Ground Water

2 Printed on pape

least 20 percen
EPA/540/R-97/504

May 1997

Proceedings of the 

Symposium on Natural Attenuation of Chlorinated Organics 

in Ground Water

Office of Research and Development

U.S. Environmental Protection Agency

Washington, DC
Disclaimer

The projects described in this document have been reviewed in accordance with the peer and

administrative review policies of the U.S. Environmental Protection Agency and the U.S. Air Force,

and have been approved for presentation and publication. Mention of trade names or commercial

products does not constitute endorsement or recommendation for use.

ii
Contents

Page

Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . viii

Executive Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ix

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . x

Where Are We Now? Moving to a Risk-Based Approach

C.H. Ward. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

Where Are We Now With Public and Regulatory Acceptance? (Resource Conservation and Recovery

Act [RCRA] and Comprehensive Environmental Response, Compensation, and Liability Act [CERCLA])

Kenneth Lovelace. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

Biotic and Abiotic Transformations of Chlorinated Solvents in Ground Water

Perry L. McCarty . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

Microbiological Aspects Relevant to Natural Attenuation of Chlorinated Ethenes

James M. Gossett and Stephen H. Zinder . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Microbial Ecology of Adaptation and Response in the Subsurface

Guy W. Sewell and Susan A. Gibson . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

Identifying Redox Conditions That Favor the Natural Attenuation of Chlorinated Ethenes in

Contaminated Ground-Water Systems

Francis H. Chapelle . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

Design and Interpretation of Microcosm Studies for Chlorinated Compounds

Barbara H. Wilson, John T. Wilson, and Darryl Luce . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

Conceptual Models for Chlorinated Solvent Plumes and Their Relevance to Intrinsic Remediation

John A. Cherry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

Site Characterization Tools: Using a Borehole Flowmeter To Locate and Characterize the 

Transmissive Zones of an Aquifer

Fred Molz and Gerald Boman . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

Overview of the Technical Protocol for Natural Attenuation of Chlorinated Aliphatic Hydrocarbons in

Ground Water Under Development for the U.S. Air Force Center for Environmental Excellence

Todd H. Wiedemeier, Matthew A. Swanson, David E. Moutoux, John T. Wilson, 

Donald H. Kampbell, Jerry E. Hansen, and Patrick Haas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

The BIOSCREEN Computer Tool

Charles J. Newell, R. Kevin McLeod, and James R. Gonzales . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

iii
Case Study: Naval Air Station Cecil Field, Florida

Francis H. Chapelle and Paul M. Bradley . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66

Case Study of Natural Attenuation of Trichloroethene at St. Joseph, Michigan

James W. Weaver, John T. Wilson, and Donald H. Kampbell . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67

Extraction of Degradation Rate Constants From the St. Joseph, Michigan, Trichloroethene Site

James W. Weaver, John T. Wilson, and Donald H. Kampbell . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71

Natural Attenuation of Chlorinated Aliphatic Hydrocarbons at Plattsburgh Air Force Base, New York

Todd H. Wiedemeier, John T. Wilson, and Donald H. Kampbell . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76

Case Study: Natural Attenuation of a Trichloroethene Plume at Picatinny Arsenal, New Jersey

Thomas E. Imbrigiotta, Theodore A. Ehlke, Barbara H. Wilson, and John T. Wilson . . . . . . . . . . . . . 85

Case Study: Plant 44, Tucson, Arizona

Hanadi S. Rifai, Philip B. Bedient, and Kristine S. Burgess . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92

Remediation Technology Development Forum Intrinsic Remediation Project at 

Dover Air Force Base, Delaware

David E. Ellis, Edward J. Lutz, Gary M. Klecka, Daniel L. Pardieck, Joseph J. Salvo, 

Michael A. Heitkamp, David J. Gannon, Charles C. Mikula, Catherine M. Vogel, 

Gregory D. Sayles, Donald H. Kampbell, John T. Wilson, Donald T. Maiers. . . . . . . . . . . . . . . . . . . . 95

Case Study: Wurtsmith Air Force Base, Michigan

Michael J. Barcelona . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

Case Study: Eielson Air Force Base, Alaska

R. Ryan Dupont, K. Gorder, D.L. Sorensen, M.W. Kemblowski, and Patrick Haas. . . . . . . . . . . . . . . 106

Considerations and Options for Regulatory Acceptance of Natural Attenuation in Ground Water

