Workshop on Monitoring

Oxidation-Reduction

Processes for Ground-water

Restoration

Workshop Summary

Dallas, Texas - April 25-27, 2000

United States

Environmental Protection

Agency
EPA/600/R-02/002

January 2002

Workshop on Monitoring OxidationReduction Processes for Ground-water

Restoration

Workshop Summary

Dallas, Texas

April 25-27, 2000

Edited by

Richard T. Wilkin

Ralph D. Ludwig

Robert G. Ford

Subsurface Protection and Remediation Division

National Risk Management Research Laboratory

Ada, OK 74820

National Risk Management Research Laboratory

Office of Research and Development

U.S. Environmental Protection Agency

Cincinnati, OH 45268
ii

Notice

The preparation of this report has been funded wholly by the U.S. Environmental

Protection Agency's Office of Research and Development. The views expressed in

this Workshop Summary are those of the individual authors and do not necessarily

reflect the views and policies of the U.S. Environmental Protection Agency (EPA).

This document has been reviewed in accordance with EPA’s peer and administrative

review policies and approved for publication. Mention of trade names or commercial

products does not constitute endorsement or recommendation for use.
iii

Foreword

The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation’s land, air, and

water resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement

actions leading to a compatible balance between human activities and the ability of natural systems to support and

nurture life. To meet this mandate, EPA’s research program is providing data and technical support for solving

environmental problems today and building a science knowledge base necessary to manage our ecological

resources wisely, understand how pollutants affect our health, and prevent or reduce environmental risks in the

future.

The National Risk Management Research Laboratory (NRMRL) is the Agency’s center for investigation of

technological and management approaches for preventing and reducing risks from pollution that threatens human

health and the environment. The focus of the Laboratory’s research program is on methods and their costeffectiveness for prevention and control of pollution to air, land, water, and subsurface resources; protection of

water quality in public water systems; remediation of contaminated sites, sediments and ground water; prevention

and control of indoor air pollution; and restoration of ecosystems. NRMRL collaborates with both public and

private sector partners to foster technologies that reduce the cost of compliance and to anticipate emerging

problems. NRMRL’s research provides solutions to environmental problems by: developing and promoting

technologies that protect and improve the environment; advancing scientific and engineering information to

support regulatory and policy decisions; and providing the technical support and information transfer to ensure

implementation of environmental regulations and strategies at the national, state, and community levels.

This publication has been produced as part of the Laboratory’s strategic long-term research plan. It is

published and made available by EPA’s Office of Research and Development to assist the user community and to

link researchers with their clients. Monitoring oxidation-reduction processes has direct application to performance

assessments of remedial measures and to risk management approaches that reduce risk through minimizing

exposure. Understanding the complexities associated with oxidation-reduction processes for remediating contaminated ground water is a major priority of research and technology transfer for the U.S. Environmental Protection

Agency’s Office of Research and Development and the National Risk Management Research Laboratory. This

report summarizes the findings of a workshop held to discuss and summarize the current state-of-the-science with

respect to methods of redox monitoring, data interpretation, and their applications to ground-water remediation.

Stephen G. Schmelling, Acting Director

Subsurface Protection and Remediation Division

National Risk Management Research Laboratory
iv
v

Contents Contents

Notice .............................................................................................................................................. ii

Foreword ........................................................................................................................................ iii

Introduction ................................................................................................................................... vii

Steering Committee ........................................................................................................................ ix

Workshop Participants .................................................................................................................... x

Agenda.......................................................................................................................................... xiii

Acknowledgments ......................................................................................................................... xv

Extended Abstracts.......................................................................................................................... 1

Role of Redox Processes and Spatial and Temporal Variability in Site Characterization ...... 2

Identifying the Distribution of Terminal Electron-Accepting Processes (TEAPS) in

Ground-water Systems.............................................................................................................. 5

Hydrogen Measurements and Redox Dynamics in Ground-water Systems .......................... 19

Geochemical Equilibrium and the Interpretation of EH ......................................................... 21

Abiotic Aquatic Redox Chemistry near the Oxic/Anoxic Boundary..................................... 25

Current State of Practice for Evaluation of Oxidation Reduction Processes

Important to the Biological and Chemical Destruction of Chlorinated Organic

Compounds in Ground Water................................................................................................. 29

Redox Processes in Petroleum Hydrocarbon Remediation or

Why I Never Want to Measure Eh Again .............................................................................. 35

Alteration of Reactive Mineral Surfaces by Ground Water .................................................. 39

Aqueous Redox Chemistry and the Behavior of Iron in Acid Mine Waters ......................... 43

Redox Processes in Inorganic Remediation ........................................................................... 49

The Relevance of Redox to Remediation............................................................................... 53

Group Discussion Summaries ....................................................................................................... 57

Redox Processes in Petroleum Hydrocarbon Site Characterization and Remediation .......... 59

Redox Processes for Remediation of Chlorinated Hydrocarbons and

Methyl tert-Butyl Ether .......................................................................................................... 67

Redox Processes in Inorganic Remediation ........................................................................... 79

Field Measurement of Geochemical Redox Parameters ........................................................ 89

Measurement of Dissolved Gases ........................................................................................ 101

Solid Phase Redox Characterization .................................................................................... 105

Appendix A ................................................................................................................................. 115
vi
vii

Introduction

Redox conditions are among the most important factors controlling contaminant transport and fate in ground-water

systems. Oxidation-reduction (redox) reactions, that is, reactions involving electron transfer, mediate the chemical

behavior of both inorganic and organic chemical constituents by affecting solubility, reactivity, and bioavailability.

In recent years, innovative technologies have emerged to meet ground water clean-up goals that take advantage of

the redox behavior of contaminant species. Remedial technologies that strategically manipulate or capitalize upon

subsurface redox conditions to achieve treatment may emphasize reductive processes (e.g., subsurface permeable

reactive barriers, monitored natural attenuation) or oxidative processes (e.g., air sparging). In this context, there is

an obvious need to provide stakeholders with a current survey of the scientific basis for understanding redox

behavior in subsurface systems within the framework of site characterization, selection of remedial technologies,

performance monitoring of remediation efforts, and site closure.

Many elements can exist in nature in more than one valence or oxidation state. Nine of the sixteen inorganic

species for which EPA has set specified maximum contaminant levels (MCLs) in drinking water exhibit multiple

oxidation states. These include antimony, arsenic, chromium, copper, lead, mercury, selenium, and nitrogen

(nitrate and nitrite). Consequently, the reactivity, solubility, and mobility in the environment of these and other

redox-sensitive elements depend, in part, on redox conditions. For example, uranium and chromium are insoluble

(immobile) under reducing (anaerobic) conditions. In contrast, iron and manganese are relatively insoluble under

oxidizing (aerobic) conditions but are quite soluble (mobile) under anaerobic conditions. Environmental mobility

of other potentially hazardous metals, such as cadmium, nickel, and zinc, is indirectly related to redox conditions

because these metals form ionic complexes and solid precipitates with redox-sensitive elements (e.g., sulfur).

Carbon may exist in several oxidation states, from +4 (most oxidized) to –4 (most reduced). Therefore, organic

contaminants in ground water can also be strongly influenced by redox conditions, especially through the metabolic

activity of microorganisms. For example, BTEX compounds are generally more biodegradable in oxidizing

(aerobic) ground water than in reducing (anaerobic) ground water. In contrast, chlorinated solvents such as

trichloroethylene and tetrachloroethylene are more biodegradable under more reducing conditions. Examples are

abundant in each of the major ground-water contaminant classes: inorganics, petroleum-derived hydrocarbons, and

chlorinated solvents, where redox processes have important implications in environmental risk assessment and risk

management.

Despite the importance of redox processes for understanding contaminant transport and fate, these processes

present recurring sources of confusion. Part of the confusion regarding redox processes no doubt relates to the

unfortunate fact that there is no single measurable quantity that uniquely defines the redox state of natural systems.

Unlike pH measurements, which correlate with the concentration (activity) of hydronium ions in aqueous solution,

there is no equivalent single measurable parameter that correlates well with redox speciation in aquifer systems. At

one point there was promise that Eh measurements (referenced potential of a platinum electrode) could fulfill the

equivalent for redox characterization that pH measurements serve for evaluating acid-base determinations (see, for

example, Garrels and Christ, 1965, Solutions, Minerals, and Equilibria, Harper and Row, New York). Subsequent

research has shown that the Eh-pH framework, while correct within the rigid constraints of thermodynamic

equilibrium, is likely to give an inaccurate model of the redox state of ground-water systems. This is primarily

because ground-water systems are rarely, if ever, in thermodynamic equilibrium. The modern view that has

emerged over the past several decades is that proper redox characterization involves a more complete assessment

of the concentrations of potential electron donors and electron acceptors, that is, species that can participate in

redox or electron transfer reactions. Standard electrode (Eh) measurements may still play a useful role in redox

characterization, but such measurements need to be interpreted within the context of other geochemical and

biochemical data.
viii

Monitoring oxidation-reduction processes in ground-water systems can be viewed as an integrated assessment of

hydrogeochemical processes, microbiological diversity, and aquifer characteristics. A complete assessment of

redox processes will involve developing a system inventory of the most important redox species. This may include

measurement of dissolved oxygen, nitrogen species, iron species, manganese species, sulfur species, carbon

species, and dissolved hydrogen. The fact that redox processes in ground water vary both spatially and temporally

amplifies the complexity of redox assessments. Yet it is recognized that site-specific redox modeling is often a

requirement for predicting future plume migration and for selecting the most successful and cost-effective

remediation approaches.

The main objective of the workshop, therefore, was to provide a forum for the presentation, discussion, and

synthesis of research on monitoring and interpreting ground-water redox conditions in the context of site

evaluation, remedy selection, and performance monitoring. The workshop was attended by researchers, practitioners, and regulators from governmental agencies and non-governmental institutions. Workshop participants

entered into a series of focused group discussions that explored redox processes within the organizational matrix of

contaminant class (petroleum hydrocarbons, chlorinated solvents, and inorganics) and measurement type (geochemical

parameters, dissolved gases, and solid-phase characterization). The principal findings of the workshop are reported

in this document as extended abstracts of the invited platform presentations and summary chapters of the focused

group discussions.

Richard T. Wilkin

Ralph D. Ludwig

Robert G. Ford

U.S. EPA

Office of Research and Development,

National Risk Management Research Laboratory,

Subsurface Protection and Remediation Division,

Ada, OK 74820
ix

Michael Barcelona

University of Michigan

Ann Arbor, MI 48109

Frank Chapelle

U.S. Geological Survey

Columbia, SC 29210

William Deutsch

RR-1 Box 1040

Sedgwick, ME 04676

Robert Ford

U.S. EPA, ORD/NRMRL

Ada, OK 74820

Ralph Ludwig

U.S. EPA, ORD/NRMRL

Ada, OK 74820

Carol Ptacek

Environment Canada

Burlington, ON L7R 4A6

Canada

Steering Committee Steering Committee

Robert Puls

U.S. EPA, ORD/NRMRL

Ada, OK 74820

Stephen Schmelling

U.S. EPA, ORD/NRMRL

Ada, OK 74820

John Westall

Oregon State University

Corvallis, OR 97331

Richard Wilkin

U.S. EPA, ORD/NRMRL

Ada, OK 74820

Richard Willey

U.S. EPA, Region 1

Boston, MA 02114

John Wilson

U.S. EPA, ORD/NRMRL

Ada, OK 74820
x

Jim Amonette

Pacific Northwest National Lab

Richland, WA 99353

Ann Azadpour-Keeley

U.S. EPA, ORD/NRMRL

Ada, OK 74820

Jim Barker

University of Waterloo

Waterloo, Ontario N2L 3G1

Shawn Benner

University of Waterloo

Waterloo, Ontario N2L 3G1

Paul Bradley

U.S. Geological Survey

Columbia, SC 29210

Chris Braun

U.S. Geological Survey

Austin, TX 78754

Kyle Cook

SAIC

Cincinnati, OH 45206

Isabelle Cozzarelli

U.S. Geological Survey

Reston, VA 20192

Andy Davis

Geomega, Inc.

Boulder, CO 80303

John Fruchter

Pacific Northwest National Lab

Richland, WA 99352

Peggy Groeber

SAIC

Cincinnati, OH 45206

Douglas Grosse

U.S. EPA, ORD/NRMRL

Cincinnati, OH 45268

Allison Hacker

SAIC

Cincinnati, OH 45206

Evelyn Hartzell

SAIC

Cincinnati, OH 45206

Janet Herman

University of Virginia

Charlottesville, VA 22903

James Ingle

Oregon State University

Corvallis, OR 97331

Sonya Jones

U.S. Geological Survey

Austin, TX 78754

Donald Kampbell

U.S. EPA, ORD/NRMRL

Ada, OK 74829

Faruque Kahn

U.S. EPA, ORD/NRMRL

Ada, OK 74820

Roger Lee

U.S. Geological Survey

Dallas, TX 75202

Patrick Longmire

Los Alamos National Lab

Los Alamos, NM 87545

Michelle Lorah

U.S. Geological Survey

Baltimore, MD 21237

Daryl Luce

U.S. EPA, Region 1

Boston, MA 02114

Donald Macalady

Colorado School of Mines

Golden, CO 80401

Workshop Par orkshop Participants ticipants
xi

Charlie Macon

Geomatrix Consultants, Inc.

Austin, TX 78703

Rick McGregor

302-285 Erb St. West

Waterloo, Ontario N21 1W5

Pat McLoughlin

Microseeps, Inc.

Pittsburgh, PA 15238

D. Kirk Nordstrom

U.S. Geological Survey

Boulder, CO 80303

Kirk O’Reilly

Chevron Research & Technology

Richmond, CA 94802

Keith Piontek

The Forrester Group

Chesterfield, MO 63017

Robert Pirkle

Microseeps, Inc.

Pittsburgh, PA 15238

Johnette Shockley

U.S. Army Corps of Engineers

Omaha, NE 68114

Tim Sivavec

GE Corp., R&D

Niskayuna, NY 12309

Chunming Su

ManTech Env.

Ada, OK 74820

Jim Szecsody

Pacific Northwest National Lab

Richland, WA 99352

Caroline Tebes Steven

U.S. EPA, ORD/NERL

Athens, GA 30605

Don Vroblesky

U.S. Geological Survey

Columbia, SC 29210

Scott Warner

Geomatrix Consultants, Inc.

Oakland, CA 94612

John Washington

U.S. EPA, ORD/NERL

Athens, GA 30605

Eric Weber

U.S. EPA, ORD/NERL

Athens, GA 30605

Todd Wiedemeier

Parsons Engineering Science

Denver, CO 80290

Barbara Wilson

Dynamac Corporation

Ada, OK 74820
xii
xiii

April 25: Plenary Presentations

Morning Session - Ground Water Redox Chemistry: Abiotic and Biotic Processes

 8:15 – Introductory Remarks Richard Wilkin

 8:30 – Role of Redox Processes and Spatial and Temporal

Variability in Site Characterization Michael Barcelona

 9:05 – Identifying the Distribution of Terminal Electron-Accepting

Processes (TEAPS) in Ground-Water Systems Frank Chapelle

 9:40 – Hydrogen Measurements and Redox Dynamics in

Ground-water Systems Don Vroblesky

10:15 – Break

10:30 – Geochemical Equilibrium and the Interpretation of Eh John Westall

11:05 – Abiotic Aquatic Redox Chemistry Near the Oxic/Anoxic

Boundary Don Macalady

11:40 – Discussion/Question & Answer

12:00 – Lunch

Afternoon Session - Redox Processes in Contaminant Remediation

1:15 – Current State of Practice for Evaluation of Oxidation Reduction

Processes Important to the Biological and Chemical Destruction

of Chlorinated Organic Compounds in Ground Water John Wilson

1:50 – Redox Processes in Petroleum Hydrocarbon Remediation Jim Barker

2:25 – Alteration of Reactive Mineral Surfaces by Ground Water Janet Herman

3:00 – Break

3:15 – Aqueous Redox Chemistry and the Behavior of Iron

in Acid Mine Drainage D.K. Nordstrom

3:50 – Redox Processes in Inorganic Remediation Bill Deutsch

4:25 – The Relevance of Redox to Remediation Andy Davis

5:00 – Discussion

Agenda
xiv

April 26: Breakout Sessions

Morning Sessions Morning Sessions – Redox and Remediation

Session 1. Redox Processes in Petroleum Hydrocarbon Remediation

Session 2. Redox Processes in Chlorinated Hydrocarbon and MTBE Remediation

Session 3. Redox Processes in Inorganic Remediation

Afternoon Sessions – Redox Measurement

Session 4. Measurement Issues: Geochemical Parameters

Session 5. Measurement Issues: Dissolved Gases

Session 6. Measurement Issues: Solid Phase

April 27: Summary Presentations

Session 1. Michael Barcelona, Ralph Ludwig

Session 2. Frank Chapelle

Session 3. Robert Puls

Session 4. Richard Wilkin

Session 5. Frank Chapelle

Session 6. Robert Ford
xv

Acknowledgments Acknowledgments

The success of the workshop and the completion of this document are in no

small way attributed to the steadfast determination of many individuals. The

contributions and efforts of each workshop participant and facilitator are greatly

appreciated. A special thanks goes to the authors of all the plenary session

presentations, Robert Pirkle (Microseeps, Inc.) for suggesting the idea of a Redox

Workshop, and Bruce Pivetz, Scott Warner, Kirk Nordstrom, Rick McGregor,

Keith Piontek, and Jim Amonette for providing technical reviews of the Session

Summary chapters. Douglas Grosse (EPA/ORD/NRMRL) provided direction

throughout the planning and development of the conference. Science Applications

International Corporation (SAIC) handled conference logistics and provided

support during and after the workshop. Martha Williams (CSC) provided support

in document preparation. Bill Lyon (MERSC) collected and assembled the

information presented in Appendix A.
1

Extended Extended Abstracts Abstracts

Invited speakers who gave oral presentations on the first day of the Workshop prepared the following extended

abstracts. The speakers and general topics of these presentations were selected by the steering committee. The

plenary presentations covered a range of topics relating to redox chemistry in ground water as they relate to contaminant

remediation. These presentations were intended to provide a foundation for discussion at the Workshop.
2

Site characterization represents the initial phase of

active monitoring undertaken as part of active or passive

organic contaminant remediation efforts. Initial

characterization work sets the stage to evaluate the

progress of transformations of contaminants. There have

been frequent observations of: parent compound

disappearance, active microbial populations with

biotransformation capabilities, and the appearance/

disappearance of organic and inorganic constituents

which provide indirect evidence of bioremediation at

contaminated sites. However, quantitative evidence for

sustained removal of toxic compounds from complex

mixtures solely by biological processes is lacking. This

is due largely to the reliance on monitoring well samples

for evidence of biological activity rather than

identification of the mass of contaminants (and total

reactive organic carbon), redox capacities, and

estimations of the net removal/transformation of reactive

compounds over time.

A dynamic approach to quantitative site

characterization is needed which recognizes intrinsic

bioremediation as an active cleanup approach. Careful

attention must be paid to the identification of the threedimensional distribution of contaminant mass. Then, the

correspondence between the contaminant distribution and

favorable physical, geochemical and microbial conditions

in the subsurface over time provide a basis for net

contaminant removal estimates. Mere adaptations of

detective ground-water monitoring networks are

insufficient for the quantitative evaluation of intrinsic

bioremediation technologies.

The practice of site characterization for remediation

of subsurface organic contaminants has evolved slowly

in the past decade. Early guidelines (Scalf et al., 1981;

Barcelona et al., 1985; U.S. EPA, 1986), 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 used widely

regardless of the physicochemical characteristics of

contaminant mixtures or the complexity of hydrogeologic

settings. For 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 led to

modified, monitoring well-based site characterization

approaches (Eklund, 1985). These approaches to site

characterization and monitoring network design often

suffer from the failure to identify the total mass of

contaminant in the subsurface. There are three main

reasons for this outcome.

First, although volatile organic compounds (VOC’s)

are mobile in ground water and frequently early indicators

of plume movement (Plumb, 1987), their detection in

vapor or well samples and apparent aqueous concentration

distribution does not identify the total mass distribution

of organic contaminant (Robbins, 1989). Secondly, efforts

to correlate observed soil vapor or ground water VOC

concentrations with those in subsurface solid cores have

often been unsuccessful. This is because previous bulk

jar collection/refrigeration at 4°C guidelines for core

samples for VOC analyses lead to gross negative errors

(U.S. EPA, 1992). Thirdly, “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,

particularly redox conditions.

More recent guidelines and recommendations on

network design and operations have led to more

comprehensive, cost-effective site characterization advice

(U.S. EPA, 1992; U.S. EPA, 1994). Also, excellent

reviews of characterization and long term monitoring

Role of Redox Pr Role of Redox Processes and Spatial and ocesses and Spatial and Temporal emporal Variability in Site

Characterization

Michael J. Barcelona The University of Michigan

Ann Arbor, MI 48109-2099
3

needs and approaches in support of in-situ remediation

efforts should guide us in this regard. Site characterization

efforts provide a basis for long term monitoring design

and actually continue throughout the life of a remediation

project.

A dynamic, ongoing site characterization effort

includes objectives to:

1) 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;

2) determine the corresponding spatial distribution of

total reactive organic matter (e.g., overall reduction

capacity; degradable normal, aliphatic and aromatic

hydrocarbon compounds, etc.). Overall microbial

biomass and activity in the subsurface, (and other

bioremediation indicators) are largely due to the total

mass of reactive organic carbon;

3) estimate the temporal stability of hydrogeologic and

geochemical conditions which favor microbial

transformations in background, source and

downgradient zones during the first year of

characterization and monitoring;

4) derive initial estimates of net microbial

transformations of contaminants, under known redox

conditions, which may be refined in an efficient longterm monitoring network design. Over long time

periods, redox conditions and rates of degradation

can vary substantially near source areas.

The first three objectives establish the environment

of major contamination and geochemical conditions under

which bioremediation may occur. The latter two

objectives are vitally important since the evaluation of

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 (Wilson, 1993) and was later

developed into technical protocols by Wiedemeier, et al.

(Wiedemeier et al., 1995; Wiedemeier et al., 1996) (See

characterization detail, Table 1).

The initial site characterization phase should be

designed to provide spatially dense coverage of critical

data over volumes corresponding to ten to one-hundred

year travel times along ground water flow paths. If the

flowpath intersects a discharge zone in less than 100 years,

then the volume should be scaled accordingly. For

example, if the flowpath discharges after 10 years, the

critical volume would be one year of travel time. The

“volume-averaged” values of the contaminants,

hydrogeologic and geochemical parameters within zones

along the flow path(s) should be derived from large

enough datasets to permit estimation of statistical

properties (e.g., mean, median, correlation distance,

variance, etc.). In general, this means that the datasets

for derived mass loadings of contaminants, aquifer

properties, and geochemical constituents (Table 1)

derived from spatial averages of data points must include

approximately 30 or more data points (Journel, 1986;

Hoeksema and Kitanidis, 1985; Gilbert and Simpson,

1985). Indeed, this minimum dataset size strictly applies

to points in a plane. Two major decisions which must be

made with regard to how spatially averaged masses of

contaminants, electron donors (e.g., organic carbon, Fe2+,

S=, NH3, etc.) and electron acceptors (e.g., O2, NO3

-

, NO2

- 

,

Fe, and Mn oxides, SO4

= , etc.) are to be estimated. Some

attention should be given to field screening methods for

free phase or high residual NAPL Phase. (Xie et al., 1999).

The dynamic approach to site characterization for

chlorinated hydrocarbons is more demanding in review

and data analysis than that for hydrocarbons. There are

very few published examples of site characterization

effects for these contaminants which have estimated mass

loadings in specific media.

Table 1. able 1. Site Specific Parameters to be Determined during

Site Characterization (Modified from Wiedemeier

et al., 1995)

FRACTIONATION AND SPATIAL EXTENT

OF CONTAMINATION

1. Extent and type of soil and ground water

contamination.

2. Location and extent of contaminant source

area(s).

(i.e., areas containing free- or residualphase product).

3. The potential for a continuing source due

to leaking tanks or pipelines.

HYDROGEOLOGIC AND GEOCHEMICAL

FRAMEWORK

4. Ground water geochemical parameter

distributions.

5. Regional hydrogeology including:

- Drinking water aquifers, and
4

- Regional confining units.

6. 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 distributions, (sand vs. silt vs.

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,

7. Definition of potential exposure pathways

and receptors.

References

Barcelona, M. J., Gibb, J. P., Helfrich, J. A., and Garske,

E. E. 1985. Practical Guide for Ground-Water

Sampling, Illinois State Water Survey, SWS Contract

Report 374, Ada, OK, U.S. Environmental Protection

Agency.

Eklund, B. 1985. Detection of Hydrocarbons in Ground

Water by Analysis of Shallow Soil Gas/Vapor, API

Publication No. 4394, Washington, D.C.

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

estimating spatial patterns of contaminants: Potential

and problems. Environmental Monitoring and

Assessment, v. 5, pp. 113-135.

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

spatial structure of properties of selected aquifers.

Water Resources Research, v. 21, pp. 563-572.

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

earth sciences. Mathematical Geology, v. 18, pp. 119-

140.

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

monitoring data from CERCLA and RCRA sites.

Ground Water Monitoring Review, v. 7, pp. 94-100.

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

partially penetrating wells on contaminant detection,

mapping and modeling. Ground Water, v. 27, pp.

155-162.

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

and Fryberger, J. S. 1981. Manual Of Groundwater

Sampling Procedures. National Water Well

Association, 93 p.

U.S. Environmental Protection Agency. 1986. RCRA

Technical Enforcement Guideline Document,

OSWER-9950.1, U.S. Government Printing Office,

Washington, D.C., 208 p.

U.S. Environmental Protection Agency. 1992. RCRA

Ground Water Monitoring: Draft Technical Guidance

Document, EPA/530-R-93-001, U.S. Government

Printing Office, Washington, D.C.

U.S. Environmental Protection Agency. 1994.

Proceedings of the Ground Water Sampling

Workshop, Dallas, TX, December 8-10, 1993, U.S.

Environmental Protection Agency, U.S. Government

Printing Office, Washington, D.C.

Wiedemeier, T. H., Wilson, J. T., Kampbell, D. H., Miller,

R. N., and Hansen, J. E. 1995. Technical protocol

for implementing intrinsic remediation with longterm monitoring for natural chemical attenuation for

fuel contamination dissolved in groundwater.

Prepared for the Air Force Center for Environmental

Excellence, Brooks Air Force Base, Texas.

Wiedemeier, T. H., Swanson, M. A., Moutoux, D. E.,

Gordon, E. K., Hansen, J. E., Hass, P., Wilson, J. T.,

Kampbell, D. H., and Chapelle, F. H. 1996. Draft

protocol for evaluating natural attenuation of

chlorinated solvents in groundwater. Prepared for

the Air Force Center for Environmental Excellence,

Brooks Air Force Base, Texas.

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

In In Situ Bioremediation – When Does It Work?,

Committee on In-situ Bioremediation, Water Science

and Technology Board National Research Council,

National Academy Press, Washington, D.C., pp. 160-

164.

Xie, G., Barcelona, M. J., and Fang, J. 1999.

Quantification and interpretation of TPH in sediment

samples by a GC-MS method and comparison with

EPA 418.1 and a rapid field method. Analytical

Chemistry, v. 71, pp. 1899-1904.
5

Introduction

The distribution of terminal electron-accepting

processes (TEAPs) in ground-water systems can be

assessed by documenting the availability of particular

electron acceptors (oxygen, nitrate, Fe(III), sulfate), by

showing the distribution of characteristic final products

(Fe(II), sulfide, methane), and measuring concentrations

of intermediate products (hydrogen) of microbial

metabolism. When applied to a gasoline-contaminated

site in South Carolina (Laurel Bay), this methodology

shows that redox conditions changed continuously since

the initial spill occurred in 1990. By 1994, oxygen had

been depleted in the contaminated zone, and Fe(III) and

sulfate-reduction were the predominant TEAPs.

Methanogenesis was not measurable in 1994. By 1996,

depletion of Fe(III) and sulfate resulted in the initiation

of a discrete methanogenic zone in the source area.

Between 1996 and 2000, this methanogenic zone

progressively grew until most of the plume core was

dominated by methanogenic metabolism. These dynamic

spatial and temporal changes in TEAPs, which have been

documented at several other sites as well, appear to be

characteristic of redox processes in contaminated groundwater systems. The methodology for deducing TEAPs,

which is illustrated in this paper by application to a

gasoline-contaminated aquifer, is generally applicable to

all ground-water systems.

Reduction-oxidation (redox) processes affect the

chemical composition of ground water in all aquifer

systems. In addition, ambient redox processes affect

biodegradation or biotransformation rates of

contaminants such as petroleum hydrocarbons,

chlorinated solvents, and metals. Thus, reliable methods

for characterizing redox processes are crucial to remediating contaminated ground-water systems. The

purpose of this paper is to describe a non-equilibrium,

kinetic framework for documenting the spatial and

temporal distribution of microbially mediated terminal

electron-accepting processes (TEAPs) in ground-water

systems.

The Equilibrium Approach

The traditional approach for characterizing redox

processes in ground water is based on conventions and

methods developed in classical physical chemistry (Sillen,

1952). In particular, Back and Barnes (1965) used

platinum electrode measurements to characterize the Eh

of ground-water samples. This approach was

systematized by Stumm and Morgan (1981), who

suggested that the theoretical concentration of electrons

in aqueous solution (pe), could be used in direct analogy

to hydrogen ion concentration (pH) as a “master variable”

to describe redox processes. In this treatment, the pe of a

water sample is a linear function of Eh (pe = 16.9 •Eh at

25o

C).

The definition of Eh, and thus pe, is given by the

Nernst equation (Equation 1), in which the Eh of a solution

is related to concentrations of aqueous redox couples at

chemical equilibrium and the voltage of a standard

hydrogen electrode (Eo

). For example, when

concentrations of aqueous Fe3+ and Fe2+ are at equilibrium,

Eh is defined as:

Eh E

R T

n F

aFe

aFe

o = +

+

+

2 303 3

2

. log (1)

Equation 1 illustrates an important point. Eh is only

uniquely defined when a system is at thermodynamic

equilibrium (Drever, 1982, p. 257). If the activity of Fe3+

and Fe2+ ions in Equation 1 are not at equilibrium,

measured Eh (based on electrode potential) represents

electrode response to all reactions active at the platinum

surface, but will not be an “Eh” by the definition Equation

1.

In the 1960’s and 1970’s, when ground-water

systems were thought to be largely sterile environments

devoid of microbial life, assuming equilibrium or nearequilibrium conditions seemed to be a reasonable

approach. However, in the early 1980’s, it became clear

that ground-water systems contained active, respiring,

reproducing microorganisms (Wilson et al., 1983;

Identifying the Distribution of Identifying the Distribution of Terminal Electr erminal Electron-accepting Pr on-accepting Processes (TEAPS) in

Ground-water ound-water Systems Systems

Francis H. Chapelle, James E. Landmeyer, and Paul M. Bradley U.S. Geological Survey

Columbia, SC 29210-7651
6

Chapelle et al., 1987). Furthermore, it gradually became

clear that many of the important redox processes

occurring in ground-water systems were catalyzed by

microorganisms (Baedecker et al. 1988; Chapelle and

Lovley, 1992; Baedecker et al., 1993). This realization

coincided with growing evidence that Eh measurements

of ground waters were problematic. In particular, it was

shown that Eh measurements with platinum electrodes

were usually not consistent with Eh’s calculated from

the Nernst equation using different redox couples

(Lindberg and Runnels, 1984). There are several

reasons for these problems. These include:

1. The actual concentrations of free electrons in water

are so low (~ 10-55 M) that they are essentially zero

(Thorstenson, 1984). Thus, while pe (and Eh) is a

thermodynamically definable quantity, it is not

practical to measure (as is pH).

 2. The pH electrode responds to aqueous concentrations

of hydrogen ions. However, the Eh electrode does

not respond to aqueous concentrations of electrons,

but rather to electron transfers between solutes

(Thorstenson, 1984). A platinum Eh electrode,

therefore, readily responds to concentrations of Fe2+

and Fe3+ because they react rapidly with platinum.

However, because CO2

 and CH4 react sluggishly on

a platinum surface, the Eh electrode is relatively

insensitive to the CO2/CH4 couple.

3. Microorganisms cannot actively respire and

reproduce unless there is available free energy to

drive their metabolism. That is, microorganisms

require that their immediate environment not be at

thermodynamic equilibrium. Thus, using Eh to

describe redox processes driven by microbial

processes violates the underlying equilibrium

assumption of Eh.

In light of these difficulties, it is not surprising that

Eh measurements in ground-water systems are so often

problematic.

The Kinetic Approach

Equilibrium considerations are not the only way to

describe redox processes. The metabolism of

microorganisms is based on the cycling of electrons from

electron donors (often organic carbon) to electron

acceptors such as molecular oxygen, nitrate, ferric iron,

sulfate, carbon dioxide, or other mineral electron

acceptors. This flow of electrons is capable of doing work.

Microorganisms capture this electrical energy, convert it

to chemical energy, and use it to support their life

functions. If it is assumed that redox processes in groundwater systems are driven predominantly by microbial

metabolism, it becomes possible to describe these

processes by the cycling of electron donors, electron

acceptors, and intermediate products of microbial

metabolism. Because this is an inherently nonequilibrium, kinetic description, it is termed the “kinetic

approach” (Lovley, Chapelle, and Woodward, 1994).

A kinetic description of redox processes in groundwater systems includes two components. These are

(1) documenting the source of electrons (electron donor)

that supports microbial metabolism, and (2) documenting

the final sink for electrons (electron acceptors) that

supports microbial metabolism. In many ground-water

systems, identifying electron donors is not a difficult

problem since particulate or dissolved organic carbon is

the most common source of electrons for subsurface

microorganisms. A more difficult problem is determining

the terminal electron accepting processes (TEAPs) that

occur in a system. This problem is made even more

difficult by the inherent heterogeneity of ground-water

systems. This heterogeneity causes both spatial (Chapelle

and Lovley, 1992) and temporal (Vroblesky and Chapelle,

1994) variations in TEAPs.

This paper describes methodology for deducing the

spatial and temporal distribution of TEAPs in groundwater systems. This methodology is illustrated using

long-term monitoring data of a gasoline spill site in

South Carolina.

Methodology for Deducing Deducing TEAPs

Microbial Competition

Studies in aquatic sediment microbiology have

clearly demonstrated that microbially mediated redox

processes tend to become segregated into discrete zones.

When this happens, the observed sequence of redox zones

follows a predicable pattern. At the sediment-water

interface, oxic metabolism predominates. This oxic zone

is underlain by zones dominated by nitrate reduction,

manganese reduction, and ferric iron reduction (Froelich

et al., 1979). In more organic rich marine sediments, it is

commonly observed that a sulfate-reducing zone overlies

a zone dominated by methanogenesis (Martens and

Berner, 1977).

For many years, the mechanisms causing the

observed segregation of redox zones were not clear.

However, studies with pure cultures of methanogens and

sulfate reducers (Lovley and Klug, 1983), followed by

studies with aquatic sediments (Lovley and Klug, 1986;
7

Lovley and Goodwin, 1987), showed that redox zonation

resulted from the ecology of aquatic sediments. In aquatic

sediments, organic matter oxidation is carried out by food

chains in which fermentative microorganisms initiate

biodegradation with the production of fermentation

products such as acetate and hydrogen (Figure 1). These

fermentation products are then consumed by terminal

electron-accepting microorganisms such as Fe(III)-

reducers or sulfate reducers. Because Fe(III) reduction

produces more energy per mole of acetate or hydrogen

oxidation, Fe(III) reducers are able to lower

environmental concentrations of these fermentation

products below levels required by less efficient sulfate

reducers. Thus, when Fe(III) is available, Fe(III)-reducers

can outcompete sulfate reducers for available hydrogen

(Figure 1), and sequester the majority (although not all)

of the available electron flow. This, in turn, leads to the

observed development of redox zones dominated by

particular redox processes.

The principle mechanism of microbial competition

is easily illustrated experimentally. If methanogenic

aquatic sediments are placed in anoxic vials, with some

vials amended with oxygen-free distilled water

(methanogenic treatment), some amended with a 5 mM

nitrate solution (nitrate-reducing treatment), some

amended with a suspension of amorphous ferric

hydroxide suspension (Fe(III)-reducing treatment), and

some amended with a 5 mM sulfate solution (sulfatereducing treatment), hydrogen concentrations are

observed to change over several days to different

characteristic levels (Figure 2). Methanogenesis, the least

efficient hydrogen-consuming TEAP, is characterized by

the highest hydrogen concentrations (5-15 nM) by the

end of the experiment. Sulfate reduction is characterized

by hydrogen concentrations in the 1-4 nM range, Fe(III)

reduction in the 0.2-0.8 nM range, and nitrate reduction

having hydrogen concentrations < 0.1 nM at the end of

the experiment. This leads to the tendency for discrete

redox zones to develop in sediments. This overall

principle, in turn, suggests a methodology for determining

the distribution of redox zones in ground-water systems.

Interpreting Concentrations of Final Products and

Intermediate Products

Ground-water systems are usually more complex

hydrologic systems than aquatic sediments. Part of this

complexity stems from the inherently heterogeneous

lithologic and geochemical nature of ground-water

systems (Cherry et al., 1995; Cozzarelli et al., 1999). Part

Figure 1. Figure 1. Development of predominant terminal electron-accepting processes due to microbial competition for electron donors.

HS, CO 2
8

of this complexity also stems from the predominance of

advective ground-water flow as the principal solute

transport mechanism, in contrast to molecular diffusiondominated aquatic sediments. Nevertheless, because

particular TEAPs require the presence of certain electron

acceptors (nitrate or sulfate, for example), and because

particular TEAPs produce characteristic final products

(methane from methanogenesis, sulfide from sulfate

reduction, Fe(II) from Fe(III) reduction), concentration

trends of potential electron acceptors and final products

through time and/or space can be used to deduce ongoing

TEAPs. Similarly, concentrations of intermediate

products such as hydrogen can provide insight into the

spatial and temporal distribution of TEAPs. A schematic

diagram showing the logic behind this methodology is

shown in Figure 3.

If, for example, sulfate concentrations are observed

to decrease with time or along a flowpath segment of a

ground-water system, sulfide concentrations are observed

to increase, and hydrogen concentrations are observed to

be in the range characteristic of sulfate reduction, the

occurrence of sulfate reduction can be identified in that

portion of the aquifer with a high degree of confidence

(Figure 3).

In practice, however, this methodology often

encounters uncertainties. For example, if methane,

sulfide, and Fe(II) concentrations are observed to increase

along a flowpath segment or with time, it can be

concluded that methanogenesis, sulfate reduction, and

Fe(III) reduction are ongoing (Baedecker et al., 1993).

However, because methane, sulfide, and Fe(II) are

actively transported by flowing ground water, it is often

difficult to determine the spatial location of each redox

zone. If methane is present in ground water at a particular

well location, it cannot be concluded that methanogenesis

is occurring in that well’s screened interval. Rather, it

can only be concluded that methanogenesis is occurring

somewhere upgradient of that well. In these cases,

concentrations of hydrogen, which is continuously cycled

by microorganisms and thus not subject to significant

advective transport, may more accurately identify redox

zones. However, it is always possible that wells can be

screened across different redox zones, leading to

intermediate, non-diagnostic concentrations of hydrogen.

Moreover, redox zones shift areally and temporally in

response to recharge events (Vroblesky and Chapelle,

1994), the availability of organic matter, or other

environmental factors. Because of this observed

variability, the terminology “predominant” TEAPs, rather

than “exclusive” TEAPs, is used. Under such dynamic

conditions, discrete redox zones may not be welldeveloped. Such possible complexities must always be

considered when using the methodology of Figure 3.

Redox Evolution, Laurel Bay, South Carolina

A gasoline-contaminated aquifer in Laurel Bay,

South Carolina can be used to illustrate the methodology

Time (hrs)

0 20 40 60 80 100

Hydrogen Concentration (nM)

0

2

4

6

8

10

12

14

16

18

20

22

24

Methanogenic

Sulfate-reducing

Fe(III)-reducing

Nitrate-reducing

Figure 2. Figure 2. Characteristic response of hydrogen concentrations to methanogenic, sulfate-reducing, Fe(III)-reducing, and

nitrate-reducing conditions.
9

shown in Figure 3. This site has been previously described

(Landmeyer et al., 1998). The aquifer consists of highly

permeable and relatively homogeneous marine beach

sands underlain by a clay confining bed. Gasoline leaked

into the aquifer from an underground storage tank in about

1990 (Figure 4). The tank was removed in 1992, and an

investigation of ground-water contamination initiated

soon after. Figure 4 shows a cross section of this site

along the direction of ground-water flow with the

locations of monitoring wells used to delineate redox

processes. Figures 5-9 show changes in redox conditions

along the axis of this cross section during the time period

1994-2000.

In 1994, concentrations of benzene in the

contaminant source area were in the 40 mg/L range, and

decreased downgradient (Figure 5a). Upgradient of the

contaminant source area, ground water contained high

concentrations of dissolved oxygen (Figure 5a) and sulfate

(Figure 5c), which decreased rapidly as ground water

moved into the contaminated zone. Fe(II) concentrations

were low upgradient, and increased in the source area

(Figure 5b). Concentrations of sulfide and methane were

below detection levels (Figure 5c).

Based on the concentrations of electron acceptors

(oxygen, sulfate) and metabolic products (methane,

sulfide, Fe(II)), it can be deduced that Fe(III) reduction

and sulfate reduction were ongoing in the source area in

1994, but that oxic metabolism predominated upgradient

of the source area and downgradient of well LB-8. There

was no evidence of methanogenesis in 1994.

Figure 3. Figure 3. Flowchart for deducing predominant terminal electron-accepting processes in ground-water systems.
10

Over the next several years, redox conditions

changed markedly. By 1996, the anoxic zone was

extending further downgradient (Figure 6a), and

methanogenesis and sulfate reduction had become

increasingly more prevalent processes. Concentrations

of methane increased markedly (Figure 6b), and the zone

of sulfate depletion which was confined to the source

area in 1994 (Figure 5c), extended 200 feet downgradient

in 1996 (Figure 6b). The initiation of methanogenesis in

and near the source area is indicated both by the

generation of methane (Figure 6b) and by the presence

of hydrogen concentrations in the range characteristic of

methanogenesis (Figure 6d).

This trend toward increasing importance of methanogenesis continued over time. By 1998, concentrations

of dissolved oxygen were below 1 mg/L for the entire

600 foot plume (Figure 7a), concentrations of methane

(Figure 7b) and sulfide (Figure 7c) continued to increase.

Hydrogen concentrations were in the methanogenic range

near the source area, decreased into the sulfate reducing

range between 200 and 400 feet along the flowpath, and

then decreased into the range characteristic of Fe(III)

reduction (Figure 7d).

By January, 2000, low dissolved oxygen concentrations continued to dominate the core of the plume

(Figure 8a), concentrations of methane remained

relatively high (Figure 8b), sulfate concentrations

remained relatively depleted (Figure 8c), and hydrogen

concentrations in the range characteristic of

methanogenesis extended from the source area to 400

feet downgradient.

The picture that emerges from this analysis at the

Laurel Bay site using the logic of Figure 3, is that of an

initially oxic aquifer (pre-1990), which became anoxic

in the contaminant source area soon after the gasoline

release (Figure 9). Once anoxic conditions were achieved

(1994), the predominant TEAP was Fe(III) reduction with

some evidence of sulfate reduction, but no evidence of

methanogenesis. By 1996, the source area had become

methanogenic, and this methanogenic zone gradually

expanded until it reached 400 feet downgradient in 2000.

Similar evolution of redox zones in petroleum

hydrocarbon-contaminated aquifers has been demonstrated previously (Baedecker et al., 1993; Cozzarelli

et al., 1999), and so this overall behavior is expected.

This analysis does illustrate, however, that a kinetic

Figure 4. Figure 4. Cross section showing location of contaminated zone, orientation of the contaminant plume, and locations of

monitoring wells at the Laurel Bay site.
11

Figure 5. Figure 5. Concentrations of redox-sensitive species at the Laurel Bay site, 1994.

