toluene, ethylbenzene, and xylene (BTEX) in ground

water beneath the site. The BTEX biodegradation that

occurs under natural conditions is being characterized

to provide for more realistic assessment of potential

risks posed by the site, and to support long-term site

management decision-making.

The technical protocol proposed by Dr. John Wilson of

the U.S. Environmental Protection Agency’s Robert S.

Kerr Environmental Research Laboratory was used as

a guide in planning the intrinsic bioremediation demonstration. Key parameters assessed included seasonal

variations in ground-water flow direction and velocity,

and changes in the concentrations of hydrocarbons and

geochemical parameters (e.g., electron acceptors)

across the site. The ground-water monitoring revealed

significant changes in ground-water geochemistry at locations upgradient, within, and downgradient of the

NAPL zone. These data, together with knowledge of the

stoichiometry of hydrocarbon biodegradation under various redox conditions, were used to characterize the extent

of naturally occurring hydrocarbon biodegradation. 

It was found that while oxygen flux into the plume occurs, the majority of contaminant mass removal occurs

under anoxic conditions. Site data suggest that sulfate

is the most significant electron acceptor in terms of

hydrocarbon mass removal, with over 90 percent of the

hydrocarbon mass removal attributable to sulfate reduction. Additional hydrocarbon mass removal is attributable to reduction of nitrate, ferris iron, and oxygen; and

methanogenic activity.

The project provided evidence that naturally occurring

contaminant biodegradation plays a significant role in

limiting contaminant migration from the site, as summarized below:

• There is a significant flux of electron acceptors, particularly sulfate, into the NAPL zone. Depressed concentrations of these electron acceptors within the NAPL

zone are indicative of contaminant biodegradation.

• Elevated levels of bicarbonate, a byproduct of contaminant biodegradation, are found in monitoring

wells located downgradient of the NAPL zone. This

indicates that ground water from the NAPL zone has

reached the monitoring wells. BTEX is not present at

detectable concentrations in these wells.

• Contaminant transport calculations indicate that, in

the absence of naturally occurring contaminant

biodegradation, benzene would likely have migrated

to the downgradient monitoring wells by this time. As

stated above, benzene has not been detected in

these monitoring wells.

• The presence of oxygen and nitrate in the downgradient wells indicates that flux of electron acceptors into

ground water downgradient of the NAPL zone provides additional capacity for contaminant biodegradation, in the event of further lateral BTEX migration.

136
Demonstrating the Feasibility of Intrinsic Bioremediation at a Former

Manufactured Gas Plant

Ian D. MacFarlane

EA Engineering, Science, and Technology, Sparks, MD

Edward J. Bouwer

The Johns Hopkins University, Baltimore, MD

Patricia J.S. Colberg

University of Wyoming, Laramie, WY

A former manufactured gas plant (MGP) in Baltimore,

Maryland, is the subject of an investigation to assess

natural, in situ biodegradation for the purposes of basing

remediation decisions on intrinsic bioremediation and

engineered, enhanced bioremediation. Tar, a byproduct

of the former gas manufacturing operation, is found in

the site’s subsurface to depths as great as 100 ft. The

tar is a dense nonaqueous-phase liquid (DNAPL) that

contains monocyclic aromatic hydrocarbons (MAHs),

such as benzene and toluene, and polycyclic aromatic

hydrocarbons (PAHs), such as naphthalene and

benzo(a)pyrene. The tar DNAPL acts as a long-term

source of aromatic hydrocarbons to ground water, and

concentrations in ground water over much of the 70-acre

site are close to theoretical effective solubilities.

Laboratory and field data were collected to evaluate

biodegradation. The first phase of the laboratory investigation, consisting of tests on 16 soil samples from

various depths, was to discern whether or not microbes

existed in the subsurface, if they could use MGP-tar as

a carbon source, and if relationships could be established between hydrogeology, contaminant distribution,

and microbiological characteristics. Aerobic and anaerobic enumerations were performed, followed by fatty acid

analyses to identify the microbes. Viable aerobic bacteria were detected in all subsurface samples. Bacteria

were grown in the samples plated under anaerobic conditions, but at counts of 10 percent to 50 percent less

than the corresponding aerobic counts. Tar-degrading

bacteria were detected in 7 of the 16 samples. 