Mary Jane Nearman. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112

Lessons Learned: Risk-Based Corrective Action

Matthew C. Small . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116

Informal Dialog on Issues of Ground-Water and Core Sampling

Donald H. Kampbell . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 118

Appropriate Opportunities for Application—Civilian Sector (RCRA and CERCLA)

Fran Kremer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

Appropriate Opportunities for Application—U.S. Air Force and Department of Defense

Patrick Haas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 124

Intrinsic Remediation in the Industrial Marketplace

David E. Ellis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 129

Contents (continued)

Page

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Environmental Chemistry and the Kinetics of Biotransformation of Chlorinated Organic 

Compounds in Ground Water

John T. Wilson, Donald H. Kampbell, and James W. Weaver . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133

Future Vision: Compounds With Potential for Natural Attenuation

Jim Spain . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137

Natural Attenuation of Chlorinated Compounds in Matrices Other Than Ground Water: 

The Future of Natural Attenuation

Robert E. Hinchee . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142

Poster Session

Anaerobic Mineralization of Vinyl Chloride in Iron(III)-Reducing Aquifer Sediments

Paul M. Bradley and Francis H. Chapelle. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

Intrinsic Biodegradation of Chlorinated Aliphatics Under Sequential Anaerobic/Co-metabolic Conditions

Evan E. Cox, David W. Major, Leo L. Lehmicke, Elizabeth A. Edwards, Richard A. Mechaber,

and Benjamin Y. Su . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 147

Analysis of Methane and Ethylene Dissolved in Ground Water

Steve Vandegrift, Bryan Newell, Jeff Hickerson, and Donald H. Kampbell . . . . . . . . . . . . . . . . . . . . . 148

Estimation of Laboratory and In Situ Degradation Rates for Trichloroethene and cis-1,2-Dichloroethene

in a Contaminated Aquifer at Picatinny Arsenal, New Jersey

Theodore A. Ehlke and Thomas E. Imbrigiotta . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149

Measurement of Dissolved Hydrogen in Ground Water

Mark Blankenship, Francis H. Chapelle, and Donald H. Kampbell . . . . . . . . . . . . . . . . . . . . . . . . . . . 151

Evidence of Natural Attenuation of Chlorinated Organics at Ft. McCoy, Wisconsin

Jason Martin. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 152

Challenges in Using Conventional Site Characterization Data To Observe Co-metabolism of

Chlorinated Organic Compounds in the Presence of an Intermingling Primary Substrate

Ian D. MacFarlane, Timothy J. Peck, and Joy E. Ligé . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 153

Development of an Intrinsic Bioremediation Program for Chlorinated Solvents at an Electronics Facility

Michael J. K. Nelson, Anne G. Udaloy, and Frank Deaver. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154

Overview of the U.S. Air Force Protocol for Remediation of Chlorinated Solvents by Natural Attenuation

Todd H. Wiedemeier, John T. Wilson, Donald H. Kampbell, Jerry E. Hansen, 

and Patrick Haas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 155

Incorporation of Biodegradability Concerns Into a Site Evaluation Protocol for Intrinsic Remediation

Robert M. Cowan, Keun-Chan Oh, Byungtae Kim, and Gauri Ranganathan . . . . . . . . . . . . . . . . . . . 156

Contents (continued)

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A Field Evaluation of Natural Attenuation of Chlorinated Ethenes in a Fractured Bedrock Environment

Peter Kunkel, Chris Vaughan, and Chris Wallen. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 157

Intrinsic Bioattenuation of Chlorinated Solvents in a Fractured Bedrock System

William R. Mahaffey and K. Lyle Dokken . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 158

Modeling Natural Attenuation of Selected Explosive Chemicals at a Department of Defense Site

Mansour Zakikhani and Chris J. McGrath. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159

Long-Term Application of Natural Attenuation at Sierra Army Depot

Jerry T. Wickham and Harry R. Kleiser . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 160

When Is Intrinsic Bioremediation Cost-Effective? Financial-Risk Cost-Benefit Analysis at Two

Chlorinated Solvent Sites

Bruce R. James, Evan E. Cox, David W. Major, Katherine Fisher, and Leo G. Lehmicke . . . . . . . . . 161

Natural Attenuation of Chlorinated Organics in Ground Water: The Dutch Situation

Lex W.A. Oosterbaan and Hans Rovers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 162

Natural Attenuation as a Cleanup Alternative for Tetrachloroethylene-Affected Ground Water

Steve Nelson . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 163

Natural Attenuation of Trichloroethene in a Sandy Unconfined Aquifer

Neale Misquitta, Dale Foster, Jeff Hale, Primo Marchesi, and Jeff Blankenship . . . . . . . . . . . . . . . . . 164