March 1994

A

0 100 200 300 400

Dissolved Oxygen (mg/L)

0

1

2

3

4

5

6

Benzene (mg/L)

0

10

20

30

40

50

Dissolved Oxygen

Benzene

B

0 100 200 300 400

Concentration (mg/L)

0

5

10

15

20

25

Dissolved iron

Methane

C

Distance Along Flowpath (feet)

0 100 200 300 400

Concentration (mg/L)

0

2

4

6

8

10

12

14

16

18

Sulfate

Sulfide
12

A

0 100 200 300 400

0

Dissolved Oxygen (mg/L)

1

2

3

4

5

Benzene (mg/L)

0

2

4

6

8

10

12

14 Dissolved Oxygen

 Benzene 

B

0 100 200 300 400

Dissolved Iron (mg/L)

0

10

20

30

40

Methane (mg/L)

0.0

0.2

0.4

0.6

0.8

1.0

Dissolved Iron

Methane

C

0 100 200 300 400

Sulfate (mg/L)

0

10

20

30

40

50

Sulfide (mg/L)

0

5

10

15

20

25

30

35

Sulfate

Sulfide

D

Distance Along Flowpath (feet)

0 100 200 300 400

Hydrogen (nM)

0.1

1

10

Methanogenesis

Sulfate Reduction

Fe(III) Reduction

June 1996

Figure 6. Figure 6. Concentrations of redox-sensitive species at the Laurel Bay site, 1996.
13

A

0 100 200 300 400 500 600 700

Dissolved Oxygen (mg/L)

0.0

0.2

0.4

0.6

0.8

1.0

Benzene (mg/L)

0

1

2

3

Dissolved Oxygen

Benzene

B

0 100 200 300 400 500 600 700

Dissolved Iron (mg/L)

0

10

20

30

40

50

Methane (mg/L)

0.6

0.8

1.0

1.2

1.4

1.6

1.8

Iron

Methane

C

0 100 200 300 400 500 600 700

Sufate (mg/L)

0

2

4

6

8

10

12

0

Sulfide (mg/L)

1

2

Sulfate

Sulfide

D

Distance Along Flowpath (feet)

0 100 200 300 400 500 600 700

Hydrogen (nM)

0.1

1

10

Hydrogen Methanogenesis

Sulfate Reduction

Fe(III) Reduction

January, 1998

Figure 7. Figure 7. Concentrations of redox-sensitive species at the Laurel Bay site, 1998.
14

January 2000

A

0 100 200 300 400 500 600 700

Dissolved Oxygen  (mg/L)

0

1

2

3

4

5

6

7

Dissolved Oxygen

Benzene 

B

0 100 200 300 400 500 600 700

Dissolved Iron (mg/L)

0

5

10

15

20

25

Methane (mg/L)

0

1

2

Dissolved Iron

Methane

C

0 100 200 300 400 500 600 700

Dissolved Sulfate (mg/L)

0

2

4

6

8

10

Sulfide

Sulfate

D

Distance Along Flowpath (feet)

0 100 200 300 400 500 600 700

Dissolved Hydrogen (nM)

0.1

1

10 Hydrogen

Methanogenesis

Sulfate Reduction

Fe(III) Reduction

Figure 8. Figure 8. Concentrations of redox-sensitive species at the Laurel Bay site, 1996.
15

approach that considers sequential utilization of electron

acceptors, production of final products, and

concentrations of intermediate products can lead to an

accurate evaluation of redox processes in ground-water

systems.

Conclusions Conclusions

The spatial and temporal distribution of redox

conditions in ground-water systems are complex, even

at relatively simple sites. For example, the Laurel Bay

site is a relatively homogeneous aquifer, but the

development of discrete redox zones occurred gradually

over many years and involved considerable spatial

variability. Nevertheless, the data show that the overall

evolution of redox processes follows certain predictable

patterns. Available electron acceptors at this site were

consumed in the order (O2

>Fe(III)>SO4>CO2) with

methanogenesis (CO2

 reduction) occurring only after

other available electron acceptors were exhausted. This

is consistent with the model (Figure 2) of sequential

electron-acceptor utilization, and the principle of

microbial competition for available electron donors.

Figure 9. Figure 9. Schematic diagram showing the evolution of redox zones at the Laurel Bay site, 1994-2000.
16

The data clearly show, however, that the

development of these redox zones is accompanied by

considerable heterogeneity. For example, in 1996

concentrations of sulfate and sulfide were highly variable

at the interface between low-sulfate ground water adjacent

to the contaminant source area (Figure 6c), and highsulfate ground water downgradient of this interface.

Clearly, the shift between Fe(III) reduction to sulfate

reduction involved interfingering, poorly defined redox

zones. The development of these spatially complex zones

probably depends on many factors including variations

in contaminant concentrations, hydraulic conductivity,

Fe(III) content of aquifer sediments, and sulfate

concentrations.

In spite of this complexity, it is possible to document

the distribution of redox processes in ground-water

systems in both time and space using a kinetic approach

based on well-known principles of microbial ecology

(Figures 2 & 3). This kinetic approach provides a much

better resolution of ongoing microbial processes than is

afforded by traditional equilibrium approaches (Eh or pe)

to redox processes. This, in turn, makes it possible to

make more precise predictions concerning the fate and

mobility of redox-sensitive contaminants in ground-water

systems.

References

Back, W. and Barnes, I. 1965. Relation of electrochemical

potentials and iron content to ground-water flow

patterns. U.S. Geological Survey Professional Paper

498-C.

Baedecker, M. J., Siegel, D. L., Bennett, P. C., and

Cozzarelli, I. M. 1988. The fate and effects of crude

oil in a shallow aquifer:1. The distribution of chemical

species and geochemical facies. In U.S. Geological

Survey Toxic Substances Hydrology Program Review,

eds. G. E. Mallard and S. E. Ragone, U.S. Geological

Survey Water-Resources Investigations Report 88-

4220, pp. 13-20.

Baedecker, M. J., Cozzarelli, I. M., Eganhouse, R. P.,

Siegel, D. I., and Bennett, P. C. 1993. Crude oil in a

shallow sand and gravel aquifer, III. Biogeochemical

reactions and mass balance modeling in anoxic

ground water. Applied Geochemistry, v. 8, pp. 569-

586.

Chapelle, F. H., Zelibor, J. L., Grimes, D. J., and Knobel,

L. L. 1987. Bacteria in deep coastal plain sediments

of Maryland: A possible source of CO2

 to ground

water. Water Resources Research, v. 23, pp. 1625-

1632.

Chapelle, F. H., and Lovley, D. R. 1992. Competitive

exclusion of sulfate-reduction by Fe(Ill)-reducing

bacteria: A mechanism for producing discrete zones

of high-iron ground water. Ground Water, v. 30, pp.

29-36.

Chapelle, F. H., McMahon, P. B., Dubrovsky, N. M.,

Fujii, R. F., Oaksford, E. T., and Vroblesky, D. A.

1995. Deducing the distribution of terminal electronaccepting processes in hydrologically diverse groundwater systems. Water Resources Research, v. 31, pp.

359-371.

Cherry, J. A., Barker, J. F., Feenstra, S., Gillham, R. W.,

Mackay, D. M., and Smyth, D. J. A. 1995. The Borden

site for groundwater contamination experiments:

1978-1995. In Groundwater and Subsurface

Remediation: Research Strategies for In-situ

Remediation, eds. H. Kolbus, B. Barczewski, and H.

P. Kosschitzky, Springer-Verlag, Berlin, pp.102-127.

Cozzarelli, I. M., Herman, J. S., Baedecker, M. J., and

Fischer, J. M. 1999. Geochemical heterogeneity of a

gasoline-contaminated aquifer. Journal of

Contaminant Hydrology, v. 40, pp. 261-284.

Drever, J. I. 1982. The Geochemistry of Natural Waters.

Prentice Hall, Englewood Cliffs, New Jersey, 388 p.

Froelich, P. N., Klinkhammer, G. P., Bender, M. L.,

Luedtke, N. A., Heath, G. R., Cullen, D., and

Dauphin, P. 1979. Early oxidation of organic matter

in pelagic sediments of the eastern equatorial Atlantic:

suboxic diagenesis. Geochimica et Cosmochimica

Acta, v. 43, pp. 1075-1090.

Landmeyer, J. E., Chapelle, F. H., Bradley, P. M.,

Pankow, J. F., Church, C. D., and Tratnyek, P. G. 1998.

Fate of MTBE relative to benzene in a gasolinecontaminated aquifer (1993-98). Ground Water

Monitoring and Remediation, v. 18, pp. 93-102.

Lindberg, R. D. and Runnells, D. D. 1984. Ground-water

redox reactions: An analysis of equilibrium state

applied to Eh measurements and geochemical

modeling. Science, v. 225, pp. 925-927.

Lovley, D. R. and Klug, M. J. 1983. Sulfate reducers can

outcompete methanogens at fresh water sulfate

concentrations. Applied and Environmental

Microbiology, v. 45, pp. 187-192.

Lovley, D. R., and M. J. Klug. 1986. Model for the

distribution of methane production and sulfate

reduction in freshwater sediments. Geochimica et

Cosmochimica Acta, v. 50, pp. 11-18.
17

Lovley, D. R. and Goodwin, S. 1988. Hydrogen

concentrations as an indicator of the predominant

terminal electron-accepting reactions in aquatic

sediments. Geochimica et Cosmochimica Acta, v. 52,

pp. 2993-3003.
18
19

The use of hydrogen, a transient intermediate

product of anaerobic microbial metabolism, as an

indicator of predominant terminal electron accepting

processes (TEAPs) was introduced by Lovley and

Goodwin (1988). The standard method used for

measuring H2 in ground water is a gas-stripping procedure

that has previously been referred to as the “bubble strip”

method (Vroblesky and Chapelle, 1994; Chapelle and

McMahon, 1991; Chapelle et al., 1995). In this method,

a standard 250 ml gas sampling bulb (Supelco) is

continuously purged with ground water at a rate of about

500 mL/min. A 20-mL bubble of H2

-free N2 gas is

introduced into the sampling bulb through the septum.

As the bubble is vigorously agitated by the stream of

inflowing water, slightly soluble gases such as H2 are

stripped from the water to the gas phase over time, and

asymptotically come into equilibrium with the water

flowing through the bulb. After equilibrium is achieved

(~20-30 min), a gas sample is withdrawn from the bulb

by using glass syringes with stopcocks and analyzed for

H2

. Analysis is done using gas chromatography with

reduction gas detection (Trace Analytical, Menlo Park,

California). Other methods that have been used to collect

H2 from ground water include a down-hole sampler, and

a diffusion sampler (Chapelle et al., 1997).

Measurement of H2

 using the gas-stripping method

can be affected by a variety of factors. A field investigation

showed that one factor affecting the H2 concentration is

the type of pumping equipment. Four different methods

for pumping ground water were evaluated. These were

(1) a peristaltic pump (Geotech Environmental

Equipment, Inc., Denver, Colorado), (2) a stainless steel

piston pump (Bennet Sample Pumps, Inc., Amarillo,

Texas), (3) a bladder pump (Well Wizard Bladder Pump,

QED Environmental Systems, Ann Arbor, Michigan), and

(4) a stainless steel submersible pump (MP-1, Grundfos

Pumps, Corp., Clovis, California). The peristaltic pump

draws water under negative pressure, the piston and

bladder pumps push water under positive pressure, and

the Grundfos pump pushes water under positive pressure

by impellers driven by a Direct Current (DC) electrical

motor.

The various pumps found that peristaltic, piston,

and bladder pumps all gave similar results when applied

to water produced from the same well (Chapelle et al.,

1997). It was observed, however, that peristaltic-pumped

water (which draws water under a negative pressure)

enhanced the gas-stripping process and equilibrated

slightly faster (~20 min) than either piston (~25 min) or

bladder pumps (~30 min), which push water under a

positive pressure (Figure 1A).

The direct current-driven submersible pump gave

much higher H2

 concentrations than the other pumps.

Because H2

 can be electrolytically produced by direct

current, it is probable the high H2 concentrations observed

using this submersible pump reflect the interaction of

direct current with water. Thus, direct current submersible

pumps appear to not be suitable for measuring H2 in

ground water (Chapelle et al., 1997).

When the peristaltic pumping rate was increased

from 500 to 1,000 mL/min, equilibration was achieved

in proportionally less time (Figure 1B). These

observations indicate that equilibration times, and thus

the accuracy and reproducibility of the gas-stripping

method, can vary depending on pumping methods and

rates. These results indicate that specific combinations

of pumping systems and rates should be individually

tested to evaluate the length of time needed for

gas-aqueous phase equilibration.

Measurements from two field sites indicate that iron

or steel well casings produce H2

 ,which masks H2

concentrations in ground water. PVC-cased wells, or

wells cased with other materials that do not produce H2,

are necessary for measuring H2

 concentrations in ground

water (Chapelle et al., 1997).

Measurements of the dissolved H2

 and other

biologically active solutes in ground water from a shallow

Hydrogen Measur ogen Measurements and Redox Dynamics in Gr ements and Redox Dynamics in Ground-water ound-water Systems Systems

Don A. Vroblesky and Francis H. Chapelle U.S. Geological Survey

Columbia, South Carolina 29210-7651
20

petroleum hydrocarbon-contaminated aquifer indicate

that the distribution of TEAPs is highly dynamic in time

and space (Vroblesky and Chapelle, 1994). During times

when little or no sulfate or oxygen is introduced to the

ground water by recharge from rainfall or alternate

sources, the available oxygen, Fe(III), and sulfate can be

depleted by respiring bacteria, leaving methanogenesis

as the predominant TEAP. Introduction of oxygen from

rainfall infiltration into the methanogenic parts of the

aquifer can cause precipitation of Fe(III) as grain coatings.

Once the oxygen is depleted, Fe(III)-reducing bacteria

can take advantage of the precipitated Fe(III) to sequester

most of the electron flow from degradation of organic

compounds. Introduction of sulfate into a methanogenic

part of the aquifer can cause the predominant TEAP to

shift to sulfate reduction. The shift from methanogenesis

to sulfate reduction was observed to range from <10 days

to about 31⁄2 months, depending on the sulfate and organic

carbon concentrations.

References

 Chapelle, F. H., Vroblesky, D. A., Woodward, J. C., and

Lovely, D. R. 1997. Practical considerations for

measuring hydrogen concentrations in groundwater.

Environmental Science and Technology, v. 31,

pp. 2873-2877.

Figure 1. Figure 1. (A) Comparison of equlibration time of the gas-stripping procedure using the bladder, piston, and peristaltic

pumps, and (B) the influence of flow rate on equilibration times for the peristaltic pump (modified from Chapelle

et al., 1997).

Chapelle, F. H., McMahon, P. B., Dubrovsky, N. M.,

Fujii, R. F., Oaksford, E. T., and Vroblesky, D. A.

1995. Deducing the distribution of terminal electronaccepting processes in hydrologically diverse

ground-water systems. Water Resources Research,

v. 31, pp. 359-371.

Chapelle, F. H. and McMahon, P. B. 1991. Geochemistry

of dissolved inorganic carbon in a coastal plain

aquifer: Sulfate from confining beds as an oxidant

in microbial CO2

 production. Journal of Hydrology,

v. 127, pp. 85-108.

Lovley, D. R. and Goodwin, S. 1988. Hydrogen

concentrations as an indicator of the predominant

terminal electron-accepting reactions in aquatic

sediments. Geochimica et Cosmochimica Acta, v. 52,

pp. 2993-3003.

Vroblesky, D. A. and Chapelle, F. H. 1994. Temporal

and spatial changes of terminal-accepting processes

in a petroleum hydrocarbon-contaminated aquifer and

the significance for contaminant biodegradation.

Water Resources Research, v. 30, pp. 1561-1570.
21

Among the most important factors controlling the

persistence, mobility, and biological effects of many

organic and inorganic pollutants are oxidation-reduction

(redox) transformations. For example, it is known that

dehalogenation of aromatic compounds occurs most

readily under anoxic conditions, while ring cleavage and

conversion to CO2

 occurs most readily under oxic

conditions. In design of in-situ remediation, control of

microbial metabolism through the proper balance between

carbon substrate and the terminal electron acceptor is

critical in obtaining the desired transformation of an

organic contaminant. In general, degradation rates and

pathways of organic compounds are known to depend

strongly on ambient redox conditions. Among inorganic

chemicals, speciation, and hence transport and biological

effects, depends strongly on redox reactions. For

example, Cr(VI) is generally mobile and toxic, while

Cr(III) is insoluble, immobile, and relatively nontoxic.

Thus, a clear understanding of the redox state and

speciation of chemicals is required in virtually all aspects

of hazardous materials management: evaluation of

disposal options, risk assessments for contaminated sites,

evaluation of clean-up options, and performance

assessment. Thus, it is useful to recast the issue of

“Geochemical Equilibrium and the Interpretation of EH”

in terms of the broader underlying questions, (i) what is

the redox state of an environment?, and (ii) does a

particular contaminant undergo a redox transformation

under these conditions?

The Problem

All too often the problem of characterizing the redox

state of an environment is approached in the following

manner: (i) measure the apparent redox potential with a

Pt electrode; (ii) calculate an equilibrium distribution of

the pollutant from the measured redox potential and pH;

(iii) remark in disgust that the calculated distribution does

not match the observed distribution, that the Pt electrode

does not really work for this application, and that the

system is probably not at equilibrium anyway.

The pitfalls of redox potential measurements in the

environment have been well documented (Morris and

Stumm, 1967; Whitfield, 1974; Lindberg and Runnells,

1984): low redox buffer capacity of the aqueous phase,

slow kinetics among the environmental redox reactions,

and slow kinetics of the electrode itself. Still, Pt

electrodes continue to be used, and in spite of these

problems, appear to give useful data under some

circumstances, as will be discussed.

Alternatives to the Pt-electrode include (i) direct

determination of concentrations of all significant

redox-active species in water samples; (ii) redox titrations

of aquifer materials to determine an effective differential

or integral “redox-capacity;” and (iii) active control of

redox conditions in laboratory “redox-stats.”

Stumm (1984) advances the concept of determining

the concentrations of major redox active species as a

means of characterizing the redox status of an

environment. Frevert (1984) used commercially available

metal and O2

 electrodes and a feedback system to attempt

to control the redox status of a system. Although he

considered slow kinetics of redox reactions in his study,

he concluded that it is impractical to characterize the

redox status of a sediment-water system in terms of a

single parameter. Related work on redox capacities was

carried out by Whitfield (1972), who investigated the

voltammetric properties of sediment-water systems.

In summary, if one chooses to advance the argument

that it is impossible to characterize the redox status of a

natural system, there is certainly evidence to support the

argument. On the other hand, if one recognizes that redox

transformations are extremely important in transport of

many elements and in degradation of many organic

compounds, that some redox reactions are coupled, and

that additional information on characterizing the redox

status of the environment would be helpful, then there is

evidence that much insight could be gained from further

experimentation.

Geochemical Equilibrium and the Interpr Geochemical Equilibrium and the Interpretation of EH

John C. Westall Oregon State University

Corvallis OR 97331
22

Approach to Characterizing the Redox Status oach to Characterizing the Redox Status

 Our general conceptual approach to

characterization of redox transformations of contaminants

in the subsurface environment involves interactions along

the three legs of a triangle, at the apices of which are: (i)

the geochemical matrix, (ii) the redox sensor, and (iii)

the redox-active contaminant. Coupling of redox

processes must be understood at two different linkages:

contaminant and matrix, and sensor and matrix. The

contaminant-matrix linkage must be investigated on a

species-by-species basis, without the a priori assumption

of equilibrium.

The Geochemical Matrix

The “geochemical matrix” drives the redox transformations. One way to approach redox conditions in

aquifers is to consider major oxidant and reductant

species. In some regards, these species can be considered

“redox poising” species.

O2

 - In the presence of significant concentrations of O2,

experimental EH values (Pt electrode vs SHE) are

relatively high (300 - 400 mV).

MnO2

(s)/Mn(II) - Mn(III) and Mn(IV) oxide/hydroxide

minerals are present in some aquifers; equilibrium

potentials of this couple are relatively high, about

500-650 mV at pH 7, although appearance of Mn(II)

is observed at lower potentials (200 to 300 mV).

Fe(OH)3

(s)/Fe(II) - Fe(III) oxide/hydroxide minerals are

found in many near surface aquifers. Iron is about

ten times more abundant than manganese, and is

therefore important in poising redox potential. Fe(III)

is a considerably weaker oxidant than either O2

 or

MnO2(s). The equilibrium redox potential is about

100-200 mV at pH 7.

SO4

2-/S0

/S2- - The sulfur system is the most important

inorganic system poising redox potentials in reducing

conditions (EH -100 to -300 mV).

Organic compounds are the ultimate reductants in

natural systems and have a major role in defining redox

conditions.

The Redox Sensor

To answer the underlying question, “will the

contaminant undergo a redox transformation under the

prevailing environmental conditions,” the ideal sensor

would obviously be based on the sensor itself. For

example, if the contaminant could be either

(i) immobilized on a suitable electrochemical or

spectrochemical transducer, or (ii) injected, withdrawn,

and chemically detected in a push-pull test, then one

would have directly the information on the extent of

conversion. As an alternative to the contaminant itself,

redox indicators have been immobilized to form such

redox sensors (Lemmon et al., 1996).

As surrogates for this direct information one can

determine the concentration of chemical species that

reflect the redox status of the environment. In

environments with significant amounts of organic carbon

and microbial activity, the redox status is probably best

defined by the terminal electron acceptor for degradation

of organic carbon (e.g., O2, Mn(IV), NO3

-

, Fe(III), SO4

2-,

CO2

, H2 , etc.). Microbial degradation of organic

compounds is generally most closely linked to this

measure of redox status. Over the last several years

Lovley, Chapelle and coworkers have advanced the idea

that H2

(g) is the best indicator of the metabolic state of a

subsurface microbial system (Lovley et al., 1988, 1994;

Chapelle et al., 1996); hence H2

(g) could be the best

indicator of the likelihood that a contaminant will undergo

a redox transformation, particularly under reducing

conditions. In environments in which transport processes

are rapid compared to redox processes, the presence,

absence, or concentration of a particular species may not

alone be sufficient to indicate the primary terminal

electron acceptor and the redox status of the environment.

Therefore, rates of redox processes relative to rates of

transport processes are important ancillary information.

From a practical point of view, the availability of

facilities, personnel, and time for chemical analysis are

important. If a method is impractical, it will not be used,

even though it might give the best answer. If a simple

method (e.g., Pt electrode) gives the same information

as a more complex method (e.g. complete chemical

analysis), is there any need for the complete chemical

analysis? Associated with the use of the Pt electrode are

an array of questions that ultimately determine its utility:

What sorts of species poise the Pt electrode at each of the

major terminal electron acceptor steps in a biologically

active medium (e.g., O2

, Mn(IV), NO3

-

, Fe(III), SO4

2-,

CO2, H2 etc.)? What is the exchange current density, i.e.,

how well is the electrode poised (Bard and Faulkner,

1980)? Does it vary dramatically with redox level? What

role is played by “mediators”, i.e., soluble molecules that

shuttle between particle surfaces and the electrode

surface? Does the performance depend strongly on

whether or not the electrode is in direct contact with the

particles?

Contrary to what one may be led to believe by a

superficial review of the literature, the Pt electrode can

be of value in determining environmental redox
23

conditions. Data from experiments of Patrick and coworkers (e.g., Masscheleyn et al., 1990, 1992) show that

redox transformations of inorganic species, and

appearance and disappearance of terminal electron

accepting species track very well Pt electrode potentials.

This discussion is not to advance the argument that the

Pt electrode even approaches an ideal indicator for

environmental redox reactions, but just to say that its

performance warrants serious consideration.

A complicating factor in almost all of the methods

is that the majority of redox buffer capacity in the

subsurface environment is associated with biotic, biogenic

or mineral particles, not with the aqueous solution. Then

one of the major issues in understanding redox

transformations of environmental contaminants is the

distinction between transformations that occur directly

at particles (biotic or abiotic) vs. transformations that

occur in solution, presumably via “mediator molecules”

that conduct between the particle surfaces and the

contaminant being transformed. This issue is important

not only in characterizing transformations of

contaminants, but also in how sensors work. Particlefree media are much more amenable to chemical analysis

such as spectrophotometry, cyclic voltammetry, etc.

A further complication is the apparent redox

“disequilibria” observed in many subsurface

environments. These conditions arise when transport

kinetics are rapid compared to reaction kinetics.

Associated with this observation is the question, how do

each of the various “redox sensor technologies” described

above track these changes?

Summary

In assessing the redox status of the subsurface

environment, it is important to pose the right question.

Previous work has indicated that the question, “what is

the EH or p∈ of a particular environment?” is not optimal;

a better approach is to pose the question, “will a certain

redox transformation take place under a certain set of

conditions, and how can this set of conditions best be

recognized?” Judging performance of redox sensors on

this criterion may lead to the best decision.

References

Bard, A. J. and Faulkner, L. R. 1980. Electrochemical

Methods. Wiley, New York, pp. 391 398.

Chapelle, F. H, Hack, S. K., Adriens, P., Henry, M. A.,

and Bradley, P. M. 1996. Comparison of EH and H2

Measurements for delineating redox processes in a

contaminated aquifer. Environmental Science and

Technology, v. 30, pp. 3565-3569.

Frevert, T. 1984. Can the redox conditions in natural

waters be predicted by a single parameter?

Schweizerische Zeitschrift Fur Hydrologie, v. 46,

pp. 269 290.

Lemmon, T. L., Westall, J. C., and Ingle, J. D., Jr. 1996.

Development of redox sensors for environmental

applications based on immobilized redox indicators.

Analytical Chemistry, v. 68, pp. 947-953.

Lindberg, R. D. and Runnells, D. D. 1984. Groundwater

redox reactions: An analysis of equilibrium state

applied to Eh measurements and geochemical

modelling. Science, v. 225, pp. 925-927.

Lovley, D. R. and Goodwin, S. 1988. Hydrogen

concentrations as an indicator of the predominant

terminal electron-accepting reactions in aquatic

sediments. Geochimica et Cosmochimica Acta, v. 52,

pp. 2993-3003.

Lovley, D. R., Chapelle, F. H., and Woodward, J. C. 1994.

Use of dissolved H2

 concentrations to determine

distribution of microbially catalyzed redox reactions

in anoxic groundwater. Environmental Science and

Technology, v. 28, pp.1205-1210.

Masscheleyn, P. H., DeLaune, R. D., and Patrick, W. H., Jr.

1990. Transformations of selenium as affected by

sediment oxidation-reduction potential and pH.

Environmental Science and Technology, v. 24, pp. 91-

96.

Masscheleyn, P. H., Pardue, J. H., DeLaune, R. D., and

Patrick, W. H., Jr. 1992. Chromium redox chemistry

in a lower Mississippi bottomland hardwood wetland.

Environmental Science and Technology, v. 26,

pp. 1217-1226.

Morris, J. C. and W. Stumm. 1967. Redox equilibria and

measurements of potentials in the aquatic

environment. In Equilibrium Concepts in Natural

Water Systems, ed. W. Stumm, Advances in

Chemistry Series No. 67, American Chemical

Society, Washington, D.C., pp. 270-285.

Stumm, W. 1984. Interpretation and measurement of

redox intensity in natural waters. Schweizerische

Zeitschrift Fur Hydrologie, v. 46, pp. 291-296.

Whitfield, M. 1972. The electrochemical characteristics

of natural redox cells. Limnology and Oceanography,

v. 17, pp. 383-393.

Whitfield, M. 1974. Thermodynamic limitations on the

use of the platinum electrode in Eh measurements.

Limnology and Oceanography, v. 19, pp. 857-865.
24
25

Considerations of ground water oxidation/reduction

chemistry are often predicated on chemical measurements

designed to determine the “redox state” of the system

under investigation. There has been a long-standing

tendency to seek one or more universal parameters that

will serve as uniformly applicable indicators of the redoxstate of any ground-water system. The desired outcome

is a set of measured characteristics that enable prediction

of redox processes. There is abundant literature to

document that such a universal mechanism for prediction

of redox processes is not forthcoming (see for example

Lyman et al., 1987 and references therein).

For any measured set of activities of the constituent

chemicals of a redox couple, there is a corresponding

Nernstian electrochemical potential. The disturbing fact

relative to the chemistry of natural waters is that parallel

determinations of electrochemical potentials calculated

from different redox couples in the same water sample

often imply distinctly different redox conditions (White

et al., 1990; Walton-Day et al., 1990). These and other

studies imply that the lack of a definable “redox state” of

an aquatic system, is a consequence of the lack of

thermodynamic equilibrium among the components of

the system.

This condition is particularly prevalent in a class of

systems that is especially important in considerations of

contaminant transformations in ground water. By this is

meant those ground-water systems that are substantially

depleted in dissolved oxygen (DO). The most commonly

used methods of analyses for DO have detection limits

greater than 2 μmolar, a concentration much greater than

the detection limits for most other redox-active species

in natural and contaminated ground waters.

Several examples from the literature will confirm

that many ground waters in the regime between truly

anoxic and 2 μmolar DO are characterized by pronounced

spatial and/or temporal disequilibria (Peiffer et al., 1999;

Walton-Day et al., 1990; Walton-Day, 1991; Eary and

Schramke, 1990). In these systems, both electrode

measurements of Eh and determinations of specific redoxactive species such as ferrous iron are likely to

erroneously imply extremely low oxygen partial

pressures.

A seldom-used colorimetric method is described that

uses a proprietary redox-sensitive dye and allows accurate

detection of DO levels as low as 0.2 μmolar. Use of this

technique illustrates the importance of this low oxygen

tension regime as far as comprehensive understandings

of ground-water redox conditions are concerned.

Detectable DO in waters containing measurable ferrous

iron and sulfide, for example, is not uncommon. Eh

measurements in such systems typically imply oxygen

partial pressures 10-20 orders of magnitude lower than

measured values (White et al., 1990).

In other words, redox disequilibrium is perhaps most

commonplace in the types of ground-water systems most

relevant to considerations of contaminant transformation.

Furthermore, the causes of such disequilibria can

commonly be the presence or absence of precisely those

chemical or biological processes that are most important

in considerations of the rates and mechanisms of

contaminant transformations.

Many of these disequilibria are “intentionally”

maintained by the selective mediation of electron transfer

processes by microbial systems. Thus, sulfate/sulfide

ratios in systems containing active sulfate reducing

bacterial populations may represent Nernstian potentials

vastly different from those calculated from “abiotic”

couples present in the same system. There is little

evidence for any universal electron transfer mediator that

rapidly reacts to assure redox equilibria among the many

redox active species in a natural or contaminated groundwater system.

Perhaps the closest approach to this notion of a

“universal” electron transfer mediator is the complex

mixture of substances commonly called natural organic

matter (NOM). NOM in both the dissolved and particulate

phases of ground-water systems is capable of mediating

Abiotic Abiotic Aquatic Redox Chemistr Aquatic Redox Chemistry near the Oxic/Anoxic Boundar the Oxic/Anoxic Boundary

Donald L. Macalady Colorado School of Mines

Golden, CO 80401
26

a wide range of electron transfer processes. NOM in this

context refers to the suite of extracellular organic

chemicals present in any natural water system. The

components of NOM can vary from molecules recently

excreted by active microorganisms to facilitate electron

transfer process of utility to the microbial community to

“abiotic” degradation products from the debris of long

inactive living systems. The ability to mediate electron

transfer processes seems to be ubiquitous among NOM

samples (Macalady and Ranville, 1998; Tratnyek and

Macalady, 2000).

NOM has been shown to be an effective mediator

of electron transfer in a wide variety of redox processes.

Included are the reduction of nitroaromatic compounds,

azo compounds, halogenated organic chemicals, metals

and metal/organic complexes (see for example, Macalady

and Ranville, 1998; Larson and Weber, 1994; Curtis and

Reinhard, 1994; Wolfe et al., 1986; Wittbrodt and Palmer,

1996; Skogerbee, 1981; Matthiessen, 1996). NOM can

also serve as an electron transfer mediator in microbial

processes, e.g. the microbial reduction of solid iron (III)

oxyhydroxides (Lovley et al., 1996). The role of NOM

in mediating the transfer of electrons in the opposite

direction has received less attention, but the oxidation of

sulfide and ferrous iron is clearly enhanced in the presence

of oxidized NOM. In fact, one of the few redox processes

that is apparently not mediated by NOM is the reduction

of molecular oxygen (Peiffer, unpublished data).

The mechanism(s) by which NOM facilitates

electron transfer processes is only partially understood.

Clearly quinone-like functional groups within NOM

structures are an important part of this reactivity.

However, in certain pH ranges, other functional groups

may be important (Dunnivant et al., 1992; Perlinger et

al., 1996; Schwarzenbach et al., 1990; Gantzer and

Wackett 1991; Schindler et al., 1976).

Attempts to determine the role of NOM in specific

ground-water redox processes are related to both the

solution phase and particulate NOM fractions in the

aquifer matrix. The potential roles of NOM include direct

participation as an electron transfer mediator and an

indirect role as a transport inhibitor (particulate NOM)

or facilitator (dissolved and/or colloidal NOM). Specific

examples of such processes serve to illustrate the

importance of NOM in such considerations.

References

Curtis, G. P. and Reinhard, M., 1994. Reductive

dehalogenation of hexachlorethane, carbon

tetrachloride, and bromoform by anthrahydroquinone

disulfonate and humic acid. Environmental Science

and Technology, v. 28, pp. 2393-2401.

Dunnivant, F. M., Schwarzenbach, R. P. and

Macalady, D. L. 1992. Reduction of substituted

nitrobenzenes in aqueous solutions containing natural

organic matter. Environmental Science and

Technology, v. 26, pp. 2133-2141.

Eary, L. E. and Schramke, J. A. 1990. Rates of inorganic

oxidation reactions involving dissolved oxygen. In

Chemical Modeling of Aqueous Systems, II, Chapter

30, eds. D. C. Melchior and R. L. Bassett, American

Chemical Society, Washington, D.C., pp. 379-396.

Gantzer, C.J. and Wackett, L.P., 1991. Reductive

dechlorination catalyzed by bacterial transition-metal

coenzymes. Environmental Science and Technology,

v. 25, pp. 715-722.

Larson, R. A. and Weber, E. J. 1994. Reaction

Mechanisms in Environmental Organic Chemistry.

Lewis Publishers, Chelsea, Michigan, pp. 169-273.

Lovley, D. R., Coates, J. D., Blunt Harris, E. L.,

Phillips, E. J. P. and Woodward, J. C. 1996. Humic

substances as electron acceptors for microbial

respiration. Nature, v. 382, pp. 445-448.

Lyman, W. J., Bodek, I., Reehl, W. F., Rosenblatt, D. H.

1987. Electron transfer reactions. In Methods for

Estimating Physicochemical Properties of Inorganic

Chemicals of Environmental Concern, Final Report,

Chapter 2, U.S. Army Medical Research and

Development Command, Contract DAMD

17-83-C-3274.

Macalady, D. L. and Ranville, J. F. 1998. The chemistry

and geochemistry of natural organic matter. In

Perspectives in Environmental Chemistry, Chapter

5, ed. D. L. Macalady, Oxford University Press,

New York, pp. 94-137.

Matthiessen, A. 1996. Kinetic aspects of the reduction of

mercury ions by humic substances. Fresenius

Journal of Analytical Chemistry, v. 354, pp. 747-749.

Perlinger, J. A., Angst, W., and Schwarzenbach, R. P. 1996.

Kinetics of the reduction of hexachloroethane by

juglone in solutions containing hydrogen sulfide.

Environmental Science and Technology, v. 30,

pp. 3408-3417.

Peiffer, S., Walton-Day, K., Macalady, D. L., 1999. The

interaction of natural organic matter with iron in a

wetland receiving acid mine drainage. Aquatic

Geochemistry, v. 5, pp. 207-223.
27

Schindler, J. E., Williams, D. J., Zimmerman, A. P. 1976.

Investigation of extracellular electron transport by

humic acids. In Environmental Biogeochemistry, Vol.

1., ed. J. O. Nriagu, Ann Arbor Science, Ann Arbor,

Michigan, pp. 109-115.

Schwarzenbach, R. P., Stierli, R., Lanz, K. and Zeyer, J.

1990. Quinone and iron porphyrin mediated

reduction of nitroaromatic compounds in

homogeneous aqueous solution. Environmental

Science and Technology, v. 24, pp. 1566-1574.

Skogerboe, R. K. 1981. Reduction of ionic species by

fulvic acid. Analytical Chemistry, v. 53, pp. 228-232.

Tratnyek, P. G. and Macalady, D. L. 2001. Oxidationreduction reactions in aquatic systems. In Estimation

of Chemical Properties for the Environmental and

Health Sciences: A Handbook of Methods,

eds. B. Boethling and D. Mackay, Ann Arbor Press,

Ann Arbor, Michigan.

Walton-Day, K., Macalady, D. L., Brooks, M. H., Tate,

V. T. 1990. Field methods for measurement of ground

water redox chemical parameters. Ground Water

Monitoring Review, v. 10, pp. 81-89.

Walton-Day, K. 1991. Hydrology and Geochemistry of a

Natural Wetland Affected by Acid Mine Drainage,

St. Kevin Gulch, Lake County, Colorado. Ph. D.

Thesis, Colorado School of Mines, Golden, CO,

300 p.

White, A. F., Peterson, M. L., Solbau, R. D. 1990.

Measurement and interpretation of low levels of

dissolved oxygen in ground water. Ground Water,

v. 28, pp. 584-589.

Wittbrodt, P. R. and Palmer, C. D. 1996. Effect of

temperature, ionic strength, background electrolytes,

and Fe(III) on the reduction of hexavalent chromium

by soil humic substances. Environmental Science and

Technology, v. 30, pp. 2470-2477.

Wolfe, N. L., Kitchens, B. E., Macalady, D. L., and

Grundl, T. J. 1986. Physical and chemical factors

that influence the anaerobic degradation of methyl

parathion in sediment systems. Environmental

Toxicology and Chemistry, v. 5, pp. 1019-1026.
28
29

The most important oxidation reduction process

involved in the destruction of chlorinated organic

compounds in ground water is biological sequential

reductive dechlorination (Vogel and McCarty, 1985). In

some cases this process does not provide benefit to the

organisms that carry it out, the process is entirely

accidental, and may be considered a form of cometabolism. In other cases the biological process yields

energy to the microorganisms and can support their

growth and proliferation (Maymo-Gatell et al., 1995).

In this circumstance the process functions as a respiration,

and has been termed halorespiration (See discussion in

Chaper 6, Wiedemeier et al., 1999). Chlorinated organic

compounds may also be destroyed by chemical reaction

in aquifers, usually involving direct chemical reaction

with sulfide or ferrous iron. Examine Butler and Hayes

(1999), and Devlin and Muller (1999) for illustrations of

recent research.

In “The Technical Protocol for Evaluating Natural

Attenuation of Chlorinated Solvents in Ground Water”,

the U.S. Environmental Protection Agency used a scoring

system to identify sites where geochemical conditions

were appropriate or not appropriate for biological

reductive dechlorination of chlorinated solvents. Weights

were given to the concentrations of important electron

acceptors, including oxygen, nitrate, and sulfate, to the

concentrations of important electron donors including

molecular hydrogen, volatile fatty acids, dissolved native

organic carbon (TOC), and petroleum derived

monoaromatic hydrocarbons (BTEX), and the

concentrations of metabolic end products including

methane, ferrous iron, sulfide, chloride and carbon

dioxide represented as increases in carbon dioxide or

alkalinity. Weights were also given to general descriptions

of the oxidation/reduction environment such as electrode

potential, pH, and temperature. Finally, weight was given

to the accumulation of metabolic daughter products

including dichloroethylene, dichloroethane, vinyl

chloride, ethene, or ethane. Weight was only given when

the daughter products were not originally present in the

material that was released to the environment. The

weighted scores were totaled, and compared to a table

that interpreted the scores as providing either inadequate

evidence, limited evidence, adequate evidence, or strong

evidence of anaerobic biodegradation of chlorinated

organic compounds.

The National Research Council (2000) in the report

Natural Attenuation for Groundwater Remediation noted

on pages 210 and 211 that

Unfortunately, this scoring system is being

widely adopted for uses that the authors never

intended. For example, many states are using it

to evaluate natural attenuation for all types of

chlorinated solvents. Tables of natural

attenuation scores are showing up in remedial

investigation reports at Superfund sites. Maps

and cross sections showing natural attenuation

scores are being included in final reports as a

key line of evidence. Some regulators are

accepting this inappropriate use of scoring.

* The method applies only to chlorinated

ethenes.

* The scores emphasize reducing

environments more than dehalogenation

reactions.

* A reduced geochemical environment does

not guarantee that natural attenuation will

occur, because geochemical environments

can be very reduced without reductive

dehalogenation of chlorocarbons occurring

(for example, if dehalogenating bacteria are

not present).

* The scoring system included items that are

of current research interest (for example,

hydrogen concentration), but that may have

limited practical impact on making

remediation decisions.

Current State of Practice for Evaluation of Oxidation Reduction Processes Important to the

Biological and Chemical Destruction of Chlorinated Organic Compounds in Ground Water

John T. Wilson U.S. Environmental Protection Agency

Ada, OK 74820
30

 Table 1. able 1. Analytical Parameters and Weighting for Preliminary Screening for Anaerobic Biodegradation processes used in

the Technical Protocol for Evaluating Natural Attenuation of Chlorinated Solvents in Ground Water, U.S. EPA,

1998

Analysis

Concentration in

Most Contaminated

Zone Interpretation Value

Oxygen* <0.5 mg/L Tolerated, suppresses the reductive pathway at higher

concentrations

3

Oxygen* >5 mg/L Not tolerated: however, VC may be oxidized aerobically -3

Nitrate* <1 mg/L At higher concentrations may compete with reductive pathway 2

Ferrous iron >1 mg/L Reductive pathway possible; VC may be oxidized under Fe(III)-

reducing conditions

3

Sulfate* <20 mg/L At higher concentrations may compete with reductive pathway 2

Sulfide* >1 mg/L Reductive pathway possible 3

Methane* <0.5 mg/L

>0.5 mg/L

VC oxidizes

Ultimate reductive daughter product, VC Accumulates

0

3

Oxidation Reduction

Potential* (ORP)

against Ag/AgCl

electrode

<50 millivolts (mV)

<-100mV

Reductive pathway possible

Reductive pathway likely

1

2

pH* 5 < pH < 9

5 > pH >9

Optimal range for reductive pathway

Outside optimal range for reductive pathway

0

-2

TOC > 20 mg/L Carbon and energy source; drives dechlorination; can be natural or

anthropogenic

2

Temperature* > 20o

C At T >20oC biochemical process is accelerated 1

Carbon Dioxide >2x background Ultimate oxidative daughter product 1

Alkalinity >2x background Results from interaction between CO2 and aquifer minerals 1

Chloride* >2x background Daughter product of organic chlorine 2

Hydrogen >1 nM Reductive pathway possible, VC may accumulate 3

Hydrogen <1 nM VC oxidized 0

Volatile Fatty Acids > 0.1 mg/L Intermediates resulting from biodegradation of aromatic compounds;

carbon and energy source

2

BTEX* > 0.1 mg/L Carbon and energy source; drives dechlorination 2

Tetrachloroethene Material Released 0

Trichloroethene* Material released

Daughter product of PCE

0

2a/

DCE* Material released

Daughter product of TCE.

If cis is > 80% of total DCE it is likely a daughter product

1,1-DCE can be chemical reaction product of TCA

0

2a/

VC* Material released

Daughter product of DCE

0

2a/

1,1,1-Trichloroethane* Material released 0

DCA Daughter product of TCA under reducing conditions 2

Carbon Tetrachloride Material Released 0

Chloroethane* Daughter product of DCA or VC under reducing conditions 2

Ethene/Ethane >0.01mg/L

>0.1 mg/L

Daughter product of VC/ethene 2

3

Chloroform Material Released

Daughter Product of Carbon Tetrachloride

0

2

Dichloromethane Material Released

Daughter Product of Chloroform

0

2

* Required analysis. a/ Points awarded only if it can be shown that the compound is a daughter product (i.e., not a constituent of the source NAPL).
31

* The system identifies interactions between

contaminants only for electron donors.

The National Research Council (2000)

recommended on page 15 of their report that national

consensus guidelines be created for evaluation of natural

attenuation, and that

* The national consensus guidelines and

future protocols should eliminate the use of

“scoring systems” for making decisions on

natural attenuation. The evaluation

methods outlined in Chapter 4 [of their

report], using conceptual and footprints of

natural attenuation, should replace scoring

systems.