The second laboratory phase consisted of more detailed

microcosm studies performed by the Johns Hopkins

University (JHU) and the University of Wyoming (UW).

JHU used 49 soil samples from five boreholes with site

ground-water and individual radiolabeled target substrates (benzene, naphthalene, phenanthrene, and acetic acid) in sealed vials to make microcosms that

mimicked in situ redox conditions (i.e., oxygenated or

unoxygenated). Viable bacteria were enumerated, and

total cell counts were performed by the acridine orange

direct-count method. Similar microcosm studies are being performed by UW under sulfidogenic, iron-reducing,

and methanogenic conditions using phenol, benzene,

toluene, naphthalene, and phenanthrene as the targeted radiolabeled compounds. Benzene, naphthalene,

and phenanthrene were observed to mineralize 6 percent

to 24 percent, 8 percent to 43 percent, and 3 percent to

31 percent, respectively, in JHU aerobic microcosms

over a 4-week incubation period. Anaerobic naphthalene mineralization (7 percent to 13 percent) was observed in two JHU samples in the presence of NO3.

Half-lives calculated from first-order degradation rates

typically ranged from tens to hundreds of days, with a

lower half-life in the initial stages of incubation followed

by a slower rate presumably indicative of oxygen- or

nutrient-limiting conditions. Under sulfate-reducing conditions, phenol mineralized 13 percent to 18 percent in

200 d, but benzene, naphthalene, and phenanthrene

showed less than 1 percent mineralization in the same

period, and toluene showed less than 1 percent mineralization in 165 d. The apparent limited transformation

may, in fact, be due to the small inoculum sizes used, a

phenomenon documented by others.

Field investigations for natural in situ biodegradation

included aqueous phase redox conditions, biogenic

product analyses, and apparent attenuation of model

137
contaminants. Ground-water quality data generally

showed reduced conditions with little or no measurable

oxygen, low redox potential (-70 mV average), high

biochemical oxygen demand in source zones (>200 mg/L),

elevated sulfate (2,200 mg/L average), and elevated

iron (570 mg/L average). Biogenic gases (CO2, H2S,

and CH4) were detected at levels greater than atmospheric in 11 of the 16 wells measured.

Apparent degradation rates were calculated using the

first-order model by regressing the natural log of constituent concentrations (adjusted for dilution) by estimated in situ travel time. Because no nonreactive,

conservative tracers are unique to the tar sources, dilution was estimated by considering dissolved carbon

(organic plus inorganic) as a tracer. While the use of

carbon as a tracer is prone to error due to its reactive

nature, its use is conservative in that observed carbon

concentrations would tend to be less than actual anthropogenic carbon, resulting in overestimates of dilution

and underestimates of degradation rates. Half-lives for

benzene, toluene, ethylbenzene, xylenes, and naphthalene were calculated to be 729, 660, 877, 855, and

2,166 d, respectively. Half-lives for the three-ring and

greater PAHs were not calculated due to the poor regression correlation coefficients. Degradation estimates

showed that toluene is the most preferred aromatic substrate studied and that naphthalene appears to degrade

the slowest as predictable from the literature. Surprisingly, the benzene rate was only slightly less than the

toluene rate.

Laboratory investigations have shown that 1) microbes

exist in the subsurface, 2) microbes are capable of using

tar as a carbon source, 3) various redox conditions can

be established with site consortia, and 4) site bacteria

can degrade selected aromatic constituents under aerobic and anaerobic conditions. Field evidence showing

various redox conditions and biogenic products of organic degradation gives clues to the possible fate of

aromatic hydrocarbons, but this indirect evidence can

only be used to support more definitive evidence in

demonstrating intrinsic bioremediation. Although estimates of in situ constituent decay are based on numerous assumptions and are fraught with uncertainty, this

evidence is needed to show real attenuation of aromatic

hydrocarbons. In this case, enough geochemical and

hydrogeologic data were available to segregate dilution

(an important “attenuation” process) from degradation

processes. The task of estimating apparent degradation

was facilitated by the relatively simple hydraulics and

the aged system to allow assumption for negligible sorption. The attenuation assessment demonstrated that

contaminant loss was observed over time along the

aqueous plume travel path (i.e., travel time, rather than

at one point over time) and degradation, probably biotic,

was measured for target contaminants.