Nyer et al. (1998) cautioned against the use of the

scoring system as a primary method to substantiate natural

attenuation, because the system failed to recognize natural

attenuation at a site where subsequent sampling and

evaluation revealed that natural attenuation was believed

to be occurring. In other words, the scoring system

produced false negatives.

To improve the site-screening methodology for

evaluating sites where natural reductive dechlorination

may be selected as a remedial option, Stiber et al. (1999)

constructed a causative model for the reductive

dechlorination process, then solicited expert knowledge

from twenty-two experts on reductive dechlorination in

ground water. He solicited expert opinion on fourteen

types of evidence in “The Technical Protocol for

Evaluating Natural Attenuation of Chlorinated Solvents

in Ground Water” (EPA, 1998). The experts were asked

to provide their estimates of the probability that reductive

dechlorination was occurring at a site based on site

conditions for each type of evidence. The expert opinions

were scaled by comparing the difference in the logarithm

of the odds ratios where

∆LOR = log [prior probability (true | evidence)/ posterior

probability (false | evidence)] - log [posterior

probability (true | evidence)/ prior probability

(false | evidence)]

The experts put the greatest weight and greatest trust

on data showing the accumulation of metabolic

Table 2. Values for the change in the log of the odds ratio ( DLOR) for the probability that anareobic degradation is occurring

by reductive dechlorination associated with findings from screening data with respect to different types of evidence.

The more positive the change with positive findings, the greater the weight the experts put in that type of evidence.

The more negative the change with negative findings, the greater the weight the experts put in that type of evidence.

These data are extrapolated from the Average Model in Figure 4 of Stiber et al., 1999.

Type of Evidence Positive Findings Negative Findings

Terminal Electron Accepting Process 0.07 -0.15

Hydrogen 0.10 -0.15

Oxidation Reduction Potential 0.07 -0.15

Dissolved Organic Carbon 0.04 -0.10

BTEX 0.02 -0.10

TPH 0.02 -0.05

Oxygen 0.03 -0.15

Temperature 0.01 -0.01

pH 0.01 -0.10

Dichloroethene 0.63 0.80

Vinyl Chloride 0.95 0.40

Ethene and Ethane 0.78 0.25

Methane 0.43 0.38

Chloride 0.50 0.30
32

transformation products, and methane and chloride (Table

2). They put much less weight on data concerning electron

acceptors, hydrogen, oxidation reduction potential,

dissolved oxygen, TPH, and BTEX. They put little weight

on data concerning pH and essentially no weight on data

concerning temperature. Stiber et al. (1999) noted that

“Simply identifying that prerequisite conditions are

adequate is not sufficient; experts want to see products

to have confidence that reductive dechlorination is

occurring.” They also noted that measuring the

concentration of molecular hydrogen was a particularly

divisive issue. “Some experts believe that, if performed

correctly, hydrogen measurements provide the best means

to assess the reducing conditions. Others feel that there

are too many variables in microbial communities for

hydrogen to be a reliable measurement.”

Measurement of hydrogen in ground water is a

particularly powerful technique to understand electron

flow in microbial communities (Lovley et al., 1994).

However, measurement of hydrogen requires careful

sampling as well as state-of-art laboratory analytical

measurements (Chapelle et al., 1997). At this point the

use of hydrogen to understand reductive dechlorination

has had mixed results. It is not clear if this mixed

performance is a result of inadequate field protocols for

collection and analysis of hydrogen, or is a result of real

variations in the concentrations of hydrogen from one

time to another in the same plume, or from one plume to

another.

Recent work has shown that reductive

dechlorination may proceed in ground water at relatively

low concentrations. Smatlak et al. (1996) observed that

the physiological threshold for the concentration of

hydrogen that allowed growth of dechlorinating

microorganisms in their laboratory cultures was less than

2 nM. Yang and MacCarty (1998) compared the presence

or absence of cis-dichloroethylene on the equilibrium

concentration of hydrogen in batch microcosms fed with

benzoate as a fermentable source of hydrogen. In the

presence of cis-dichloroethylene, reductive dechlorination

to ethene dominated electron flow, and the concentration

of hydrogen was poised at 2.2 +/- 0.9 nM. In the cultures

that did not have cis-dichloroethylene, the cultures were

methanogenic and the equilibrium concentration of

hydrogen was 10.9 +/- 3.3 nM. The dechlorinating

microorganisms out-competed methanogenic microorganisms, and poised the concentration of hydrogen gas

in the ground water.

Jakobsen et al. (1998) showed from theoretical

calculations and field data from Denmark, that the

equilibrium concentration of hydrogen in cold ground

waters (near 10o C) should be approximately half of the

equilibrium concentration of hydrogen expected in

warmer ground waters and laboratory studies (near 20o C

to 25o C). A concentration hydrogen of at least 1.0 nM

would be required to sustain dechlorinating

microorganisms in colder ground waters.

One of the most problematic issues in evaluation of

natural attenuation is the tendency of reductive

dechlorination to “stall” in many plumes at the level of

dichloroethylene, presumably because they have depleted

their sources of reducing power. In these plumes the

dichloroethylene persists without further metabolism to

vinyl chloride, ethene, or ethane. Concentrations of

hydrogen significantly below 1.0 nM may indicate that

further reductive dechlorination should not be expected.

In other plumes dichloroethylene and vinyl chloride

do not accumulate because these compounds are directly

oxidized by aerobic microorganisms (Davis and

Carpenter, 1990; Bradley and Chapelle, 2000; Hartmans

and de Bont, 1992; Klier et al., 1998), or iron reducing

bacteria ( Bradley and Chapelle, 1996, 1997; Bradley et

al., 1997) or manganese reducing bacteria (Bradley et

al., 1998), or by bacteria using native humic material as

the electron acceptor (Bradley et al., 1998). A

concentration of hydrogen above 1.0 nM may indicate

that reductive dechlorination is on-going in these plumes,

and the reduction transformation products are not

accumulating because they are destroyed by the oxidative

processes described above. Analysis for hydrogen could

be used to distinguish between active plumes and inactive

plumes that have low concentrations of vinyl chloride,

ethene, and ethane because reductive dechlorination is

limited and these compounds were not produced in the

first place.

Suggestions for Consideration by the Consideration by the Work

Groups

The use of geochemical parameters to determine

the “footprint” of a plume can be useful in determining

whether an apparent attenuation in the field is real, or

whether it is an artifact of an inadequate network of

monitoring wells. The primary use of information on

the distribution of oxidation/reduction parameters should

shift. Instead of using the oxidation/reduction parameters

to identify plumes or regions of plumes where conditions
33

for natural attenuation are favorable, the parameters

should be used for establishing “footprints” of plumes.

The most direct limitation on microbial activity is

the supply of the requisite substrates for metabolism.

However, the parameters that address the requisites for

reductive dechlorination (such as hydrogen, BTEX,

volatile fatty acids, and dissolved native organic matter)

have little weight in the evaluation of natural attenuation

by subject matter experts. Data use and data quality

objectives should be developed for the geochemical

parameters that measure the requisites for metabolism

of contaminants in aquifers. Field protocols should be

developed for sampling, preservation, and analysis of

these parameters.

The most convincing evidence of natural attenuation

is direct evidence of transformation of the contaminants.

The “Technical Protocol for Evaluating Natural

Attenuation of Chlorinated Solvents in Ground Water”

(U.S. EPA, 1998) relied on the accumulation of

transformation products as evidence of transformation.

Biodegradation of organic compounds may also result in

isotopic fractionation of the contaminants remaining.

Natural attenuation due to biodegradation could be

distinguished from simple dilution or dispersion because

the isotopic ratio of the contaminant remaining in the

plume would become progressively heavier as

biodegradation proceeded. Data on stable isotope ratios

could be used as a second line of evidence to support an

evaluation of natural attenuation (see Hunkeler et al., 1999

for a recent application of this approach to a PCE spill in

Canada). Data use and data quality objectives should be

developed for the use of stable isotope ratios to document

contaminant biodegradation in ground water.

References

Bradley, P. M. and Chapelle, F. H. 1996. Anaerobic

mineralization of vinyl chloride in Fe(III)-reducing

aquifer sediments. Environmental Science and

Technology, v. 30, pp. 2084-2092.

Bradley, P. M. and Chapelle, F. H. 1997. Kinetics of DCE

and VC mineralization under methanogenic and

Fe(III)-reducing conditions. Environmental Science

and Technology, v. 31, pp. 2692-2696.

Bradley, P. M. and Chapelle, F. H. 2000. Aerobic microbial

mineralization of dichloroethene as sole carbon

substrate. Environmental Science and Technology,

v. 34, pp. 221-223.

Bradley, P. M., Chapelle, F. H, and Lovley, D. R. 1998.

Humic acids as electron acceptors for anaerobic

microbial oxidation of vinyl chloride and

dichlorethene. Applied and Environmental

Microbiology, v. 64, pp. 3102-3105.

Bradley, P. M., Chapelle, F. H., and Wilson, J. T. 1997.

Field and laboratory evidence of intrinsic

biodegradation of vinyl chloride contamination in a

Fe(III)-reducing aquifer. Journal of Contaminant

Hydrology, v. 31, pp. 111-127.

Bradley, P. M., Landmeyer, J.E., and Dinicola, R. S.

1998. Anaerobic oxidation of [1,2-

14C]dichloroethene under Mn(IV)-reducing

conditions. Applied and Environmental Microbiology,

v. 64, pp. 1560-1562.

Butler, E.C. and Hayes, K. F. 1999. Kinetics of the

transformation of trichloroethylene and

tetrachloroethylene by iron sulfide. Environmental

Science and Technology, v. 33, pp. 2021-2027.

Butler, E.C. and Hayes, K. F. 2000. Kinetics of the

transformation of halogenated aliphatic compounds

by iron sulfide. Environmental Science and

Technology, v. 33, pp. 422-425.

Chapelle, F. H., Vroblesky, D. A., Woodward, J. C., and

Lovely, D. R. 1997. Practical considerations for

measuring hydrogen concentrations in groundwater.

Environmental Science and Technology, v. 31,

pp. 2873-2877.

Davis, J. W. and Carpenter, C. L. 1990. Aerobic

biodegradation of vinyl chloride in ground water

samples. Applied and Environmental Microbiology,

v. 56, pp. 3878-3880.

Devlin, J. F., and Muller, D. 1999. Field and laboratory

studies of carbon tetrachloride transformation in a

sandy aquifer under sulfate reducing conditions.

Environmental Science and Technology, v. 33,

pp. 1021-1027.

Hartmans, S., and de Bont, J. A. M. 1992. Aerobic vinyl

chloride metabolism in mycobacterium aurum Li.

Applied and Environmental Microbiology, v. 58,

pp. 1220-1226.

Jakobsen, R., Albrechtsen, H-J., Rasmussen, M., Bay, H.,

Berg, P. L., and Christensen, T. H. 1998. H2

concentrations in a landfill leachate plume

(Grindsted, Denmark): In situ energetics of terminal

electron acceptor processes. Environmental Science

and Technology, v. 32, pp. 2142-2148.

Klier, N. J., West, R. J., and Donberg, P. A. 1998. Aerobic

biodegradation of dichloroethylenes in surface and

subsurface soils. Chemosphere, v. 38, pp. 1175-1188.
34

Lovley, D. R., Chapelle, F. H., and Woodward, J. C. 1994.

Use of dissolved H2

 concentrations to determine

distribution of microbially catalyzed redox reactions

in anoxic groundwater. Environmental Science and

Technology, v. 28, pp.1205-1210.

Maymo-Gatell, X., Tandoi, V., Gossett, J. M., and

Zinder, S. H. 1995. Characterization of an H2

-utilizing

enrichment culture that reductively dechlorinates

tetrachloroethylene to vinyl chloride and ethene in

the absence of methanogenesis and acetogenesis.

Applied and Environmental Microbiology, v. 61,

pp. 3928-3933.

National Research Council. 2000. Natural Attenuation

for Groundwater Remediation, 219 p.

Nyer, E., Mayfield, P., and Hughes, J. 1998. Beyond the

AFCEE protocol for natural attenuation. Ground

Water Monitoring Review, Summer, pp. 70-77.

Smatlak, C. R., Gossett, J. M., and Zinder, S. H. 1996.

Comparative kinetics of hydrogen utilization for

reductive dechlorination of tetrachloroethene and

methanogenesis in an anaerobic enrichment culture.

Environmental Science and Technology, v. 30,

pp. 2850-2858.

Stiber, N. A., Pantazidou, M., and Small, M. J. 1999.

Expert system methodology for evaluating reductive

dechlorination at TCE sites. Environmental Science

and Technology, v. 33, pp. 3012-3020.

U.S. Environmental Protection Agency. 1998. The

Technical Protocol for Evaluating Natural

Attenuation of Chlorinated Solvents in Ground Water,

EPA/600/R-98/128, U.S. Government Printing

Office, Washington, D.C.

Vogel, T. M., and McCarty, P. L. 1985. Biotransformation

of tetrachloroethylene to trichloroethylene,

dichloroethylene, vinyl chloride, and carbon dioxide

under methanogenic conditions. Applied and

Environmental Microbiology, v. 49, pp. 1080-1083.

Wiedemeier, T. H., Rafai, H. S., Newell, C. J., and

Wilson, J. T. 1999. Natural Attenuation of Fuels and

Chlorinated Solvents in the Subsurface,

John Wiley & Sons, New York, 617 p.

Yang, Y. and McCarty, P. L. 1998. Competition for

hydrogen within a chlorinated solvent dehalogenating

anaerobic mixed culture. Environmental Science and

Technology, v. 32, pp. 3591-3587.
35

The Context for Redox Processes in Monitored

Natural Natural Attenuation of Petroleum

Hydrocarbons ocarbons

The study of transport and fate of petroleum

hydrocarbons in ground water is “mature” and has

evolved into a remedial technology – monitored natural

attenuation (MNA). For example, the US EPA OSWER

Directive 9200.4-17P (April, 1999) states: “Natural

attenuation processes, particularly degradation, are

currently best documented at petroleum fuel spill sites.

Under appropriate field conditions, the regulated

compounds benzene, toluene, ethylbenzene, and xylene

(BTEX) may naturally degrade through microbial activity

and ultimately produce non-toxic end products (e.g.,

carbon dioxide and water).” The natural BTEX

degradation is “through microbial activity”, so microbial

activity is the key process. Microbial activity must be

considered in the context of the field conditions.

Classification of subsurface microbial activity is

commonly via the dominant terminal electron accepting

process (TEAP). Thus we can identify sulfate reduction

as being the dominant TEAP and can discuss BTEX

biodegradation by a community in which sulfate is the

dominant electron acceptor (EA). This is the link between

petroleum hydrocarbon MNA and redox processes.

Geochemically, a TEAP is considered a redox process.

The OSWER Directive requires monitoring

programs that, among other things, demonstrate that

natural attenuation is occurring and that deleterious

“changes in environmental conditions” are not. Collection

of site-specific data is required “to estimate with an

acceptable level of confidence both the rate of attenuation

processes and the anticipated time required to achieve

remediation objectives.” The current distribution of

contaminants must be adequately defined and it must be

sufficiently clear, from field monitoring data, that

potential receptors are protected. A three-tiered approach

is recommended, each providing successively more

detailed information. These “three lines of evidence” are:

1. historical data that demonstrate loss of contaminant

mass and/or concentrations over time,

2. hydrogeologic and geochemical data to demonstrate

indirectly the type(s) of NA processes active at the

site and their rate(s),

3. data from field or microcosm studies which directly

demonstrate the occurrence of a particular NA process

at the site.

For selecting MNA for petroleum hydrocarbons,

demonstrating an understanding of how the current

contamination was produced and how it will likely evolve

is useful, if not essential. While contaminant distribution

is the primary evidence, the need for a credible story

should not be dismissed. It both relies upon and

demonstrates the scientific underpinning for MNA.

Not being microbial ecologists, hydrogeologists and

geochemists have had to simplify the evaluation of

microbial activity. Since microbial ecologists seemed to

be willing to discuss biodegradation in terms of TEAPs,

that compromise has been accepted and subsurface

environments and geochemical processes have been

grouped in terms of TEAPs. Under current guidance (e.g.,

Wiedemeier et al., 1999), we are directed to examine the

apparent utilization of EAs and to define the distribution

of dominant TEAPs in plumes. That has led us to redox

processes.

Thus, redox processes and measurements are not

of primary importance in MNA for petroleum

hydrocarbons. They are only important as they affect the

supporting “story”.

Challenges for MNA for Petroleum

Hydrocarbons ocarbons

Site assessment Site assessment

The critical need at research and “real” sites is

insight into what has occurred and what is occurring in

the subsurface. A noted contaminant hydrogeologist

characterised MNA, especially for chlorinated solvents

as “1990’s microbiology; 1970’s hydrogeology”.

Redox Pr Redox Processes in Petr ocesses in Petroleum Hydr oleum Hydrocarbon Remediation or ocarbon Remediation or Why I Never Why I Never Want to

Measure Eh Again

J. F. Barker University of Waterloo

Waterloo, ON, Canada N2L3G1
36

Chapelle (1999) states: “The most important lesson that

can be gleaned from the history of petroleum hydrocarbon

bioremediation is the critical importance of accurately

assessing the hydrogeology of each site.” It is clear to

hydrogeologists that the critical weakness in the

application of MNA lies within their discipline.

Quantitative EA-Organic Mass Balances Don’ Quantitative EA-Organic Mass Balances Don’t

Add Up

If MNA is really mature, shouldn’t we expect

quantitative balances between EAs consumed and

petroleum hydrocarbons biodegraded, at least at “research

sites”? A controlled, in place source experiment at CFB

Borden has been recently reported (King and Barker,

1999). About 70 kg of coal tar creosote (CTC) was

emplaced below the water table and the plumes of

dissolved chemicals that developed were monitored with

six “snapshot” samplings over four years, each using 450

to 2800 samples from multilevel monitoring wells. The

analysis of background/plume redox-sensitive species

indicated that the major TEAPs were aerobic respiration

and sulfate reduction. Mass balance calculations

suggested the apparent consumption of EAs would

support biodegradation of about 300 g of CTC organics

in the plumes. However, more than 1200 g of CTC

organics was apparently biodegraded in the plume.

Usually, it is expected that more EAs will appear to

be utilized than can be balanced by organic contaminant

loss. This is because of the presence of other, non-target

organics that exert a demand for EAs. In the CTC case,

the likely problem is the incomplete transformation of

organics. The above mass balance assumed complete

mineralization. Significant concentrations (up to 18 mg/L)

of aromatic acids, the probable metabolites, were found

in a very superficial sampling of nine points. In this, the

biodegradation reactions would need to be re-written

based on actual transformations; not a quantitative process

at present.

While depletion of electron acceptors and donors

and the appearance of reaction products can be

qualitatively supported at most sites, acceptable balance

between EAs and organics consumed is unlikely to be

demonstrated, especially when “real world” monitoring

(5 – 40 points, typically) is the basis. That is not a very

secure scientific underpinning for MNA. Some credible

quantitative mass balances need to be produced.

Transport of EA Utilization Indicators Utilization Indicators

Delineation of redox environments or TEAPs

appears useful/essential in developing the story

supporting MNA of petroleum hydrocarbons.

Albrechtsen et al. (1999) captured the problem very well:

“Delineations of redox environments have been based

mainly on the concentrations of redox sensitive

components in ground water. However, since most of the

reactive components may be transported by ground water,

for example, methane and Fe(II), their presence in a

sample does not necessarily reflect redox conditions at

the sampling location but perhaps conditions at an

upgradient location.”

The promise of H2

Hydrogen appears to this desperate researcher as

potentially the “ideal” indicator of the TEAP active at

the monitored location. While H2 can be transported, it is

sensitive to local microbial activity and so its

concentration is adjusted to reflect local conditions

(Lovley et al., 1994). As such, it overcomes the transport

problem of the other geochemical indicators. I sense a

healthy reluctance to accept this hypothesis and look

forward to further demonstration such as reported by

Albrechtsen et al. (1999). The key is to compare H2

concentrations directly to microbially-significant aspects

of TEAP and hydrocarbon degradation, not to flawed

(transported, reactive) aqueous geochemical data.

Research Needs ch Needs

MNA for petroleum hydrocarbons is underpinned

by a demonstrated understanding of the controlling

processes. The key is the transport and fate of

contaminants, but EAs and TEAPs (perhaps defined as

redox processes) are involved.

One aspect is to determine the distribution, in space

and perhaps in time, of TEAPs. The researchers from the

Groundwater Research Centre at the Technical University

of Denmark (TUD) have provided the most

comprehensive characterization of ground-water redox

environments, not in BTEX plumes, but in landfill

leachate plumes. They have used essentially all the

techniques, many of which are referenced by Albrechtsen

et al. (1999).

To demonstrate understanding of interactions of

inorganics/organics/minerals/microbial communities

insightful geochemical interpretation must be employed.

Since transport is significant, I think we are admitting

that the processes might be more complex than can be

handled by “typical” site monitoring, supported by simple

spreadsheet models. We need at least a few detailed field

studies in which all important parameters are measured

and interpretation is quantitative. If the problem is as
37

complex as I have suggested, then qualitative

demonstration of process understanding will require

sophisticated flow-transport-reaction models.

Using reactive transport modeling, certain key redox

features can be reproduced, but it is clear that matching

all features, especially pH in poorly buffered systems,

strains our conceptual understanding. At actual field sites

major hydrogeological uncertainties, unknown source

functions, and temporal variability will be added. If we

are to demonstrate a comprehensive understanding of the

redox processes involved in biodegradation of petroleum

hydrocarbons, we must do so through matching results

from such models with detailed field observations from

at least a few sites.

Do we need this for MNA of petroleum

hydrocarbons? We needed it five years ago to provide

the underpinning for current guidance. Do we still need

it? MNA is being selected, errors will be made, and the

blame shouldn’t reside with bad scientific underpinning.

So, I think this research is still essential.

Is it being done? The modeling tools appear to be

sufficiently developed (e.g., BIORXNTRN, Hunter et al.,

1998; Mayer, 1999) and some models are being employed

in this area (e.g., Curtis et al., http://wwwmn.cr.usgs.gov/

bemidji/). Certainly, TUD researchers have the field

studies and experience for landfill leachate plumes. The

USGS is apparently well along at the Bemidji site

(http://wwwmn.cr.usgs.gov/bemidji) and other

hydrocarbon-contaminated sites (e.g., Wurtsmith) have

major components of this in place. All of these sites

(perhaps less so at Bemidji) suffer from uncertainties

about the source and/or complex hydrogeology. Too bad

the controlled field experiments in simple

hydrogeological settings (e.g., Barker et al., 1987; King

and Barker, 1999) were not more comprehensive. With

simple hydrogeology, simple and known sources, mass

balances, etc. would have been a reasonable expectation

and the concept would have been quantitatively evaluated.

The Role of Redox Processes in Practical MNA

for Petroleum Hydrocarbons ocarbons

Is proper redox characterization required to

implement MNA for petroleum hydrocarbons?

No. The primary information is the distribution of

contaminants. Redox characterization only contributes

to the credibility of the story of how the plume got to the

current status.

Is proper redox characterization required for remedy

selection? Also, no. I might reject MNA for BTEX plumes

in naturally anaerobic aquifers, since I don’t feel benzene

degrades reliably under such conditions. However, I

would acknowledge others would not share this

pessimism and so my redox-based view would not be

compelling. We would, I am sure, need to rely on the

primary line of evidence - the distribution of BTEX and

interpretation of apparent attenuation.

Is proper redox characterization required for

performance assessment? Again, no. While, for example,

assessment of deleterious “changes in environmental

conditions” might be tracked through redox

measurements, direct contaminant measurements will

likely be the assessment criteria. Perhaps redox

contributes to the credibility of the assessment. For

example, if benzene is found beyond the predicted point,

perhaps noting the unexpected extent of reduced

conditions would be useful.

Redox processes are part of the supporting story

for MNA of petroleum hydrocarbons. As such, it is

probably more important that research-level studies

demonstrate understanding of redox-related processes.

Our goal should be to demonstrate this. We should then

assure hydrogeologists that they must rely primarily on

hydrogeological studies, not assessments of redox

processes, in applying MNA to BTEX-contaminated

groundwater.

References

Albrechtsen, H. J., Bjerg, P. L., Ludvigsen, L., Rugge, K.

and Christensen, T. H.. 1999. An anaerobic field

injection experiment in a landfill leachate plume,

Grindsted, Denmark 2. Deduction of anaerobic

(methanogenic, sulfate-, and Fe(III)-reducing) redox

conditions. Water Resources Research, v. 35,

pp.1247-1256.

Barker, J. F., Patrick, G. C., and Major, D. 1987. Natural

attenuation of aromatic hydrocarbons in a shallow

sand aquifer. Ground Water Monitoring Review, v. 7,

pp. 64-71.

Chapelle, F. H. 1999. Bioremediation of petroleum

hydrocarbon-contaminated ground water: The

perspectives of history and hydrology. Ground Water,

v. 37, pp. 122-132.

Curtis, G. P., Cozzarelli, I. M., Baedecker, M. J., and

Bekins, B. A. Coupled biogeochemical modeling of

ground water contamination at the Bemidji Minnesota

crude oil spill site, from http://wwwmn.cr.usgs.gov/

bemidji/.
38

Hunter, K. S., Wang, Y., and Van Cappellen, P. 1998.

Kinetic modeling of microbially-driven chemistry of

subsurface environments: coupling transport,

microbial metabolism and geochemistry. Journal of

Hydrology, v. 209, pp. 53-80.

King, M. W. G., Barker, J. F., Devlin, J. F., and B. J. Butler.

1999. Migration and natural fate of a coal tar creosote

plume; 2. Mass balance and biodegradation

indicators. Journal of Contaminant Hydrology, v. 39,

pp. 281-307.

Lovley, D. R., Chapelle, F. H., and Woodward, J. C. 1994.

Use of dissolved H2

 concentrations to determine

distribution of microbially catalyzed redox reactions

in anoxic groundwater. Environmental Science and

Technology, v. 28, pp. 1205-1210.

Mayer, K. U. 1999. A numerical model for

multicomponent reactive transport in variably

saturated porous media, Ph.D. thesis, University of

Waterloo, Waterloo, Ontario, Canada.

Wiedemeier, T. H., Rafai, H. S., Newell, C. J., and Wilson,

J. T. 1999. Natural Attenuation of Fuels and

Chlorinated Solvents in the Subsurface, John Wiley

& Sons, New York, 617 p.
39

Two different investigations of shallow sandy

aquifers inform our thinking about the role of reactive

iron minerals in hydrogeological systems. Ground water

in a number of settings has been described as having

elevated concentrations of dissolved iron in anoxic

portions of contaminant plumes (e.g. Baedecker et al.,

1993; Lyngkilde and Christensen, 1992), and the coupling

of microbial reduction of Fe(III) to the degradation of

organic compounds links the study of the biogeochemistry

of iron to our desire to understand processes influencing

contaminated aquifers. Questions about the transport and

fate of reactive constituents in ground water are linked

to the study of iron through recognition of the importance

of sorption processes on sesquioxide mineral surfaces

common in sedimentary aquifers limiting the migration

of metals, anions, and bacteria. We have been grappling

with questions related to the stability of iron mineral

phases in contaminated aquifers and the role of Fe(III)-

bearing minerals and grain coatings as sorption substrates.

In our ongoing study of the transport of bacteria

through an unconsolidated sandy aquifer in the Coastal

Plain of Virginia, we address a number of questions

related to the potential for bacterial attachment to mineral

solids. In laboratory and field-scale experiments, the

presence of these positively charged metal-oxide coatings

on quartz sand has been shown to greatly affect the

sorption and transport of reactive constituents. The

mineralogy of the metal oxide affects the affinity its

surface has for ion sorption. The presence of more than

one oxide in a system may affect the sorption

characteristics as well. In addition to metal-oxide

abundance, surface area is commonly employed as an

estimator of total sediment reactivity, and the result can

be inconsistent with an assessment of that portion of the

surface that actually participates in reactions.

Our investigation was undertaken on the Eastern

Shore near the village of Oyster where depositionally

similar sediments have been exposed to chemically

distinct ground-water conditions. In a narrow, organicrich leachate plume derived from vegetable waste pits,

dissolved oxygen was only 0 to 0.9 mg L-1, whereas it

was 5.0 to 11 mg L-1 in the regionally extensive aerobic,

uncontaminated ground water. In addition, the amount

of dissolved iron in the aerobic ground water was 0.001

to 0.01 mg/L-1, but it was 12 to 42 mg/L-1 in the anaerobic

zone. Samples of sediments from cores were subjected

to metal-oxide extractions and determination of

operationally defined fractions of iron (ferrous and ferric)

and aluminum. The amount of extractable iron was an

order of magnitude higher for the aerobic sediments than

for the anaerobic sediments indicating that reductive

dissolution removed the oxide coatings. The total iron

in the HCl extracts of the aerobic sediments ranged from

3.2 - 52 μmol/g whereas it was only 0.9 - 2.4 μmol/g in

the anaerobic sediments (Table 1). There was very little

ferrous iron (Fe(II)) extracted from either of the two

zones, and the values from each zone only ranged from

0.01 - 0.5 μmol/g of Fe(II) extracted. It is likely that

microbial oxidation of organic contaminants is linked to

the reduction of iron from the surfaces of sediments in

the polluted portion of the aquifer. Although the majority

of the iron had been removed from the surfaces of the

anaerobic sediments, there was still an appreciable

amount of surface area measured.

We sought a means to characterize the aquifer solids

for their potential for bacterial attachment, and we

developed a single-point sorption determination for a

reactive anion. In attempting to evaluate the reactivity

of mineral surfaces in an aquifer setting, sorption of

solutes directly probes that portion of the solid surface

that participates in reactions involving other negatively

charged reactive constituents, including bacteria. The

reactivity of the sediment surfaces, as indicated by the

sorption of 35SO4

2-, was an order of magnitude higher in

the aerobic vs. anaerobic sediments. The sulfate sorbed

to the aerobic sediments (2.2 to 50 μmol/g) was an order

of magnitude greater than for the anaerobic sediments

(0.007 to 0.25 μmol/g). The presence of anaerobic

Alteration of Reactive Mineral Surfaces by Gr Alteration of Reactive Mineral Surfaces by Ground Water

Janet S. Herman1 1University of Virginia

Aaron L. Mills1 Charlottesville, VA 22904

Isabelle M. Cozzarelli2 2

U. S. Geological Survey

Reston, VA 20192
40

conditions did not significantly alter the amount of

extractable aluminum oxides on the surface of the

sediments, and those coatings helped to maintain a high

surface area for the anaerobic sediments. It appears that

the removal of the iron oxides from the surfaces under

anaerobic conditions was solely responsible for the

significant reduction of sediment reactivity observed.

The iron-oxide content was the only one of the

surface properties to be significantly altered by the

anaerobic conditions with a concomitant alteration in the

ability of the sediments to retain the anion sulfate (Knapp

et al., in review). It appeared, therefore, the presence or

absence of iron oxides was the dominant factor controlling

sulfate sorption. These results would indicate an

alteration in the reactive surfaces of minerals when

exposed to anaerobic ground water and the potential for

greater movement of reactive constituents, including

bacteria, through these aquifer sediments.

As redox conditions in a contaminant plume change

with the depletion of oxygen and nitrate as electron

acceptors, ferric iron is consumed as an electron acceptor

coupled to the oxidation of organic matter (e.g. Lovley

et al., 1989). The glacial-outwash, sandy aquifer at the

USGS Toxic Substances Hydrology site near Bemidji,

MN, became contaminated by crude oil in 1979.

Observations since nearly the time of the pipeline burst

show accumulations of Fe(II) in the site ground water

over time, whereas no ferrous iron was found in the

uncontaminated background water (e.g. Baedecker et al.,

1993). Previous field and microcosm results (Baedecker

et al., 1993, Lovley et al., 1989) indicate the degradation

of toluene may proceed according to

( )( )

2

C H 36Fe OH 65H 7HCO 36Fe 87H O 7 8 3 S 3 2

+ −+

+ +→ + +

In this study, we examined the distribution of Fe in

the sediments of the aquifer in order to assess the

magnitude of impact of Fe(III) reduction in the

contaminant plume on the geochemistry of the sediments.

Sampling and extraction of sediments revealed that the

average HCl-extractable Fe(III) concentration in the most

contaminated portion of the aquifer was 16.2 μmol/g, a

30% reduction from the value in background sediments

of 23.8 μmol/g (Tuccillo et al., 1999). Comparison

between HCl extractions that should capture poorly

crystalline Fe(III) solids and Ti-citrate-EDTA-bicarbonate

extractions that should reductively dissolve amorphous

and crystalline Fe(III) oxides indicates that the bulk of

the microbially mediated iron reduction is dissolving

amorphous or poorly crystalline oxides.

In contrast to the study in Virginia in which virtually

no Fe(II) was extracted from aquifer solids, at Bemidji

we found 19.2 μmol/g Fe(II), as much as 4 times the

background sediments with 4.6 μmol/g Fe(II) (Figure 1).

Scanning electron microscopy detected authigenic ferroan

calcite in the anoxic sediments (Baedecker et al., 1992;

Tuccillo et al., 1999). Likely the contrast can be attributed

to the availability of minerals that buffer pH, giving the

contaminated ground water at Bemidji at pH near 6.9

whereas at the Oyster, VA, site it is closer to 6.3 in the

anaerobic ground water.

One of the most striking features of the distribution

of iron content of the aquifer sediments was the extreme

concentration at the anoxic/oxic transition zone. HClextractable Fe(III) reached values as great as 49.8 μmol/g

at this boundary (Figure 1). This 70% increase in total

extractable Fe at the interface indicates the reoxidation

and precipitation of Fe mobilized from aquifer sediments

Aerobic Core Anaerobic Core 

Depth (m) Total Fe 

(μmol/g) 

Depth 

(m) 

Total Fe 

(μmol/g) 

 -0.31 3.35 

 -0.81 18.4 

 -1.31 5.49 

-1.36 2.10 

-1.44 1.54 

-1.53 1.88 

-1.61 2.16 

-1.69 2.12 

-1.77 2.32 

 -1.81 10.5 

-1.88 2.45 

-2.32 2.40 

-2.80 1.27 

 -2.31 52.4 

 -2.81 10.5 

 -3.31 4.26 

 -3.81 3.16 

-3.88 1.97 

-4.26 1.34 

 -4.31 10.9 

Table 1. able 1. Concentration of total extractable iron with depth

(below the water table) for samples from aerobic

and anaerobic cores.
41

located upgradient in the anoxic, contaminated

groundwater plume as that water flows downgradient into

an oxic background ground water. Scanning electron

microscopy confirmed the abundant Fe(III)

oxyhydroxides at the anoxic/oxic boundary (Tuccillo et

al., 1999).

The alteration of sediment chemistry has

implications for the availability of electron acceptors and

the biodegradation of contaminant organics. As

availability of Fe(III) decreases, alternate hydrocarbon

degradation processes linked to methanogenesis will be

less efficient and intermediate metabolic products of

biodegradation will persist and be transported in the

ground water. The Fe(II) that is mobilized out of the

sediments in the most contaminated portion of the aquifer

will be reoxidized at some location downgradient and

precipitate as an “iron curtain” that can alter not only the

reactivity of the solid phase but can also change the

permeability as mineral grains are cemented together and

the interstitial spaces filled with Fe(III) precipitates. The

high Fe(III) content of that anoxic/oxic boundary gives

rise to a sorptive capacity of the transition-zone sediments

that is enhanced above both background and anoxic

sediments. The presence of the Fe(III) oxyhydroxides

has implications for the reactivity of aquifer sediments,

altering the nature and extent of reaction between sorbing

species and mineral surfaces. The contaminants that may

be retarded in their transport are not just those of the

original crude oil spill but potentially include bacteria,

anions, and metals. Improved understanding of the

biogeochemistry of iron in contaminated sedimentary

aquifers likely will lead to a better ability to anticipate

the rate of transport of a variety of solutes and colloids.

References

Baedecker, M. J., Cozzarelli, I. M., Evans, J. R. and

Hearn, P. P. 1992. Authigenic mineral formation in

aquifers rich in organic material. In Water-Rock

Interaction: Proceedings of the 7th International

Symposium on Water-Rock Interaction, eds. Y. K.,

Kharaka, and A. S. Maest, Balkema, pp. 257-261.

Baedecker, M. J., Cozzarelli, I. M., Eganhouse, R. P.,

Siegel, D. I., and Bennett, P. C. 1993. Crude oil in a

shallow sand and gravel aquifer - III. Biogeochemical

reactions and mass balance modeling in anoxic

groundwater. Applied. Geochemistry, v. 8,

pp. 569-586.

Knapp, E. P., Herman, J. S., Mills, A. L., and

Hornberger, G. M. 2001. Redox alternation of

reactive mineral surfaces in biologically active

groundwater. Applied Geochemistry, in review.

Lovley, D. R., Baedecker, M. J., Lonergan, D. J.,

Cozzarelli, I. M., Phillips, E. J. P., and Siegel, D. I.

1989. Oxidation of aromatic contaminants coupled

to microbial iron reduction. Nature, v. 339,

pp. 297-299.

Figure 1. Graphs of various fractions of iron extracted from

the sediments and of aqueous Fe2+ in pore waters

vs. distance from the center of the oil body. Circles

represent the mean of samples from all depths for

each core. Error bars represent standard errors

(s.d./n0.5) which gives an indication of within-core

variability.
42

Lyngkilde, J. and Christensen, T. H. 1992. Redox zones

of a landfill leachate pollution plume (Vejen,

Denmark). Journal of Contaminant Hydrology, v. 10,

pp. 273-289.

Tuccillo, M. E., Cozzarelli, I. M., and Herman, J. S. 1999.

Iron reduction in the sediments of a hydrocarboncontaminated aquifer. Applied. Geochemistry, v. 14,

pp. 71-83.
43

Introduction

The earth’s surface is in a steady struggle between

oxidation, driven by oxygen in the atmosphere, and

reduction, driven by organisms and organic matter. The

redox cycle established about 600 million years ago by

photosynthesizing and respiring organisms, organic

burial, and weathering processes maintains the

biogeochemical conditions we enjoy today. Without this

constant redox disequilibria in the air and in waters and

in rocks, life would not exist. Plants and petroleum would

spontaneously combust and the atmosphere would

become a sea of nitrogen oxides.

Confusion regarding redox (reduction-oxidation)

potentials and redox chemistry of aquatic systems stems

from misunderstandings of how aqueous redox reactions

occur and from misunderstandings of the assumptions

that enter into any calculations. Specific conceptual

difficulties arise in the following three areas: (1) the

existence and representation of the aqueous electron,

(2) the kinetics of electron transfer reactions, and (3) the

measurement and interpretation of redox in water bodies.

This paper discusses these various issues, shows examples

of equilibria and disequilibria of redox chemistry in

natural waters with particular attention to the ferrousferric redox couple, and explains the challenges of

modeling redox geochemistry.

The Aqueous Electron

Free aqueous electrons have been synthesized in

solution at micromolar concentrations by intense

radiation, either by flash photolysis or pulse radiolysis

(Hart and Anbar, 1970). These hydrated electrons are

extremely ephemeral, having half-lifes of 230 μs at pH

of 7. They quickly form hydrogen atoms and hydrogen

gas. Their concentrations in natural waters would be

negligible to nonexistent (Hostetler, 1984; Thorstenson,

1984). Hence, discussions of the activity of aqueous

electrons, designated as ae- , do not relate to a physically

meaningful quantity. This fact distinguishes aqueous

electrons from aqueous protons because aH+ is a

physically meaningful quantity. The mathematical entity

“pΗ” is physically measurable and has chemical

significance whereas the entity “p∈,” or - log a e-, does

not have physicochemical significance when applied to

aqueous electrons. There is no analogy between pH and

p∈ in a chemical sense. To draw the analogy in a

mathematical sense can be very misleading.

Kinetics of Electron Transfer Reactions Reactions

Reactions involving aqueous solutions are divided

into homogeneous (occurring only within the solution

phase) and heterogeneous (occurring between a solid or

gas surface and the solution phase). Homogeneous redox

reactions can go quickly if only one electron is being

transferred at a time and if the activation barrier is not

too large. However, numerous examples show that large

activation barriers do exist – otherwise nitrogen, oxygen,

methane, and carbon dioxide would not all coexist in the

atmosphere as they do. Eary and Schramke (1990) showed

clearly the wide range of oxidation rates for several

inorganic reactions as a function of pH and dissolved

oxygen. Redox reactions involving two electrons (more

than two per reaction step is generally forbidden) or

involving heterogeneous reactions are usually slower

unless catalysts or high temperatures are provided.

Empirical homogeneous redox reactions are

described by the usual kinetic equations (zero-order,

first-order, second-order, Michaelis-Menten for

enzyme catalysis, etc.) but electrode kinetics are

described by the current-overpotential equation:

ii e e [ ] C t

C

nF RT C t

C

o nF RT

o

r

r

= − − −− 0

0 0 1 ( ,) / ( ,) ()/ * *

αη α η (1)

where the exchange current, i, is dependent on the

equilibrium exchange current, i 0 , the concentration

gradient for the oxidized species between the surface,

Co(0,t), and the bulk solution, Co

* , the concentration

gradient for the reduced species, Cr (0,t), and the bulk

solution, Cr

*, the symmetry factor, a, the overpotential,

h, the Faraday constant, F, the ideal gas constant, R, and

Aqueous Redox Chemistr Aqueous Redox Chemistry and the Behavior y and the Behavior of Iron in Acid Mine Acid Mine Waters

D. Kirk Nordstrom U.S. Geological Survey

Boulder, CO 80303
44

the Kelvin temperature, T (Bard and Faulkner, 1980). If

the surface concentrations do not differ significantly from

the bulk solution concentrations then equation (1) reduces

to the Butler-Volmer equation developed in the 1920’s:

i ie e o [ ] nF RT nF RT = − − −− αη α η / ()/ 1 (2)

Equation (2) shows how the current across a metalsolution interface depends on the difference in potential

between the actual non-equilibrium and equilibrium

potential differences, η = E - Eeq. When this difference

(the overpotential) goes to zero, the equation reduces to

the well-known Nernst equation of equilibrium

thermodynamics written here for a half-cell with one

reduced and one oxidized species:

E E o R T

n F

a

a

oxidized

reduced = + ln ν

ν (3)

where E o is the standard electrode potential based on the

standard hydrogen electrode, the natural logarithm

contains the ratio of the activity of the oxidized species

to the activity of the reduced species, each taken to the

exponent represented by the stoichiometric coefficient, v.

The equation is also commonly written with a negative

sign before the log term with reduced species activity on

the top and the oxidized underneath. These equations

remind us that the Nernst equation is a special case of the

more general equation of electrode kinetics. The challenge

is knowing when a system is in equilibrium and when it

is not.

There are two systems being considered with respect

to equilibrium conditions. One system is the electrodesolution interface and the other is the environmental

system composed of gases, waters, rocks, and biological

activity. This section addresses the electrode-solution

system and the next section addresses the environmental

system.

The ease with which electrons are transferred across

the metal-solution interface depends on the nature of the

metal, the nature of the aqueous redox-sensitive species,

and the solvent properties. Metal electrodes need to be

non-reactive and electrically conductive. Thus, gold,

platinum, and graphite have worked well and among these

platinum is most frequently used. The metal is not usually

a limiting factor in reaching equilibrium as long as the

surface is free of electroactive coatings and other adsorbed

impurities. For aqueous redox species, there are some

serious limitations: (1) the net exchange current across

the interface must be effectively zero, i.e. electrochemical

reversibility must be maintained, (2) to maintain

reversibility, redox ions must have individual exchange

currents greater than 10-7 amp cm-2 (Morris and Stumm,

1967), which translates into concentrations of about 10-5 m

or greater for two redox constituents in natural waters:

Fe(II/III), S(II-), and possibly U(IV/VI). Several

examples in the literature have shown the equilibrium

maintained between the Pt electrode and the Fe(II/III)

redox couple in laboratory studies (Morris and Stumm,

1967; Macalady et al., 1990; Stipp, 1990). Peschanski

and Valensi (1949) showed that the Pt electrode responded

quantitatively and reversibly to changes in sulfide ion

activity in the S(II-/0) system in the laboratory.

Experiments designed to examine the electrochemical

reversibility of the As(III/V) and Se(IV/VI) redox couples

showed that they do not reach equilibrium (Kempton et

al., 1990; Runnells et al., 1987; Runnells and Skoda,

1990). Hence, only iron and reduced sulfur species

demonstrate electrochemical reversibility for aqueous

conditions likely to be found in the environment.

Measurement and Interpretation of Redox in

Waters

Two facets of redox chemistry need to be considered

when examining redox in aquatic systems: the attainment

of equilibrium for a redox couple at a Pt electrode surface

and the attainment of equilibrium between different redox

couples in ground water or surface water. Field

measurements and their interpretation are not simple and

proper precautions should be exercised (Langmuir, 1971;

Nordstrom and Wilde, 1998). Laboratory and theoretical

considerations would lead to the view that only rarely

would redox potential measurements in waters reflect

equilibrium and even then it should only be observed for

iron and reduced sulfur. Field work confirms this view.

For the iron system, an update of the original diagram by

Nordstrom et al. (1979) that compares measured vs.

calculated Eh for more than 60 acid mine waters is shown

below in Figure 1. The calculated Eh values were based

on ferrous and ferric iron determinations (To et al., 1999),

complete water analyses, and speciation based on output

from the WATEQ4F program (Ball and Nordstrom, 1990).

The agreement is generally excellent with most Eh

values agreeing to within 25 mV except at low Eh where

a strong positive deviation occurs. Figure 2 plots the

difference between calculated and measured Eh as a

function of total dissolved iron concentrations. Marked

deviations occur below 10-5 m as predicted from lab

studies. These results reflect the mixed potential effect

as the iron concentrations become too low to be

electroactive and oxygen begins to be sensed by the
45

electrode. However, the O2/H2O redox equilibrium

potential is never reached because oxygen is not

sufficiently electroactive. Hence, field studies have been

able to demonstrate the same equilibrium response of a

platinum electrode to the Fe(II/III) redox couple as in

the lab. This study also shows the reliability of the

speciation code. Similar success has been shown with

over 50 measurements of asulfide compared to Eh electrode

measurements by Berner (1963) for anoxic marine

sediments. Boulegue and Michard (1979) made similar

measurements in sulfide-rich springs using acid-base

titrations. The difficulty with these field measurements

is that the slightest trace of oxygen can produce

polysulfides, thiosulfate, and elemental sulfur which will

cause deviations in the Eh measurement. Likewise, for

anoxic ground waters high in iron, the introduction of a

well, pumps, and sampling lines causes oxygen to enter

the sampling system and instantaneously oxidizes any

reduced ferrous iron if the pH is circumneutral. The result

is the precipitation of hydrated ferric oxides which will

tend to cause a Fe(II)/Fe(OH)3

 redox couple to occur

regardless of whether or not it actually occurs in the

unaltered ground water. The only other example that

might be approaching equilibrium is the U(IV/VI) redox

equilibrium for 6 Finnish ground-water samples from the

work of Ahonen et al. (1994). No other redox couples in

natural aquatic systems are known to produce an

equilibrium potential at an electrode surface.

Furthermore, when two or more redox couples have been

measured on the same water sample, they do not give the

same Eh value (Nordstrom et al., 1979; Lindberg and

Runnells, 1984), and they should not be expected to give

the same value because they tend to react at different

rates with different mechanisms. If redox species all came

to equilibrium it would destroy life. Life depends on redox

gradients and redox disequilibria. Geochemical codes that

force redox equilibrium for all elements are grossly

oversimplified and cannot be expected to reflect realistic

redox conditions (unless temperatures > 100o

C). There

are only three approaches that will permit successful

application of geochemical models to redox reactions:

(1) determine individual redox species analytically,

(2) apply independent speciation to reduced and oxidized

forms of the species, and (3) apply mass balances

(Parkhurst and Plummer, 1993). Changes in pH that

reflect pyrite oxidation and aqueous iron oxidation can

be simulated effectively by speciation and mass balances

if a judicious choice of equilibrium constants is made

(Nordstrom, 2000). The only way to determine the redox

conditions of a water is to analyze the sample for those

redox species of specific concern. There are several

examples of the same water containing two or more of

the following constituents O2

, Fe(II), Fe(III), N2, NO3,

NH3

, H2S, SO4, CH4, CO2, and H2. There can be no such

thing as redox equilibrium or a single redox potential in

these waters. The way to understand redox chemistry is

to learn the rates and mechanisms of reactions, the main

catalysts in the environment, and the main sources and

sinks of redox active species.

Conclusions Conclusions

Redox studies on aquatic systems in the lab and in

natural waters lead to the following conclusions:

1. The absence of aqueous electrons in natural waters

means they cannot be measured nor defined by

analogy with aqueous protons.

2. Ground waters and surface waters do not have a

“redox potential,” p∈, or Eh but they do have a pH.

To speak of a redox potential of an aqueous solution

or natural water does not have any meaning.

3. Redox species in waters do not reach an equilibrium

state readily, if at all. Redox disequilibria is the rule.

 Leviathan samples

 Iron Mountain samples

 Summitville samples

0 0.2 0.4 0.6 0.8 1.0

Eh calculated from Fe (II/III), volts

0

0.2

0.4

0.6

0.8

1.0

-8 -6 -4 -2 0

-0.2

0

0.2

0.4

0.6

log total dissolved iron, molality

Figure 1. Figure 1. Comparison of calculated with measured Eh.

Figure 2. Figure 2. Difference in Eh as a function of Fe molarity.
46

4. Redox potential measurements with platinum (or

similar) electrodes are generally useless except to

estimate ferrous-ferric activity ratios or sulfide

activities when the concentrations are greater than

about 10-5 m. Otherwise the potential may drift

according to electrokinetic phenomena, mixed

potentials, or impurities at the metal electrode surface.

It is more reliable to measure such concentrations

directly, where possible, and calculate the activities

with a speciation model.

5. To determine the redox chemistry of a water it is

necessary to determine all the relevant redox species

directly. These species can be expected to react in

different ways and at different rates. Homogeneous

redox reactions can be rather slow and heterogeneous

redox reactions can be even slower.

6. Redox species can be very sensitive to atmospheric

contamination. Contamination and interferences

begin with drilling a well, continue with the manner

and rate of pumping, and end with a set of data that

is often inconsistent or discrepant.

7. A judicious choice of redox parameters depends on

site characteristics, budget, and objectives.

8. A monitoring schedule depends on site-specific

factors such as flow rates, types of redox species

present, rates of redox reactions, redox capacity of

the system, and rates and mechanisms of processes

in the vadose zone.

9. Research knowledge is in the process of finding out

how redox works and how fast. The knowledge is

not at a point where we can say definitively what the

protocol for redox is nor can we model any specific

site to predict just what will happen over time except

in a general, but not necessarily helpful, sense.

References

Ahonen, L., Ervanne, H., Jaakola, T., and Blomqvist, R.

1994. Redox chemistry in uranium-rich groundwater

of Palmottu uranium deposit, Finland. Radiochimica

Acta, v. 20, pp. 1-7.

Ball, J. W. and Nordstrom, D. K. 1990. User’s manual

for WATEQ4F, with revised thermodynamic data base

and test cases for calculating speciation of major,

trace, and redox elements in natural waters

U.S. Geological Survey Open-File Report,

91-183, 189 p.

Bard, A. J. and Faulkner, L. R. 1980. Electrochemical

Methods. John Wiley and Sons, New York, 718 p.

Berner, R. A. 1963. Electrode studies of hydrogen sulfide

in marine sediments. Geochimica et Cosmochimica

Acta, v. 27, pp. 563-575.

Boulegue, J. and Michard, G. 1979. Sulfur speciation and

redox processes in reducing environments. In

Chemical Modeling in Aqueous Systems, Chapter 2,

ed. E. A. Jenne, American Chemical Society

Symposium Series 93, American Chemical Society,

Washington, D.C., pp. 25-50.

Eary, L. E. and Schramke, J. A. 1990. Rates of inorganic

oxidation reactions involving dissolved oxygen. In

Chemical Modeling of Aqueous Systems, II, Chapter

30, eds. D. C. Melchior and R. L. Bassett, American

Chemical Society, Washington, D.C., pp. 379-396.

Hart, E. J. and Anbar, M. 1970. The Hydrated Electron.

Wiley Interscience, New York, 267 p.

Hostetler, J. D. 1984. Electrode electrons, aqueous

electrons, and redox potentials in natural waters.

American Journal of Science, v. 284, pp. 734-759.

Kempton, J. H., Lindberg, R. D., and Runnells, D. D.

1990. Numerical modeling of platinum Eh

measurements by using heterogeneous electrontransfer kinetics. In Chemical Modeling in Aqueous

Systems II, Chapter 27, eds. D. C. Melchior and R.

L. Bassett, American Chemical Society Symposium

Series 416, American Chemical Society,

Washington, D.C., pp. 339-349.

Langmuir, D. 1971. Eh-pH determination. In Procedures

in Sedimentary Petrology, Chapter 26, ed.

R. E. Carver, Wiley-Interscience, New York, pp 597-

635.

Lindberg, R. D. and Runnells, D. D. 1984. Ground water

redox reactions: An analysis of equilibrium state

applied to Eh measurements and geochemical

modeling. Science, v. 225, pp. 925- 927.

Macalady, D. L., Langmuir, D., Grundl, T., and

Elzerman, A. 1990. Use of model-generated Fe3+ ion

activities to compute Eh and ferric oxyhydroxide

solubilities in anaerobic systems. In Chemical

Modeling in Aqueous Systems, Chapter 28,

ed. E. A. Jenne, American Chemical Society

Symposium Series 93, American Chemical Society,

Washington, D.C., pp. 350-367.

Morris, J. C. and Stumm, W. 1967. Redox equilibria

and measurements of potentials in the aquatic

environment. In Equilibrium Concepts in Natural

Water Systems, Chapter 13, ed. W. Stumm, American
47

Chemical Society Advances in Chemistry Series 67,

American Chemical Society, Washington, D.C.,

pp. 270-285.

Nordstrom, D. K. 2000. Advances in the

hydrogeochemistry and microbiology of acid mine

waters. International Geology Review, v. 42,

pp. 499-515.

Nordstrom, D. K. and Wilde, F. D. 1998. Reductionoxidation potential (electrode method). In National

Field Manual for the Collection of Water-Quality

Data, eds. F. D. Wilde, D. B. Radtke, J. Gibs, and

R. T. Iwatsubo, U.S. Geological Survey Techniques

in Water-Resources Investigations Book 9,

Handbooks for Water-Resources Investigations, A6.5.

Nordstrom, D. K., Jenne, E. A., and Ball, J. W. 1979.

Redox equilibria of iron in acid mine waters. In

Chemical Modeling in Aqueous Systems, Chapter 3,

ed. E. A. Jenne, American Chemical Society

Symposium Series 93, American Chemical Society,

Washington, D.C., pp. 51-79.

Parkhurst, D. L. and Plummer, L. N. 1993. Geochemical

models. In Regional Ground-Water Quality, Chapter

9, ed. W. M. Alley, Van Nostrand Reinhold, New York,

pp. 199-226.

Peschanski, D. and Valensi, G. 1949. Contribution a

l’electrochemie des solutions aqueuses de

polysulfures. Journal de Chimie Physique, v. 46,

pp. 602-619.

Runnells, D. D. and Skoda, R. E. 1990. Redox modeling

of arsenic in the presence of iron: Applications to

equilibrium computer modeling. In Proceedings:

Environmental Research Conference on Groundwater

Quality and Waste Disposal, EPRI Report EN-6749,

Palo Alto, California.

Runnells, D. D., Lindberg, R. D., and Kempton, J. H.

1987. Irreversibility of Se(VI)/Se(IV) redox couple

in synthetic basaltic ground water at 25o

C and 75oC.

Materials Research Society Symposium Proceedings,

v. 84, pp. 723-733.

Stipp, S. L. 1990. Speciation in the Fe(II)-Fe(III)-SO4

-

H2

O system at 25oC and low pH: Sensitivity of an

equilibrium model to uncertainties. Environmental

Science and Technology, v. 24, pp. 699-706.

Thorstenson, D. C. 1984. The concept of electron activity

and its relation to redox potentials in aqueous

geochemical systems. U.S. Geological Survey OpenFile Report 84-072, 45 p.

To, T. B., Nordstrom, D. K., Cunningham, K. M., Ball, J.

W., and McCleskey, R. B. 1999. New method for

the direct determination of dissolved Fe(III)

concentration in acid mine waters. Environmental

Science and Technology, v. 33, pp. 807-813.
48
49

Redox processes can play a role in subsurface

remediation by either removing inorganic contaminants

from ground water and immobilizing them in the solid

phase or enhancing mobility in the aquifer so that the

contaminants can be removed from the system. The redox

process may directly or indirectly lead to remediation.

An example of direct remediation is the reduction of

Cr(VI) to Cr(III) and precipitation of a less soluble Cr(III)

hydroxide solid. Indirect redox remediation would occur

when dissolved ferrous iron is added to ground water

followed by iron oxidation, precipitation of hydrous ferric

oxide (HFO) and adsorption and removal of arsenic or

other inorganic contaminants from ground water.

Conversely, arsenic might be mobilized and removed

from an aquifer by injecting a reductant to dissolve

arsenic-adsorbing solids and release arsenic to the ground

water, which would then be pumped from the aquifer.

The purpose of this paper is to:

♦ Identify redox-sensitive contaminants

♦ Discuss natural redox conditions that attenuate

movement

♦ Describe applied redox methods that immobilize

contaminants

♦ Describe applied redox methods that mobilize

contaminants

Additional activities at the Redox Workshop related

to inorganic contaminants will generate site

characterization requirements and methods of measuring

important parameters. Data interpretation and use will

also be discussed separately.

Redox-Sensitive Elements Redox-Sensitive Elements

Elements that occur in more than one valence state

in natural environments and may be present as

contaminants are listed in Table 1. These elements may

be amenable to direct redox manipulation methods to

produce remediation or at least contaminant concentration

reduction. These elements may also be affected by

indirect redox remediation. Elements that are not redoxsensitive (such as Pb, Zn and Al) may also be affected by

indirect redox methods.

Table 1. able 1. Redox-Sensitive Elements Potentially Present as

Contaminants in the Environment

 StatesE

Elements Redox State

Antimony (Sb) Antimony (Sb) III, V

Mercury (Hg) 0, I, II

Arsenic (As) Arsenic (As) III, V

Nitrogen (N) ogen (N) -III, 0, III, V

Carbon (C) Carbon (C) -IV to IV

Molybdenum (Mo) Molybdenum (Mo) III, IV, V, VI

Chromium (Cr) omium (Cr) III, VI

Selenium (Se) Selenium (Se) -II, 0, IV, VI

Copper (Cu) Copper (Cu) I, II

Sulfur (S) Sulfur (S) -II to VI

Iron (Fe) on (Fe) II, III

Vanadium (V) anadium (V) III, IV, V

Manganese (Mn) Manganese (Mn) II, III, IV

Uranium (U) Uranium (U) IV, VI

Many applied remediation methods for inorganic

compounds are based on natural processes that attenuate

or immobilize these compounds in the subsurface. This

discussion of remediation methods will start by looking

at these natural environments that concentrate inorganic

compounds by redox processes.

Redox Barriers to Inorganic Compounds Redox Barriers to Inorganic Compounds

The concept of a geochemical barrier in the natural

system was first suggested by Perel’man in 1961

(Perel’man, 1986, Geochemical barriers: theory and

practical applications. Applied Geochemistry, 1:669-680).

A geochemical barrier is a zone in the subsurface

characterized by a sharp change in physical or chemical

environment along the flowpath and is often associated

Redox Pr Redox Processes in Inorganic Remediation

Bill Deutsch Sedgwick, ME 04676
50

with the precipitation of elements from solution. Redox

barriers are classified as a type of physico-chemical

barrier, which also includes alkaline, acidic, adsorption,

evaporation and thermodynamic barriers. Three types

of redox barriers have been described: oxidizing, reducing

with sulfide, and reducing without sulfide. Because

barriers represent a change in environmental condition,

each type of barrier must be considered in the context of

the three redox types of water that might enter the barrier.

For example, weakly oxidizing water may encounter a

strongly oxidizing barrier or a reducing barrier with or

without sulfide. Finally, the pH of the aquifer will affect

the type of element immobilized by the geochemical

barrier. Table 2 provides the potential encounters

developed by Perel’man and the elements that might be

concentrated at the three redox barriers.

Examples of the geochemical barriers listed in Table

2 include oxidation and precipitation of iron from mildly

oxidizing or reducing waters without sulfide

downgradient of a landfill; presence of U, As, and Mo in

roll front deposits formed by reduction of oxidizing water;

and precipitation of metal sulfides when oxidizing water

becomes reduced by organic matter in the presence of

sulfide. Table 2 shows the types of direct redox

remediation that might occur in geochemical barriers.

The barrier will lower concentrations of inorganic

contaminants, however the table does not provide

information on whether the resulting concentration will

be below a remediation cleanup level. The equilibrium

solution concentration will be the solubility of the mineral

in the barrier environment. This can be estimated by

proper characterization of the environment and

calculating the solubility of minerals in that environment.

An indirect natural redox attenuation process is the

precipitation of iron and manganese hydroxides followed

by adsorption of other contaminants. These solids have

a predominantly positive surface charge at neutral to

acidic pH, and in this pH range will preferentially adsorb

elements that are anionic in groundwater such as arsenic

(AsO4

3-, AsO33-), chromium (CrO42-), selenium (SeO42-,

SeO3

2-), and molybdenum (MoO42-). At pH higher than 7

to 8, the mineral surfaces are predominantly negatively

charged and they will be better adsorbents for cations

such as Cu2+, Cd2+, Zn2+ and Ba2+.

Nitrate is an inorganic contaminant that does not

form minerals and thus would not be removed from

solution and concentrated in a geochemical barrier. The

fate of nitrate is usually tied to biological redox processes

that reduce nitrogen to N2

 or NH4

+.

Methods of Immobilizing Inorganic

Contaminants Contaminants

Applied techniques for immobilizing inorganic

contaminants are generally based on the well-documented

natural processes that impact element mobility,

particularly at geochemical barriers. If manganese

concentrations in mildly oxidizing water downgradient

from a landfill must be lowered, Table 2 suggests either

Table 2. able 2. Elements Concentrated at Geochemical Redox Barriers (after Perel’man 1986)

Elements in parentheses are less commonly found in the association

Barrier Type pH < 3 pH 3 – 6.5 pH 6.5 – 8.5 pH > 8.5 

Oxidizing Water Entering Barrier 

Oxidizing Fe Fe, Mn, Co Mn - 

Reducing w/ H2S Tl, Cu, Hg, Pb. 

Cd, Sn, As, Sb, 

Mo, U 

Tl, Mn, Co, Ni, Cu, 

Zn, Pb, Cd, Hg, Sn, 

Cr, Mo, U 

Tl, Cr, Mo, U, Se, 

V 

Cu, Ag, Zn, Cr, 

Mo, U, V, As 

Reducing w/o H2S Cu, U, Mo Cu, U, Mo Cu, Cr, U, Mo, 

Se, V 

Cu, Ag, Cr, Mo, U, 

Se, V, As 

Reducing H2S Water Into Barrier 

Oxidizing S, Se, (Fe) S, Se S, Se S, Se 

Reducing w/ H2S - - - - 

Reducing w/o H2S - - - - 

Reducing Water w/o H2

S Into Barrier 

Oxidizing Fe Fe, Mn, Co (Fe), Mn, Co (Mn) 

Reducing w/ H2S Tl, Pb, Cd, Sn Tl, Fe, Co, Ni, Pb, 

Cu, Zn, Cd, Hg, U 

Tl, Fe, Co, Ni, 

Cu, Zn, Cd, Hg, 

(Mo, U) 

Tl, Cu, Zn, Cd, 

Hg, Mn, (Fe, Co, 

Ni, U) 

Reducing w/o H2S Cu, U, Mo Cu, U, Mo Mo, U Mo, U 
51

increasing the redox potential to precipitate a less soluble

Mn(IV) oxide or reducing the redox potential and adding

sulfide to form a less soluble sulfide mineral. Some of

the issues to consider in the applied redox approach are

the type and amount of reactant, means of emplacement,

reaction kinetics, unwanted byproducts, solubility of

contaminant-containing mineral, and geochemical

stability of the imposed barrier environment.

The types of oxidizing and reducing compounds that

have been tested or suggested to lower inorganic

contaminant concentrations in an aquifer are listed in

Table 3. Most of these compounds can be dissolved in

water and injected into the aquifer. Air and ozone are

added to the subsurface in a vapor phase. Elemental iron

(Feo

) is a solid that is emplaced in a trench that intercepts

the contaminated water. It has been shown to be effective

not only at reducing the concentrations of inorganic

contaminants (particularly chromate and uranium) but

also chlorinated volatile organic compounds such as PCE

and TCE. As shown in Table 3, the current list of reagents

is fairly limited and the elements tested are not extensive.

This type of remediation is expected to expand as its

effectiveness becomes more widely known and the

injection of reagents to achieve remediation becomes

more commonly accepted by regulatory agencies.

In general, the amount of oxidant or reductant that

must be added to achieve remediation will not exceed

100 mg/L. For example to reduce 50 mg/L Cr(VI) in

CrO4

2- to Cr(III) would require only 12 mg/L sulfide,

assuming all the sulfide is oxidized to sulfate. An

important consideration for some reactions is the rate of

oxidation or reduction, which may be slow in relation to

a remediation timeframe. For example, the half-life for

chromium reduction may be several years at pHs greater

than 7 unless a strong reductant is applied. The solubility

of the reaction product containing the contaminant can

be estimated by ion speciation calculation using the

projected environmental conditions during reagent

injection. The long-term solubility of the mineral must

also be estimated using the aquifer geochemical

conditions that are expected to prevail after active

remediation has ceased.

Methods of Mobilizing Inorganic Contaminants Methods of Mobilizing Inorganic Contaminants

It is preferable to leave a contaminant in place if its

mobile concentration can be reduced to below a cleanup

level, however in some instances immobilization may

not be possible and removal is necessary. Removal of an

inorganic contaminant requires not just extraction of

contaminated ground water but also dissolution of

contaminant-containing minerals and perhaps desorption

of the contaminant from the surface of solids. Table 2

can be used to develop redox methods of removing

contaminants from the solid phase. For example, if

Table 2 shows that an element is concentrated under

reducing conditions (e.g., Cr, U, Mo, Se), then creating

oxidizing conditions in the aquifer should mobilize the

contaminant. This is exactly the process used to mine

uranium by in-situ leach methods. Oxygen is added to

the subsurface where uranium has been concentrated by

natural processes in the reduced U(IV) mineral uraninite

(UO2

). Uraninite is soluble under the imposed oxidizing

conditions leading to its dissolution and the removal of

dissolved uranium in extraction wells.

Table 3. able 3. Potential Oxidants and Reductants for Inorganic Remediation

Oxidants Elements Tested 

Dissolved Oxygen Fe 

Air U 

Ozone 

Oxygen Release Compound (ORC ) 

KMnO4

Fenton’s Reagent 

Reductants Elements Tested 

Feo Cr, U, Tc, NO3

-

, Mo 

Dithionite (S2O4

2

-) Cr, U, Tc, Pu 

Na2 S U, Cr (Ag, As, Cd, Hg, Ni, Pb, 

Zn) 

H2 S Cr 

FeSO4

.

xH2O Cr 

Hydrogen Release Compound (HRC) Cr 

Organic Material SO4

2-, NO3-

, Se, Fe(S), Ni(S) 
52

Table 4 shows imposed redox conditions that are

expected to mobilize contaminants concentrated in three

types of subsurface environments. The methods of

inducing redox conditions are the same as those given in

Table 3. Mobilization of contaminants can be complicated

by the strong adsorption and high capacity of iron and

manganese oxyhydroxides for other metals. If oxidizing

conditions are generated in a reducing environment to

mobilize metals, iron and manganese solids may

precipitate and remove from solution some of the metals

that are the target of remediation. This occurrence may

require a stepwise mobilization process to first oxidize

the elements in the minerals and then reduce and dissolve

iron and manganese oxyhydroxides to release adsorbed

contaminants.

Table 4. able 4. Redox Manipulation to Mobilize Inorganic Contaminants

Natural and applied redox processes can be used to

achieve remediation goals at sites contaminated with

inorganic compounds. This has been documented for

many elements by the existence of natural geochemical

barriers, and recent applied techniques have shown that

the natural redox processes can be enhanced by chemical

addition. A goal of the Redox Workshop has been to

expand the use of redox processes for inorganic

remediation by providing examples of natural attenuation

analogues and the application of these natural processes

to enhance applied restoration techniques.

Mobilizing 

Environments 

pH < 3 pH 3 – 6.5 pH 6.5 – 8.5 pH > 8.5 

Initial Oxidizing Environment

Reducing w/o 

Sulfide 

Fe, Se, S Fe, Mn, Co, Se, S Fe, Mn, Co, 

Se, S 

Mn, Se, S 

Initial Reducing w/ S2- Environment

Oxidizing Tl, Cu, Hg, 

Pb. Cd, Sn, 

As, Sb, Mo, 

U 

Tl, Mn, Co, Ni, 

Cu, Zn, Pb, Cd, 

Hg, Sn, Cr, Mo, U 

Tl, Cr, Mo, U, 

Se, V 

Cu, Ag, Zn, 

Cr, Mo, U, 

V, As 

Initial Reducing w/o S2- Environment

Oxidizing Cu, U, Mo Cu, U, Mo Cu, Cr, U, Mo, 

Se, V 

Cu, Ag, Cr, 

Mo, U, Se, 

V, As 
53

Introduction

Redox conditions represent an important control on

the speciation of metals including As (III/V), Cr (III/VI),

Sb (III/V), and Se (IV/VI). Although the measurement

and environmental geochemistry of these constituents is

well understood, application of this knowledge in setting

regulatory standards for soils, ground water, surface water

and sediments is rare. Hence, bulk chemistry is usually

considered in setting standards rather than evaluating the

ameliorating effect of metal valence on toxicity and

mobility. Given the escalating costs of environmental

restitution, it is critical that reasonable options, including

the influence of valence, be considered in setting realistic,

risk-based cleanup goals.

This paper provides examples of the value of

understanding redox conditions and species in the

environment. Arsenic is less bioavailable in the As(III)

form in mine wastes due to sparing dissolution of these

less soluble forms in the short GI tract residence times,

while selenium form in sediment is important because

Se(IV) is less toxic than Se(VI). AVS/SEM analysis on

both metalloids has demonstrated that they are

incorporated in framboidal pyrite, rendering them nonbioavailable.

In addition, novel approaches to constrain metal

solubility by modifying the valence state are realistic

goals. For example, the mitigating influence of

precipitation reactions in sediments is demonstrated

through sulfate/sulfide reduction/precipitation and

coincident encapsulation of metals in a biologically active

system in Richmond, California. For example, the

solubility of Cr(III) in acid ground waters can be reduced

by careful pH control while remediation of acid pit lakes

can be effected by modification of the pH with

concomitant precipitation of amorphous ferric hydroxide

as the Fe(II)/Fe(III) ratios change, and sequestration of

previously soluble ions.

The practical application of redox conditions to

remedial actions is of interest, first from the perspective

The Relevance of Redox to Remediation

Andy Davis Geomega

Boulder, CO 80303

of “does it matter,” and secondly in defining how to

implement closure in the event that redox conditions are

relevant. From the closure/remediation perspective, the

importance of redox is closely related to the implied risk

associated with different valence states. For example,

As(III) is more toxic in groundwater whereas As(V) may

be more toxic in an estuarine setting, depending on the

exposure pathway; Cr(VI) is generally more toxic than

Cr(III), while Se(VI) is more toxic than Se(IV).

From a pragmatic perspective, although the relative

risk associated with different valence states may be

known, translating theory into practical closure that meets

regulatory standards is rare. For example, contaminated

environmental systems are complex and heterogeneous,

frequently consisting of multiple compartments and

contaminants, while translating theory into practical

reality is fraught with uncertainty. In addition, the

significance of the issue and a reasonable benchmark

representing success is generally not understood by the

decision maker, while with few exceptions, the remedial

actions do not finesse the benefit to be accrued by altering

valence states in the environment.

Despite such recidivism, where viable, manipulation

of redox conditions may portend significant financial

benefit and should be included as a weapon in the

remedial arsenal. The application of redox can be divided

into the following areas:

1) Active manipulation (e.g., redox control/response)

2) Constraints on risk (e.g., definition of valence in land

application, metal bioavailability), and

3) Passive long-term (e.g., intrinsic remediation by

dechlorination)

This paper reviews the genesis of the redox-based

standards, describes a few examples where redox states

have been incorporated into closure/remedial measures,

forecasts what may work and whether success or failure

can be predicted, and discusses issues that are likely to

become important in the future.
54

The Evolution of The Evolution of Valence-Based Risk

Promulgation of the Aquatic Water Quality Control

(AWQC) Standards may represent the initial foray that

led to a description of the relative toxicities of metal

valences in the “Gold Book” (EPA, 1986). The later

document identified a range of conditions that affect

different trophic levels. For example, the alga,

selenastrum capricornutum was reported 45 times more

sensitive to As(V) than As(III), although other data present

conflicting information. In general, the resiliency to

arsenic is species and salinity specific.

Chromium toxicity appears more straightforward,

in that there is general concurrence that Cr(III) is less

toxic than Cr(VI). For example the AWQC standard for

Cr(VI) is 100 μg/l, while for Cr(III) it is 3,433 μg/l. In

freshwater, the acute toxicity for Se(IV) is reported as

260 μg/l, while for Se(VI) it is 760 μg/l. However, there

has been no directives that clearly identify the relationship

between valence species and the total dissolved

concentration in aqueous or sediment systems, and the

ramifications to risk assessments and, ultimately, to

selection of cleanup standards.

Precipitation of Chromium from Ground Water

At a former chemical facility in Wilmington, a

historical release of chromium sulfate has generated an

acidic plume (pH~3) of Dense Aqueous Phase Liquid

(DAPL) with complex chemical characteristics. A Release

Abatement Measure (RAM) is being instituted to assess

the effectiveness of in-situ precipitation of ground water

solutes using various chemical agents applied in solution

or slurry through existing wells with the objective of

raising the pH, while avoiding oxidation of the Cr(III) to

soluble Cr(VI). The effectiveness will be determined by

monitoring the areal extent of DAPL precipitation. The

goal of this treatment is to mitigate the flux of solutes

from the DAPL to the overlying ground water that

discharges into the on-site ditch system, hence reducing

precipitates forming in the ditch as ferrous iron in solution

discharges and precipitates in surface water.

In the target area, the local ground water/diffuse

layer/DAPL profile consists of approximately 6 feet of

DAPL overlain by a 3 foot thick diffuse layer. Comparison

of ground water surface elevations from co-located wells

in the region of the ditches demonstrates that ground water

discharges to the ditch system resulting in the formation

of floc. It is anticipated that the DAPL RAM will inter

DAPL, reducing the solute flux to the South Ditch, and

eventually floc formation.

Bench-scale studies observed decreases in metal

concentrations attributed both to precipitation reactions

and to changes in solubility due to the more neutral pH

achieved through treatment. For example, chromium

decreased from 2,000 mg/l to < 0.1mg/l, while similar

decreases were observed for Al and Fe. Additionally,

sulfate concentrations decreased by up to 30%. Hence,

neutralization of acidic DAPL to circumneutral (~6)

conditions should result in a decrease in DAPL solubility

by a precipitation reaction of the form:

9 Fe+2 + Cr2

[SO4]

3 + 24 OH- ⇒ 8 Fe0.75Cr0.25(OH)3

 +

3FeSO4

0

 + 6 ewith the relevant redox reaction controlled by oxidation

of ferrous to ferric iron and coprecipitation of Cr(III) in

the amorphous ferric hydroxide matrix. The optimal

liquor concentration determined from the bench scale

testing will be applied to the target area. The neutralizing

liquor and the sodium bromide tracer (at 100 mg/L

effective bromide concentration) will be initially

delivered into the injection well by passive head

differential. As delivery slows, the liquor may be pumped

under pressure if feasible.

Minimizing risk by recognizing inert

sedimentary phases sedimentary phases

Castro Creek, an industrial drainage ditch that

historically received disposed process fluids, contains

residual arsenic in the sediments. This project initially

identified solubility-constraining reactions and then

sought to enhance the natural conditions to further

influence pore water arsenic.

The molar SEM/AVS ratios for arsenic are well

below 1, ranging from 0.0003 at B2 to 0.62 at F. Although

SEM/AVS data for arsenic has not been directly coupled

with arsenic toxicity, other studies (Davis et al., 1996)

have demonstrated that when the SEM/AVS ratio is <1,

bioaccumulation of arsenic does not occur in higher

trophic level receptors (fish). Therefore, based on these

studies (Davis et al., 1996), the ambient geochemical

conditions of the creek sediments facilitates sequestration

of arsenic in insoluble mineral phases such as Fe(As)S,

Fe(As)SO4

, and AsS2 that are likely to reduce As

bioavailability in a manner akin to the metal-sulfide

complexes described by DiToro (1990 and 1992).

 Encouraging Encouraging Arsenic Reduction in Sediments Arsenic Reduction in Sediments

Recognizing the benefits to be realized from a

reducing sediment environment resulted in research

(currently ongoing) to facilitate additional removal of
55

arsenic from pore water at the site. Currently a remedial

design scheme intended to remove arsenic (As) and

control remnant As mobility in marsh sediment using a

Peat-Iron-Gypsum (PIG) blanket to minimize As

solubility by reducing As(V) to As(III) and facilitating

co-precipitation of As in an inert FeS2 (e.g., FexAs1-xS2

and FexAs1-xSO4) phase is being investigated.

Iron oxide and sulfates containing up to 3,000 mg/kg

As have been found precipitating in creek sediments

suggesting that natural conditions in these areas are

conducive to immobilization of incorporated As.

Theoretically, addition of appropriate amendments would

attenuate most of the remaining soluble pore-water As

over the ambient range of redox conditions, also reducing

the flux of As into the creek.

A bench scale study was run for 24 hours (thought

to be an adequate period for pyrite precipitation; Rickard

1995) under zero-head space conditions in a nitrogen

atmosphere and each bottle sacrificed after 7 days. The

aqueous and solid phases were separated by centrifugation

and samples of each collected and analyzed. Measured

geochemical parameters include SO4

2-, S2-, Fe(II), Fe(III),

pH, Eh, and specific conductivity. A sample was also

analyzed for inorganic arsenic species [As (III)/(V)]. The

results were surprising in that the reaction kinetics were

much slower than has been reported in the literature using

pure materials.

Demonstrating pesticide dechlorination as part Demonstrating pesticide dechlorination as part

of intrinsic remediation

An investigation was undertaken to ascertain the

potential and actual degradation of chlorinated pesticides

(α-, β-, and γ-BHC, DDT, and DDD), and the purgeable

aromatic compounds (ethylbenzene and xylene) at the

Marzone Superfund Site, Tifton, Georgia. The objective

was to test the hypothesis that Monitored Natural

Attenuation (MNA) would provide an effective groundwater polishing strategy to augment the remedial collector

trench recently installed downgradient of the site. The

trench effectively intercepted over 90% of the mass of

all Compounds of Concern (COCs) while concentrations

downgradient of the trench are at, or close to, the relevant

numerical standards.

The geochemical conditions were characterized by

analyzing redox chemistry (Eh, iron and sulfur valence

states), gas concentrations (hydrogen, methane, ethane,

ethene), and other indicators (chloride, alkalinity, etc).

These data demonstrated that upgradient ground water

that was unaffected by the historical Marzone Site

operations is moderately aerobic and oxidizing, becoming

moderately reducing proximal to the operating facility.

In the core of the plume, conditions become strongly

reducing based on the increase in the relative ratio of

hydrogen, the presence of dissolved iron exclusively in

the ferrous form, and on the concomitant sag in the

oxygen ratio. Downgradient of the plume core, redox

conditions are moderately reducing and methanogenic,

and again become more oxidizing at the furthest

downgradient location, where site ground-water oxygen

concentrations are highest.

The historical COC data demonstrated that in most

wells, COC concentrations have decreased over the 8

years since monitoring began, while a suite of degradation

products provided conclusive evidence for dechlorination.

The results provide evidence that the geochemical

environment is conducive for mineralization, while the

analytical chemical data demonstrate that MNA

represents a viable alternative to reach the final remedial

goals.

Closure of Bald Mountain Heap #1

Placer Dome is in the process of obtaining approval

from the Nevada Division of Environmental Protection

(NDEP) to close heap leach Pad #1 at the Bald Mountain

Mine (BMM), located in White Pine County, Nevada.

Specifically, the closure includes:

♦ Applying heap effluent to site soils and possibly

increasing the application rate by enhancing

evaporation,

♦ Re-grading the pad to stabilize the pad slopes for

erosion control and to control runoff, and

♦ Covering the heap with site soils and revegetating to

minimize runoff and infiltration of meteoric waters,

while meeting risk objectives for plants, livestock,

and wildlife.

Groundwater in the vicinity of the site is

approximately 500 feet below the surface (Hydro-Search,

1981), and is therefore unlikely to be impacted at

reasonable application rates.

Water quality in the barren and pregnant ponds

associated with Pad #1 are shown for constituents that

have on occasion exceeded standards since 1998.

Concentrations of a few constituents have decreased since

1998 to below or near NDEP standards (CN, Hg, Ni),

while others have consistently remained above the

standards (Sb, As, NO3

-NO2, Se, SO4, pH).
56

A batch attenuation test indicated Langmuirean

solute behavior, from which n and Kd

 were obtained. For

some compounds (e.g., SO4) sorption is minimal, while

for others (e.g., As) sorption will effectively attenuate

solutes in the soil horizon.

Results from the speciation analyses for As, Se, and

Sb show that the predominant species are As(V), Se(VI),

and Sb(V). For As and Sb, the predominant valence state

is the less toxic of the forms, while Se (VI) is of similar

toxicity to Se (IV). The predominant species detected for

As, Se, and Sb are consistent with the generally oxidizing

and alkaline geochemical conditions measured for the

heap effluent. For example, As(V) constituted >80% of

the total As, Se(VI) was >90% of the total Se, and Sb(V)

was >90% of the total Sb analyzed. The total solute

concentration based on the sum of the two valence states

is also in good agreement with the total solute

concentration reported by SVL Analytical. The effect of

attenuation on solute concentration where solute lost from

solution by sorption to soils is illustrated by the difference

between the sorption data and the theoretical 1:1

equivalence curve.

Conclusions Conclusions

Redox can play a significant role in modifying metal

solubility. The applications are far-reaching, ranging from

incorporation in human health and ecological risk

assessments, the development of novel methods to

remediate sites, to demonstration of natural attenuation.

However, while understanding and applying redox

strategies is important, the evolving technology must be

applied within a regulatory framework that will allow

their incorporation into the panoply of environmental data

interpretation and restitution. At the same time, it is

incumbent on industry to demonstrate that redox

strategies really do provide meaningful and realistic

solutions to environmental problems.

References

Davis, A. and Olsen, R. L. 1995. The geochemistry of

chromium migration and remediation in the

subsurface. Ground Water, v. 33, pp. 759-768.

DiToro, D., Mahoney, J., Hansen, D., Scott, K., Hicks, M.,

Mayr, S., and Redmond, M. 1990. Toxicity of

cadmium in sediments: The role of acid volatile

sulfide. Environmental Toxicology and Chemistry,

v. 9, pp. 1487-1502.

DiToro, D., Mahoney, J., Hansen, D., Scott, K.,

Carlson, A., and Ankley, G. 1992. Acid volatile

sulfide predicts acute toxicity of cadmium and nickel

in sediments. Environmental Science and

Technology, v. 26, pp. 96-101.

PTI. 1997. Evaluation of Reactive Barrier Effectiveness

Batch and Column Tests at the Wilmington,

Massachusetts Site. December, 1997.

Rickard, D. T. 1995. Kinetics of FeS precipitation: Part

1. Competing reaction mechanisms. Geochimica et

Cosmochimica Acta, v. 59, pp. 4367-4379.
57

Group Discussion Summaries oup Discussion Summaries

Six breakout sessions were organized to summarize and evaluate key considerations during assessment and

implementation of remedial strategies for redox sensitive contaminants. Three concurrent sessions took place

in the morning and afternoon of the second day of the Workshop. The focus of the three morning sessions

centered on identification of approaches and parameters necessary to evaluate remediation of petroleum

hydrocarbons, chlorinated hydrocarbons, MTBE, and inorganics. The focus of the three afternoon sessions

centered on aqueous, gas phase, and solid phase measurement issues. Items emphasized during the afternoon

sessions, in part, depended on the outcome of the morning sessions.

Before arriving at the workshop participants were delivered a set of questions that along with the Plenary

Presentations, served as a launch point for the small group discussions. These questions are provided below.

Session 1. Redox processes in petroleum hydrocarbon remediation

(R. Ludwig, M. Barcelona, A. Azadpour-Keeley, J. Barker, P. Bradley, I. Cozzarelli, R. Ford, S. Warner,

E. Hartzell, D. Macalady, K. O’Reilly, K. Piontek, D. Vroblesky, D. Kampbell)

What remediation processes specific to petroleum hydrocarbons require the need for redox measurements

during site characterization and/or remediation (e.g., MNA, electron acceptor addition, thermal treatment,

surfactant flushing, etc)? Do we need to distinguish between categories of petroleum hydrocarbons (e.g.,

aromatics, PAHs, alkenes, alkanes, etc.)? What specific redox measurements are required (e.g., Eh, pH, DO,

Fe2+, etc) and how are the measurements used (e.g., determining feasibility vs. non-feasibility of specific

remedial processes, predicting rates of clean-up, determination of enhancements that could potentially be used

to accelerate clean-up, etc.)?

Session 2. Redox processes in chlorinated hydrocarbon and MTBE remediation

(J. Wilson, F. Chapelle, J. Amonette, C. Braun, J. Ingle, S. Jones, R. Lee, M. Lorah, C. Macon, P. Groeber,

R. Pirkle, T. Sivavec, J. Szecsody, C. Tebes Steven, J. Washington, E. Weber, T. Wiedemeier, B. Wilson)

What remediation processes specific to chlorinated hydrocarbons and MTBE require the need for redox

measurements during site characterization and/or remediation (e.g., MNA, enhanced bioremediation, zero

valent metal treatment, in-situ oxidation, in-situ redox manipulation, etc.)? Do we need to distinguish between

categories of chlorinated hydrocarbons (e.g., chlorinated methanes, chlorinated ethenes, chlorinated ethanes,

etc.)? What specific redox measurements are required (e.g., Eh, pH, DO, Fe2+, etc.) and how are the

measurements used (e.g., determining feasibility vs. non-feasibility of specific remedial processes, predicting

rates of clean-up, determination of enhancements that could potentially be used to accelerate clean-up, etc.)?

Session 3. Redox processes in inorganic remediation

(R. Puls, B. Deutsch, J. Westall, S. Benner, A. Davis, J. Fruchter, J. Herman, P. Longmire, D. Luce,

R. McGregor, D.K. Nordstrom, C. Ptacek, S. Schmelling, J. Shockley, C. Su, R. Wilkin, D. Willey, K. Cook)

What remediation processes specific to inorganic contaminants require the need for redox measurements

during site characterization and/or remediation (e.g., MNA, zero valent metal treatment, in-situ redox

manipulation, etc.)? What parameters must be measured during site assessment and remediation? Are the

required parameter measurements universal or contaminant-specific? Is there a universal set of parameters that

is sufficient for all remedial strategies? What specific redox measurements are required (e.g., Eh, pH, DO, Fe2+,

etc) and how are the measurements used? Is there a need to evaluate redox characteristics of the solid matrix

in addition to the mobile phase?
58

Session 4. Measurement Issues: Geochemical Parameters

(R. Wilkin, C. Ptacek, A. Azadpour-Keeley, B. Deutsch, J. Fruchter, J. Ingle, P. Longmire, M. Lorah, D. Luce,

K. Cook, D. Macalady, D.K. Nordstrom, T. Sivavec, C. Su, J. Westall, B. Wilson, J. Wilson, T. Wiedemeier,

K. Piontek)

Participants in this breakout group will identify and discuss the techniques used to determine the concentrations

of key redox indicator species, such as Dissolved Oxygen, Ferrous Iron, and Hydrogen Sulfide, as well as the

Oxidation-Reduction Potential (ORP/Eh). Group discussion will focus on identifying the best practices for

sample collection, sample handling, documentation, and measurement techniques for these key parameters.

We will attempt to identify rigorous QA/QC procedures that can be adopted in the field, and determine how

these parameters may be practically applied to understand the migration and evolution of contaminant plumes.