The combined laboratory and field evidence point to

natural in situ biodegradation as an active process in the

site’s low oxygen, subsurface environment. Intrinsic

bioremediation, albeit slow due to mass transfer limitations from the tar to the aqueous phase, may be technically viable for controlling aqueous aromatic hydrocarbon

contamination emanating from MGP-tar sources. Laboratory studies are continuing to explore compound-specific biodegradation under various conditions, and plans

are being formulated now for an in situ biodegradation

pilot study.

138
Natural and Enhanced Bioremediation of Aromatic Hydrocarbons at Seal Beach,

California: Laboratory and Field Investigations

Harold A. Ball, Gary D. Hopkins, Eva Orwin, and Martin Reinhard

Western Region Hazardous Substance Research Center, Stanford, CA

Introduction

The objective of this study was to develop our understanding of environmental factors that are important for

intrinsic anaerobic biodegradation of aromatic hydrocarbons in contaminated ground-water aquifers, and to

determine methods to enhance this process. The focus

of the investigation was a site at the Seal Beach Naval

Weapons Station in southern California, where a significant gasoline spill resulted in contamination of the

ground-water aquifer (1). In the field, nitrate was present

at about 5 mg/L in background wells and approached

detection levels ( mg/L) in the contaminated wells. There

was a high natural background sulfate concentration of

about 85 mg/L in the ground water, and methane was

detected in the contaminated well headspace. The distribution of aromatics present in the contaminated ground

water differs from that expected from dissolution of pure

gasoline (2). This suggests that natural biotransformation of several organic species is occurring at the site.

The project was divided into laboratory and field components, which were interrelated. The goals of both the

laboratory and field experiments were to determine the

capability of the native aquifer microbial community to

transform aromatic hydrocarbon compounds under anaerobic conditions and to understand the effect of alternate environmental conditions on the transformation

processes. Field experiments were carried out on site at

Seal Beach. At the field site, experimental monitoring of

biotransformation under natural (presumed sulfatereducing) and nitrate-enhanced conditions have been

carried out. 

Approach and Results

Laboratory Study

In a laboratory microcosm experiment (3), we evaluated

several factors which were hypothesized to influence in

situ biotransformation processes. Individual monoaromatic

compounds (e.g., benzene, toluene, ethylbenzene, and

m-, p-, and o- xylene) were the primary substrates. In

replicate bottles during the first 52 d of the study, toluene

and m+p-xylene (here, m-xylene and p-xylene were

measured as a summed parameter) were biotransformed in the unamended ground-water samples under

presumed sulfate-reducing conditions. Addition of nitrate to the ground water increased rates of toluene

biotransformation coupled to nitrate reduction, stimulated biotransformation of ethylbenzene, and inhibited

the complete loss of m+p-xylene that was observed

when nitrate was not added and sulfate-reducing conditions prevailed. Addition of the nutrients ammonia and

phosphate had no effect on either the rate of aromatics

transformation or the distribution of aromatics transformed. When Seal Beach sediment was placed into

nitrate-reducing media, ethylbenzene was transformed

first, followed by toluene. When the sediment was

placed into sulfate-reducing media, lag times were increased, but toluene and m-xylene were ultimately

transformed just as in the microcosms with ground water

alone. Although methane had been detected in the field,

there appeared to be no transformation of aromatic

compounds in the methanogenic microcosms during the

period of the experiment.