Session 5. Measurement Issues: Dissolved Gases

(D. Kampbell, F. Chapelle, J. Barker, P. Bradley, C. Braun, S. Jones, R. Lee, R. Ludwig, P. McLoughlin,

K. O’Reilly, P. Groeber, R. Pirkle, J. Shockley, D. Vroblesky, S. Warner)

Participants in this breakout group will identify the acceptable methods for determining the concentrations of

dissolved gases in ground water. Discussion will be primarily focused on hydrogen, ethene/ethane, and

methane. Group discussion will center on the best practices for sample collection, sample handling, documentation, and measurement techniques for these gas species. A specific question to be discussed is how is gas

phase characterization used to recognize the footprint and dynamics of a contaminant plume?

Session 6. Measurement Issues: Solid Phase

(R. Ford, J. Herman, J. Amonette, M. Barcelona, S. Benner, I. Cozzarelli, A. Davis, C. Macon, R. McGregor,

R. Puls, S. Schmelling, J. Szecsody, C. Tebes Steven, J. Washington, E. Weber, E. Hartzell)

Participants in this breakout group will identify the key solid phase parameters relevant to redox-controlled

processes active during in-situ remediation of ground water. In addition, we will identify acceptable methods

for sample collection and characterization. Emphasis will be placed on solid-phase iron. An example question

to be addressed is what is the most appropriate method(s) for quantifying the fraction of Fe-bearing minerals

that function as electron acceptors during microbially mediated redox transformation of a contaminant?

Discussion will also address analytical characterization and the role of solid phase sulfur and carbon during

remedial assessment and implementation.
59

Session 1 Summary: Session 1 Summary:

Redox Pr Redox Processes in Petr ocesses in Petroleum Hydr oleum Hydrocarbon Site Characterization and

Remediation

Ralph Ludwig, Michael Barcelona, and Keith Piontek

1.1 Introduction

The objective of this session was to evaluate the value and potential role of redox parameter measurements in

petroleum hydrocarbon site characterization and remediation. (Redox parameter measurements are herein

defined as any parameter measurement that provides insight into the electron transfer status of an aquifer

system or the propensity of an aquifer system to donate or accept electrons.) General conclusions of the session

included:

1. The need to measure redox parameters at petroleum hydrocarbon impacted sites will depend on

site-specific questions that need to be answered and the remedial strategies that are under

consideration. Knowledge of the electron transfer (redox) status within an aquifer will not always

be required.

2. Collection of at least some redox parameters such as dissolved oxygen, dissolved (ferrous) iron,

and dissolved manganese can be very beneficial in the initial stages of site characterization to help

develop a conceptual model for the site and to aid in screening potential remedial strategies. The

success or failure of some remedial strategies can be impacted by the electron transfer status of the

aquifer.

3. Additional redox parameter measurements including dissolved nitrate/nitrite, dissolved sulfate,

methane, dissolved carbon dioxide (as measured by acidimetric titration), and oxidation-reduction

potential (e.g., electrode-based ORP measurements) may be useful particularly in cases where a

natural or enhanced bioremediation strategy is under consideration.

4. In cases where an electrode-based redox potential (ORP) measurement is made, the measurements

should be used for qualitative purposes only to support other redox parameter measurements in the

development of the site conceptual model. No direct effort should be made to link electrode-based

redox potential (ORP) measurements to specific electron acceptor pathways or to the speciation of

redox sensitive species in the aquifer.

5. Redox parameter measurements should not be used as surrogate measurements to reduce the

number of hydrocarbon monitoring locations or reduce the frequency of hydrocarbon sample

collection. Redox parameter measurements should be used strictly in support of direct hydrocarbon

measurements to better understand (conceptualize) subsurface processes and plume behavior at the

site.

1.2 Role of Redox Parameter Measurements

Extensive research addressing petroleum hydrocarbon behavior and fate in ground water has been conducted

over the past twenty years. Various in situ and ex situ technologies for remediating petroleum hydrocarbon

impacted sites have been developed and utilized with varying degrees of success. Evaluation of the feasibility

of using one or more remedial technologies at a particular site may in some cases benefit significantly from

knowledge of the redox conditions in the subsurface. Knowledge of specific redox parameters and conditions

at a site can aid in establishing baseline conditions from which the existing behavior or anticipated future

behavior of a hydrocarbon plume can be evaluated. In some cases, redox parameters may also be used as a

cost-effective means of evaluating the performance of specific remedial technologies.
60

Three key questions will generally need to be answered in assessing the role of redox parameter measurements

at petroleum hydrocarbon impacted sites.

1. Will knowledge of the redox status in the aquifer be of any value in evaluating the feasibility of

potential remedial strategies at the site?

2. If so, which redox parameter measurements might be useful?

3. Will the redox parameter measurements be used for qualitative or quantitative purposes?

Knowledge of the redox status of a hydrocarbon impacted aquifer system could certainly be important should

in-situ bioremediation technologies be under consideration. Specifically, the presence/absence and distribution

of electron acceptors can have important implications with regard to biodegradation pathways and plume

behavior. Petroleum hydrocarbons (aromatics, PAHs) as well as coal-derived hydrocarbons (e.g., cresols,

phenols) generally undergo much more rapid and efficient biodegradation under aerobic (oxidizing) conditions. Thus, knowledge of the presence/absence of oxygen (and other potential electron acceptors) provides

insight into existing electron and biological activity and the means by which bioactivity can potentially be

enhanced. Knowledge of redox conditions can also be important in evaluating the biodegradation potential of

gasoline additives such as methyl-tert butyl ether (see Session 2 Summary).

The potential use of redox parameter measurements in the evaluation of technologies other than bioremediation

for hydrocarbon impacted aquifer systems will not be so obvious and will generally need to be considered on

a case by case basis. For example, in the case where a thermal based remediation technology might be the

preferred remedy, redox parameter measurements are not likely to be of any value. Alternatively, if a remedy

involving delivery of oxygen to the subsurface is to be considered, then knowledge of specific redox

parameters (e.g., dissolved iron and manganese concentrations) in the ground water and their potential impact

on system performance can be important.

The use of redox data for qualitative or quantitative purposes will depend on how the redox data are to be

utilized. If the objective is simply to determine which redox processes might be dominant at a particular

location, then establishment of the relative presence/absence of selected electron acceptors within and outside

a hydrocarbon plume will generally suffice. However, if an effort is being made to answer questions regarding

electron acceptor capacity and longevity (e.g., for a monitored natural attenuation remedy), then a quantitative

approach may be needed. This would require a more rigorous and defensible (and generally more sophisticated) data collection and analyses process than would be required for a qualitative analysis. The significant

spatial and temporal variability of many redox parameters can make quantitative analyses at sites extremely

challenging.

1.3 Site Characterization 1.3 Site Characterization Applications Applications

The need to collect redox parameter measurements at petroleum hydrocarbon impacted sites will depend on the

site-specific questions that need to be answered. The most important information regarding petroleum

hydrocarbon plume migration and behavior will generally be gained from direct knowledge of both the

temporal and spatial distribution of specific contaminants of concern (e.g., BTEX compounds). Redox

parameter measurements are best used as secondary lines of evidence for locating residual source areas and for

better understanding, predicting, and/or confirming dissolved hydrocarbon plume migration and behavior. It is

known that a strong correlation generally exists between petroleum hydrocarbon concentrations and the

relative distribution of various electron acceptors and reduction products including dissolved oxygen, ferrous

iron, manganese, nitrate, sulfate, and methane. This correlation results from the utilization of electron

acceptors in a favored sequential order by microorganisms during oxidation (metabolism) of hydrocarbon

compounds. The use of oxygen as an electron acceptor allows for the greatest energy gain (and thus the most

efficient and rapid degradation) during oxidation of hydrocarbons by microorganisms. Thus, as long as oxygen

is available, oxygen will be the predominant and preferred electron acceptor utilized by microorganisms.

However, once oxygen is depleted, microorganisms must turn to less favorable electron acceptors, which

results in the oxidation process becoming increasingly less efficient. The relative relationships between

petroleum hydrocarbon concentrations, electron acceptors, and redox potential are illustrated in Figure 1.1. It

should be noted that the relative relationships are equilibrium tendencies and may or may not be correlated

directly.

Collection of selected redox parameter measurements in the initial stages of site characterization may aid in

establishing the feasibility of one or more technologies at a given site as well as evaluating the stability of the
61

area of affected ground water. For example, the absence of oxygen, presence of high levels of ferrous iron, and

low ORP readings in the core of a BTEX plume at a site (relative to background locations) can suggest strong

bioactivity and therefore indicate conditions that might be conducive to an aerobic bioremediation-based

remedial strategy. Low Eh ground water with a high dissolved iron and manganese content (which can lead to

clogging of pumping wells and above ground treatment systems) could signify greater maintenance attention

should an ex-situ remedial strategy such as pump and treat be considered. Low Eh ground water with a high

dissolved iron and manganese content could also present problems for in situ systems such as air sparging. The

collection of redox information can thus be important in contributing to the initial screening and evaluation of

candidate remedial technologies.

Redox parameters that should be measured both within and outside the plume in the initial stages of site

characterization should at a minimum include dissolved oxygen, ferrous iron, and manganese. ORP measurements may also be considered, but should be used solely as a secondary line of evidence in support of these

more reliable redox parameter measurements. If bioremediation is to be considered as a candidate remedy,

then the redox parameter measurements should be expanded to include nitrate/nitrite, sulfate, and in some

cases, alkalinity.

Measurement of dissolved oxygen, dissolved (ferrous) iron, and manganese within a hydrocarbon plume

relative to background concentrations at the site will provide an indication of whether biodegradation processes

are in effect and consequently whether the aquifer system is amenable to natural or enhanced bioattenuation.

Measurement of these parameters will also provide an indication of design features that may need to be

incorporated into candidate in situ and ex situ treatment systems to avoid potentially adverse reactions (e.g.,

precipitation). Measurement of nitrate/nitrite and sulfate/sulfide will provide an indication of additional

electron acceptor availability/capacity within the aquifer while alkalinity measurements can be used as a

TPHBTEX 

Oxygen

O2→H2O

 Nitrate 

NO3

-

→N2

 Manganese

Mn4+→Mn2+

 Iron

Fe3+→Fe2+

Sulfate

SO4

2-→H2S

CO2

CO2→ CH4

Redox 

Potential 

 

Low High High Low Low High Low High 

 O2 NO3

- Mn2+ Fe2+ SO42- CH4

High Low Low High High Low High Low 

Order of electron acceptor use during biodegradation 

 Relative aqueous-phase concentrations 

Figure 1.1 Relative relationships between TPH-BTEX concentrations and electron acceptor/reduction product

concentrations and redox potential in ground water within a petroleum hydrocarbon plume.
62

measure of microbial CO2 production during biodegradation in the absence of other potential sources of

alkalinity. In addition to these parameters, field measurements should generally include temperature, pH, and

conductivity to further determine the extent to which the ground water may or may not be favorable to selected

treatments.

Specific examples where redox parameter measurements may play a useful role in site characterization

activities at petroleum hydrocarbon contaminated sites include the following.

Presence/Absence of Biodegradation Presence/Absence of Biodegradation Activity

The magnitude of biodegradation activity in ground water may be an important factor in the selection of

remedial approaches such as enhanced bioremediation or monitored natural attenuation. Biodegradation byproducts (e.g., degraded hydrocarbons, enzymes, residual inorganic constituents) of petroleum hydrocarbons

are not easily quantified in ground water. Thus, there is a need to rely on alternative evidence such as the

presence or absence of electron acceptors and reduction products to support the presence/absence of biodegradation activity at sites. For example, the absence of oxygen and presence of dissolved (ferrous) iron in a zone

of ground-water contamination provides evidence that biodegradation is occurring. Increases in alkalinity may

be indicative of carbon dioxide production associated with microbial respiration.

Petroleum Hydrocarbon Source Petroleum Hydrocarbon Source Area Delineation

Decreasing dissolved oxygen, nitrate, and sulfate concentrations and increasing manganese and ferrous iron

concentrations generally coincide with increasing proximity to hydrocarbon source areas. These measurements, as well as methane concentrations and ORP measurements, can be used as secondary lines of evidence

to complement TPH and BTEX concentration measurements in verifying the locations of petroleum hydrocarbon source areas.

Dissolved Plume Delineation

Higher dissolved oxygen and nitrate concentrations and lower dissolved (ferrous) iron and manganese

concentrations will generally be more characteristic of ground water outside the limits of a dissolved petroleum

hydrocarbon plume. Within the limits of a dissolved plume, dissolved oxygen and nitrate concentrations will

generally decrease and dissolved iron and manganese will increase with increasing proximity to the dissolved

contaminant center of mass. These measurements can potentially be used to complement TPH and BTEX

measurements in verifying the limits of dissolved plumes and the locations of dissolved contaminant center of

masses.

Dissolved Contaminant Plume Behavior Dissolved Contaminant Plume Behavior

Shrinking dissolved contaminant plumes would be expected to show gradually increasing oxygen and nitrate

concentrations and decreasing dissolved (ferrous) iron and manganese concentrations around their peripheries

with time. In contrast, expanding dissolved contaminant plumes would be expected to show gradually

decreasing oxygen and nitrate concentrations and increasing dissolved (ferrous) iron and manganese concentrations at their peripheries with time. These measurements can potentially be used to complement TPH and

BTEX measurements in determining whether source areas and dissolved plumes are shrinking, expanding, or at

steady state.

1.4 Remedial 1.4 Remedial Technology echnology Applications Applications

Measurement of selected redox parameters in ground water may have application in determining the feasibility

of utilizing specific in-situ and ex-situ remedial treatments and in evaluating the performance of certain

remedial systems once implemented. Technologies for remediating petroleum hydrocarbon impacted ground

water currently include ex-situ treatments such as pump-and-treat and in-situ treatments focused on enhanced

biodegradation or monitored natural attenuation. Technologies often considered for remediating residual phase

and non-aqueous phase petroleum hydrocarbon contamination in the subsurface currently include in-situ

surfactant/co-solvent flushing, in-situ thermal treatment, and in-situ oxidation. Examples of the potential

applications of redox parameter measurements in determining the feasibility and evaluating the performance of

remedial systems are presented below.

Determination of Natural Bioattenuation Potential Determination of Natural Bioattenuation Potential

Dissolved oxygen, iron, nitrate, and sulfate concentrations can be used to indicate the presence/effect of natural

bioattenuation processes on the ambient hydrochemical system. High dissolved (ferrous) iron concentrations
63

(e.g., > 1 ppm) coupled with low dissolved oxygen levels (e.g., < 1 ppm) near a source area or within a

dissolved plume would generally indicate active bioattenuation. In contrast, sustained elevated dissolved

oxygen levels and low dissolved (ferrous) iron levels near a source area or within a dissolved hydrocarbon

plume could indicate the absence of natural bioattenuation processes. Estimation of the long-term capacity of

an aquifer system to naturally attenuate petroleum hydrocarbons could be attempted by determining average

up-gradient electron acceptor fluxes into the impacted zone and quantifying the aquifer solid phase “bioavailable”

iron present within the impacted zone. The instantaneous BTEX biodegradation capacity of an aquifer system

can be approximated based on the differences in up-gradient dissolved electron acceptor concentrations relative

to dissolved electron acceptors and by-products near a petroleum hydrocarbon source area using the following

relationship (U.S. EPA, 1996).

Biodegradation capacity (mg/L) =

[(O2 concentration up-gradient - O2

 concentration in source area)/3.14)] +

[(NO3

- concentration up-gradient - NO3-

 concentration in source area)/4.9)] +

[(SO4

2- concentration up-gradient - SO42- concentration in source area)/4.7)] +

[Fe2+ concentration in source area/21.8] +

[CH4

 concentration in source area/0.78]

This relationship for BTEX biodegradation is the basis of EPA’s BIOSCREEN model used as a screening tool

to assess the potential role of natural bioattenuation at petroleum hydrocarbon contaminated sites.

It should be noted that although solid-phase “bioavailable” iron may serve as a major long-term electron

acceptor source, it may become depleted with time. If the natural attenuation process is highly dependent on

the presence of solid-phase ferric iron as an electron acceptor source, the eventual depletion of the iron source

could result in further expansion of a hydrocarbon plume that might have initially been determined to be stable

or receding. Quantification of “bioavailable” solid phase iron is addressed in the Session 6 Summary.

Performance Monitoring of Remedial Treatment Systems reatment Systems

Redox parameters such as dissolved oxygen measurements can be used to determine zones of influence for in

situ bioenhancement treatments involving delivery of oxygen into an aquifer. Dissolved oxygen measurements

coupled with pre-treatment and post-treatment redox potential and dissolved (ferrous) iron measurements, for

example, can provide evidence of whether oxidizing conditions favorable for petroleum hydrocarbon biodegradation are being established and maintained. In cases where a monitored natural attenuation process has been

implemented, redox parameter measurements (e.g., dissolved oxygen, dissolved (ferrous) iron) within and

around the periphery of the plume can be used as supporting evidence to determine whether the plume is stable

or is shrinking. This should be supplemented by solid phase ferric iron analysis to ensure adequate long-term

electron acceptor capacity/longevity within the impacted aquifer.

Evaluation of Potential Impacts on Ex-situ Treatment Systems reatment Systems

Some ex-situ ground-water treatment systems may be adversely impacted by the presence of iron, manganese,

and/or carbonates. This could be the case for any treatment system involving a significant change in redox

potential, pH, or temperature during treatment of the ground water. Vulnerable components of treatment

systems could include filters, air strippers, bioreactors, and re-injection systems. For example, an increase in

redox potential of influent water during air stripping as a result of contact with oxygen can induce precipitation

of dissolved (ferrous) iron and manganese, which may ultimately clog the stripper. Measurement of dissolved

(ferrous) iron, manganese, alkalinity, and dissolved oxygen in ground water supplemented by redox potential

measurements can provide essential information needed in the selection and proper design of ex-situ treatment

systems.

Evaluation of Potential Impacts on In-situ Treatment Systems reatment Systems

Some in-situ treatment systems may be adversely impacted by various precipitation reactions or dissolution

reactions that occur in response to changes in redox potential and pH. For example, the hydraulic conductivity

of in-situ aquifer treatment zones that involve a change in redox potential or pH may decrease significantly over

time as a result of specific precipitation or mineralization reactions that occur. Air sparging systems, for

example, may induce precipitation of iron and manganese as a result of an increase in redox potential brought

about by the introduction of oxygen into the aquifer. Measurement of parameters (see Session 4 Summary)
64

1 Only saturated zone remedial technologies are considered. Surface soil and vadose zone remedial technologies (e.g., bioventing) are not

addressed.

2 Includes data needed for remediation technology assessment and/or design.

3 Characterization of ground-water geochemistry is important to the design of virtually any remedial system involving extraction of ground

water. Fouling of ground-water treatment and handling systems with iron precipitates, scale, and other solids is a significant design

consideration. Potentially vulnerable systems include filters, strippers, bioreactors, re-injection systems (e.g., wells, french drains, etc.) and

ancillary components (e.g., in-line monitors such as flow meters are particularly vulnerable).

4 Total iron loading is important, but it is also important to quantify dissolved (ferrous) iron to assess how much iron precipitation can be

avoided by maintaining reduced conditions in the treatment system.

5 “Additional parameters” refers to important ground-water geochemical parameters not typically considered redox parameters.

6 Performance monitoring requirements will depend on system design, but will typically include analyses for one or more of the parameters

specified under site characterization. However, performance monitoring is typically performed on extracted ground water rather than the in

situ ground-water system.

Table 1.1 Ground-water Redox Monitoring Guidance for Petroleum Hydrocarbon Site Characterization and

Remediation

Ground-water Redox Monitoring Requirements Remediation Technology1

Site Characterization2

 Performance Monitoring 

Ground-water extraction with 

treatment and/or re-injection3

 

Iron speciation4

, manganese. 

Additional parameters5

: TSS, 

hardness 

Variable6

Monitored natural attenuation 

(MNA) 

Protocol parameters7 (e.g., ASTM 

E-1943-98) Variable8 

Air sparging (mass removal in 

source zone)9 Variable10 DO and ferrous iron11 

In-situ aquifer treatment zones 

(e.g., permeable reactive 

barriers)12

Ferrous iron and manganese Variable depending on specific 

technology 

Enhanced bioremediation – 

oxygen addition 

Protocol parameters13

Additional parameters: solid 

phase oxygen demand 

DO, ferrous iron, TIC 

Enhanced bioremediation – 

addition of other electron 

acceptors 

Protocol parameters13 TIC, others14

In-situ chemical oxidation 

Protocol parameters 

Additional parameters: TPH, 

solid phase COD, and solid 

phase iron mineralogy15 

DO, ferrous iron, TIC 

Additional parameters: vadose 

zone CO2 

Surfactant flooding16 

None 

Other parameters: TDS, major 

ions17

Variable18

Solvent flooding16 Variable19 Variable 

Thermal methods (e.g., steam 

injection) None None 

Phytoremediation Research needed20 Research needed 

such as dissolved oxygen, dissolved (ferrous) iron, and manganese, and alkalinity in ground water can be used

in conjunction with geochemical modeling to predict the nature and extent of precipitation/dissolution

reactions that can be expected to occur during treatment.

1.5 Summary

The need for redox parameter measurements in the characterization and remediation of petroleum hydrocarbon

impacted aquifer systems will vary based on the site-specific questions that need to be answered. A summary

of potential site characterization and performance monitoring redox parameter needs for a number of different

remediation technologies is presented in Table 1.1.
65

7 A variety of protocols describe the range of redox parameters that should be included in characterization of natural attenuation at petroleum

hydrocarbon sites. The first round of ground-water monitoring should typically involve monitoring for a relatively complete spectrum of redox

parameters. The suite of analytical parameters used in subsequent monitoring can be optimized based on the initial findings.

8 This should be assessed on a site-specific basis.

9 Air sparging can also be used to create a “treatment wall” (with stripping and/or aerobic biodegradation providing the treatments) to

minimize volatile hydrocarbon plume migration. For this air sparging application, also refer to the monitoring requirements associated with

“in-situ aquifer treatment zones” .

10 If the intent of the application is stripping of volatile constituents of concern, no redox characterization is required. If the intent is to promote

aerobic biodegradation (for non-volatile constituents that will not be “stripped” from the ground water), it is prudent to confirm the premise

of oxygen-limited conditions through analyses for MNA protocol parameters.

11 Performance monitoring will typically include assessment of the zone of sparging influence. This can be assessed through determination of

spatial and temporal changes in dissolved oxygen (DO) and ferrous iron (a decrease in ferrous iron can be used as a supplemental indicator

of aerobic conditions).

12 Includes any treatment technology relying on ground-water flow through a subsurface zone in which ground-water composition/conditions are

altered (e.g., continuous wall systems, funnel and gate systems, sparging barriers, ground-water circulation wells, etc.)

13 When careful “mapping” of naturally occurring biological activity is critical to system design, use of hydrogen analyses in site characterization should be considered.

14 Performance monitoring requirements will be a function of the specific electron acceptor used. For example, performance monitoring for

nitrate addition will typically include nitrogen speciation. See Session 2 Summary for definition of total inorganic carbon (TIC).

15 TPH measurement is used in calculating the upper boundary on aqueous phase oxygen demand. COD = chemical oxygen demand.

16 Focuses on the data needs associated with the in situ aspects of the technology. This technology typically includes ground-water extraction,

treatment, and/or re-injection; also refer to the data needs specified for this category of remediation technologies.

17 Surfactant system design/effectiveness (e.g., interfacial tension reduction, potential surfactant precipitation) is a function of ground-water

geochemistry. TDS = total dissolved solids.

18 Driven by the chemistry of the specific surfactant system being used.

19 There may be no need for redox data to assess the feasibility or design of a solvent flushing system. However, MNA will typically be used as a

“polishing step” to deal with residual constituents of concern and the solvent. Also, use of biodegradable solvents may drastically lower the

ground-water redox potential.

20 Given the focus of this document on ground-water geochemistry, the phytoremediation application of interest is the use of trees or other

vegetation to serve as solar driven ground-water extraction systems. From a practical perspective, since this phytoremediation application is

most often implemented as an “incremental boost” to natural attenuation processes, the ground-water redox will already have been characterized as part of the natural attenuation assessment. However, research is needed to allow interpretation of the effects of ground-water redox

conditions on potential phytoremediation effectiveness.

References

American Society for Testing and Materials (ASTM). 1995. Standard guide for risk-based corrective action

applied at petroleum release sites, ASTM E 1739-95. West Conshohocken, Pennsylvania.

American Society for Testing and Materials (ASTM). 1998. Standard guide for remediation of ground water by

Natural Attenuation at petroleum release sites, ASTM E 1943-98. West Conshohocken, Pennsylvania.

Chapelle, F. H. 1999. Bioremediation of petroleum hydrocarbon-contaminated ground water: The perspectives

of history and hydrology. Ground Water, v. 37, pp. 122-132.

U. S. Environmental Protection Agency. 1988. Cleanup of Releases from Petroleum USTs: Selected Technologies,

EPA/530/UST-88/001.

U. S. Environmental Protection Agency. 1996. BIOSCREEN – Natural Attenuation Decision Support System,

User’s Manual Version 1.3, EPA/600/R-96/087.
66
67

Frank Chapelle, John Wilson, and Ralph Ludwig

2.1 Introduction

The purpose of this session was to develop recommendations concerning redox measurement parameters for

remediation of ground water contaminated with chlorinated solvents and MTBE. Information on redox

processes is necessary for site managers to plan and implement site characterization, select remedies, and

optimize and monitor the performance of the selected remedies. The participants in the session made

recommendations concerning the redox parameters that should be evaluated for various treatment technologies,

and on the appropriate methods to collect the data.

The technologies include various techniques for destruction of the contaminants, including physical removal,

chemical destruction, enhanced biological degradation, and monitored natural attenuation. The information

needs for each technology are discussed within these categories.

2.2 Critical Redox Parameters Required for the Evaluation and Optimization of Dif the Evaluation and Optimization of Different

Treatment eatment Technologies for Chlorinated Hydrocarbons ocarbons

The information needed to characterize a site and select a remedy is often different from information needed to

optimize a technology or monitor its performance. The tables in this section present a brief description of the

major issues associated with redox parameters, grouped by major category of technology.

The "Technical Protocol for Evaluating Natural Attenuation of Chlorinated Solvents in Ground Water"

(Wiedemeier et al., 1998) presents a comprehensive list of redox parameters that are useful for evaluating

monitored natural attenuation in ground water. The parameters are listed in Table 2.1 of the Technical

Protocol, and include alkalinity, arsenic, chloride, conductivity, ferrous iron, hydrogen, manganese, methane,

nitrate, oxidation reduction potential (ORP), oxygen, pH, sulfate, sulfide, temperature, and total organic carbon

in water.

Some of the parameters are useful for assessment of the system redox status, including dissolved oxygen,

hydrogen, methane, sulfide, oxidation reduction potential, pH, and temperature. Some of the parameters are

useful for assessment of the potential for microbial metabolism of contaminants. Oxygen can support

microbial metabolism of certain chlorinated hydrocarbons including cis-dichloroethylene and vinyl chloride.

In contrast, the presence of sulfate, nitrate, or oxygen can inhibit anaerobic biological processes such as

reductive dechlorination of the chlorinated hydrocarbons. During reductive dechlorination, the contaminant

serves as the electron acceptor in microbial metabolism. If sulfate, nitrate or oxygen are present they can

compete with the contaminant for supplies of electron donors. Total organic carbon in water is a bulk

parameter that can be used to estimate the supply of electron donors for reductive dechlorination.

Other redox parameters in the EPA Technical Protocol are measures of the degradation process. These include

chloride and alkalinity. Chloride is an ultimate biodegradation product of chlorinated hydrocarbons. The

analysis for alkalinity in the Technical Protocol was originally intended as an estimate of the amount of carbon

dioxide that has been produced in the ground water by biological processes. In many aquifers, carbon dioxide

produced by biological processes cannot be measured directly due to reactions with carbonate minerals in the

aquifer matrix. Alkalinity measurements may also provide an overestimate of microbial CO2

 production in

contaminant plumes with high concentrations of titrable organic acids. To address this later issue, a better

estimate of the amount of carbon dioxide that has been produced is the change in total inorganic carbon (TIC)

in ground water. The TIC can be measured directly, or calculated from measurements of alkalinity, pH, and

dissolved carbon dioxide (see section 2.8).

Session 2 Summary: Session 2 Summary:

Redox Pr Redox Processes for ocesses for Remediation of Chlorinated Hydr Remediation of Chlorinated Hydrocarbons ocarbons and

Methyl tert-Butyl Ether Butyl Ether
68

Certain of the redox parameters can be used to establish the “footprint” of the contaminated water (National

Research Council, 2000). If the geochemical “footprints” are similar between contaminated ground water in

source areas and water produced from monitoring wells farther downgradient, they can be used to establish a

hydrological connection between the source and the monitoring well. If the geochemical “footprints” in down

gradient monitoring wells are distinctly different from the geochemistry of ground water at the source, they are

not in the flow path from the source area of the plume. Geochemical “footprints” can also reveal the major

biological and chemical oxidation/reduction processes that are responsible for removing chlorinated hydrocarbons from ground water. Chloride, methane, specific conductance, and TIC are particularly useful as

“footprints.”

Some of the redox parameters are required to evaluate risk from materials that were not the original

contaminants of concern. Arsenic and manganese are included because the geochemical conditions that allow

natural biodegradation of chlorinated hydrocarbons also allow biological reduction of natural aquifer materials

that contain arsenic and manganese. Under conditions that promote natural anaerobic biodegradation of

chlorinated hydrocarbons, the concentration of total arsenic and manganese produced from natural sources may

exceed drinking water standards.

The appropriate regulatory authorities will determine on a site specific basis when it is necessary to analyze for

arsenic and manganese. The circumstances where analysis for hydrogen is useful are discussed in Section 2.7.

The following redox parameters should be routinely applicable to most sites, and will be referred to as the

Routine EPA Protocol Parameters: dissolved oxygen, nitrate, sulfate, chloride, ferrous iron, sulfide, hydrogen,

methane, oxidation/reduction potential, pH, temperature, specific conductance, total inorganic carbon, and

total organic carbon.

Chlorinated hydrocarbons can be removed from aquifers by air sparging, bioventing or phytoremediation (see

Table 2.1). They can be chemically destroyed by dithionite reduction, oxidation by the Fenton reaction,

oxidation by permanganate, or by reduction by zero valence iron in a permeable reactive barrier (see Table 2.2).

They can be biologically degraded through substrate additions, phytoremediation in the rhizosphere, and

through monitored natural attenuation (see Table 2.3).

Air sparging is a process of stripping volatile compounds from ground water through the injection of air.

Redox chemistry is not involved in this purely physical process. In theory precipitation of Fe (III) by oxidation

of Fe (II) in ground water could plug the pore space in the aquifer. In practice this is not an important problem.

The success of air sparging is controlled by the physics of flow of the injected air as controlled by the geology

of the site.

Bioventing is technology applied to the unsaturated zone that is intended to treat ground water. It treats the

residual hydrocarbons present as a NAPL that act as a long term source of contamination to ground water. It

is successful at sites where the water table moves down during the time period of treatment to expose the oily

phase at residual saturation to air in the unsaturated zone. Treatment is a combination of volatilization and in

situ biodegradation. The recommended redox parameters are important to understand the biodegradation

processes.

Phytoremediation has particular application to chlorinated hydrocarbons in shallow ground water in fine

textured geological materials. Water is removed by evapotranspiration. The chlorinated hydrocarbons can be

transpired with the water, or they can volatilize into air spaces created when the water is extracted. The relevant

geochemical parameters determine whether plant roots can grow in the contaminated ground water (Pivetz,

2001).

In conventional practice for site characterization, the major emphasis is on sampling and analysis of ground

water. Much less attention has been paid to characterization of aquifer solids. An analysis of aquifer solids is

critical to the successful use of non-biological transformation technologies such as dithionite reduction,

permanganate oxidation or the Fenton reaction to treat organic contaminants. The natural aquifer solids play an

important role in the chemistry of the treatment process. The availability and reactivity of iron minerals in the

aquifer are particularly important.

This is less true for passive reactive barriers using zero-valent iron in the treatment process. Only ground water

is treated, and it is treated in the reactive wall as opposed to the aquifer. Passive reactive barriers are designed

to allow adequate residence time of contaminated ground water to reduce the contaminant concentrations to the

design goals. The design is constrained by the kinetics of reaction of the contaminants of concern with the

reactive medium in the barrier, and with the seepage velocity of ground water through the barrier. In theory,
69

the geochemistry of the ground water should directly affect the long term capacity of a barrier to treat water. As

an example, reaction with oxygen in ground water converts zero valent iron to higher oxidation states. The

reaction of zero valence iron with oxygen and contaminants raises the pH. A variety of reaction products can

form precipitates that may plug the pore space. These concerns are discussed in more detail in Powell et al.

(1998) on pages 25 and 26. At present there is a research program coordinated between private industry and

several federal agencies to evaluate the life cycle of passive reactive barriers. However, the present state of

practice is to design them to work well, but not necessarily to design them to last for a predetermined time. As

a consequence, geochemical considerations are less important to their design and implementation.

In technologies that add substrates (such as vegetable oil, lactate, molasses, or hydrogen) to support reductive

dechlorination of chlorinated hydrocarbons, the success of the technology may be related to the availability of

mineral nutrients for growth of bacteria, in particular phosphate and reduced fixed nitrogen in the form of

ammonium ion. The primary source of reducing power is hydrogen, which is provided directly as dihydrogen,

or as organic substrates that can be fermented to form dihydrogen. Organisms that use nitrate and nitrite as

electron acceptors can effectively compete with organisms using chlorinated hydrocarbons for the available

pool of dihydrogen. The presence of measurable nitrate or nitrite will preclude or strongly inhibit reductive

dechlorination.

Chlorinated hydrocarbons can be biochemically degraded in soil or aquifer material in direct proximity to plant

roots (the rhizosphere). The oxidation parameters are used to predict microbiological activity in the

rhizosphere. In this application, total organic carbon is an indirect measure of plant root exudates that stimulate

microbial activity against the chlorinated hydrocarbons.

Monitored Natural Attenuation is an important remedial technique for chlorinated hydrocarbons. The EPA

protocol parameters have proved themselves useful as a qualitative tool to recognize geochemical environments where reductive dechlorination is possible (Wiedemeier et al., 1998). However, the role of alkalinity in

the original EPA protocol should be replaced with an analysis or calculation of TIC (see section 2.8).

Table 2.1 Redox Parameters for Evaluation of Treatment by Physical Removal

Treatment Technology Site Characterization Performance Monitoring/ 

Remediation Optimization 

Air Sparging None None 

Bioventing to remove 

oily phase residuals of 

chlorinated 

hydrocarbons mixed 

with petroleum 

hydrocarbons. 

Concentrations of oxygen and carbon 

dioxide in soil gas. Concentrations of 

methane in soil gas to define background 

conditions. An in situ respiration test to 

measure the instantaneous rate of oxygen 

consumption. 

Same 

Phytoremediation by 

evapo-transpiration of 

water. 

Will plant roots grow in the contaminated 

ground water? 

Conductivity, pH, Dissolved Oxygen, 

Sulfide. 

None 
70

Table 2.2 Redox Parameters for Evaluation of Chemical Destruction

Treatment Technology Site Characterization Performance Monitoring/ 

Remediation Optimization 

Dithionite Inherent stability of Dithionite in the 

aquifer matrix. Dissolved Fe(III) 

extractable with 1 N HCl. Mineralogy of 

Iron. 

ORP, reducing capacity of 

sediments measured as the 

demand for dissolved oxygen. 

Other Oxidation 

Processes 

Fenton=s reaction 

Permanganate 

Routine EPA Protocol Parameters, 

Estimate volume of aquifer material in 

source area contaminated with organic 

materials. Permanganate demand of 

contaminated material. 

Routine EPA Protocol 

Parameters, carbon dioxide in 

soil gas, Residual permanganate 

using colorimetric test. 

Zero Valent Iron Not Critical Dissolved Fe (II), dissolved 

oxygen, sulfide, sulfate, ORP, 

conductivity, pH 

(measured either quarterly or 

semi-annually) 

Table 2.3 Redox Parameters for Evaluation of Biological Degradation

Treatment Technology Site Characterization Performance Monitoring/ 

Remediation Optimization 

Substrate Additions 

Vegetable Oil 

Lactate 

Molasses 

Hydrogen 

Routine EPA Protocol Parameters 

hydrogen, nitrite, ammonium ion, 

phosphate. 

(Have the redox processes 

changed as a result of the 

substrate addition?) 

Routine EPA Protocol 

Parameters, acetate, formate, 

propionate, hydrogen, nitrite, 

ammonium ion, phosphate. 

Phytoremediation 

through 

biodegradation in the 

Rhizosphere 

Routine EPA Protocol Parameters with 

emphasis on Total Organic Carbon. 

Routine EPA Protocol 

Parameters with Emphasis on 

change in oxidation/reduction 

status. 

Monitored Natural 

Attenuation of 

Chlorinated Ethenes 

Chlorinated Ethanes 

Chlorinated Benzenes 

Routine EPA Protocol Parameters plus 

Potential Daughter Products. Hydrogen if 

necessary to interpret biodegradation 

processes. 

Routine EPA Protocol 

Parameters plus Potential 

Daughter Products. Hydrogen 

if necessary to confirm 

continued biological activity. 
71

2.3 Critical Redox Parameters for MTBE

Methyl-tert-butyl-ether (MTBE), an oxygenate added to petroleum hydrocarbon fuels to produce cleaner

engine emissions, has been increasingly detected in ground water throughout the United States since the early

1990s. The increased presence of MTBE in ground water has been of particular concern due to the high

mobility, relative recalcitrance, and suspected carcinogenic properties of this compound.

Although initially believed to undergo negligible or no biodegradation in the subsurface, there is an increasing

body of evidence indicating that aerobic biodegradation can be used to remove MTBE from ground water.

Salinitro et al. (2000) used bioaugmentation and aerobic in situ bioremediation to reduce MTBE in an aquifer

to concentrations below 10 μg/liter. Redox parameters that are important for design of in-situ aerobic MTBE

bioremediation systems include dissolved oxygen, sulfide and ferrous iron in ground water. Oxygen

enhancement will stimulate biodegradation of MTBE if the original ambient oxygen concentration is less than

1 mg/liter. The presence of high concentrations of ferrous iron can cause problems with precipitates and

plugging of pipes and distribution systems. The presence of ferrous iron or sulfide in ground water is often

indicative of a high non-biological demand for oxygen associated with precipitates of iron and sulfide in the

aquifer solids. Redox parameters important for monitoring in-situ aerobic bioremediation include dissolved

oxygen, ferrous iron, and TIC. Dissolved oxygen concentrations should exceed a minimum concentration of

2 mg/liter to support degradation of MTBE, and should exceed 5 mg/liter for optimum performance. The

increase in TIC over time can be used to monitor oxidation of the organic contaminants in the aquifer. In

particular, redox parameters will be useful in monitoring the performance (e.g., zone of influence) of any

in-situ enhanced bioremediation treatment system.

Bradley et al. (2001) have reported significant aerobic mineralization of MTBE in microcosm studies using

MTBE and non-MTBE contaminated stream and lake bed sediments from eleven sites throughout the

United States. The study indicated that even microbial communities indigenous to newly contaminated

sediments exhibited some innate ability to mineralize MTBE under aerobic conditions. The magnitude of

MTBE mineralization in the study was observed to vary inversely with the sediment grain size distribution.

Landmeyer et al. (2001) showed that aerobic metabolism of MTBE in the bed sediments of a stream prevented

transfer of MTBE from an anaerobic plume in ground water to aerobic surface water. Natural aerobic

biodegradation of MTBE in surface waters and sediments probably make a substantial contribution to surface

water quality.

Natural aerobic biodegradation of MTBE is probably not important for plumes of MTBE contamination in

ground water. Kolhatkar et al. (2000) examined seventy-four MTBE plumes from spills of gasoline in the

eastern United States. The ground water that contained MTBE contamination was strongly reducing. The most

contaminated wells at forty-three of the seventy-four sites had no detectable sulfate and an accumulation of at

least 0.5 mg/liter of methane. Only eighteen of the seventy-four sites had detectable concentrations of sulfate

and methane concentrations less than 0.5 mg/liter. Nitrate was depleted in contaminated wells at every site. In

the survey reported by Kolhatkar et al. (2000), water samples were collected according to the standard practice

for sampling monitoring wells at underground storage tanks as approved by the local regulatory authority. In

almost every case, the water samples were collected with bailers (see warnings in section 2.6). The water as

sampled from the contaminated wells contained oxygen, even though the same water contained substantial

concentrations of ferrous iron, sulfide, or methane. When water samples are collected with bailers, it is almost

impossible to prevent contamination of the sample with oxygen. The widespread use of bailers for sampling

wells at UST spills has created a mistaken impression at some MTBE spills that the ground water is aerobic,

which can lead to a mistaken conclusion that aerobic natural degradation of MTBE in that particular aquifer is

possible.

Other studies have indicated some biodegradation of MTBE under oxygen limiting conditions including

methanogenic conditions (Wilson et al., 2000), iron reducing conditions, or nitrate reducing conditions

(Bradley et al., 2001). However, it is unlikely that monitored natural attenuation will be an acceptable solution

for MTBE in ground water at many sites. Kolhatkar et al. (2000) applied an objective test for natural

attenuation of MTBE to the seventy-four stations they sampled in the eastern United States. Based on one

round of sampling from existing monitoring wells, natural attenuation of MTBE was shown to be statistically

significant at the 90% confidence level at only four of the seventy-four stations.

In many cases, an ex-situ approach may be preferred for addressing MTBE in ground water that is produced

and distributed for water supply (Keller et al., 2000). As with any potential ex-situ ground-water treatment
72

Table 2.4 Appropriate Field and Laboratory Methods for Critical Redox Parameters

Parameter Field Method Laboratory Method 

Dissolved 

Oxygen 

Colorimetric methods are 

recommended. Electrodes can be used 

successfully in a flow-through or 

overflow cell. See discussion in 

section 2.4 below and Session 4 

Summary. 

Analysis on samples shipped to the laboratory is 

not recommended. 

Ferrous Iron 

Fe (II) 

Colorimetric analysis, use of field 

spectrophotometer is preferred. (See 

Session 4 Summary) 

Analysis on samples shipped to the laboratory is 

not recommended. 

Manganese Colorimetric analysis, use of field 

spectrophotometer is preferred. 

Analysis on samples shipped to the laboratory is 

not recommended. 

Total 

Inorganic 

Carbon 

(TIC) 

Determine alkalinity (endpoint 

titration to pH 4.5) and pH. For water 

with pH < 6.5, also use a colorimetric 

field kit for free carbon dioxide. 

Calculate TIC from the sum of 

carbonate and bicarbonate alkalinity 

plus free carbon dioxide. (See section 

2.8) 

The samples should be shipped without 

preservative, and sealed into the container 

without headspace. Laboratory analysis by 

purging TIC from an acidified sample and 

measuring CO2 by infrared absorption. 

Sulfide Colorimetric analysis, use of field 

spectrophotometer is preferred. 

Analysis on samples shipped to the laboratory is 

not recommended. If analysis on samples 

shipped to the laboratory is necessary, the 

samples should be shipped with an appropriate 

preservative, such as zinc acetate. 

Methane 

Preserve in the field, ship to the 

laboratory for analysis. (See Session 5 

Summary) 

Headspace analysis using gas chromatography 

with a flame ionization detector (GC/FID) 

Oxidation/ 

Reduction 

Potential 

Platinum electrode in a flow- through 

cell or an overflow cell. Readings 

should be corrected to the standard 

hydrogen electrode. (See Session 4 

Summary) 

Analysis on samples shipped to the laboratory is 

not recommended. 

Specific 

Conductance 

Electrode in a flow-through cell or an 

overflow cell. The electrode may not 

automatically correct for temperature; 

measure and report temperature. 

Electrode in a water sample that has not been 

preserved. The samples should be sealed into the 

container without headspace to prevent the 

oxidation and precipitation of iron. 

pH 

Electrode in a flow-through cell or an 

overflow cell. The electrode should 

be calibrated in the field prior to use. 

Electrode in a water sample that has not been 

preserved. The samples should be sealed into the 

container without headspace. 

Temperature Thermometer in a flow-through cell or 

an overflow cell. 