Bioreactor Study

A pilot-scale facility consisting of 90-L reactors was constructed at the Seal Beach site (4-6). The facility was

designed for the operation of three anaerobic in situ

bioreactors. The reactors consisted of aquifer-sedimentfilled, stainless steel cylindrical vessels with the capability to control and monitor both hydrodynamic flow and

supplements to the composition of the native groundwater influent. Initial operation of the three anoxic/anaerobic reactors focused on evaluating anaerobic bioremediation strategies for aromatic hydrocarbons under

natural (presumed sulfate-reducing) and enhanced denitrifying conditions. 

139
Bioreactor results were consistent with the laboratory

microcosm experiments. Toluene and m+p-xylene were

degraded in both the unamended and nitrate-amended

bioreactors. Degradation of ethylbenzene was stimulated by nitrate addition. There was no evidence that

benzene or o-xylene was transformed in either reactor.

The final percentage removal efficiency appeared to be

higher in the unamended bioreactor, where flow was

slower.

Field Study

Field experiments have been conducted to assess aromatic biotransformation in a test zone within the contaminated aquifer at the Seal Beach site. Initial work

focused on evaluation of intrinsic bioremediation as evidenced by the distribution of aromatic species in background wells. Subsequent experiments to determine our

ability to enhance this biotransformation have been conducted using a slug test experimental design in which a

single well was used for the injection of the “slug” or test

pulse and the same well was used to extract the test

pulse. Since the native ground water contained a variety

of electron acceptors and the water used for the injected

pulses was water that had previously been extracted

from the test zone, the ground water was treated to

control the concentration of all electron acceptors and

organics during the injection of the test pulse. Before

injection, the desired salts were added back to the deoxygenated injection stream, and the stream was metered into the injection well. Sodium bromide was added

as a conservative tracer. Under this scenario, the different electron acceptors investigated (e.g., nitrate and

sulfate) could be added as desired. During initial tracer

studies, the injection water was organics free, and thus

the source of the organics was desorption from the in

situ aquifer solids. In subsequent and ongoing bioremediation studies, benzene, toluene, ethylbenzene, mxylene, and o-xylene were added with the injection pulse

at a concentration of approximately 200 μg/L each.

The initial bromide tracer data showed stable tracer

concentrations and indicated no substantial encroachment of native ground water detected in the first 0.4 pore

volumes. There was a very small hydraulic gradient at

the site, hence recovery of the bromide mass from the

test wells ranged from 93 percent to 99 percent with the

extraction of three pore volumes over a 103-d period.

During the tracer test, the equilibrium desorption concentrations for the aromatic hydrocarbons when the

electron acceptors nitrate and sulfate were absent from

the ground water were evaluated. Benzene, ethylbenzene, and o-xylene concentrations remained relatively

stable and thus appeared to be at an equilibrium. The

toluene and m+p-xylene concentrations had a downward trend relative to benzene once the native ground

water encroached after approximately 0.4 pore volumes, suggesting that the nitrate and sulfate concentrations available in the native ground water supported

some intrinsic biological activity in the latter part of the

experiment for toluene and m+p-xylene removal. 

In a nitrate augmentation experiment, nitrate and aromatics were added to the injection pulse, resulting in

complete consumption of toluene and m-xylene followed

by ethylbenzene within the first 2 wk. o-Xylene was

slowly degraded, and its concentration approached zero

by day 60. There was no apparent loss of benzene when

compared with the inert tracer. The addition of nitrate to

the test region appeared to enhance the natural anaerobic denitrifying population. This would confirm that there

was already an active nitrate-reducing population in the

aquifer whose activity was enhanced by the addition of

nitrate. With the exception of o-xylene transformation,

these results were comparable with those from the nitrate-amended microcosm and bioreactor experiments,

wherein toluene, ethylbenzene, and m-xylene were

transformed under denitrifying conditions.

During the tracer study, methane was detected in the

test wells. With the encroachment of the native ground

water and associated increase in nitrate and sulfate

concentrations, the methane concentration decreased

to values close to zero, suggesting that nitrate and

sulfate inhibit methanogenesis at this site. 

Acknowledgment

Funding for this study was provided by the U.S. Environmental Protection Agency’s Office of Research and

Development, under agreement R-815738-01 through

the Western Region Hazardous Substance Research

Center. The content of this study does not necessarily

represent the views of the Agency. Additional funding

was obtained from the Chevron Research and Technology Company, Richmond, California.

References

1. Schroeder, R.A. 1991. Delineation of a hydrocarbon

(weathered gasoline) plume in shallow deposits at

the U.S. Naval Weapons Station, Seal Beach, California. Water Res. Invest. Rep. 89-4203. 

2. Cline, P.V., J.J. Delfino, and P.S.C. Rao. 1991. Partitioning of aromatic constituents into water from

gasoline and other complex solvent mixtures. Environ. Sci. Technol. 25(5):914-920.

3. Ball, H.A., and M. Reinhard. 1994. Laboratory study

of monoaromatic hydrocarbon degradation under anaerobic conditions at Seal Beach, California. In

preparation.

4. Ball, H.A., and M. Reinhard, M. 1994. Pilot-scale

study of monoaromatic hydrocarbon degradation under anaerobic conditions at Seal Beach, California.

In preparation. 

140
5. Huxley, M.P., C. Lebron, M. Reinhard, H. Ball, H.F.

Ridgway, and D. Phipps. 1992. Anaerobic and aerobic degradation of aromatic hydrocarbons using in

situ bioreactors at an unleaded gasoline spill site.

Presented at the 18th Environmental Symposium of

the American Defense Preparedness Association,

Alexandria, VA.

6. Reinhard, M., L.E. Wills, H.A. Ball, T. Harmon, D.W.

Phipps, H.F. Ridgway, and M.P. Eisman. 1991. A field

experiment for the anaerobic biotransformation of

aromatic hydrocarbon compounds at Seal Beach,

California. In: Hinchee, R.E., and R.F. Olfenbuttel,

eds. In situ bioreclamation: Applications and investigations for hydrocarbon and contaminated site remediation. Boston, MA: Butterworth-Heinemann.

141
The Complete Dechlorination of Trichloroethene to Ethene Under Natural

Conditions in a Shallow Bedrock Aquifer Located in New York State

David Major, Evan Cox, and Elizabeth Edwards

Beak Consultants Limited, Guelph, Ontario

Paul W. Hare

General Electric Company, Corporate Environmental Programs, Albany, NY

Introduction

In anaerobic environments, chlorinated ethenes can act

as electron acceptors in a process called reductive dehalogenation (specifically, reductive dechlorination).

The extent of dechlorination, however, has been shown

to vary in anaerobic dechlorination studies depending

upon the flow and availability of electrons within the

anaerobic microbial community. For example, dechlorinated intermediates such as dichloroethene (DCE) and

vinyl chloride (VC) were found to accumulate during the

dechlorination of tetrachloroethene (PCE) and trichloroethene (TCE) (1-3). Freedman and Gossett (4), however, were the first to observe the complete dechlorination

of PCE to ethene in a methanogenic enrichment culture,

and DeBruin et al. (5) showed that PCE could be reduced to ethane. The first field observation of the complete dechlorination of PCE to ethene was documented

by Major et al. (6). Their laboratory and field study

showed that anaerobic microorganisms in a low-permeability aquifer were capable of naturally dechlorinating

PCE in the presence of methanol. This paper documents that microorganisms in a bedrock aquifer unit are

also capable of completely dechlorinating TCE to

ethene.

Study Site Conditions

The study site is located in the Finger Lakes region of

central New York. The property was used from 1951 to

1990 as a manufacturing plant for a variety of electrical

components, including high-voltage semiconductors. In

the early to mid-1960s waste solvents were disposed of

in an unlined evaporation pit. TCE, which was often

mixed with acetone or methanol, was among the solvents disposed of in the unlined evaporation pit. As a

result, these chemicals are now found in the overburden

and bedrock units beneath the study site.