Not Possible. 

Hydrogen 

Collect using bubble Stripping 

Technique. Analysis in a field 

laboratory is preferred, using gas 

chromatography with a reducing gas 

detector. (See Session 5 Summary) 

Analysis using gas chromatography with a 

reducing gas detector. A flame ionization 

detector is not sufficiently sensitive to hydrogen. 

Nitrate 

Sulfate 

Chloride 

Not recommended. Adequate field 

methods exist for these parameters; 

however, they are not cost-effective 

compared to analysis in a laboratory. 

Ion Chromatography. The following preservation 

and holding times are recommended: nitrate, 4 °C 

with 48 hour holding time; sulfate, 4 °C with 28 

day holding time; chloride, ambient temperature 

with 28 day holding time (see U.S. EPA, 1999). 
73

process, the geochemistry of the ground water (e.g., iron and manganese content) can be an important factor in

determining which ex situ treatment process is likely to be most cost-effective. Low redox ground waters

containing high levels of dissolved iron and manganese may pose problems (e.g., precipitate clogging or

coating) for treatment systems in which increases in redox potential occur during treatment. In such cases, a

pre-treatment or modified treatment process may be required to achieve effective overall treatment. Thus, in

cases where an ex-situ treatment process is under consideration, the measurement of redox parameters such as

dissolved iron/manganese may be beneficial in evaluating the feasibility of using selected above ground

treatment technologies.

2.4 Relationship of Redox Processes to Phytoremediation

In the root zone (rhizosphere) of plants, especially trees, it has been shown that exudates from roots supply

organic carbon, especially organic acids, that stimulate bacterial growth (Lee et al., 2000; Schnoor et al., 1995).

In addition, the decay of root biomass can also add organic carbon to the vadose zone and shallow ground

water. Where near-reducing conditions have been established in ground water near phreatophytes such as

cottonwood trees, reductive dechlorination of TCE has been observed in the rhizosphere (Lee et al., 2000).

Alternatively, some plants can provide some oxygen to roots, or create avenues for atmospheric oxygen to enter

ground water or the deep vadose zone through root holes, perhaps enhancing the aerobic environment in

shallow ground water.

Thus, biodegradative processes can occur in the rhizosphere of plants, especially phreatophytes in addition to

the well known removal of contaminated ground water by transpiration and degradation of the contaminants

within the plant biomass (Schnoor et al., 1995).

2.5 Appropriate Field and Laboratory Methods for Critical Redox Parameters Critical Redox Parameters

Many redox parameters are unstable; they start to change as soon as the ground water is exposed to the

atmosphere (see Session 4 Summary and Appendix A). As a consequence, many of these redox analyses must

be done at the well head. The following recommendations are based on the combined experience of the

participants at the session. They represent a group consensus on methods and techniques that can be expected

to provide data of adequate quality at hazardous waste sites.

2.6 General Considerations for Ground-Water Sampling

The method used to collect a sample of ground water is often more important for data quality than the method

used to analyze the sample. Recommendations for improved practice are discussed in detail in Puls and

Barcelona (1996). If ground water is exposed to the atmosphere while it is being collected into the sampling

well, or when it is pumped from the well, the sample can be seriously compromised. The ground water may

gain oxygen from the atmosphere. The oxygen introduced into the sample may chemically react with ferrous

iron or sulfide and transform these materials. Oxygen will also disturb measurements of oxidation reduction

potential.

The standing water in a well is exposed to the atmosphere; it is usually oxygenated with respect to the native

ground water in the aquifer. In most applications, the standing water within the well is purged and replaced

with ground water that has not been altered. To avoid aeration, water should be pumped from the well using a

gentle device such as a peristaltic pump or bladder pump. Bailers should only be used when there is no other

option for geochemical sampling. If a bailer must be used, the bailer should be slowly immersed in the standing

column of water in the well to minimize aeration. After sample collection, the water should be drained from the

bottom of the bailer through tubing into the sampling container.

Ground water can also be aerated if the well is pumped down too far. The ground water will enter the well, then

cascade down the inside of the screen as the water level in the well drops. To avoid aeration, wells that are

screened below the water table should not be pumped at a rate which lowers the standing water in the well to

a level below the top of the screen. If practical, wells screened across the water table should be pumped at a rate

that lowers the total height of the water column in the well no more than five to ten percent.

An overflow cell or a flow-through cell should be used for the measurement of well-head parameters such as

pH, temperature, specific conductance, dissolved oxygen, and oxidation-reduction potential. Experts in the

area of field sampling of ground water differ in their preference for flow-through cells or overflow cells. Flowthrough cells effectively prevent contact of a water sample with oxygen, particularly when they don’t trap a

bubble in the cell. Depending on the position of the outlet tube and flow restrictions in the outlet tube, water in
74

the flow-through cell can be under positive pressure. This will preclude the proper use of certain electrodes that

depend on gravity flow of electrolyte to a salt bridge in the electrode. The user’s manual for an electrode should

be consulted to determine if this may be a problem. Overflow cells can function as well as flow-through cells

when the flow from the pump is presented directly to the sensor surface of the electrodes, and then flows up past

the body of the electrode before it overflows. The position of an electrode can be adjusted in the overflow cell

to make the elevation of the reservoir of electrolyte higher than the overflow of water being monitored.

The "Technical Protocol for Evaluating Natural Attenuation of Chlorinated Solvents in Ground Water"

(Wiedemeir et al., 1998, EPA/600/R-98/128) depicts a flow-through cell on page A4-17. The authors of the

Technical Protocol used an Erlenmeyer flask as an overflow cell because the wide base of the flask kept the

leads to the electrodes from tipping the flask over. However, this configuration provides a reservoir for mixing

of the ground water in the flask and does not provide directed flow of ground water across the sensor surface of

the electrodes. The overflow cell should be a cylinder that is only slightly larger than the space needed to

contain the electrodes. A graduated cylinder works well. It can be held in place with clamps or a lead

doughnut.

2.7 When to Measure Hydrogen in Ground Water

Measurement of hydrogen in ground water should be considered when the redox processes cannot be

interpreted from the more conventional Protocol Parameters. As an example, a trichlorethylene plume may

show a marked reduction in the concentration of trichloroethylene downgradient of the source, with accumulation of minor concentrations of the first reduction transformation product, cis-dichloroethylene, and no

detectable vinyl chloride. This distribution can be easily interpreted in two ways. Reductive dechlorination

may be weak, resulting in transformation of a minor amount of trichloroethylene to cis-dichloroetheylene. The

reduction in concentrations of trichloroethylene are caused by physical processes such as dilution and

dispersion. In the second interpretation, reductive dechlorination is strong, and is the cause of the loss of

trichloroethylene. The transformation product cis-dichloroethylene does not accumulate because it is further

transformed to carbon dioxide, without going through vinyl chloride as an intermediate. If concentrations of

hydrogen are low, less than 1 nanomolar, the second interpretation is excluded and the first interpretation is

supported. If hydrogen concentrations are high, greater than 1 nanomolar, the second interpretation is

supported, although not necessarily proven. Hydrogen concentrations should be measured when they are

necessary to understand the redox processes. When the redox processes can be interpreted from the more

conventional parameters, measurement of hydrogen is not necessary. A recent publication compares two

techniques for sampling hydrogen (McInnes and Kampbell, 2000).

2.8 Determination or Calculation of Calculation of Total Inorganic Carbon (TIC) otal Inorganic Carbon (TIC)

Microorganisms degrade chlorinated hydrocarbons through a variety of mechanisms including aerobic

respiration, iron and manganese reduction, or use as the electron acceptor in reductive dechlorination. MTBE

has been shown to be degraded under aerobic, nitrate reducing, iron reducing, sulfate reducing and methanogenic

conditions. The most universal end product of biodegradation of chlorinated hydrocarbons or MTBE is carbon

dioxide that is produced from the contaminant itself, or from the compound used as an electron donor in

reductive dechlorination. This biogenic carbon dioxide can be seen in plumes of contaminated ground water as

an increase in the TIC above the background TIC in uncontaminated water in the aquifer. It is often one of the

most revealing “footprints” of ground water that has been remediated by natural biological processes.

This geochemical parameter is termed TIC because it is measured on water samples that have not been filtered.

In most cases the measured TIC is equivalent to the dissolved inorganic carbon (DIC) in the ground water. It

is best not to filter ground water samples before analysis for inorganic carbon. Often ground water is not in

equilibrium with the atmosphere, and it is difficult to avoid exchange of carbon dioxide between the water

samples and the atmosphere during filtering.

The concentration of TIC can be determined directly by adding acid to the water sample, partitioning carbon

dioxide to a head space and measuring the carbon dioxide directly with an infrared detector (see methods for

total inorganic carbon in U.S. EPA, 1999). This is the preferred method. The concentration of TIC can also be

calculated from measurements of alkalinity, free carbon dioxide, and pH as described later in this section.

Methods for alkalinity are described in U.S. EPA (1999), and for free carbon dioxide in APHA (1999) and

ASTM (1996).

These calculations are generally accurate; however, there are important categories of plumes where these

calculations can produce major errors. The calculations assume that there is no hydroxide alkalinity associated
75

with strong bases. Occasionally an unnaturally high pH is caused by disposal of caustic waste materials in the

subsurface. A high pH can also be caused by improper grouting or cementing of monitoring wells. The

calculations also assume that there is no alkalinity associated with other weak acids. Ground waters often

contain organic acids that contribute to alkalinity (Hemond, 1990). Plumes of leachate from solid municipal

waste landfills often have large amounts of alkalinity associated with organic acids such as acetate and

butyrate.

The correspondence between carbon dioxide produced during metabolism and the increase in TIC is not onefor-one. In acid ground water, some of the carbon dioxide may be lost across the water table. This effect is

probably not important in most plumes, but it should be considered and evaluated. In aquifers that contain

carbonate minerals in the aquifer matrix, the biogenic carbon dioxide can ultimately react with calcium or

magnesium carbonate minerals to dissolve them as calcium or magnesium hydrogen carbonate. In this reaction

there are two moles of TIC produced in ground water for each mole of biogenic carbon dioxide consumed.

Once formed, TIC can precipitate with ferrous iron as siderite, or sorb to the anion exchange complex. In many

ground waters these reactions do not appear to be in equilibrium. As a consequence, it is usually impossible

to directly convert a change in TIC to an amount of carbon dioxide produced by biological activity. However,

it is often possible to calculate a range of carbon dioxide concentrations that could have produced the measured

TIC.

There are two calculations that can be used to estimate the total inorganic carbon (TIC) from pH, alkalinity,

and free carbon dioxide.

Method 1: Calculate TIC from the sum of measured alkalinity and measured free CO2

Dissolved inorganic carbon is the sum of dissolved carbon dioxide, carbonic acid, bicarbonate, and carbonate.

For convenience we will express the sum of dissolved carbon dioxide and carbonic acid as carbonic acid

[H2

CO3

*

].

[TIC] = [H2

CO3

*

] + [HCO3

-1] + [CO3-2]

Alkalinity is conventionally determined as the amount of acid required to titrate water to a colorimetric end

point near pH 4.8. It is conventionally expressed as the amount of CaCO3

 that would be consumed by that

amount of acid. The meq/liter of acid required is the alkalinity in mg/liter of CaCO3

 divided by 50.

Alkalinity as [H+

] consumed = [HCO3

-1] + 2[CO3-2]

Each 100 mg/liter of CaCO3

 alkalinity accepts 2 meq/liter of H+.

CaCO3

 alkalinity (mg/liter)/50 = 0.5 [HCO3

-1] + [CO3-2]

Free carbon dioxide is calculated from the amount of base needed to convert all the H2

CO3* and HCO3

-1 to

CO3

-2. It is usually expressed in mg/liter of CO2

.

Free carbon dioxide as [H+] provided = 0.5 [HCO3

-1] + [H2

CO3

*

]

Each 44 mg/liter of Carbon Dioxide provides 2 meq/liter of H+

Free Carbon Dioxide (mg/liter)/ 22 = 0.5 [HCO3

-1] + [H2

CO3

*

]

CaCO3

 alkalinity/50 + free CO2 /22 = [H2CO3

*

] + [HCO3

-1] + [CO3-2] = [TIC]

TIC (mg Carbon/liter) = 12 *(CaCO3

 alkalinity/50 + Free CO2/22)

Method 2. Calculation of TIC from Method 2. Calculation of TIC from Alkalinity and pH

This method should only be applied to waters with pH > 6.5 and easily measurable CaCO3 alkalinity.

[TIC] = [H2

CO3

*

] + [HCO3

-1] + [CO3-2]
76

The calculation proceeds in three steps:

(1) Calculate [HCO3

-1] from measured alkalinity and pH.

(2) Then calculate [H2

CO3

*

] and [CO3

-2] from the pH, the [HCO3-1] as calculated previously, and the

governing pKa.

(3) Then add the calculated [H2

CO3

*

] + [HCO3

-1] + [CO3-2] and multiply by 12,000 to convert to TIC

(mg C/liter).

Alkalinity is conventionally determined as the amount of acid required to titrate water to a colorimetric end

point near pH 4.8. It is conventionally expressed as the amount of CaCO3

 that would be consumed by that

amount of acid. The meq/liter of acid required is the alkalinity in mg/liter of CaCO3

 divided by 50.

CaCO3

 alkalinity as [H+] consumed = [HCO3

-1] + 2[CO3-2]

CaCO3

 alkalinity (mg/liter)/50 = [HCO3

-1] + 2[CO3-2]

Calculation of [HCO3

-1]

The ratio of [HCO3

-1] and [CO3-2] can be predicted from the pKa of [HCO3-1] and the pH.

The pKa of [HCO3

-1] to [CO3-2] and [H+] is 10.49 at 10 °C.

In other words: ([H+

][CO3

-2] )/[HCO3-1] = 10 -10.49 and [CO3-2] = [HCO3-1]/ 10(10.49-pH)

Substituting [HCO3

-1]/ 10(10.49-pH) for [CO3-2] in equation (1) yields

CaCO3

 alkalinity (mg/liter)/50 = [HCO3

-1] +2([HCO3-1]/10(10.49-pH)).

Solving for [HCO3

-1] yields

[HCO3

-1] = (mg/liter CaCO3

/50)* 10(10.49-pH) / (2+ 10(10.49-pH)).

Calculation of [CO3

-2]

Again the pKa of [HCO3

-1] to [CO3-2] and [H+

] is 10.49 at 10 °C.

In other words: ([H+

][CO3

-2] )/[HCO3-1] = 10 -10.49 and [CO3-2] = [HCO3-1] * 10 -10.49 + pH

Calculation of [H2CO3

*]

The pKa of [H2CO3

*

] to [HCO3

-1] and [H+

] is 6.42 at 10 °C.

In other words: ([H+

][HCO3

-1] )/[H2

CO3

*

] = 10 -6.42 and [H2CO3

*

] = [HCO3

-1]/ 10 -6.42 + pH

References

American Public Health Association, American Water Works Association,Water Environment Federation (APHA).

1999. Standard Methods for the Examination of Water and Wastewater, Titrimetric method for free carbon

dioxide 4500- CO2 C, American Public Health Association, Washington, D.C.

American Society for Testing and Materials (ASTM). 1996. Standard Test Methods for Total and Dissolved

Carbon Dioxide in Water, ASTM D 513-92. Philadelphia, Pennsylvania.

Bradley, P. M., Landmeyer, J. E., and Chapelle, F. H. 2001. Widespread potential for microbial MTBE degradation

in surface-water sediments. Environmental Science and Technology, v. 35, pp. 658-662.

Hemond, H. F. 1990. Acid neutralizing capacity, alkalinity, and acid-base status of natural waters containing

organic acids. Environmental Science and Technology, v. 24, p. 1486-1489.

Keller, A. A., Sandall, O. C., Rinker, R. G., Mitani, M. M., Bierwagen, B., and Snodgrass, M. J. 2000. Evaluation

of physicochemical treatment technologies for water contaminated with MTBE. Ground Water Monitoring

and Remediation, v. 20, pp. 114-126.

Kolhatkar, R., Wilson, J., and Dunlap, L. 2000. Evaluating natural attenuation of MTBE at multiple UST Sites.
77

In Proceedings of NGWA/API Conference and Exposition on Petroleum and Organic Chemicals in Ground

Water: Prevention, Detection and Remediation. Anaheim, California, November 15-17, 2000, pp. 32-49.

Landmeyer, J. E., Chapelle, F. H., Herlong, H. H., and Bradley, P. M. 2001. Methyl tert-butyl ether biodegradation

by indigenous aquifer microorganisms under natural and artificial oxic conditions. Environmental Science

amd Technology, v. 35, pp. 1118-1126.

Lee, R. W., Jones, S. A., Kuniansky, E. J., Harvey, G., Sherwood-Lollar, B, and Slater, G. F. 2000. Phreatophyte

influence on reductive dechlorination in a shallow aquifer contaminated with trichloroethene (TCE).

International Journal of Phytoremediation, v. 2, pp. 193-211.

McInnes, D. M. and Kampbell, D. 2000. The bubble stripping method for determining dissolved hydrogen (H2

)

in well water. Field and Analytical Chemistry and Technology, v. 4, pp. 283-296.

National Research Council. 2000. Natural Attenuation for Groundwater Remediation, National Academy Press,

p. 11.

Pivetz, B. 2001. Phytoremediation of Contaminated Soil and Ground Water at Hazardous Waste Sites, EPA/540/

S-01/500. U. S. Environmental Protection Agency, Cincinnati, OH. [http://www.epa.gov/ada/download/

issue/epa_540_s01_500.pdf].

Powell, R. M., Blowes, D. W., Gilham, R. W., Schult, T., Puls, R. W., Vogan, J. L., Powell, P. D., and Landis, R.

1998. Permeable Reactive Barrier Technologies for Contaminant Remediation, EPA/600/R-98/125.

U. S. Environmental Protection Agency, Cincinnati, OH. [http://www.epa.gov/ada/download/reports/

reactbar.pdf].

Puls, R .W. and Barcelona, M. J. 1996. Low-Flow (Minimal Drawdown) Ground-Water Sampling Procedures,

EPA/540/S-95/504. U. S. Environmental Protection Agency, Cincinnati, OH. [http://www.epa.gov/ada/

download/issue/lwflw2a.pdf].

Salanitro, J. P., Johnson, P.C., Spinnler, G. E., Maner, P. M., Wisniewski, H. L., and Bruce, C. 2000. Field-scale

demonstration of enhanced MTBE bioremediation through aquifer bioaugmentation and oxygenation.

Environmental Science and Technology, v. 34, pp. 4152-4162.

Schnoor, J. L., Licht, L. A., McCutcheon, S. C., Wolfe, N. L., and Carreira, L. H. 1995. Phytoremediation of

organic and nutrient contaminants. Environmental Science and Technology, v. 29, pp. 318-323A.

U. S. Environmental Protection Agency. 1999. EPA Methods and Guidance for Analysis of Water, June 1999,

Office of Water, Washington, D.C, Version 2.0, EPA 821-C-99-004.

Wiedemeier, T. H., Swanson, M. A., Moutoux, D. E., Gordon, E. K., Wilson, J. T., Wilson, B. H., Kampbell, D. H.,

Haas, P. E., Miller, R. N., Hansen, J. E., and Chapelle, F. H. 1998. Technical Protocol for Evaluating Natural

Attenuation of Chlorinated Solvents in Ground Water. EPA/600/R-98/128. U. S. Environmental Protection

Agency, Cincinnati, OH. [http://www.epa.gov/ada/download/reports/protocol.pdf].

Wilson, J. T., Cho, J. S., Wilson, B. H., and Vardy, J. A. 2000. Natural Attenuation of MTBE in the Subsurface

under Methanogenic Conditions, EPA/600/R-00/006. U. S. Environmental Protection Agency, Cincinnati,

OH. [www.epa.gov/ada/reports].
78
79

Robert W. Puls and William J. Deutsch

3.1 Introduction

The purpose of this session was to develop recommendations concerning reduction/oxidation (redox) measurement parameters for inorganic hazardous waste site remediation. It is hoped that the information provided will

be useful to site managers to guide site characterization, data collection efforts, and, in particular, performance

evaluation of in-situ remedial approaches. The following issues and/or questions were addressed in the session:

1) What remediation processes specific to inorganic contaminants require the need for redox

measurements during site characterization and/or remediation (e.g., MNA, zero valent metal

treatment, in-situ redox manipulation, etc.)?

2) What parameters must be measured during site assessment and remediation? Are the required

parameter measurements universal or contaminant-specific? Is there a universal set of parameters

that is sufficient for all remedial strategies?

3) What specific redox measurements are required (e.g., Eh, DO, Fe, etc.), how are redox measurements

best made, and how are the measurements used? Is there a need to evaluate redox characteristics of

the solid matrix in addition to the mobile phase?

The overall findings and conclusions from this session were:

• Geochemical models are recommended for use at all sites. Site data are needed to support the geochemical

model.

• A set of core aqueous-phase redox measurement parameters are recommended for initial characterization

at all sites to understand the redox status of the aquifer. Supplemental measurement parameters may be

needed on a site-specific basis.

• There are commonalities for site characterization across treatment types and contaminant types.

• Eh measurement is included in the recommended core parameters, even though this measurement is

controversial and may only provide qualitative information.

• Site diversity and differences in treatment applications preclude a “hard” list of redox measurement

parameters applicable to every site. Measurement parameters will change from site-to-site.

• Solid-phase redox measurements are generally not performed. The solid-phase should be tested for metals

to determine the form of the metal contaminant before and after treatment and to look at metal stability.

• Solid-phase sample preparation and analytical methods are lagging behind aqueous-phase methods and

need more development.

• We need to be able to predict the performance of a selected site remedy (treatment) and its influence on the

risk assessment. A decision tree for guidance on site evaluation would be useful.

• We need to define parameters/criteria for site closure (post-treatment monitoring).

3.2 Redox Processes Impacting Inorganic Contaminants ocesses Impacting Inorganic Contaminants

Redox processes are chemical reactions that include a transfer of electrons and consequently a change in

valence state of elements that are either oxidized to a higher valence state or reduced to a lower valence state.

This change in valence state of the elements impacts remediation through three mechanisms:

• Change in speciation to a lower/higher toxicity

• Solubility decrease to immobilize contaminants or increase in solubility to mobilize contaminants

• Adsorption/desorption impacts on contaminant mobility

Session 3 Summary: Session 3 Summary:

Redox Pr Redox Processes in Inorganic Remediation
80

Change in Valence and Toxicity

By definition, redox processes involve a change in valence state of the elements whereby one or more elements

are oxidized and one or more are reduced. If the inorganic contaminant participates in these electron transfer

reactions, then its valence state will change. The toxicity of an element is a function of its valence state because

different valence states of the elements react to a greater or lesser degree with biological functions. For

example, the acute toxicity of As(III) is substantially greater than that of As(V) (Franke and Moxon, 1936).

Furthermore, As(III) is the initial substrate in the methylation of arsenic (Thomas, 1994); thus, changes in the

arsenic valence state will have an effect on the production of the methylated arsenic species.

The relative toxicities of the different valence states of other redox-sensitive elements appear to be related to

both availability and toxic effects. For example, the trivalent form of chromium [Cr(III)] is considered much

less toxic than Cr(VI), however, this may be due as much to the lower solubility of Cr(III) solids in aquifer

environments compared to Cr(VI) solids as to the inherent toxicity of the two valence states of chromium

(Losi et al., 1994). In the case of selenium, availability and toxicity appears to be related to the presence of the

more oxidized, soluble selenate [Se(VI)] and selenite [Se(IV)] redox states compared to the reduced, less

soluble forms of elemental selenium (Seo

) and selenide [Se(-II)] (Neal, 1990).

Mineral Solubility Impacts on Remediation

Under equilibrium conditions, the solubility of a mineral limits the dissolved concentration of one of its

constituent elements. The solubility of minerals containing the redox-sensitive elements can vary by orders of

magnitude depending on the valence state of the element in the mineral, the redox status of the aquifer and in

particular, the pH of the ground water. For example, the solubility of the Cr(III) mineral Cr(OH)3

 under typical

ground-water conditions is in the part per billion range whereas the solubilities of Cr(VI) minerals under these

same conditions are in the part per million range (Richard and Bourg, 1991). As a consequence, reducing

Cr(VI) to Cr(III) either by adding a reductant to the aquifer or by allowing the natural system sufficient time to

reduce the chromium valence state and reach equilibrium has the potential for achieving a ground-water

cleanup level by precipitation of the relatively insoluble Cr(OH)3

 mineral.

The use of precipitation of low solubility minerals as a remediation measure is directly related to the concept of

geochemical barriers in natural systems (Perelman, 1986). A geochemical barrier is a zone in the subsurface

characterized by a sharp change in physical or chemical environment along the flowpath and is often associated

with the precipitation of elements from solution. Redox barriers are classified as a type of physico-chemical

barrier, which also includes alkaline, acidic, and adsorption barriers. Because barriers represent a change in

environmental condition, each type of barrier must be considered in the context of the types of water with

different redox characteristics that might enter the barrier. In addition to being a major control on solubility, the

pH of the aquifer will affect the type of element immobilized by the geochemical barrier.

Methods of Immobilizing Inorganic Contaminants Methods of Immobilizing Inorganic Contaminants

Applied techniques for immobilizing inorganic contaminants are generally based on the well-documented

natural processes that impact element mobility, particularly at geochemical barriers. If manganese concentrations in mildly oxidizing water downgradient from a landfill must be lowered, then either increasing the redox

potential to precipitate a less soluble Mn(IV) oxide or reducing the redox potential and adding sulfide to form

a less soluble sulfide mineral might be considered. Some of the issues to consider in the applied redox

approach are the type and amount of reactant, means of emplacement, reaction kinetics, unwanted byproducts,

solubility of contaminant-containing mineral, and geochemical stability of the imposed barrier environment.

Some of the complexity of the situation has been shown by Seaman et al. (1999) who found that in situ

immobilization of Cr(VI) by using Fe(II) solutions was complicated by the hydrolysis of Fe(III) which lowered

the pH and mobilized Cr(VI) in column experiments.

The types of oxidizing and reducing compounds that have been tested or suggested to lower inorganic

contaminant concentrations in an aquifer are listed in Table 3.1. Most of these compounds can be dissolved in

water and injected into the aquifer. Air and ozone are added to the subsurface in a vapor phase. Elemental iron

(Feo

) is a solid that can be emplaced in a trench or as a slurry that intercepts the contaminated water. It has been

shown to be effective not only at reducing the concentrations of inorganic contaminants (particularly chromate

and uranium), but also chlorinated volatile organic compounds such as PCE and TCE. This type of remediation

is expected to expand as its effectiveness becomes more widely known and the injection of reagents to achieve

remediation becomes more commonly accepted by regulatory agencies.
81

Table 3.1 Potential Oxidants and Reductants for Inorganic Remediation

Oxidants/elements treated* Reductants/elements treated* 

Dissolved oxygen / Fe Fe0

(PRB) / Cr, As, U, Tc, Se, Mo 

Air / U, As Colloidal Fe0

 / Cr, As, U, Tc 

Oxygen Release Compound / As Dithionite / Cr, U, Tc 

Potassium permanganate Sodium sulfide / Cr, U 

Fenton=s reagent Hydrogen sulfide / Cr 

Fe2+ / Cr 

Ferrous sulfate / Cr 

Hydrogen Release Compound / Cr 

Organic materials / sulfate, nitrate, ARD 

* known elements treated or proposed following successful lab testing 

The adsorption capacity of an aquifer can be increased to immobilize contaminants by adding dissolved ferrous

[Fe(II)] iron to an aquifer and then oxidizing the ferrous iron to precipitate ferric (oxy)hydroxide. The large

surface area of this solid and its strong affinity for both dissolved cations and anions has the potential for

substantially lowering dissolved contaminant concentrations.

Methods of Mobilizing Inorganic Contaminants Methods of Mobilizing Inorganic Contaminants

It is preferable to leave a contaminant in place if its mobile concentration can be reduced to below a cleanup

level and rendered non-toxic; however, in some instances immobilization may not be possible and removal is

necessary. Removal of an inorganic contaminant requires not just extraction of contaminated ground water, but

also dissolution of contaminant-containing minerals and perhaps desorption of the contaminant from the

surface of solids. If an element is concentrated under reducing conditions (e.g., Cr, U, Mo, Se), then creating

oxidizing conditions in the aquifer should mobilize the contaminant. This is exactly the process used to mine

uranium by in-situ leach methods. Oxygen is added to the subsurface where uranium has been concentrated by

natural processes in the reduced U(IV) mineral uraninite (UO2

). Uraninite is soluble under the imposed

oxidizing conditions leading to its dissolution and the removal of dissolved uranium in production wells.

Mobilization of contaminants can be complicated by the strong adsorption and high capacity of iron and

manganese oxyhydroxides for other metals. If oxidizing conditions are generated in a reducing environment to

mobilize metals, iron and manganese solids may precipitate and remove from solution some of the metals that

are the target of remediation. This occurrence may require a stepwise mobilization process to first oxidize the

elements in the minerals and then reduce and dissolve iron and manganese oxyhydroxides to release adsorbed

contaminants. The kinetics of these reactions will also control the oxidation/reduction of the elements and can

greatly impact any remedial scheme.

Adsorption/Desorption Impacts on Contaminant Mobility

Adsorption is the attachment of dissolved ground-water constituents to the surfaces of the aquifer solids and

desorption is the release of the constituents back into the water. Redox reactions that affect the affinity or the

capacity of the solid surfaces for dissolved constituents can impact contaminant mobility. Adsorption is very

pH dependent because cations (or anions) of the same charge will adsorb at very different pH values depending

on the element (Jain et al., 1999; Langmuir, 1997). Any remedial scheme which involves a redox manipulation

that affects the pH of the system or alters the nature of sorbent surfaces can have profound impacts on element

adsorption/desorption reactions and thus contaminant mobilization/immobilization.
82

3.3 Site Characterization Requirements

It is necessary to collect data during site characterization on the aqueous and solid phase geochemistry of the

system, model the data, and construct a site conceptual model. This is an iterative, evolutionary process.

Aqueous Chemistry

Numerous redox measurement parameters required to characterize site geochemistry and to evaluate treatment

were discussed in the session. Table 3.2 shows which parameters were considered to be core measurements that

are recommended for all sites, as well as additional site-specific parameters that may be useful.

There was considerable debate about the usefulness and necessity of the Eh measurement. It is considered by

some to be more of a qualitative measurement because it can give inconclusive results. The electrode will often

not stabilize in oxygenated waters, or in reduced waters without sulfur being present. Many members of the

panel felt that Eh determinations using electrodes were of limited usefulness in mixed waters (different groundwater sources). Discrepancies between Pt electrode measurements and Eh calculated from redox couples have

been well documented (Lindberg and Runnells, 1984). The time required for stabilization of the electrode can

also make the measurement impractical in the field. It was noted that because of this, Eh probably should not

be a stabilization parameter for low-flow ground-water sampling. Others felt that in some settings the

measurement has been shown to be of value (acid rock drainage) and research involving long-term performance of permeable reactive barriers has focused on Eh as a potential indicator of system performance. Eh can

be helpful in context with other data, such as iron speciation. Some participants felt that it is a good idea to

collect Eh data even though it may be a questionable measurement.

Finally, several members felt that the most valid approach to evaluate redox was collection of multiple

parameters, including Pt electrode measurements, such as iron speciation and other site-specific redox couples

(e.g., arsenic, sulfur). Table 3.3 lists some important characteristics of many redox-sensitive elements.

 

Core Parameters Site-Specific Parameters

Ground Water Ground Water Solid-Phase

Temperature Dissolved Inorganic Carbon Oxidizing Capacity 

Eh, pH, alkalinity Total Dissolved Solids Reducing Capacity 

Turbidity Fluoride Neutralization Capacity 

Total and Ferrous Iron Hydrogen Adsorption Capacity 

Dissolved Oxygen Methane Cation Exchange Capacity 

Specific Conductance Sulfide Acid Generation Capacity 

Dissolved Organic Carbon Priority Metal Contaminants 

(As, Cd, Cr, Cu, Hg, etc.) 

Total Metal(s) Concentration

Major ions (Ca2+, Mg2+, Na+, 

K+, Cl-

, S2-, SO4

 2-, NO3 -, 

Ammonium, Phosphate ) 

Iron/Aluminum Extraction 

Aluminum, Silica, Manganese Mineralogy (XRD) 

Total Organic Carbon 

Sulfur Extraction

Table 3.2 Core and Site-Specific Redox Measurement Parameters
83

Collection of turbidity data was seen as important for a variety of reasons including its usefulness as a sampling

equilibration parameter using low-flow techniques where metals are contaminants of concern, and to explain

data anomalies (e.g., high total iron concentrations under oxidizing conditions) and evaluate the influence of

particulates on metals concentrations.

The collection of major anion and cation data is important in order to properly evaluate the aqueous

geochemistry of the system and perform accurate geochemical modeling in support of a site conceptual model.

Additional information on this topic is covered in the next section. Fluoride was seen as important in some

hydrogeologic settings because of its high complexing capacity and thus, its importance in adequate determination of ion balances for geochemical modeling purposes.

Solid Phase Geochemistry

There was a general consensus among the group that more solid phase characterization was required for site

characterization and performance assessment of in-situ technologies for remediation of inorganic-contaminated sites. Some of these are listed in Table 3.2. The solid phase should be tested to confirm the form of the

metal associated with the solid phase (mineralogy) to determine its stability or mobilization potential; this is

particularly important for monitored natural attenuation (MNA) assessments. If immobilization is invoked for

MNA, then it is imperative to know what the immobilization mechanism is and under what conditions it may

be reversed (i.e., mobilized). Assessment of the oxidiation and reduction capacities of sediments can be

important for several in-situ remedial technologies including monitored natural attenuation and the creation of

‘redox barriers’ using injections of liquid reductants (Palmer and Puls, 1994; Fruchter et al., 2000). Sulfur

extractions are useful to assess the nature of the redox environment and also the stability of precipitated solid

phases. Similarly, extractions for amorphous and crystalline iron oxides are important in terms of metal

stability and iron bioavailablity.

Field versus Laboratory Field versus Laboratory Analyses Analyses

There was consensus among the group that certain parameters must be measured in the field due to stability

issues and other measurements are recommended for the field but could be done in the laboratory. Table 4

Parameter Comments

Field Measurements -

Temperature Must be done in the field; downhole (monitoring 

wells) preferred 

 pH Must be done in the field, robust, easily measured, 

flow cell recommended 

Eh Must be done in the field, variable, more standard 

checks recommended, flow cell recommended 

Dissolved Oxygen Must be done in the field, variable, accuracy of 

probes limited at less than 1 mg/L 

Turbidity Must be done in the field 

Specific Conductance Recommended for field 

Alkalinity Recommended for field 

Redox Indicators S2-, H2

 

Laboratory Analyses – Water

Major Cations Ca, Mg, Na and K 

Major Anions HCO3/CO3, Cl, SO4, NO3/NO2

Redox Indicators Mn, Fe, S2-, As(III)/As(V), CH4, DOC, 

Minor/Trace Constituents Al, Si, PO4, Br, F 

Metals of Potential Concern As, Cd, Co, Cr, Cu, Hg, Mo, Ni, Pb, Tl, Zn 

Table 3.3 Table 3.3 Sampling Parameters for Inorganic Characterization
84

attempts to address these concerns and provides at least a partial listing of some of the more commonly

measured parameters.

Site Conceptual Model Development Site Conceptual Model Development

The development of a site conceptual model is imperative for any site characterization effort, particularly as a

project moves closer toward remedial selection. It involves the incorporation of a comprehensive assessment of

physical, chemical, and biological processes operative within a defined geologic framework at a site. The group

felt strongly that an iterative evolutionary process be employed to develop and refine a site conceptual model

regarding ground-water flow, hydrogeochemical setting or definition, and contaminant fate and transport. The

developing model must continually be tested against new data and be consistent with data collected throughout

the site characterization process.

3.4 Geochemical Modeling

There are two primary reasons for using geochemical computer models at a site. The first is to better

understand the processes active at the site and be able to explain observed compositions of the various phases

and trends in composition across the site. The second reason is to aid in the remedial design for cleanup of a

contaminated site. Accurate geochemical modeling of a contaminated site can be used to estimate the outcome

of various remedial actions. The model can be used to select the best remediation technologies and to optimize

application of the technology. This applies to the evaluation of both natural and applied redox processes as well

as to other geochemical-based remediation measures such as pH neutralization.

The development of a geochemical model of a system starts with the compilation of available information on

water composition, presence of solid phases, hydrogeology, etc. The existing site conditions are evaluated in

terms of the known types of geochemical and physical processes that occur in the environment. These

processes include mineral dissolution/precipitation, oxidation/reduction, adsorption/desorption, ground-water

recharge, and mixing of ground waters from two or more aquifers. The combination of site specific conditions

and known, common geochemical/physical processes leads to the development of a conceptual model of how

the site operates. The conceptual model can be combined with a computer code to calculate the impacts of the

various processes and provide quantitative estimates of these impacts on the composition of the various phases.

The result is a geochemical computer model consisting of the conceptual model and the computer code that

together can simulate existing conditions and predict the impact on the system of changes in environmental

conditions.

3.4 Modeling Capabilities and 3.4 Modeling Capabilities and Applications Applications

The presence of dissolved constituents in water is a result of chemical interactions between water and the vapor

and solid phases in contact with the water. A competent geochemical model of a site must be able to simulate

the processes that occur between these phases. They include gas dissolution and exsolution from ground water,

aqueous speciation, mineral dissolution/precipitation, oxidation/reduction, and adsorption/desorption. Two of

the most commonly used equilibrium mass transfer codes are MINTEQA2 (Allison et al., 1991) and

PHREEQC (Parkhurst, 1995). MINTEQA2 is available on the EPA website at ftp://ftp.epa.gov/epa_ceam/

wwwhtml/softwdos.htm and PHREEQC at the USGS website http://h2o.usgs.gov/software/geochemical.html.

The ability to simulate changing environmental conditions opens up the entire realm of remediation design to

geochemical modeling. Once an accurate model of an existing contaminated site has been developed, it is

relatively simple to vary site parameters and evaluate the impact of the new condition on contaminant

concentrations in the solution and solid phases. The parameters that might be varied include pH, redox

potential, ionic strength, and concentrations of complexing compounds. For remediation purposes, it is

possible to evaluate both in-situ fixation and immobilization of contaminants as well as enhanced mobility and

removal of contaminants from a system. For example, arsenic is strongly adsorbed onto iron oxyhydroxide

minerals in shallow, oxidizing aquifers. If the dissolved arsenic concentration must be reduced in order to

attain a cleanup goal, then adding additional iron minerals to the aquifer might enhance adsorption of arsenic

and its removal from ground water. Alternatively, arsenic might be removed from the aquifer by circulating a

reducing solution through the aquifer that dissolves the iron mineral adsorbents and releases arsenic to ground

water so that it can be flushed more easily from the aquifer. Both of these remediation alternatives can be

simulated with a geochemical model.
85

Element Redox States Water Mobility 

Antimony III, V Sb solids of both valence states are very 

soluble. 

Arsenic III, V As (III) sulfide minerals insoluble at low 

Eh, As(V) minerals soluble. 

Chromium III, VI Relatively immobile due to low solubility 

of Cr (III) hydroxides. Cr(VI) minerals 

highly soluble. 

Copper I, II Cu(OH)2 formation limits ground-water 

concentration; Cu solubility limited by 

hydroxysulfate, hydroxycarbonate. 

Iron II, III Fe oxyhydroxides (HFO) form at pH > 

2.5 under oxidizing conditions and Fe (II) 

is limited by sulfides and carbonate under 

reducing conditions depending on pH and 

solution composition. 

Manganese II, III, IV Mn (III, IV) oxides and oxyhydroxides 

relatively insoluble when pH > 5 and 

oxidizing; Mn (II) carbonate more 

soluble but stable in reducing conditions. 

Mercury 0, I, II Soluble solids when pH > 4 and 

oxidizing; relatively insoluble sulfide 

minerals under reducing conditions. 

Nitrogen -III, 0, III, V No concentration limiting solid phases; 

dentrification may limit nitrate. 

Selenium -II, 0, IV, VI Soluble selenite/selenate minerals; 

selenide less soluble than selenite which 

is less soluble than selenate. 

Sulfur -II to VI Soluble sulfate minerals; insoluble sulfide 

minerals. 

Vanadium III, IV, V May substitute for Fe in ferric oxide 

lattices; mobile in reducing conditions. 

Uranium IV, VI Soluble VI minerals; insoluble IV oxides. 

Table 3.4 Characteristics of Redox-Sensitive Inorganic Contaminants

One of the additional benefits of modeling remediation is that secondary reactions that might not be

immediately obvious to the investigator may become clear in reviewing the modeling results. For example,

adding iron to an aquifer to precipitate as iron oxyhydroxide and provide additional adsorption sites may

change the pH of the solution because iron precipitation is a hydrolysis reaction that produces hydrogen ions.

If the system is not well buffered, the pH will decrease, which, in turn, may mobilize other metals. The aqueous

speciation computer codes used to develop the geochemical models can take into account the formation of

hydrogen ions during iron precipitation and will allow for buffering by carbonate/bicarbonate ions in solution.

If carbonate minerals are also present in the system, the modeler may want to allow them to interact with the

solution as iron carbonate precipitates.

Limitations Limitations

The utility of a geochemical model for a particular site is limited by errors introduced during development of

the model and our ability to model certain processes. The first opportunity for introducing error into the

process is data collection. The model developed is highly dependent on accurate physical and chemical data

from the site. We need to know flow direction to determine whether concentrations are increasing or

decreasing along the flow path and we need accurate measurements of temperature, pH, Eh, and concentrations
86

of dissolved constituents. For example, an incorrect Eh will affect the speciation of all redox-sensitive species

that are allowed by the modeler to equilibrate with the measured value. If nonequilibrium conditions exist for

some of the allowed redox pairs, then the model calculations will be incorrect. This will, in turn, affect

calculations of mineral equilibrium and solubility. Therefore it is important to collect speciation data on iron

and other redox-sensitive elements which may be present in the aquifer to increase confidence in the

approximate redox poising of the system.

Most of the major inorganic geochemical processes impacting natural systems are well known and incorporated into the modeling codes. However, certain systems may be relatively unique and some of the processes

occurring may not be standardized in the available computer codes. For example, some of the elements that

may be of interest for evaluating radionuclide migration (such as plutonium and americium) are not in the

standard databases of MINTEQA2 or PHREEQC. Also, a rare mineral or amorphous solid may limit solution

concentrations of a contaminant that does not normally reach high concentrations in the natural environment. If

this solid phase is not in the code database, its formation cannot be simulated. Most of the available computer

codes allow the user to input new thermodynamic data for elements, solution complexes and solid phases. This

allows the user to develop an appropriate model if concentration and thermodynamic data are available.

The computer code used to make the calculations required by the model has its own limitations. Many of the

equilibrium thermodynamic codes do not have a method of calculating reaction rates. If a mineral forms or

dissolves slowly in a system, the model developed from these codes will not account for these kinetic effects.

This is not a major limitation for most aquifer systems where residence times are measured in years; however,

kinetic effects can become more important in modeling surface water systems or reactions anticipated to occur

during applied remediation methods such as the injection of reactants into an aquifer. PHREEQC is one of the

few models which does have kinetic data as part of the code.

These limitations should be considered in developing a geochemical model, however they will rarely be

sufficient reason for not developing a model for a site. Considering the number of useful applications for a

geochemical model in site characterization, contaminant transport and remediation design, it is often beneficial

to collect sufficient data for modeling and develop a geochemical model for each site.

3.5 Applications to Remediation Methods Applications to Remediation Methods

A number of in-situ remedial methods were discussed where redox processes are key to implementation and

success. These included the following: physical emplacement (permeable reactive barriers), reagent injection

(potassium permanganate, Fenton's reagent, colloidal iron, sodium dithionite, sodium sulfide), and natural

attenuation. The group chose to focus on permeable reactive barriers due to a preponderance of experience

with this in-situ technology by members of the group. Additional discussions involved the issue of acid rock

drainage (ARD) because of the magnitude of the problem and recognition of ARD as an emerging environmental isssue.