Results

Our study involved collecting representative groundwater samples from 21 existing ground-water monitoring

wells, mostly in the shallow bedrock unit, for geochemical and microbiological analyses. 1,2-DCE and VC were

detected in ground-water samples, which indicated that

TCE was being biodegraded in the subsurface at the

site. These TCE degradation products were not used or

produced at the site, and thus their presence can only

be attributed to the dechlorination of TCE. In addition,

the detection of ethene provides evidence that VC is

being dechlorinated at the site. Three observations of

the relative distribution of TCE and its dechlorination

products suggested that the migration of the volatile

organic compounds (VOCs) in the shallow bedrock unit

is being controlled by biodegradation. First, the distribution of TCE is much less extensive than the observed

distributions for 1,2-DCE, VC, and ethene. Second, the

distribution of VC, which should be greater than that of

1,2-DCE as predicted by its mobility in ground water,

was essentially the same as the distribution of 1,2-DCE.

Third, VC and ethene migrate at similar rates relative to

ground-water velocity and should have had similar distributions, but the distribution of VC was less than

ethene. The distribution of acetone and methanol is

considerably limited in comparison to the distribution of

the chlorinated VOCs and ethene. The mobility of acetone and methanol should be approximately the same

as the average linear ground-water flow; however, their

distribution was less than the distribution of the VOCs.

This suggests that acetone and methanol are also being

biodegraded. 

In addition to the distributions of VOCs, the distributions

of inorganic anions, methane and methane isotopes,

and volatile fatty acids (e.g., acetate) were used as

indirect measures of the activity of functional groups of

142
microorganisms in the bedrock. The depletion of sulfate

and the production of methane and acetate indicated

that sulfate-reducing, methanogenic, and acetogenic

bacteria were active in the bedrock aquifer. Isotopic

analysis of methane indicated that the methane was

produced biotically. Furthermore, the distribution of

methane and methane isotopes clearly showed that the

microorganisms were active in the bedrock. 

Microbial biomass, composition, and nutritional status

were assessed by extracting and analyzing phospholipid fatty acids (PLFAs) and respiratory quinones

from microorganisms that were trapped onto 0.22-μm

membranes. Analysis of respiratory quinones indicated

that the microbial populations at the site are strictly

anaerobic. The microbial biomass in the ground water

ranged from 1.6 x 102

 cells/mL to 4.2 x 104 cell/mL. The

total biomass appeared generally to correlate with the

presence of VOCs and other nutrients, and was found

to be higher in the areas containing acetone and methanol. The total biomass was orders of magnitude higher

near VOC source areas, as well as in areas with measurable concentrations of acetone and methanol, than at

downgradient or background (upgradient) locations. The

microbial biomass distribution suggested that a biologically active zone (BAZ) has developed in response to

the presence of VOCs, acetone, and methanol. The

microbial populations in the samples generally demonstrated nutritional or environmental stress, as indicated

by the ratio of specific PLFAs. Stress may be due to an

inadequate supply of nitrogen and phosphorous to support ideal growth. Cluster analysis of the PLFA data

showed that three population groups exist at the site.

The population groups appear to coincide with observed

changes in the concentration and types of VOCs and

other geochemical parameters.

Conclusions

This study provides evidence that microbial populations

can exist and function in bedrock. Furthermore, these

populations possess an intrinsic capability to anaerobically dechlorinate TCE to ethene when suitable substrates are present to support their growth. At this study

site, an active and diverse anaerobic microbial community, consisting of sulfate-reducing, methanogenic, and

acetogenic bacteria, has been established and is being

maintained by acetone and methanol. This anaerobic

microbial community is affecting the distribution and

migration of TCE, TCE biodegradation products, and

other chemicals at the site. 

References

1. Bouwer, E.J., and P.L. McCarty. 1983. Transformation of 1- and 2-carbon halogenated aliphatic organic

compounds under methanogenic conditions. Appl.

Environ. Microbiol. 45(4):1,286-1,294.

2. Parsons, F., P.R. Wood, and J. DeMarco. 1984.

Transformations of tetrachloroethylene and trichloroethylene in microcosms and ground water. J. Am.

Water. Assoc. 76:56-59.

3. Wilson, B.H., G.B. Smith, and J.J. Rees. 1986.

Biotransformations of selected alkylbenzenes and

halogenated aliphatic hydrocarbons in methanogenic aquifer material: A microcosm study. Environ.

Sci. Technol. 20(10):997-1,002.

4. Freedman, D.L., and J.M. Gossett. 1989. Biological

reductive dechlorination of tetrachloroethylene and

trichloroethylene to ethylene under methanogenic

conditions. Appl. Environ. Microbiol. 55(9):2,144-

2,151.

5. De Bruin, W.P., M.J.J. Kotterman, M.A. Posthumus,

G. Schraa, and A.J.B. Zehnder. 1992. Complete biological reductive transformation of tetrachloroethene

to ethane. Appl. Environ. Microbiol. 58(6):1,996-

2,000.

6. Major, D.M., E.W. Hodgins, and B.J. Butler. 1991.

Field and laboratory evidence of in situ biotransformation of tetrachloroethene to ethene and ethane at

a chemical transfer facility in North Toronto. In:

Hinchee, R.E., and R.F. Olfenbuttel, eds. On-site

bioremediation, pp. 147-171. Boston, MA: Butterworth-Heinemann.

143
Appendix A:

Regulatory Meeting Participants

145
U.S. Environmental Protection Agency

Office of Research and Development

Symposium on Natural Attenuation of Ground Water

Regulatory Meeting on State and Federal Issues 

Impacting Natural Attenuation

Hyatt Regency Denver

Denver, CO

August 29, 1994

Final Participant List

Michael Barden

Hydrogeologist

Wisconsin Department of Natural Resources

101 South Webster Street

P.O. Box 7921 (SW/3-ERR)

Madison, WI 53707

608-264-6007

Fax: 608-267-2768

Dolloff Bishop

Chief, Biosystems Branch

Water & Hazardous Waste Research Division

U.S. Environmental Protection Agency

26 West Martin Luther King Drive

Cincinnati, OH 45268

513-569-7629

Fax: 513-569-7105

Francis Chapelle

Hydrologist

U.S. Geological Survey

720 Gracern Road - Suite 128

Columbia, SC 29210

803-750-6116

Fax: 803-750-6181

Matthew Charsky 

Environmental Scientist

Hazardous Site Control Division

U.S. Environmental Protection Agency

401 M Street, SW (5203-G)

Washington, DC 20460

703-603-8777

Fax: 703-603-9100

Daniel Cozza 

Remedial Project Manager

U.S. Environmental Protection Agency

77 West Jackson Boulevard (HSRW-6J)

Chicago, IL 60604

312-886-7252

Fax: 312-353-5541

Marty Faile 

Environmental Engineer

Air Force Center for Environmental Excellence

U.S. Air Force

8001 Arnold Drive

Brooks AFB, TX 78235-5357

210-536-4342

Fax: 210-536-4330

Felix Flechas

Environmental Engineer

U.S. Environmental Protection Agency

999 18th Street - Suite 500 (HWM-HW)

Denver, CO 80202

303-293-1524

Fax: 303-293-1724

Rutherford Hayes

Remedial Project Manager

South Carolina Section

North Superfund Branch

U.S. Environmental Protection Agency

345 Courtland Street, NE (4WD-NSRB)

Atlanta, GA 30365

404-347-4103

Fax: 404-347-1695

146
Michael Jawson 

Chief, Subsurface Processes Branch

Robert S. Kerr Environmental Research Laboratory

U.S. Environmental Protection Agency

P.O. Box 1198

Ada, OK 74820

405-436-8560

Fax: 405-436-8703

Robin Jenkins 

Hydrogeologist

Division of Environmental Response & Remediation

Utah Department of Environmental Quality

168 North 1950, W - 1st Floor

Salt Lake City, UT 84116

801-536-4100

Fax: 801-359-8853

Fran Kremer 

Chairperson, Biosystems Technology 

Development Program

U.S. Environmental Protection Agency

26 West Martin Luther King Drive

Cincinnati, OH 45268

513-569-7346

Fax: 513-569-7585

John Kuhns

Superfund Project Manager

U.S. Environmental Protection Agency

77 West Jackson Boulevard (HSRW-6J)

Chicago, IL 60604

312-353-6556

Fax: 312-353-5541

Tim Larson 

Engineer

Bureau of Waste Cleanup

Florida Department of Environmental Protection

2600 Blair Stone Road

Tallahassee, FL 32399-2400

904-488-3935

Fax: 904-922-4368

Darryl Luce 

Remedial Project Manager

U.S. Environmental Protection Agency

JFK Federal Building (HSN-CAN5)