Permeable Reactive Barriers Permeable Reactive Barriers

In terms of initial site assessment for permeable reactive barrier application, the core list referenced above in

Table 3.2 was deemed adequate by the group. The following additional measurement parameters specific to

PRBs were also recommended by the participants:

• verify site ground water/barrier interaction - rate of loading of contaminant (e.g., Cr) onto reactive media

(e.g., iron)

• microbiological activity (biomass estimates and populations [e.g., iron oxidizers, sulfate reducers, etc.])

upgradient, downgradient and within the PRB

• longevity (examination of precipitation and clogging within reactive zone over time)

• metal speciation

• solid-phase analysis (see Table 3.2)

• hydrology

• conservative tracer to track movement into, and through barrier (bromide, chloride, iodide, etc.)

It is imperative to know the extent of source area contamination to design an appropriate remedial system.

Will/can the source area be removed/treated? If left in place, how long will the source area contribute to
87

contamination of the ground water? Identification of the plume boundaries in three-dimensional space as well

as temporal variation is important in siting and designing the PRB. Verification monitoring to demonstrate

treatment effectiveness is also needed. For long-term usage, one needs to determine how quickly mineral

precipitates are accumulating on the iron surface, which will decrease porosity over time and will also passivate

the reactive surfaces. Coatings may accumulate faster in some geochemical environments compared to others

(e.g., high sulfate or high alkalinity waters). Regulators want to know how long the PRB is going to work.

In the specific case of chromium remediation using PRBs, the following list of parameters has been measured:

• Cr speciation, aqueous

• Cr speciation, solid phase (sequential extractions, surface analyses)

• Oxidation capacity, adsorption capacity of aquifer solids

• Permeability assessment of PRB over time (water levels, flow meters, tracer tests)

• Microbiological assessments (phospholipid fatty acid analyses)

Cost-effective methods for long-term performance monitoring of these systems are currently being evaluated

by a number of groups. It is hoped that some direct and inexpensive methods may be employed to accurately

and safely predict performance and longevity.

3.6 Acid Rock Drainage

This group discussed numerous technologies used to remediate Acid Rock Drainage (ARD) and many more

that are proposed but not used. Many of these techniques, however, apply to cleanup of surface waters, not

ground waters. Interactions between surface waters and ground waters are important, but the nature of these

interactions and their environmental impacts are not well understood. Redox chemistry is what ARD

production, remediation, and performance assessment is all about. All remediation technologies need to

address neutralization and many utilize reductive processes. ARD can be categorized into six types:

• Underground workings

• Tailings

• Waste Rock

• Contaminated Aquifers

• Heap leach pads

• Pit lakes/Open pits (not necessarily acid, but redox is definitely important)

• Slag

• Backfill

Necessary redox parameters are the same as those outlined for other inorganic contaminant situations. Iron

speciation [Fe(II)/Fe(III)] is necessary instead of optional and prioritization of other parameters will be sitespecific. For example, some sites require measurement of As(III)/As(V), but not H2

S and vice versa.

Three groups of mineral assemblages are needed to undergo solid phase analyses: oxidizable (sulfide) minerals,

gangue minerals, and country rock minerals. Solid phase analyses are needed to evaluate 1) the amounts/types

of acid-generating sulfide minerals, 2) the amounts/types of acid-neutralizing minerals (carbonates, aluminosilicates), and 3) the reactivity of these minerals (function of crystal form, size, and composition).

An important aspect of remedial technologies for ARD is that there are very few, clear-cut success stories with

cleanup, long-term maintenance, and with monitoring for decades demonstrating compliance. Hence, it is

difficult to look to case histories for guidance.

3.7 Summary

The members of this group were acknowledged experts in the area of environmental assessment and

remediation for inorganics in the environment. Members were employed in industry, universities and the

government sector. Many had extensive field experience in monitoring and remediation. Strong consensus

was reached on a number of issues including the following:

• Use of geochemical models is strongly recommended for all sites and sufficient data for model usage needs

to be collected.
88

• Certain core geochemical parameters need to be measured at all sites.

• Supplemental parameters are also needed for all sites.

• The diversity of contaminant behavior and potential remedial technologies precludes a one-size-fits-all list

of parameters.

• Metals are persistent in the environment. If immobilization technologies are considered, the solid phase

must be analyzed to determine/verify the form of the contaminant and evaluate whether conditions will

persist to contain the contaminant.

• Development and application of analytical methods (extractions) for solid phase analyses lags those for the

aqueous phase and more research is needed in this area.

References

Allison, J. K., Brown, D. S., and Novo-Gradac, K. J. 1991. MINTEQA2/PRODEFA2, A Geochemical Assessment

Model for Environmental Systems, U.S. Environmental Protection Agency.

Franke, K. W. and Moxon, A. L. 1936. A comparison of the minimum fatal dose of selenium, tellurium, arsenic

and vanadium. Journal of Pharmacology and Experimental Therapeutics, v. 58, pp. 454-459.

Fruchter, J. S., Cole, C. R., Williams, M. D., Vermeul, V. R., Amonette, J. E., Szecsody, J. E., Istok, J. D., and

M. D. Humphrey. 2000. Creation of a subsurface permeable treatment zone for aqueous chromate

contamination using in situ redox manipulation. Ground Water Monitoring & Remediation, v. 20, pp. 66-77.

Jain, A., K. P. Raven and R. H. Loeppert. 1999. Arsenite and arsenate adsorption on ferrihydrite: Surface charge

reduction and net OH-

 release stoichiometry. Environmental Science and Technology, v. 33, pp.1179-1184.

Langmuir, D. Aqueous Environmental Geochemistry. Prentice Hall, 600p.

Lindberg, R. D. and Runnells, D. D. 1984. Ground water redox reactions: an analysis of equilibrium state

applied to Eh measurements and geochemical modeling. Science, v. 225, pp. 925-927.

Losi, M. E., Amrhein, C., and Frankenberger, W. T. Jr. 1994. Factors affecting chemical and biological reduction

of hexavalent chromium in soil. Environmental Toxicology and Chemistry, v. 13, pp. 1727-1735.

Neal, R. H. 1990. Selenium. In Heavy Metals in Soils, ed. B. J. Alloway. Blackie, Glasgow, pp. 237-270.

Palmer, C. D. and Puls, R. W. 1994. Natural Attenuation of Hexavalent Chromium in Groundwater and Soils,

EPA/540/S-94/505. U. S. Environmental Protection Agency, Cincinnati, OH.

Parkhurst, D. L. 1995. User’s guide to PHREEQC, a computer model for speciation, reaction path, advective

transport and inverse geochemical calculations. U.S. Geological Survey.

Perelman, A.I. 1986. Geochemical barriers: theory and practical applications. Applied Geochemistry, v. 1,

pp. 669-680.

Richard, F. C. and Bourg, A. C. M. 1991. Aqueous geochemistry of chromium: a review. Water Resources, v. 25,

pp. 807-816.

Seaman J.C., P.M. Bertsch and L. Schwallie. 1999. In situ Cr(VI) reduction within coarse-textured, oxidecoated soil and aquifer systems using Fe(II) solutions. Environmental Science and Technology, v. 33,

pp. 938-944.

Thomas, D. J. 1994. Arsenic toxicity in humans: research problems and prospects. Environmental Geochemistry

and Health, v. 16, pp. 107-111.
89

Session 4 Summary: Session 4 Summary:

Field Measur Field Measurement of Geochemical Redox Parameters ement of Geochemical Redox Parameters

Richard T. Wilkin and Carol J. Ptacek

4.1 Introduction

Conceptual models of contaminant transport and fate in ground water critically depend on geochemical measurements and their interpretation. Transport and fate processes are often dependent on redox processes because in

many cases oxidation-reduction reactions cause changes in speciation that have a marked effect on contaminant

mobility and toxicity. The importance of reliable field characterization is especially significant when evaluating

redox-sensitive speciation because of the potential for sample alteration (oxidation) during sample collection and

handling.

This group considered methods used by environmental scientists to characterize the redox chemistry of ground

water in support of site characterization and remedial performance monitoring. The group recognized that a wide

variety of analytical techniques are available to quantitate redox-sensitive elements (e.g., Fe, S, N). Selection of

the appropriate technique for a given parameter will depend on specific project objectives. It was recognized that

project objectives and principal data uses collected in monitoring programs change with time as more site-specific

information is obtained and conceptual models evolve. Although a wide variety of techniques are available for

redox characterization (see Appendix A), members of this group agreed that standards for improved field practices

are needed (see, for example, U.S. Geological Survey, 1997). Methods of analysis identified during the group

discussion are briefly reviewed in this chapter, and include Eh, dissolved oxygen, iron speciation, sulfur speciation,

nitrogen speciation, and alkalinity.

4.2 Sampling and Measurement Objectives ement Objectives

Sample collection programs are designed around goals associated with specific project objectives. Data Quality

Objectives (DQOs) define the types, quality, and quantity of data that are required by the various aspects of a

project. With DQOs in place, appropriate sampling methodologies, analytical protocols, and specific methods may

be considered and selected. For example, DQOs may be different if project goals put emphasis on detection and

monitoring of contaminant concentrations or whether geochemical speciation and the development of site

conceptual models are targeted goals.

The principal objective of geochemical characterization in site assessment and remedial performance monitoring is

to obtain water quality information, with no alteration of water chemistry, from a chosen sampling point. Such

uncompromised data are critical in developing accurate conceptual and quantitative models of contaminant

transport and fate. There was strong agreement among the group that sample collection practices and measurement

techniques are among the greatest challenges to developing reliable site conceptual models. Several categories of

ground-water sampling methods were recognized: accumulation or diffusion samplers, purge techniques with flowthrough cells, and bailing. The generally recommended and most commonly practiced method is low-flow,

minimal draw-down purging and sampling. Low-flow methods minimize chemical and hydrological disturbances

in and around the well. In some situations, multiple-tube bundle piezometers or diffusion samplers may provide the

best method for collecting ground-water samples without chemical alteration. Bailing techniques were not

recommended for purposes of redox characterization because of the likelihood of changing water chemistry by

unavoidable reaction with air.

Measurements of geochemical parameters can be divided into three categories: laboratory, in situ, and purging or

field measurements. Laboratory analyses are carried out on preserved samples for characterization of total or

selected metals, anions, and organic constituents. In some cases redox integrity of dissolved components can be
90

preserved by acidification (e.g., for NO3

-

/NH4

+

) because oxidation rates generally decrease substantially with

decreasing pH. In situ measurements such as down hole probes or contained sampling devices are desirable for the

measurement of unstable parameters that are ideally made at well conditions. In-situ measurements are potentially

advantageous for the determination of dissolved oxygen, temperature, and oxidation-reduction potential. Project

objectives may or may not require the extra cost and effort required to collect data using in-situ techniques.

Measurements carried out in the field include electrode measurements of DO, pH, specific conductance, Eh, and

other ions using ion-selective electrodes. Field-deployable techniques for ions also include UV-Vis spectrometry

and ion chromatography. Field analyses are made to gather information in the field to guide activities and/or

because the measured parameters are considered to be too unstable to transport samples without compromising

their chemical integrity. There was general agreement that unstable parameters, best measured in the field at the

time of sample collection, include pH, turbidity, DO, ferrous iron, alkalinity, sulfide, and oxidation-reduction

potential. Measurements of these parameters are ideally made during well purging. Members of the group reported

a variety of demonstrated techniques for preserving the redox integrity of samples pumped to the surface, such as:

minimizing tubing length, use of high-quality tubing with low oxygen diffusion coefficients, use of syringes with

luer-lok fittings, preventing direct exposure of tubing and all electronic equipment to sunlight, and ensuring airtight seals. Specific methods are discussed below for the measurement of Eh, DO, and speciation of iron, sulfur,

and nitrogen (Table 4.1).

Table 4.1 Redox Parameters and their Common Methods of Measurement

 Parameter Common Field Methods 

Standards for 

Performance 

Assessments 

References 

Oxidation-Reduction 

Potential (ORP) 

Combination Platinum 

electrode with Ag/AgCl 

reference electrode; KCl filling 

solution 

Zobell’s solution; 

Light’s solution 

Langmuir (1971) 

Nordstrom (1977) 

Standard Methods for 

the Examination of 

Water and Wastewater 

(1999) 

Dissolved Oxygen (DO) 

Membrane-covered electrode 

Colorimetric (high range, 

indigo carmine) 

Colorimetric (low range, 

rhodazine D) 

Modified Winkler titration 

Air-saturated 

water 

Hitchman (1978) 

Gilbert et al. (1982) 

White et al. (1990) 

Standard Methods for 

the Examination of 

Water and Wastewater 

(1999) 

Iron Speciation 

Ferrous iron colorimetric 

indicators (1,10 

phenanthroline; ferrozine) 

Prepared ferrous 

solutions 

Tamura et al. (1974) 

Stookey (1970) 

Sulfur Speciation 

Sulfate (turbidimetric); 

Sulfide (methylene blue 

colorimetric method) 

Prepared sulfate 

and sulfide 

solutions 

Standard Methods for 

the Examination of 

Water and Wastewater 

(1999) 

Cline (1969) 

Nitrogen Speciation 

Nitrate+Nitrite (cadmium 

reduction) 

Ammonia (Nessler method; 

Salicylate method) 

Prepared nitrate, 

nitrite, or 

ammonia 

solutions 

Standard Methods for 

the Examination of 

Water and Wastewater 

(1999) 

Alkalinity Acid titration 

Prepared 

bicarbonate or 

carbonate 

solutions 

Standard Methods for 

the Examination of 

Water and Wastewater 

(1999) 
91

4.3 Geochemical Parameters Geochemical Parameters

Eh

The Eh measurement is the most commonly used technique to characterize the oxidation-reduction state of ground

water. Eh is the measured potential of a platinum electrode corrected to the Standard Hydrogen Electrode (SHE).

Eh is not the same as the oxidation-reduction potential (ORP), which is the direct, uncorrected potential reading of

the reference electrode. It was recognized during the group discussion that although measured Eh values usually do

not correspond to Eh values calculated using the Nernst Equation with measured concentrations of aqueous redox

pairs, e.g., N(V)/N(-III) or S(-II)/S(VI), there is benefit to measuring this parameter in the field. Eh measurements

are a relatively easy approach for determining overall redox conditions. Eh measurements should be thought of as

qualitative indicators of the state of oxidation or reduction of a natural system (Langmuir, 1971). In rare situations,

such as acid-mine waters, high activities of both Fe(III) and Fe(II) provide a dominant and reversible couple that

poises Eh measurements (Nordstrom et al., 1979). Many ground waters are not well poised and speciation

calculations for ground waters have shown that redox couples typically exist in a state of disequilibrium (e.g.,

Stumm, 1966; Lindberg and Runnells, 1984). Such systems in disequilibrium yield mixed potentials that may or

may not represent the redox distribution of any specific redox couple. The reasons for this include disequilibria in

low-temperature ground water (Lindberg and Runnells, 1984; Thorstenson, 1984), poisoning or coating of Pt

electrodes (Jackson and Patterson, 1982; Whitfield, 1974), potential measurements reflect mixed potentials with

little thermodynamic significance (Morris and Stumm, 1967).

The recommended procedure for measurement of Eh involves the use of a small-volume flow-through cell into

which a temperature and Eh electrode is inserted. The type of electrode recommended for ease of use is a highquality combination Pt electrode typically with a Ag/AgCl reference electrode. After purging the flow-through cell

with several volumes of ground-water, Eh readings are recorded until they stabilize or a minimum value is

obtained. Often 30 minutes or longer is required for electrode stabilization, but the group generally agreed that

waiting >30 minutes for electrode stabilization was not a time-effective practice. Proper preparation and maintenance of the Pt electrode is required, including frequent polishing of the electrode surface, changing of the internal

filling solution, and frequent checks with standard solutions that are maintained at ground-water temperature

(American Public Health Association et al., 1999). The most common standard solution for reference checks is

Zobell’s solution (Nordstrom, 1977). Additional check solutions include Light’s solution (Light, 1972), and others

recommended by electrode manufacturers (i.e., Orion ORP standard). The importance of correcting field measurements of ORP to reference the values to the Standard Hydrogen Electrode was emphasized. To properly reference

measured values, correction factors must be available as a function of temperature for the specific type of reference

electrode and reference solution used in the measurement assembly. To determine Eh of a sample relative to the

Standard Hydrogen Electrode, measure the ORP of both sample and standard solution (at the same temperature).

Eh of the sample may then be calculated:

Eh = ORP (mV) – ORP reference solution (mV) + Eh reference solution (mV) [1]

The group discussed the limitations on the use of corrected Eh readings. It was recommended that the readings be

used in three ways: 1) as a general guide to interpret other redox data, 2) for speciation of Fe(II) and Fe(III) in acidic

mine drainage waters, and 3) for geochemical speciation calculations to determine mineral stability for selected Fe

and Mn phases that are stable under intermediate Eh conditions. Views from the session participants varied, but it

was recognized that in the absence of a field Eh reading, there was no widely available method of calculating

saturation indices for phases such as Fe(oxy)hydroxides, Mn oxides, and Mn and Fe carbonates. It was further

recommended that Eh values should not be used to determine saturation indices for phases such as FeS, which

incorporate species (HS-

) that are not generally responsive to measurement by Pt electrodes (Whitfield, 1974; WaltonDay et al., 1990).

Dissolved Oxygen

Dissolved oxygen (DO) is often the principal oxidizing chemical component in ground water; consequently, DO

can be a key parameter governing the mobility of redox-sensitive contaminants, including organic compounds,

transition metals, and transuranic metals. DO is also important in governing the nature and level of microbial

activity. It is therefore directly tied to the fate of dissolved nutrients and organic contaminants consumed through

metabolic processes (e.g., Baedecker and Back, 1979). The concentration of DO in ground water is controlled by

local inputs of oxygen-rich meteoric water, microbial respiration, biodegradation of organic matter, and reaction

with reduced mineral phases in the aquifer (Champ et al., 1979; White et al., 1990).
92

The group recognized four major methods for determining DO concentrations: 1) the Winkler titrimetric method,

2) the membrane-covered electrode method, 3) the indigo carmine colorimetric method, and 4) the rhodazine D

colorimetric method. Of these methods, the Winkler method (and its modifications) was regarded as the most

accurate and precise technique for determining DO. This method has served as a benchmark for laboratory

comparisons and calibration of the electrochemical and colorimetric techniques (e.g., Reynolds, 1969; Hitchman,

1978; Gilbert et al., 1982). From the field perspective, the Winkler method is cumbersome and it uses reagents that

can be hazardous to manage in the field. Seven different Winkler modifications exist and selection of the

appropriate modification requires prior knowledge of a sample that is often not available during field investigations.

Because of the difficulties associated with carrying out Winkler titrations in the field, alternative methods such as

membrane-covered polarographic electrodes are widely used (Hitchman, 1978). Membrane-covered electrodes are

simple to use and calibration checks are usually limited to determining oxygen concentrations in water-saturated air

(American Society for Testing Materials, 1992). Electrodes are ideally suited for in-situ measurements of DO

concentrations and for continuous monitoring of oxygen levels. Dissolved oxygen electrodes do not function at

temperatures greater than about 50 °C, which is not a limitation for most ground waters. Membrane fouling can be

a common cause of difficulty, and electrode measurements can be inaccurate without any indication that poor

results are being obtained. Hydrogen sulfide, thio-organic, and other organic compounds are the most problematic

to continued reliable electrode performance. These species react irreversibly at the electrode surface and

detrimentally affect electrode response and accuracy.

Several colorimetric methods have been developed for the measurement of DO. The indigo carmine (blue) and

rhodazine D (reddish violet) methods are the most prominently used techniques (Gilbert et al., 1982; White et al.,

1990; American Society for Testing Materials, 1995a). These colorimetric procedures provide quick and

convenient methods for accurate field measurements of dissolved oxygen. Colorimetric reagents utilize oxidationreduction indicators that upon reaction with dissolved oxygen in water transform from reduced, colorless forms to

oxidized, colored forms. The extent of color formation is proportional to the concentration of dissolved oxygen and

can be measured by visual comparison to sets of liquid color standards or with a spectrophotometer. Color

development by reaction with dissolved oxygen is independent of salinity so that a wide range of sample types from

seawater to fresh water can be analyzed with these methods without correction (Gilbert et al., 1982).

Because the colorimetric reagents involve oxidation-reduction reactions to indicate the concentration of DO, other

redox species in ground water can influence the result of colorimetric determinations (Gilbert et al., 1982).

Concentrations of the easily reduced species Fe(III), Cr(VI), and Cu(II) can lead to erroneously high DO values

when rhodazine D is used (White et al., 1990). A potential cause of false-positive readings using the rhodazine-D

reagent in ground-water studies may be the presence of very fine-grained ferric hydroxide colloids. Both Fe(II) and

Fe(III) as well as nitrite were shown to lead to inaccurate determinations of DO using the indigo carmine reagent

(Gilbert et al., 1982). Hydrogen sulfide does not interfere with either colorimetric technique. The effects of

dissolved organic carbon (DOC) on the use of the rhodazine D and indigo carmine reagents are incompletely

understood.

Winkler, electrode, and colorimetric methods for determining DO have different strengths and weaknesses. For

projects requiring the highest available quality (in terms of accuracy and precision over a wide range of DO

concentrations), the extra effort associated with the Winkler methods is warranted. Where data quality objectives

are not as strict, both electrode and colorimetric methods are acceptable. Electrodes are generally reliable at DO

levels above 1 mg/L, but are less accurate at levels below 1 mg/L. Multiple methods are preferred if high quality

data are required to meet project objectives (Wilkin et al., 2001). Figure 4.1 shows a decision tree discussed by the

group that may be useful for selecting an appropriate method or methods for DO determination.

The most common and convenient quality check for field measurements of DO is air-saturated water. Airsaturation is achieved using an aerator or through manual agitation. Members of the group pointed out that using

such methods can cause water to become supersaturated with respect to air. Air-saturated DO values are mainly

dependent on temperature and less so on salinity (Figure 4.2). Thus, when quality check values are reported the

temperature must be specified for the check to be meaningful.

Iron Speciation

Iron is present in ground water as dissolved ferrous iron (Fe(II)), colloidal particles containing either ferric or

ferrous iron, and as dissolved ferric iron (Fe(III)). The relative and absolute abundances of these iron forms vary

widely depending mainly on pH and redox state of ground water. Dissolved forms may be unassociated or
93

Is 

qualitative data

acceptable?

Is 

quantitative data

required?

Use either

electrodes or

colorimetric methods

Winkler titration not

recommended

Is DO expected

to be >1 ppm?

Winkler (azide) and

HR colorimetric

(indigo carmine)

DO Electrode

Winkler (azide) and

LR colorimetric

(rhodazine D)

Is ferrous iron

expected to

be >1ppm?

Winkler

(permanganate)

LR colorimetric

(rhodazine D)

yes

yes yes

yes yes

yes

no

no

no

Does Dissolved 

Oxygen need

to be measured?

Figure 4.1 Decision tree for selecting method for determining concentration of DO.

Figure 4.2 Air-saturated DO concentration in water.

Air-saturated, 0.001 m NaCl

(TDS=67 mg/L)

Air-saturated, 0.1 m NaCl

(TDS=5818 mg/L)

Most ground waters
94

complexed with organic or inorganic ligands. Dissolved iron is generally considered to be that fraction passing

through a 0.45 μm filter paper, however, water filtered in this way may still contain colloidal iron that will dissolve

upon addition of an acid.

Ground waters containing dissolved oxygen at concentrations above 1 mg/L will seldom contain concentrations of

dissolved iron above 1 mg/L. Exceptions to this general rule are more acidic ground waters (pH<4) in which the

solubility of ferric oxyhydroxides is sufficiently high to be detected using standard methods of analysis. The

solubility of ferrous iron increases under reducing conditions, where oxygen is depleted and Eh values are low

(<0 mV). In zones of intense iron reduction, concentrations of ferrous iron may reach levels as high as several

hundred mg/L.

Dissolved Fe(II) concentrations in ground waters are commonly measured using colorimetric techniques with

complexing reagents, such as 1,10 phenanthroline (Tamura et al., 1974), ferrozine (Stookey, 1970 as modified by

Gibbs, 1976; To et al., 1999), and bipyridine (e.g., Baedecker and Cozzarelli, 1992). There was general agreement

among the group that ferrous iron measurements using colorimetric reagents should be completed soon after

sample collection. Minimizing the amount of time between sample collection and Fe(II) measurement is necessary

because of the relatively fast oxidation rate of ferrous iron at near-neutral pH. Ferrous iron oxidizes following the

overall reaction

Fe2+ + 5/2H2

O + 1/4O2 = Fe(OH)3 + 2H+

The abiotic rate of ferrous iron oxidation is principally controlled by the partial pressure of O2, pH, and alkalinity.

At pH<3.5, the oxidation rate of ferrous iron is slow and independent of pH. At pH below about 5, however, ironoxidizing bacteria greatly increase the rate of iron oxidation. For purposes of sample preservation biotic ferrous

iron oxidation is avoided by filtration and acidification. At pH>3.5, a pH-dependent rate equation is given by Eary

and Schramke (1990):

-d[Fe2+]/dt = k[Fe2+]P O2

[OH-]2

Figure 4.3 shows the rate of ferrous iron consumption via oxidation as a function of pH (6-8) and P O2 (0.02 and

0.2 bars) estimated using this equation. Note that at pH 7, ferrous iron concentrations will decrease by a factor of

two in < 15 minutes at air-saturated conditions. A sample recovered from a well and brought to the surface does not

necessarily experience the full atmospheric P O2

 of 0.2 bars unless it is continuously agitated. A more realistic

value of P O2

 = 0.02 bars still imparts a fast loss of Fe(II), especially at pH > 7 (Figure 4.3). These calculations

illustrate the importance of carrying out Fe(II) measurements immediately after sample collection.

A common practice for estimating Fe(III) is to measure total iron (ΣFe) on filtered and acidified samples and

subtract from that quantity the concentration of ferrous iron from measurements made in the field. The group

agreed that no assumptions should be made about the speciation of the iron fraction represented by ΣFe-Fe(II). At

near-neutral pH and in the absence of significant concentrations of DOC there should be reasonable agreement

between total iron and Fe(II). However, ΣFe on filtered samples should not be assumed to be equal to Fe(II) due to

the potential formation of soluble Fe(III) - DOC complexes. Measurement of total iron can in many instances be

a useful quality check for Fe(II) measurements made in the field.

Acid mine drainage represents a special case with respect to iron measurements because at low pH the solubility of

Fe(III) increases significantly. An accurate determination of the abundances of Fe(II) and Fe(III) is especially

important because such waters are typically rich in iron and charge balance calculations strongly depend on the

Fe(II)/Fe(III) ratio (see To et al., 1999). When dealing with low pH waters it is desirable to directly measure Fe(III)

concentrations (To et al., 1999).

Field quality checks can be carried out using purchased or prepared iron solutions. Mohr’s salt (ferrous ammonium

sulfate) can be used as a standard ferrous solution. Standard solutions should be prepared daily and ideally should

be made from deoxygenated water, prepared by purging with nitrogen gas.

Sulfur Speciation

Dissolved sulfur may be present in ground water as sulfate (most oxidized form), sulfide (most reduced form), or

as species with oxidation states intermediate between sulfate and sulfide including polysulfides, sulfite, and

thiosulfate. Concentrations of the intermediate sulfur species are typically low in ground water and their

occurrence is believed to be restricted to redox interfaces that separate reducing hydrogen sulfide-bearing waters

from oxygenated waters. Such interfaces can occur within aquifers or in mixing zones between ground and surface

waters. Sampling techniques and methods for determining the concentration of intermediate sulfoxyanions can be
95

Figure 4.3 Ferrous iron oxidation kinetics at pH 6-8 and PO2

 at 0.02 and 0.2 bars.

0

2

4

6

8

10

0 2 4 6 8 10

Ferrous iron, ppm

Sulfide, ppm

pH 6

pH 8

pH 7

Figure 4.4 Iron monosulfide solubility at pH 6-8 (Ionic strength = 0.0).

pH 7, O =0.2 P 2

pH 8, O =0.02 P 2

pH 8, O =0.2 P 2

pH 7, O =0.02 P 2

pH 6
96

found in Moses et al. (1984). Sulfate concentrations in ground water vary widely from below 1 to several thousand

milligrams per liter. The highest concentrations of sulfate are typically found associated with acidic surface

drainage produced from the oxidative weathering of metal sulfides. Sulfate concentrations in ground water are

measured using gravimetric methods (precipitation as barium sulfate), the barium dichloride turbidimetric method,

or using laboratory techniques such as ion chromatography or capillary electrophoresis. Preservation is generally

not a problem for sulfate unless samples are especially rich in DOC that can be respired by sulfate reducing

bacteria. Because samples collected for anions are typically not acidified, ferrous iron-rich samples may be

problematic if sulfate becomes incorporated into iron hydroxide precipitates that form prior to sample analysis.

Hydrogen sulfide is produced in anoxic systems through the bacterial reduction of sulfate. Measurements of total

sulfide in ground water include H2

S(aq), HS-, and acid-soluble metal sulfides present in colloidal suspended

materials. These species can react rapidly in the presence of oxygen so samples must be collected with a minimum

of aeration. Methods for determining hydrogen sulfide concentrations include electrode, gravimetric, iodometric,

and colorimetric techniques.

The handling and storage methods are critical for water samples containing concentrations of dissolved sulfide.

Ideally, measurements are carried out immediately after sample collection. Preservation of samples will otherwise

be necessary to prevent losses through volatilization or oxidation. At pH less than ~6, volatilization of H2

S can be

a problem, i.e., H2

S(aq)→H2S(gas). Sulfide anti-oxidant buffers (SAOB) are alkaline solutions that contain

complexing agents such as EDTA and a reductant such as ascorbate. Hydrogen sulfide is then released from the

alkaline solutions by acidification and purging with an inert gas such as nitrogen or argon.

The most commonly used methods for determining dissolved sulfide concentrations are modifications of the

methylene blue technique described by Cline (1969). In this method, reduced sulfur species (H2

S, HS-, S2-) react

with diamine (N,N-dimethylphenyl-1,4-diamine) in the presence of ferric chloride to form a blue complex that is

measured spectrophotometrically at a wavelength of 670 nm. Visual colorimetric comparison kits are also

available. Detection limits down to 0.01 ppm sulfide are obtainable. Standard solutions can be prepared by

dissolving into deoxygenated water quantities of sodium sulfide (Na2

S × 9H2O). The concentration of sulfide is

then measured by adding excess iodine and back-titrating with thiosulfate. To minimize oxidation of sulfide

species, reagents should be added to samples immediately after sample collection followed by spectrophotometric

measurement.

Sulfide electrodes that use silver sulfide as the sensing element are generally not used in field studies. Specific ion

electrodes are better suited for laboratory studies conducted under controlled conditions. Long response times and

poisoning of the reference electrode by sulfide have limited the use of sulfide-selective electrodes (Lawrence et al.,

2000). Gravimetric determinations of sulfide concentrations are appropriate in situations where concentrations of

dissolved sulfide are > 5 mg/L. Solutions can be directly collected into bottles containing quantities of Zn acetate

or Cd acetate solutions. A 1 L sample of solution containing 5 mg/L dissolved sulfide will produce a mass of

15.2 mg or 22.5 mg of ZnS or CdS precipitate, respectively.

In ground-water systems containing dissolved sulfide, dissolved oxygen concentrations will be generally 0 or

< 1 mg/L. If hydrogen sulfide is present in ground water above 1 mg/L, concentrations of ferrous iron will generally

be negligible. The concentrations of ferrous iron and hydrogen sulfide in surface and subsurface environments are

generally governed by the solubility of iron monosulfide minerals, such as mackinawite, i.e.,

FeS + H+

 = Fe2+ + HSThe equilibrium constant (Ksp ) for this reaction is 10-3.1 where FeS is taken to be crystalline mackinawite (Davison,

1991). In Figure 4.4 the solubility curves of mackinawite are plotted at three pH values from 6 to 8. Note that

solubility decreases as pH rises, and that when concentrations of either iron or sulfide are low the other is

necessarily high.

Nitrogen Speciation

The common nitrogen species present in ground water include nitrate (NO3

-

), nitrite (NO2

-

), and ammonia (NH3 ).

At near-neutral pH typical of most ground water, ammonia is present predominantly as the ammonium ion (NH4

+

).

Although colorimetric methods are available to quantitate nitrate, nitrite, and ammonia, field determination using

these methods is not generally needed as long as laboratory measurements can be made within specified holding

times.
97

The cadmium-reduction method can be used to determine the amount of nitrate+nitrite in ground water. In this

method, cadmium metal is used to reduce nitrate to nitrite. The nitrite ions then react with sulfanilic acid and

gentistic acid to form an amber-colored compound (American Society for Testing Materials, 1995b). Strong

oxidizing and reducing substances are possible interfering components; ferric iron, in particular, will cause high

results.

Ammonia is produced during the microbiological decay of organic matter. Two colorimetric methods for ammonia

are the Nessler method and the Salicylate method. The Nessler method is probably the most commonly used

method and develops a yellow color proportional to the concentration of ammonia. High concentrations of calcium

and magnesium may cause interferences. Ferrous iron and dissolved sulfide can also interfere with ammonia

determinations using this method.

Samples collected for nitrogen speciation should be preserved by keeping cold (4°C) and measured within specified

holding times, generally < 48 h. Longer holding times are possible by acidifying samples with sulfuric acid and

keeping them cold. Before analysis the samples should be brought back to room temperature and neutralized by

adding base.

Alkalinity

Alkalinity is a measure of the acid neutralizing capacity of dissolved solutes in a water sample and is reported as

equivalents per liter or as mg/L of CaCO3

. Alkalinity is not typically included as a redox parameter. The group

included this parameter because alkalinity production often accompanies oxidation-reduction reactions, e.g., in

sulfate or iron reduction. The alkalinity concentration consists of the sum of titratable carbonate and titratable noncarbonate species in a water sample, including proton-accepting organic compounds. Proton-accepting species

include CO3

2-, HCO3-

, OH- (and metal-OH complexes), HS-, PO4

3-, ammonium, silicate, and borate. The alkalinity

measurement is important for evaluating charge balance of a solution and is a critical component of modeling

approaches using geochemical speciation computer packages. Because particulate materials can be an important

sink for acid, alkalinity determinations should be made in filtered or low-turbidity waters (turbidity < 5 NTU).

The most accurate determinations of alkalinity are made in the field at the time of sample collection. Fixed endpoint titrations to pH 4.3 are suitable for ground waters with negligible concentrations of dissolved organic matter

or titratable species other than inorganic carbon. For more complex ground waters, especially those rich in

dissolved organic material, fixed end-point titrations will give inaccurate results (see Hemond, 1990). Total

alkalinity should then be determined by measuring pH as a function of titrated acid to a pH of 3.0. Using this

method, alkalinity is calculated from the volume of acid added to reach the end-point pH. The end-point is

determined graphically by plotting pH as a function of titrant volume and locating the inflection point. The

importance of protecting the sample from oxidation was also discussed. In high Fe(II) waters, the oxidation of

Fe(II) to Fe(III) can produce sufficient acid to bias the titration, resulting in an underestimation of actual alkalinity.

Rapid analysis is recommended, and protection of the sample from aeration is needed for accurate measurements of

alkalinity in iron-rich water.

4.4 Summary

A variety of measurement options are available for monitoring the most important redox parameters in ground

water. The core list of parameters identified by this group was Eh, dissolved oxygen, iron speciation, sulfur

speciation, and nitrogen speciation. Selection of the appropriate technique for a given parameter will depend on

specific project objectives. It was recognized that project objectives and principal data uses collected in monitoring

programs change with time as more site-specific information is obtained and conceptual models evolve. There was

general agreement that there exists a need for improved standards for routine field practices. There was strong

agreement among the group that sample collection practices and measurement techniques are among the greatest

challenges to developing reliable site conceptual models.

References

American Public Health Association (APHA), American Water Works Association, Water Environment Federation.

1999. Standard Methods for the Examination of Water and Wastewater, 20th Edition. American Public Health

Association, Washington, D.C.

American Society for Testing Materials (ASTM). 1992. Standard test methods for dissolved oxygen in water, ASTM

D 888-92. Philadelphia, Pennsylvania.
98

American Society for Testing Materials (ASTM). 1995a. Standard test methods for low-level dissolved oxygen in

water, ASTM D 5543-94. Philadelphia, Pennsylvania.

American Society for Testing Materials (ASTM). 1995b. Nitrate-Nitrite in water, Test method B, ASTM D 3867-90.

Philadelphia, Pennsylvania.

Baedecker, M. J. and Back, W. 1979. Hydrogeological processes and chemical reactions at a landfill site.

Ground Water, v. 17, pp. 429-437.

Baedecker, M. J. and Cozzarelli, I. M. 1992. The determination and fate of unstable constituents of contaminated

groundwater. In Groundwater contamination and analysis at hazardous waste sites, eds. S. Lesage and R.

Jackson, Marcel-Dekker, New York, pp. 425-461.

Cline, J. D. 1969. Spectrophotometric determination of hydrogen sulfide in natural waters. Limnology and

Oceanography, v. 14, pp. 454-458.

Champ, D. R., Gulens, J., and Jackson, R. E. 1979. Oxidation-reduction sequences in ground water flow systems.

Canadian Journal of Earth Sciences, v. 16, pp. 12-23.

Davison, W. 1991. The solubility of iron sulphides in synthetic and natural waters at ambient temperature.

Aquatic Sciences, v. 53, pp. 309-329.

Eary, L. E. and Schramke, J. A. 1990. Rates of inorganic oxidation reactions involving dissolved oxygen. In Chemical

Modeling of Aqueous Systems, II, Chapter 30, eds. D. C. Melchior and R. L. Bassett, American Chemical

Society, Washington, D.C., pp. 379-396.

Gibbs, C. R. 1976. Characterization and application of ferrozine iron reagent as a ferrous iron indicator.

Analytical Chemistry, v. 48, pp. 1197-1200.

Gilbert, T. W., Behymer, T. D., and Castaneda, H. B. 1982. Determination of dissolved oxygen in natural and

wastewaters. American Laboratory, v. 14, March, pp. 119-134.

Hemond, H. F. 1990. Acid neutralizing capacity, alkalinity, and acid-base status of natural waters contaning organic

acids. Environmental Science and Technology, v. 24, pp. 1486-1489.

Hitchman, M. L. 1978. Measurement of Dissolved Oxygen, John Wiley & Sons, New York, 255 p.

Jackson, R. E. and Patterson, R. J. 1982. Interpretation of Eh and pH in a fluvial-sand aquifer. Water Resources

Research, v. 18, pp. 1255-1268.

Lawrence, N. S., Davis, J., and Compton, R. G. 2000. Analytical strategies for the detection of sulfide: a review.

Talanta, v. 52, pp. 771-784.

Langmuir, D. 1971. Eh-pH determination. In Procedures in Sedimentary Petrology, Chapter 26, ed. R. E. Carver,

Wiley-Interscience, New York, pp 597-635.

Light, T. S. 1972. Standard solution for redox potential measurements. Analytical Chemistry, v. 44, pp. 1038-1039.

Lindberg, R. D. and Runnells, D. D. 1984. Ground water redox reactions: An analysis of equilibrium state applied

to Eh measurements and geochemical modeling. Science, v. 225, pp. 925-927.

Morris, J. C. and Stumm, W. 1967. Redox equilibria and measurements of potentials in the aquatic environment. In

Equilibrium Concepts in Natural Water Systems, ed. R. F. Gould, American Chemical Society, Washington, D.C.,

pp. 270-285

Moses, C. O., Nordstrom, D. K., and Mills, A. L. 1984. Sampling and analyzing mixtures of sulphate, sulphite,

thiosulphate, and polythionate. Talanta, v. 31, pp. 331-339.

Nordstrom, D. K. 1977. Thermochemical redox equilibria of Zobell’s solution. Geochimica et Cosmochimica Acta,

v. 41, pp. 1835-1841.

Nordstrom, D. K., Jenne, E. A., and Ball, J. W. 1979. Redox equilibria of iron in acid mine waters. In ACS Symposium

Series 93, Chapter 3, ed. E. A. Jenne, American Chemical Society, Washington, D.C., pp. 51-79.

Reynolds, J. F. 1969. Comparison studies of Winkler vs oxygen sensor. Journal Water Pollution Federation, v. 41,

pp. 2002-2009.
99

Stookey, L. L. 1970. Ferrozine-A new spectrophotometric reagent for iron. Analytical Chemistry, v. 42, pp. 779-781.

Stumm, W. 1966. Redox potential as an environmental parameter; conceptual significance and operational limitation.

In 3rd International Conference on Water Pollution Research, Section 1, No. 13, Water Pollution Control Federation,

Washington D.C., pp. 1-16.

Tamura, H., Goto, K., Yotsuyanagi, T., and Nagayama, M. 1974. Spectrophotometric determination of iron (II) with

1,10 phenanthroline in the presence of large amounts of iron (III). Talanta, v. 21, pp. 314-318.

Thorstenson, D. C. 1984. The concept of electron activity and its relation to redox potentials in aqueous geochemical

systems. U.S. Geological Survey Open-File Report 84-072, 45 p.

To, T. B., Nordstrom, D. K., Cunningham, K. M., Ball, J. W., and McCleskey, R. B. 1999. New method for the direct

determination of dissolved Fe(III) concentration in acid mine waters. Environmental Science and Technology,

v. 33, pp.807-813.

U.S. Geological Survey. 1997. National Field Manual for the Collection of Water-Quality Data. U.S. Geological

Survey Techniques of Water-Resources Investigations, book 9. [http://water.usgs.gov/owq/FieldManual/

mastererrata.html]

Walton-Day, K., Macalady, D. L., Brooks, M. H., and Tate, V. T. 1990. Field methods for measurement of ground

water redox chemical parameters. Ground Water Monitoring Review, v. 10, pp. 81-89.

White, A. F., Peterson, M. L., and Solbau, R. D. 1990. Measurement and interpretation of low levels of dissolved

oxygen in ground water. Ground Water, v. 28, pp. 584-590.

Wilkin, R. T., McNeil, M. S., Adair, C. J., and Wilson, J. T. 2001. Field Measurement of Dissolved Oxygen: A

Comparison of Methods. Ground Water Monitoring and Remediation, v. 21, pp. 124-132.

Whitfield, M. 1974. Thermodynamic limitations on the use of the platinum electrode in Eh measurements. Limnology

and Oceanography, v. 19, pp. 857-865.
100
101

Donald Kampbell and Frank Chapelle

5.1 Introduction

A dissolved gas refers to a chemical element in a gaseous phase at ambient temperature. The dissolved gas is

absorbed by a liquid, in this case water, without a chemical interaction occurring. Dissolved gases such as oxygen,

hydrogen, and aliphatic hydrocarbons such as methane, are important components to determine the status of

biodegradation processes in ground water. For example, when dissolved oxygen is present, microorganisms will

preferentially use oxygen as the primary electron acceptor and aerobic metabolism will tend to dominate the

system. When dissolved oxygen is not present, however, the microbial population must use alternative electron

acceptors such as nitrate, ferric iron, sulfate, carbon dioxide, or, in some contaminated aquifers, chlorinated ethenes

in order to carry on respiration. When carbon dioxide serves as the primary electron acceptor (methanogenic

respiration) methane is produced. Thus, concentrations of dissolved oxygen and methane can be useful in

determining the kinds of microbial metabolism occurring in a ground-water system.

In general, decreasing concentrations of oxygen along ground-water flow paths are indicative of aerobic metabolism. Similarly, increasing concentrations of methane along ground-water flow paths indicate the presence of

active methanogenesis in an aquifer. Frequently, the anaerobic process becomes functional after oxygen becomes

depleted. Then nitrates and sulfates can become electron acceptors. Finally, methane can become a major

component characteristic of an anaerobic medium. Although the assay of dissolved gases is important to

characterize biodegradation, other conditions need to be favorable for a given xenobiotic such as the presence of

compatible microorganisms and favorable environmental settings. Plume ground-water that has a decrease in

oxygen and an increase in methane is an important characteristic that should be monitored.

Hydrogen is an important intermediate product of anaerobic microbial metabolism. Under anaerobic conditions,

the oxidation of organic matter requires food chains of symbiotic microorganisms. First, complex organic matter

is partially oxidized by fermentative bacteria with the production of intermediate compounds such as acetate and

molecular hydrogen. Secondly, these fermentative products are fully oxidized by respirative microorganisms

coupled to reduction of mineral electron acceptors such as nitrate, sulfate, ferric iron, and carbon dioxide.