Boston, MA 02203

617-573-5767

Fax: 617-573-9662

Sriram Madabhushi

Hydrologist

Division of Ground Water Protection

South Carolina Department of Health & Environmental

Control

2600 Bull Street

Columbia, SC 29201

803-734-0723

Fax: 803-734-3604

Gail Mallard

Chief, Branch of Toxic Substances Hydrology

U.S. Geological Survey

12201 Sunrise Valley Drive - 412 National Center

Reston, VA 22092

703-648-6872

Fax: 703-648-5295

Robert Menzer

Laboratory Director

Gulf Breeze Environmental Research Laboratory

U.S. Environmental Protection Agency

One Sabine Island Drive

Gulf Breeze, FL 32561-5299

904-934-9208

Fax: 904-934-9201

Amy Mills

Chief, Superfund Technical Liaison Program

Office of Research & Development

U.S. Environmental Protection Agency

401 M Street, SW (8105)

Washington, DC 20460

202-260-0569

Fax: 202-260-0507

Read Miner

Hydrogeologist

South Carolina Department of Health & Environmental

Control

2600 Bull Street

Columbia, SC 29201

803-734-5335

Fax: 803-734-3604

Roland Ramsey

Soil Scientist

Petroleum Storage Tank Division

Texas Natural Resource Conservation Commission

P.O. Box 13087

Austin, TX 78711-3087

512-239-6276

Fax: 512-239-2216

147
John Rogers

Research Microbiologist

Athens Environmental Research Laboratory

U.S. Environmental Protection Agency

860 College Station Road

Athens, GA 30605-2720

706-546-3592

Fax: 706-546-3252

Alison Thomas

Research Project Engineer

Armstrong Laboratory

U.S. Air Force

139 Barnes Drive - Suite 2 (AL/EQW)

Tyndall AFB, FL 32403-5323

904-283-6303

Fax: 904-283-6286

Guy Tomassoni

Hydrogeologist

Office of Solid Waste

U.S. Environmental Protection Agency

401 M Street, SW (5303-W)

Washington, DC 20460

703-308-8622

Fax: 703-308-8617

Katrina Varner

Environmental Scientist

Analytical Science Division

Environmental Monitoring Systems Laboratory

U.S. Environmental Protection Agency

P.O. Box 93478

Las Vegas, NV 89193-3478

702-798-2645

Fax: 702-798-2692

John Wadhams

Environmental Health Scientist

Montana Department of Health

2209 Phoenix Avenue - P.O. Box 200901

Helena, MT 59620-0901

406-444-1420

Fax: 406-444-1901

Mark Walker

Industrial Hygienist

Colorado Department of 

Public Health & Environment

4300 Cherry Creek Drive, S (HMWMD-SWIM-B2)

Denver, CO 80222-1530

303-692-3449

Fax: 303-759-5355

Lisa Weers

Engineer

Colorado Department of Public Health & Environment

4300 Cherry Creek Drive, S (HMWMD)

Denver, CO 80222-1530

303-692-3451

Fax: 303-759-5355

Candida West

Research Environmental Scientist

Robert S. Kerr Environmental Research Laboratory

U.S. Environmental Protection Agency

P.O. Box 1198

Ada, OK 74820

405-436-8551

Fax: 405-436-8703

John Wilson

Senior Research Microbiologist

Robert S. Kerr Environmental Research Laboratory

U.S. Environmental Protection Agency

P.O. Box 1198

Ada, OK 74820

405-436-8532

Fax: 405-436-8703

148