Importantly, chlorinated ethenes such as perchloroethene (PCE) and trichloroethene (TCE) can also serve as

electron acceptors during the metabolism of hydrogen. Concentrations of dissolved hydrogen can be an indication

of the efficiency of reductive dechlorination in ground-water systems. Generally, high concentrations of dissolved

hydrogen favor efficient reductive dechlorination. The rate of reductive dechlorination at particular field sites is

useful in assessing the viability of employing monitored natural attenuation. Because of this, concentrations of

hydrogen can be a useful indicator parameter.

The complete reductive dechlorination of PCE and TCE results in the formation of ethene. Furthermore, ethene can

then be further reduced to ethane under strongly reducing methanogenic conditions. Thus, the presence of ethene

and ethane can be indicators that reductive dechlorination of PCE and TCE is occurring in a ground-water system.

For these reasons, it is important to measure concentrations of dissolved ethene and ethane at sites where reductive

dechlorination may be occurring.

Abiotic and biological transformations are important mechanisms controlling natural attenuation as a plume

management strategy. Plume management requires monitoring of the transformation processes to confirm

protection of the environment and human health. Thus, measuring concentrations of oxygen, H2

, ethene, ethane,

and methane provide important information about the kinds of microbial processes occurring in ground-water

systems.

Session 5 Summary: Session 5 Summary:

Measurement of Dissolved Gases ement of Dissolved Gases
102

5.2 Measurement Methods ement Methods

The measurement of dissolved gases in ground water can aid in understanding what biochemical processes are

active, or have been active, in ground-water systems. If oxygen has been depleted, then most likely anaerobic

processes have predominated. The presence of hydrogen is useful to stimulate dechlorination of chlorinated

hydrocarbon pollutants.

Gases such as oxygen, hydrogen, and methane are slightly soluble in water. Thus, their concentrations are directly

proportional to their partial pressure as formalized by Henry’s Law. Because the Henry’s Law constant for each gas

is highly temperature-dependent, the solubility of each may be used to calculate the amount present at saturation for

given temperatures. The solubility varies greatly with temperature. For example, the solubility of oxygen in fresh

water at 1 atm. pressure ranges from 14.6 mg/L and 7 mg/L at 0o

C and 35oC, respectively. The solubility of H2 and

methane are similarly variable with temperature. At high temperature conditions where oxygen is least soluble,

biological oxidation increases along with oxygen demand. These critical conditions can limit the purification

capacity to remove waste pollutants.

Because oxygen, hydrogen, and methane are so slightly soluble in water, sampling methods must rigorously

exclude exposure to the air. This is particularly problematic with oxygen, since air is about 21 percent oxygen.

Measurements of dissolved oxygen must always be done in a closed air-free environment, and measurements

should be made as soon after sample collection as possible. In general, all measurements of dissolved oxygen

should be made in the field at the time of sample collection. Similarly, hydrogen measurements are most accurate

when made in the field at the time of sample collection. However, storage methods are now available for

transporting hydrogen samples to the laboratory. Because methane is relatively stable, and because methane is not

a major component of air, methane samples are relatively easy to collect and store for later analysis in the

laboratory. Care must be used when collecting water samples for dissolved gases, especially oxygen and hydrogen.

Exposure to the air will lead to a potential loss or gain of these components. Water sampling and analyses should

be done in a closed air-free system and as soon as possible. Because air is about 21 percent oxygen, it can be a

critical factor of contamination when collecting low level oxygen water samples. Hydrogen and methane are

constituents of air, but in the ppm range. Generally, exchange of these components with air can be overcome by

care during sampling. The expected concentration ranges of dissolved methane and hydrogen in contaminated

ground water are in the ppm range.

The primary factor for quality assurance, besides careful sampling to avoid sample contamination by air, is

instrument calibration. Pressurized cylinders containing known hydrogen or methane, ethane, and ethene can be

used. Concentrations of oxygen in water can be obtained from standard oxygen solubility versus temperature

tables. Generally, 10 percent of the collected samples should be duplicated to measure precision.

5.3 Hydrogen

A routine technique developed to analyze water samples for dissolved hydrogen has been described by McInnes

and Kampbell (2000). This is an adaptation of a method originally developed by Chapelle et al. (1997) and is called

the “bubble strip” technique. This method is an application of Henry’s Law. After the sample is collected in a

20 mL bubble inside a 250 mL gas sample vessel, a 2 mL aliquot of the bubble is removed for analysis using a

reduction gas analyzer/chromograph. A typical calibration range for hydrogen is 0.5 to 10.0 ppm.

5.4 Oxygen

Several decades ago, the dominant procedure for determining dissolved oxygen in water was the Winkler, Full

Bottle technique (U.S. EPA, 1979b; Method 360.2). The test depended on oxygen oxidizing divalent manganese to

a higher state of valence under alkaline conditions. The higher valence state manganese would then oxidize iodine

to free iodine in an acid solution. The free iodine released was equivalent to dissolved oxygen. Standard sodium

thiosulfate then measured the free iodine. Interferences from nitrates were reduced by use of sodium azide.

Dissolved oxygen can also be analyzed by electrode probe (U.S. EPA, 1979a; Method 360.1, see Appendix A). On

the end of the probe is a thin organic film over two metallic electrodes in a media of liquid electrolyte. The oxygen

that diffuses through the film is reduced electrochemically. A current is generated proportional to oxygen in the

sample. These electrodes are well suited for analyses in water because their portability, ease of operation, and

maintenance make them convenient for field applications. They are less amenable to interferences than the

Winkler test. The principal disadvantage of electrodes for measuring dissolved oxygen relates to the diffusiondependent mode of operation. When concentrations of dissolved oxygen are low, oxygen diffusion is very low and
103

the current generated by the electrode is correspondingly low. It is thus difficult for electrodes to accurately

measure dissolved oxygen concentrations below 0.5 mg/L. Because many ground-water systems contain low

concentrations of dissolved oxygen, the use of electrodes can be problematic in some systems (see Session 4

Summary).

More recent chemical methods for measuring dissolved oxygen concentrations were developed by CHEMetrics,

Inc. One method, the indigo carmine method (ASTM D 888-87), is used for dissolved oxygen concentrations in the

1-10 ppm range. The chemistry of this method relies on the reaction of reduced indigo carmine with oxygen to

produce a blue product. Another method, the rhodazine D method (ASTM D 5543-94), is used to measure

dissolved oxygen concentrations in the 0.1 to 1.0 ppm range. The chemistry of this method relies on the reaction

of rhodazine D with oxygen to form a bright pink reaction product. Neither of these methods is subject to

interferences from temperature, salinity, or dissolved gases, which often do interfere with dissolved oxygen

electrodes. For both the high range (indigo carmine) and low range (rhodazine D), the reagents are sealed in

ampules designed to fit into a plastic beaker. The beaker is attached to a stream of water pumped from the well

(sealed from the air) and is allowed to overflow. After a brief period of flushing (~30 seconds), an ampule is

inserted into the stream of water and the tip cracked. This draws water into the ampule where it reacts with the

reagents. The intensity of the color development is directly proportional to the dissolved oxygen concentration,

which can be compared visually to standards supplied with the reagents. This method is fast, reproducible, and

minimizes sample exposure to atmospheric oxygen.

5.5 Methane, Ethane, and Ethene

Glass serum bottles are filled with ground-water samples at the field site, preserved with diluted sulfuric acid, and

capped with air-tight septa. Ten percent of the water is then removed to create a gaseous headspace. After

equilibrium by shaking the bottles, an aliquot of the headspace is removed by a gas-tight syringe for analysis by gas

chromatography. Dissolved methane, ethane, and ethene in the water sample can then be completed using Henry’s

Law as explained by Kampbell and Vandegrift (1998). Quantitation limits for methane, ethane, and ethene by this

method were 0.001, 0.002, and 0.003 mg/L, respectively.

5.6 Summary

Methods used for field site characterization studies to determine the status of biodegradation processes in ground

water as related to dissolved gases were discussed. Emphasis was directed toward dissolved oxygen, hydrogen, and

aliphatic hydrocarbons such as methane, ethane, and ethene. A method for each of these parameters is listed that

has been used over a number of years to aid in identifying processes of natural attenuation.

References

Chapelle, F. H., Vroblesky, D. A., Woodward, J. C., and Lovley, D. R. 1997. Practical considerations for measuring

hydrogen concentrations in groundwater. Environmental Science and Technology, v. 31, pp. 2873-2877.

Kampbell, D. H. and Vandegrift, S. A. 1998. Analysis of methane, ethane, and ethylene in ground-water by a

standard gas chromatographic technique. Journal of Chromatographic Science, v. 36, pp. 253-256.

McInnes, D. M. and Kampbell, D. H. 2000. The bubble stripping method for determining dissolved hydrogen (H2

)

in well water. Field Analytical Chemistry and Technology, v. 4, pp. 283-296.

U.S. Environmental Protection Agency. 1979a. Dissolved Oxygen by the Membrane Electrode Technique. Methods

for Chemical Analysis of Water and Wastes, EPA Method 360.1, EPA 600/4-79-020. U.S. EPA, Government

Printing Office, Washington, DC.

U.S. Environmental Protection Agency. 1979b. Dissolved Oxygen by the Modified Winkler Full-Bottle Technique.

Methods for Chemical Analysis of Water and Wastes, EPA Method 360.2, EPA 600/4-79-020. U.S. EPA,

Government Printing Office, Washington, DC.
104
105

Robert G. Ford

6.1 Introduction

Monitoring oxidation-reduction (redox) processes for ground-water remediation must be applied both to the liquid

and solid phases in the remediation zone. Many of the redox sensitive elements that participate in chemical

reactions during subsurface remedial processes are cycled between these two phases during the reaction lifecycle.

Examples of important redox sensitive elements include carbon, iron, nitrogen, and sulfur. Site characterization

and monitoring may require the development of a site-specific database for the composition of the aquifer solid

matrix and the capacity of the matrix with respect to remedial performance measures.

The level of effort for solid phase characterization will depend on the specific remedial technology employed. Two

examples include in-situ redox manipulation of the aquifer matrix versus emplacement of a new solid matrix to

serve as a reactive barrier. Both treatment technologies require knowledge of the mechanisms and capacity of

specific redox reactions to achieve the treatment goals. For the first example, one must identify the mass and

distribution of the reactive components naturally occurring within the contaminated aquifer. This may require an

extensive level of effort to characterize both the abundance and distribution of redox-sensitive phases as well as

their reactivity during the remedial process. For the second example, characteristics of the reactive solid

components are typically known prior to barrier installation. However, it may be necessary to carry out field

measurements to characterize changes in the reactive phase(s) during remedial monitoring activities.

6.2 Session Objectives Session Objectives

The overall goal of this breakout session was to identify the key solid phase parameters relevant to redox-controlled

processes active during in-situ remediation of ground water. As part of this discussion, the group was to identify:

1) a list of solid phase parameters that addressed a variety of remedial technologies, 2) appropriate methods of

sample collection specific to characterization needs, and 3) available analytical methods for estimating solid phase

parameters. Throughout this discussion, the group was tasked with identifying characterization issues with the

greatest level of uncertainty and making recommendations concerning technical areas where the U. S. Environmental

Protection Agency should focus investment of resources.

6.3 Case-Specific Evaluation

A remedial case study was evaluated to provide a specific frame of reference for discussion of solid phase

characterization analyses that may be required as part of remedial design and monitoring. The case study was the

remedial effort implemented by the U. S. Department of Energy to treat chromate-contaminated oxic ground water

via in-situ redox manipulation of aquifer sediments at the Hanford Site in southeastern Washington (Williams et al.,

2000). The goal of this remedial effort was to reduce naturally occurring ferric iron, Fe(III), in aquifer sediments

to soluble and solid forms of ferrous iron via injection of a buffered solution of a strong reductant, dithionite. The

ferrous iron, Fe(II), subsequently reacted with dissolved chromium; reducing Cr(VI) to insoluble Cr(III)-bearing

precipitate phases (Amonette, 2000). While the reaction of Fe(II) with Cr(VI) was the specific process controlling

chromium stabilization, the overall success of the remedial process was dependent on the efficiency of the initial

reduction of native Fe(III) via dithionite injection. Thus, solid phase characterization was focused towards

evaluating the production and longevity of Fe(II) within the treatment zone.

Design of the remedial process required analysis of the mass and distribution of reactive iron in aquifer sediments

and the rates of iron reduction/oxidation and non-beneficial reactions involving dithionite. Non-beneficial

reactions included those that consumed dithionite without production of Fe(II) and the release of potentially toxic

elements during dissolution of iron oxide minerals. Sediment cores were collected within the proposed treatment

zone for mineralogical characterization and to evaluate rates and products for beneficial and non-beneficial

reactions. All tests were carried out using the <4 mm sediment fraction, with the assumption that the largest

Session 6 Summary: Session 6 Summary:

Solid Phase Redox Characterization
106

proportion of reactive iron would be contained in this size fraction. Following is a list with test description and

purpose used for site characterization and remedial performance evaluation.

Site Geology and Hydrology

Procedures that were performed included establishing the uniformity of site geology through well-site boring logs

and characterization of physical properties on split tube samples, including porosity, bulk density, and particle size

distribution. In addition, pre- and post-dithionite injection hydraulic tests were performed to establish treatment

design criteria and to evaluate potential changes in formation hydraulic properties following dithionite injection.

Sediment Iron Mineralogy

The distribution of iron in sediment minerals was examined for the <4 mm size fraction for pre- and post-treatment

samples. The predominant phases identified for pre-treatment samples included iron adsorbed to mineral surfaces,

amorphous and crystalline Fe(III) oxides, Fe(II) carbonate, and Fe-bearing clay minerals. Post-treatment samples

had a decreased fraction of amorphous Fe(III) oxide and increased fractions of adsorbed Fe(II) and Fe(II)

carbonate. Mass balance calculations suggested that dithionite reduction primarily converted amorphous Fe(III)

oxides to adsorbed Fe(II). Separate experiments indicated that adsorbed Fe(II) provided the greatest reactivity for

Cr(VI) reduction. Thus, the observed changes in iron mineralogy provided verification of the success of dithionite

reduction.

Reagent (Dithionite) Stability

Ferric iron reduction in the presence of dithionite (S2O4

2-) occurs via a chemical reaction between Fe(III) and a

sulfoxyl free radical (SO2

-

*) that exists in equilibrium with S2O4

2- (Amonette, 2000). However, dithionite is also

subject to pH-dependent disproportionation reactions resulting in other dissolved forms of sulfur (S2

O3

2- and SO32-)

that reduce Fe(III) at a much slower rate. These latter reactions are non-beneficial and occur in solution or may be

catalyzed at mineral surfaces. Tests were carried out to evaluate the optimal reaction pH and the rate at which

aquifer sediments catalyzed dithionite disproportionation. These parameters were required to establish the extent

to which the injection solution pH had to be buffered and the optimal rate of injection into the aquifer in order to

minimize the impact of dithionite disproportionation.

Sediment Oxidation Capacity

Oxidation capacity was defined for this study as the degree of dithionite consumption during the reduction of

Fe(III) in aquifer sediments. These results were corrected for disproportionation reactions that resulted in

dithionite oxidation without concurrent Fe(III) reduction. An operational definition for oxidation capacity was

required, since not all Fe(III) was accessible to reduction due to physical constraints or slow rates of reduction for

certain mineral phases. This result highlights the need to tailor characterization methods for consistency with the

remedial technology.

Sediment Reduction Capacity

Sediment reduction capacity for post-treatment sediment samples was characterized by exposure of reduced

sediments to oxygen-saturated synthetic ground water. Rates of Fe(II) oxidation in 1-D column studies, calculated

from oxygen consumption, were used to estimate the lifetime of Fe(II) within the in-situ reaction zone. These tests

were run on sediments collected prior to and following dithionite injection. Reduction capacity for pre-treatment

sediments was evaluated in the laboratory following dithionite treatment. Comparable reduction capacities were

measured for pre- and post-treatment sediment samples, providing a performance measure for success of the

reduction stage of the remedial effort. The rate of Fe(II) oxidation was also used to provide an estimate of in-situ

reaction zone longevity.

Contaminant Stabilization and Trace Metal Mobilization

The stability of reduced chromium was evaluated during sediment re-oxidation with oxygen-saturated water to

simulate field conditions (Fruchter et al., 2000; Williams et al., 2000). No leaching of chromium occurred as a

result of re-oxidation. In addition, the potential mobilization of native trace metals in the aquifer sediments that are

sensitive to changes in ground-water redox was examined during reduction with dithionite. Release of metals such

as arsenic did occur, but soluble concentrations were elevated only in reduced sediments. Soluble metal

concentrations decreased significantly during sediment re-oxidation. These results indicated that sediment

reduction effectively stabilized chromium and did not negatively impact water quality due to concurrent trace metal

leaching. In order to minimize soluble trace metal concentrations in the treated aquifer, the spent dithionite
107

injection reagent and several additional injection volumes of ground water were withdrawn following the

predetermined iron reduction period.

This example illustrates the rationale and procedures that may be a necessary component of remedial evaluation

and monitoring. Procedures such as assessment of site geology may be common to a wide range of remedial

strategies. Assessment of aquifer sediment capacity parameters may also be common, but may require tailored

procedures that are more consistent with remedial objectives. In summary, it is clear that solid phase characterization procedures will likely be required at all stages of remedial design, implementation and monitoring. In order to

provide a broader scope to encompass other remedial technologies, the discussion panel developed a more

universal list of potential characterization procedures and evaluated their utility for a wider range of remedial

technologies. These observations are discussed in the following section.

6.4 Solid Phase Characterization: Evaluation of Information Needs Solid Phase Characterization: Evaluation of Information Needs

The discussion group developed a list of potential solid phase measurements to be applied as part of site assessment

and monitoring. Some measurements would be required as part of the initial remedial assessment, while other

measurements would primarily be implemented as part of remedial monitoring. Measurement of the listed

parameters may require field- and/or laboratory-based tests. A general description is provided in Table 6.1 for each

of the discussed parameters. There may be additional parameters that require characterization for specialized

remedial technologies, but the list in Table 6.1 is generally comprehensive.

The list in Table 6.1 was then used as a reference point for evaluating data requirements for assessment, design, and

monitoring of a select number of remedial technologies. The outcome of this activity is illustrated in Table 6.2.

The remedial technologies that were evaluated included oxidant injection (permanganate), reductant injection

(dithionite), reactive barrier installation (zero valent iron), enhanced biological degradation, and monitored natural

attenuation (MNA). Several general trends were observed:

1) The reactive barrier technology requires the least amount of solid phase characterization, since it was

assumed that properties of the barrier material would be fully characterized. This also assumes

sufficient knowledge of the influence of ground-water geochemical parameters on the (bio)chemical

reactions controlling contaminant remediation within the barrier. However, some characterization of

the native aquifer material may be required to address potential deleterious effects from alteration of

the ground-water redox state.

2) Injection (oxidant/reductant), enhanced biological, and MNA technologies require more extensive

solid phase characterization, since the aquifer matrix (mineralogy and microbial population) is an

integral reactant in the remedial technology.

3) Use of MNA as a remedial technology for an inorganic contaminant will necessitate the greatest extent

of solid phase characterization. This observation is due to the importance of inorganic contaminant

immobilization onto aquifer solids as the remedial endpoint.

Analysis of this select group of remedial technologies indicates that a single list of required solid phase

characterization parameters will not apply for all technologies. Of the listed parameters, only solubilized nontarget contaminant and aquifer permeability were considered universal for all technologies. In addition, a

determination must be made whether solid phase characterization is performed on the entire sample or a specific

size fraction. Parameters that pertain to a reaction capacity, such as acid neutralizing capacity, should be assessed

for the entire sample. However, determinations of mineralogy may only be feasible following initial isolation of a

specific size fraction, e.g., determination of the clay mineral content. Thus, it was generally recommended that a

preliminary case-specific assessment should be made to identify solid phase parameter measurements critical to

successful implementation of the proposed remedial technology.

6.5 Recommended Procedures: Sample Collection and Characterization

Collection and Preservation of Solid Phase Material Collection and Preservation of Solid Phase Material

Recommendations for collection and preservation of solid phase material and its derived components were solicited

from the discussion group. In general, collection and preservation methods were grouped based on the desired

parameter to be characterized. The recommendations listed below were specific to the retrieval of subsurface

material from the saturated zone. This type of sample will generally display the greatest instability during aboveground processing.
108

Mineralogy. The greatest concerns for preservation of mineralogy in aquifer materials were the prevention of

contact with oxygen and removal of pore water. For example, iron- and manganese-bearing minerals may exist in

a reduced state within the saturated zone. Minerals that are stable under a reducing environment are subject to

significant alteration upon exposure to oxygen. In addition, solid phase structural and chemical transformations are

commonly mediated or facilitated by the pore water. Thus, removal of pore water may act to retard or impede

transformation. If dewatering is required to minimize sample alteration, it was recommended that non-aggressive

techniques, such as draining by gravity, be employed for removal of pore water.

General methods for sample handling were grouped into those employed in the field and laboratory. In the field,

sample cores should be immediately capped and frozen. Sample freezing can be accomplished either by

submersing in liquid nitrogen or placement in a portable freezer located in the field. Following transport to a

laboratory setting, frozen materials should be thawed under an oxygen-free or inert atmosphere, e.g. within a glove

box. Residual pore water may then be displaced using a non-reactive solvent such as acetone. However, while this

treatment will leave the mineralogy intact, it will likely alter the solid organic matter. Drying at ambient

temperature in air or an inert atmosphere should be employed in order to minimize alteration of amorphous

minerals. Significant changes in reactivity may result from changes in mineral structure and surface area due to

drying at even slightly elevated temperatures (e.g., Stanjek and Weidler, 1992). Some of the group participants

have had success storing materials in de-oxygenated water. However, extended periods of storage should be

avoided, since sample mineralogy will alter with time during approach to a new equilibrium with the fresh deoxygenated water.

Microbiology. The method used to preserve the microbial community within the core depends on the goal of

characterization. If characterization requires quantification of bacterial counts, then the solids should be

maintained at a reduced temperature, e.g. 4 °C, but not frozen. However, if there is no need to maintain a viable

microbial population, then samples can be frozen in the field. No other recommendations were made with regard

to microbial characterization, since the general group consensus was that this is a less problematic issue than

mineralogical preservation.

Pore Water. It was recommended that pore water be isolated from the core material in the field prior to

preservation of core solids. This may be accomplished for coarse-grained material by allowing the pore water to

drain into a collection container. Pore water may be extracted from fine-grained material by manual pressure

filtration using a syringe. Alternatively, pore water may be isolated by centrifugation if the location for sample

collection is close to laboratory facilities. Pore water should not be isolated in the laboratory after the core has been

subjected to freezing and thawing. Preservation of isolated pore water should be consistent with the desired

analytical measurements.

Characterization Methods Characterization Methods

Insufficient time was available to make recommendations for methods to be employed to characterize all of the

solid phase parameters listed in Table 6.1. Discussion was thus directed towards a select number of parameters. In

all cases, the subsurface material bulk density should be measured so that the chemical parameters listed in

Table 6.1 can be reported on a mass basis. In addition, any testing protocol must include evaluation of the

performance of the analytical procedures. Methods referenced in Tables 6.3 and 6.4 should be initially evaluated

to assess their appropriateness for site materials. The capacities of materials derived from impacted environments

may exceed the analytical range for which the methods were originally developed.

Oxidation Capacity. Methods for the determination of an ‘oxidation capacity’ for solids are listed in Table 6.3. The

most aggressive approach involves digestion of the solid with a strong chemical reductant, i.e., aqueous Cr(II).

This approach may overestimate the effective oxidation capacity of the solid material. A more targeted approach

may be achieved via dithionite titration, but this may only be useful for evaluation of dithionite injection as a

remedial technology. Likewise, methods designed to target iron oxides will only be valid for aquifer sediments in

which the oxidation capacity is dominated by these minerals. Development of a tailored method may prove most

useful, but this will necessitate additional research and development.

Reduction Capacity. Several methods are documented for the determination of a ‘reduction capacity’ for the solid

material. A selection of these methods is listed in Table 6.4. Methods for determining chemical oxygen demand

(COD) and organic matter content were not developed for the purpose of collecting data relevant towards

implementation of a remedial technology. The reducing capacity may be under- or overestimated relative to the

remedial target. However, these methods are well documented and have been commonly employed for solid phase

characterization in soil and waste treatment studies. The method for measuring dissolved oxygen demand
109

developed at the Pacific Northwest National Laboratory was designed to provide data specific to the remedial

technology employed (Williams et al., 2000). This method involves measuring oxygen demand in a continuous

Table 6.1 Solid Phase Parameters Requiring Measurement as Part of Remedial Assessment, Design, and Monitoring

Solid Phase Parameter Parameter Description 

Oxidation Capacity Capacity of sediment to oxidize a reduced chemical (contaminant or 

introduced remedial reagent) 

Reduction Capacity Capacity of sediment to reduce an oxidized chemical (contaminant or 

introduced remedial reagent) 

Neutralization Capacity Capacity of solid phase to buffer change in pH (acid or alkaline) 

Contaminant Sorption 

Capacity 

Total mass of contaminant that can be partitioned to solid phase by 

various mechanisms 

Ion Exchange Capacity Total mass of charged ions that can be partitioned to solid phase via 

an electrostatic mechanism 

Sorbed Contaminant Mass of contaminant that is partitioned to solid phase 

Solubilized Non-Target 

Contaminant 

Mass of non-target contaminant associated with solid phase that may 

be solubilized as a result of remediation 

Reagent Stability Identification of undesirable by-products produced during reaction 

between remedial reagent and contaminant/ solid phase 

Mineralogy (Bulk and Trace) Identity of mineral phases present in various size fractions of solid 

phase 

Non-Target Solid Phase 

Contaminant 

Non-target contaminant that may negatively interact with remedial 

technology 

Extractable Fe/Al/Mn Mass of Fe/Al/Mn extracted from solid phase using reagents designed 

to attack specific mineral phases 

Extractable Sulfide Mass of sulfide extracted from solid phase using reagents designed to 

attack specific sulfidic mineral fractions (see Appendix A) 

Total Organic Carbon Mass of carbon associated with organic solid phases in sediment (see 

Appendix A) 

Total Inorganic Carbon Mass of carbon associated with inorganic solid phases in sediment 

Reduction/Oxidation Reaction 

Rate 

The rate at which solid phase will reduce/oxidize (consume) 

oxidizing/reducing reagent 

Microbial Activity or 

Physiology 

Characterization of the microbial processes/characteristics controlling 

contaminant degradation/stabilization 

Microbial Population Identification of the species of microbes that inhabit contaminated 

solid phase 

Aquifer Permeability The ability of aquifer material to transmit ground water based on 

measurement of porosity 
110

Table 6.2 Solid Phase Measurements Applied as Part of Site Assessment and Monitoring for a Selected Number

of Remedial Technologies.

Remedial Technology 

Measured 

Parameter 

Oxidant 

Injection 

(Permanganate) 

Reductant 

Injection 

(Dithionite) 

Reactive 

Barrier 

(ZVI) 

Enhanced 

Biological 

Degradation 

MNA 

Inorganic 

Contaminant 

Oxidation Capacity X X X X 

Reduction Capacity X X X X 

Neutralization 

Capacity X X X X 

Contaminant 

Sorption Capacity (X) X 

Ion Exchange 

Capacity X 

Sorbed Contaminant X 

Solubilized NonTarget Contaminant X X X X X 

Reagent Stability X X 

Mineralogy (Bulk 

and Trace) X (X) X 

Solid Phase 

Contaminant X X X 

Extractable Fe/Al/Mn X X X 

Extractable Sulfide (X) X 

Total Organic Carbon X X X 

Total Inorganic 

Carbon X 

Reduction/Oxidation 

Reaction Rate X X 

Microbial Activity or 

Physiology X [X] 

Microbial Population [X] [X] [X] 

Aquifer Permeability X X X X X 

The symbols used to indicate required technology-specific measurements are as follows: X = site assessment and monitoring, (X) = postremediation characterization only, and [X] = measurement not always required. ZVI = Zero Valent Iron, MNA = Monitored Natural Attenuation.
111

Table 6.3 Methods That May Be Employed for Estimating an Oxidation Capacity for Solid Materials

Table 6.4 List of Methods That May Be Employed for Estimating a Reduction Capacity for Solid Materials

Method Source Comments 

Cr(II) 

(Barcelona and Holm, 1991a) 

(Barcelona and Holm, 1991b) 

(Barcelona and Holm, 1992) 

Most aggressive but oxygen-free 

atmosphere recommended; high 

estimate 

Digestion with Ti(III)-EDTA (Ryan and Gschwend, 1991) 

Developed for extraction of Fe 

oxides; applicability limited to iron 

oxide dominated solids 

Titration/digestion with dithionite 

solution 

(Loeppert and Inskeep, 1996) 

(Williams et al., 2000) 

Valid for only specific remedial 

technology; targets Fe oxides 

Exposure to treatment reagent Not Documented 

Case specific evaluation; requires 

preliminary research and 

development effort 

Estimate based on mineralogy Not Documented No consensus values for reference 

minerals; unreliable value 

Method Source Comments 

Chemical Oxygen Demand by 

digestion with acid dichromate 

(U.S. EPA, 1979) 

(Barcelona and Holm, 1991a) 

(Barcelona and Holm, 1991b) 

(Barcelona and Holm, 1992) 

Precipitate coatings if pH not 

buffered; high estimate 

Digestion in hydrogen peroxide 

solution (Nelson and Sommers, 1996) Developed for quantifying organic matter content 

Dissolved oxygen consumption in 

air-saturated water (Williams et al., 2000) 

Dynamic column test with 

mathematical simulation; test design 

must minimize gas diffusion from 

external sources; time-consuming, 

but realistic 

Exposure to treatment reagent Not Documented 

Case specific evaluation; requires 

preliminary research and 

development effort 

Estimate based on mineralogy Not Documented No consensus values for reference 

minerals; unreliable value 
112

flow column configuration. Variation of column flow parameters and mathematical simulation may be required to

arrive at a reliable estimate. Implementation of this method would strongly depend on the level of heterogeneity of

the aquifer solids within the treatment zone and the relative value of the data for successful implementation of the

remedial technology. As with recommendations for oxidizing capacity measurements, method development or

modification may be required to optimize the usefulness of the derived data.

Acid-Base Neutralizing Capacity. The ability of the aquifer solids to buffer changes in ground water pH is a critical

parameter for the success of a remedial technology dependent on redox reactions. Many redox reactions involve

the production/consumption of protons/hydroxyl, and the reaction rates may be optimal only within a restrictive pH

range. Thus, the acid-base neutralizing capacity of the aquifer material must be known in order to evaluate the

efficiency of the desired redox reaction(s). It may prove necessary to supplement the aquifer neutralizing capacity

depending on the remedial technology that is employed (e.g., Williams et al., 2000). Determination of the acidbase neutralizing capacity may be readily achieved through use of a standardized acid/base solution. The primary

decision with regard to quantification of the capacity is the time for equilibration at the desired pH endpoint, since

part of the neutralizing capacity of a soil is derived from rate-dependent mineral dissolution reactions. A reactionbased determination consistent with anticipated site conditions provides the most reliable estimate. It is

inappropriate to base neutralizing capacity estimates on bulk mineralogical content (e.g., total carbonate analysis),

since a fraction of the mineral content of the aquifer may not be reactive under site conditions.

Microbially Reducible Iron. The recognition that degradation of organic contaminants in ground water can be

coupled to the microbial reduction of ferric iron in aquifer sediment minerals has prompted the development of

screening methods for the estimation of solid phase microbially reducible iron (e.g., Lovley and Phillips, 1987).

Quantification of the reducible ferric iron fraction by this type of method typically involves the determination of

dissolved iron. The inherent assumption behind such a test is that solid phase ferric iron that is easily reduced under

mild conditions by a chemical reductant can be correlated to the fraction of microbially reducible iron. However,

there was no consensus within the group concerning the iron mineralogy associated with this type of extraction. In

addition, there is evidence that iron bound in clay minerals may serve as a terminal electron acceptor in the absence

of iron dissolution (Ernstsen et al., 1998). Thus, estimates from this type of extraction should be viewed with

caution. There was group consensus that observed changes in the ‘reducible’ solid phase iron fraction within an

aquifer sediment may be used as an indicator of iron serving as a terminal electron acceptor. However, this

conclusion is only valid if the absence of abiotic iron reduction reactions can be verified. There is also evidence for

reduction of solid phase iron in soils through biologically controlled reactions for which no terminal electron

acceptor process is active (Schwertmann, 1991).

6.6 Session Conclusions Session Conclusions

It was clear from this discussion that a single, clear-cut approach towards characterization of solid phase redox

parameters is not feasible. While technical consensus may be reached for a select number of analytical approaches,

a complete characterization effort will require a high degree of technical expertise. However, there was consensus

that this type of analysis is critical towards the successful implementation of many in-situ remedial technologies.

Thus, the discussion group arrived at the following overall conclusions:

1. Solid phase characterization will play a significant role for some remedial technologies.

2. More solid phase characterization will be required for technologies that rely on the inherent or

enhanced treatment capacity of the aquifer matrix versus installation of a barrier technology.

3. Evaluation of several solid phase characteristics may be of common importance to a variety of

remedial technologies, but specificity based on site, contaminant, or remediation technique prevents a

universal approach.

4. Recommendations can be made for best practices relative to characterization of solid phase redox

parameters, but it is not technically feasible to issue specific (restrictive) requirements.

5. Synthesis and documentation of existing methods and development of new methods is required to

fulfill data needs.

Solid phase characterization issues have generally been avoided from the standpoint of regulatory guidance due to

the complexity of the technical issues. However, the need for a consistent framework for evaluating solid phase

redox processes is significant with increased implementation of remedial technologies that manipulate or monitor

solid phase processes.
113

References

Amonette, J. E. 2000. Iron Redox Chemistry of Clays and Oxides: Environmental Applications. In Electrochemistry

of Clays, ed. A. Fitch, Clay Minerals Society, Aurora, Colorado.

Barcelona, M. J. and Holm, T. R. 1991a. Oxidation-reduction capacities of aquifer solids. Environmental Science

and Technology, v. 25, pp. 1565-1572.

Barcelona, M. J. and Holm, T. R. 1991b. Oxidation-reduction capacities of aquifer solids. Environmental Science

and Technology, v. 25, 2540.

Barcelona, M. J. and Holm, T. R., 1992. Oxidation-reduction capacities of aquifer solids. Environmental Science

and Technology, v. 26, pp. 2538-2539.

Ernstsen, V., Gates, W. P. and Stucki, J. W. 1998. Microbial reduction of structural iron in clays - A renewable source

of reduction capacity. Journal of Environmental Quality, v. 27, pp. 761-766.

Fruchter, J. S., Cole, C. R., Williams, M. D., Vermeul, V. R., Amonette, J. E., Szecsody, J. E., Istok, J. D., and M. D.

Humphrey. 2000. Creation of a subsurface permeable treatment zone for aqueous chromate contamination using

in situ redox manipulation. Ground Water Monitoring & Remediation, v. 20, pp. 66-77.

Loeppert, R. H. and Inskeep, W. P. 1996. Iron. In Methods of Soil Analysis: Part 3 - Chemical Methods, ed. D. L.

Sparks, Soil Science Society of America Book Series, Soil Science Society of America, Inc., Madison, Wisconsin,

pp. 639-664.

Lovley, D. R. and Phillips, E. J. P. 1987. Rapid assay for microbially reducible ferric iron in aquatic sediments.

Applied and Environmental Microbiology, v. 58, pp. 1536-1540.

Nelson, D. W. and Sommers, L. E. 1996. Total Carbon, Organic Carbon, and Organic Matter. In Methods of

Soil Analysis: Part 3 - Chemical Methods, ed. D. L. Sparks, Soil Science Society of America Book Series, Soil

Science Society of America, Inc., Madison, Wisconsin, pp. 961-1010.

Ryan, J. N. and Gschwend, P. M. 1991. Extraction of iron oxides from sediments using reductive dissolution by

titanium(III). Clays and Clay Minerals, v. 39, pp. 509-518.

Schwertmann, U. 1991. Solubility and dissolution of iron oxides. Plant and Soil, v. 130, pp. 1-25.

Stanjek, H. and Weidler, P. G. 1992. The effect of dry heating on the chemistry, surface area, and oxalate solubility

of synthetic 2-line and 6-line ferrihydrites. Clay Minerals, v. 27, pp. 397-412.

U.S. Environmental Protection Agency. 1979. Chemical Oxygen Demand. Methods for Chemical Analysis of Water

and Wastes, EPA Method 410.4, EPA 600/4-79-020. U.S. EPA, Government Printing Office, Washington, DC.

 Williams, M. D., Vermeul, V. R., Szecsody, J. E. and Fruchter, J. S. 2000. 100-D Area In Situ Redox Treatability Test

for Chromate-Contaminated Groundwater, PNNL-13349, Pacific Northwest National Laboratory,

Richland, Washington.
114
115

Appendix Appendix A

A list of EPA documented analytical methods for redox sensitive elements is included in Table A1. Abbreviated

descriptions for these methods are provided in Table A2. The full references in Table A1 should be consulted for

details on these methods. There are no EPA documented methods for the determination of ferrous iron or

oxidation-reduction potential in aqueous solutions. Summary Chapter 4 should be consulted for these methods.

Included in Tables A1 and A2 are methods for the determination of total organic carbon, total inorganic carbon and

acid-volatile sulfur in solids. There are no EPA documented methods for total sulfur and acid-digestible iron in

solids. Multiple procedures listed in Tables A1 and A2 for a given element are due to either differences in

analytical range or method of detection. The most appropriate method will be based on the total analyte mass and

specific matrix characteristics of the sample. The listed methods were developed for implementation in a

laboratory setting, though many of the colorimetric and potentiometric methods for aqueous solutions have been

adapted to field-based procedures by several commercial vendors.
116

Table A1. Table A1. EPA Methods for Selected Redox Parameters

Chemical or Method Name Number EPA ISSUE DATE

Alkalinity - Colorimetric, Automated 0310.2 600/4-79-020

Alkalinity - Titrimetric, pH 4.5 0310.1 600/4-79-020

Ammonia in Estuarine by Gas Segmented CF/CA 0349.0 600/R-97-072 09/01/1997

Ammonia-N / automated colorimetry / prelim distill 1690 02/01/1999

Ammonia-N /ion-selective electrode potentiometry 1689 02/01/1999

Ammonium in Wet Deposition by Electrometric 0350.6 600/4-86-024 03/01/1986

Ammonium in Wet Deposition/Automated Colorimetric 0350.7 600/4-86-024 03/01/1986

Bomb Preparation Method for Solid Waste/Metallic 5050 SW-846 Ch 5 09/01/1994

Carbon & Nitrogen in sediments & particulates 0440.0 600/R-97-072 09/01/1997

Extractable Sulfides 9031 SW-846 Ch 5 07/01/1992

Inorganic Anions by Ion Chromatography 0300.0 600/R-93-100 08/01/1993

Inorganic Anions by Ion Chromatography 9056 SW-846 Ch 5 09/01/1994

Inorganic Anions by Ion Chromatography 9056A SW-846 Ch 5

Inorganic Anions in Aqueous Matrices - CIE 6500 SW-846 Ch 3.3 01/01/1998

Nitrate - Ion-Selective Electrode 9210 SW-846 Ch 5 12/01/1996

Nitrate - Ion-Selective Electrode 9210A SW-846 Ch 5

Nitrate & Nitrite by Gas Segmented CF/CA 0353.4 600/R-97-072 09/01/1997

Nitrate/nitrite-N in biosolids - auto photometry 1685 02/01/1999

Nitrate/nitrite-N in biosolids/ manual colorimetry 1686 02/01/1999

Nitrate-N by IC 0300.0 600/4-79-020

Nitrate-Nitrate by Automated Colorimetry 0353.2 600/R-93-100 08/01/1993

Nitrate-Nitrite by Automated Colorimetric 0353.6 600/4-86-024 03/01/1986

Nitrite by IC 0300.0 600/4-79-020

Nitrogen, Ammonia - Colorimetric 0350.1 600/4-79-020

Nitrogen, Ammonia - Colorimetric, Titrimetric 0350.2 600/4-79-020

Nitrogen, Ammonia - Potentiometric 0350.3 600/4-79-020

Nitrogen, Kjeldahl, Total - Colorimetric 0351.2 600/4-79-020

Nitrogen, Kjeldahl, Total - Colorimetric/Automated 0351.1 600/4-79-020

Nitrogen, Kjeldahl, Total - Colorimetric/Titrimetric 0351.3 600/4-79-020

Nitrogen, Kjeldahl, Total - Potentiometric 0351.4 600/4-79-020

Nitrogen, Nitrate - Colorimetric, Brucine 0352.1 600/4-79-020

Nitrogen, Nitrate-Nitrite - Colorimetric/Cadmium 0353.2 600/4-79-020

Nitrogen, Nitrate-Nitrite - Colorimetric/Hydrazine 0353.1 600/4-79-020

Nitrogen, Nitrate-Nitrite - Manual Cadmium 0353.3 600/4-79-020

Nitrogen, Nitrite - Spectrophotometric 0354.1 600/4-79-020

Organic Carbon, Total - Combustion or Oxidation 0415.1 600/4-79-020

Organic Carbon, Total - UV Promoted 0415.2 600/4-79-020

Oxygen, Dissolved - Membrane Electrode (Probe) 0360.1 600/4-79-020

Oxygen, Dissolved - Modified Winkler 0360.2 600/4-79-020

Sulfate 0300.0 600/4-79-020

Sulfate - Colorimetric, Automated, Chloranilate 0375.1 600/4-79-020

Sulfate - Colorimetric, Automated, Chloranilate 9035 SW-846 Ch 5 09/01/1986

Sulfate - Colorimetric, Automated, Methylthymol 0375.2 600/4-79-020

Sulfate - Colorimetric, Automated, Methylthymol 9036 SW-846 Ch 5 09/01/1986

Sulfate - Colorimetry, Automated 0375.2 600/R-93-100 08/01/1993

Sulfate - Gravimetric 0375.3 600/4-79-020

Sulfate - Turbidimetric 0375.4 600/4-79-020

Sulfate - Turbidimetric 9038 SW-846 Ch 5 09/01/1986

Sulfide - Colorimetric, Methylene Blue 0376.2 600/4-79-020

Sulfide - Ion-Selective Electrode 9215 SW-846 Ch 5 12/01/1996

Sulfide - Titrimetric, Iodine 0376.1 600/4-79-020

Sulfides, Acid-Soluble & Insoluble - Titrimetric 9034 SW-846 Ch 5 12/01/1996

Sulfides, Acid-Soluble & Insoluble -Distillation 9030B SW-846 Ch 5 12/01/1996

Total Kjeldahl Nitrogen - Semi-Automated 0351.2 600/R-93-100 08/01/1993

Total Kjeldahl nitrogen / preliminary 1687 02/01/1999

Total Kjeldahl nitrogen / preliminary 1688 02/01/1999

Total Organic Carbon 9060 SW-846 Ch 5 09/01/1986
117

SOURCES OF EPA TEST METHODS

Index to EPA Test Methods, May 2000 Revised Edition, U.S. EPA Region 1 Library at http://www.epa.gov/region01/

oarm/testmeth.pdf

EPA 530/SW-846 Test Methods for Evaluating Solid Waste: Physical/Chemical Methods 3rd ed - 4 vols. November

1986. (Subscription from GPO / SN955-001-00000-1; one-time print purchase from NTIS; also on CD ROM

from NTIS and Solutions) All methods including Update IVA and Update IVB and other pre-releases online at

www.epa.gov/epaoswer/hazwaste/test/sw846.htm

EPA 600/R-93-100 Methods for the Determination of Inorganic Substances in Environmental Samples. August 1993

(NTIS / PB94-120821)

EPA 600/R-97-072 Methods for the Determination of Chemical Substances in Marine and Estuarine Environmental

Matrices. 2nd edition. September 1997. (NSCEP or NTIS /PB97-127326 or www.epa.gov:80/nerlcwww/

ordmeth.htm or CD ROM)

EPA 600/4-79-020 Methods for Chemical Analysis of Water and Wastes. Revised March 1983 (NTIS / PB84-128677

for paper; CD ROM; NEPIS/ http://www.epa.gov/cincl)

EPA 600/4-86-024 Development of Standard Methods for the Collection and Analysis of Precipitation. May 1986

(NTIS / PB86-201365)
118

 	
 

 




 	

 
   

  

   




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129

 	
 

 




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130

 	
 

 




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Table A1. Table A1. continued
131
United States

Environmental Protection

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January 2002

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 Research Laboratory

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