Analysis of Intrinsic Bioremediation of Trichloroethene-Contaminated Ground Water at 

Eielson Air Force Base, Alaska

Kyle A. Gorder, R. Ryan Dupont, Darwin L. Sorensen, Maria W. Kemblowski, and 

Jane E. McLean . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 165

Involvement of Dichloromethane in the Intrinsic Biodegradation of Chlorinated Ethenes and Ethanes

Leo L. Lehmicke, Evan E. Cox, and David W. Major . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 166

Intrinsic Bioremediation of 1,2-Dichloroethane

Michael D. Lee, Lily S. Sehayek, and Terry D. Vandell . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167

A Practical Evaluation of Intrinsic Biodegradation of Chlorinated Volatile Organic Compounds

Frederick W. Blickle, Patrick N. McGuire, Gerald Leone, and Douglas D. Macauley . . . . . . . . . . . . . 168

New Jersey’s Natural Remediation Compliance Program: Practical Experience at a Site Containing

Chlorinated Solvents and Aromatic Hydrocarbons

James Peterson and Martha Mackie . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 169

Field and Laboratory Evaluations of Natural Attenuation of Chlorinated Organics at a 

Complex Industrial Site

M. Alexandra De, Julia Klens, Gary Gaillot, and Duane Graves . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 170

Contents (continued)

Page

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Assessment of Intrinsic Bioremediation of Chlorinated Aliphatic Hydrocarbons at Industrial Facilities

Marleen A. Troy and C. Michael Swindoll . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171

Natural Attenuation as Remedial Action: A Case Study

Andrea Putscher and Betty Martinovich . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172

Patterns of Natural Attenuation of Chlorinated Aliphatic Hydrocarbons at Cape Canaveral 

Air Station, Florida

Matt Swanson, Todd H. Wiedemeier, David E. Moutoux, Donald H. Kampbell, and 

Jerry E. Hansen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 173

Applying Natural Attenuation of Chlorinated Organics in Conjunction With Ground-Water Extraction 

for Aquifer Restoration

W. Lance Turley and Andrew Rawnsley . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 174

A Modular Computer Model for Simulating Natural Attenuation of Chlorinated Organics in Saturated

Ground-Water Aquifers

Yunwei Sun, James N. Petersen, T. Prabhakar Clement, and Brian S. Hooker . . . . . . . . . . . . . . . . . 175

State and Federal Regulatory Issues

Federal and State Meeting on Issues Impacting the Use of Natural Attenuation for Chlorinated Solvents

in Ground-Water—An Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 179

Summary of Roundtable Discussion on Regulatory Issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183

Appendix A

Symposium Agenda. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187

Contents (continued)

Page

vii
Acknowledgments

The papers abstracted in this book were presented at the Symposium on Natural Attenuation of

Chlorinated Organics in Ground Water, held September 11-13, 1996, in Dallas, Texas. The symposium was a joint effort of the U.S. Environmental Protection Agency’s (EPA’s) Biosystems Technology Development Program, the U.S. Air Force Armstrong Laboratory’s Environics Directorate (USAF

AL/EQ) at Tyndall Air Force Base, Florida, and the U.S. Air Force Center for Environmental

Excellence (AFCEE) at Brooks Air Force Base, Texas. Fran Kremer and John Wilson of EPA’s Office

of Research and Development, Cathy Vogel of USAF AL/EQ, and Marty Faile and Patrick Haas of

USAF AFCEE served as co-organizers of the symposium.

viii
Executive Summary

The U.S. Environmental Protection Agency (EPA), the

U.S. Air Force Armstrong Laboratory’s Environics Directorate, and the U.S. Air Force Center for Environmental

Excellence hosted a conference entitled “Symposium

on Natural Attenuation of Chlorinated Organics in

Ground Water” in Dallas, Texas, September 11-13,

1996. Approximately 650 people attended, including

researchers; field personnel from federal, state, and

local agencies; and representatives from industry and

academia. Four speakers opened the symposium

with welcoming remarks and introductory talks on the

role of natural attenuation in remediating contaminated

sites.

Fran Kremer from EPA’s Office of Research and Development (ORD) opened the symposium by welcoming

the symposium participants to Dallas. She noted that the

use of natural attenuation as a remediation strategy has

increased exponentially over the last few years. She

attributed much of the progress made in adopting natural attenuation at appropriate sites to the collaboration

of researchers and agencies like ORD and the U.S. Air

Force.

Patricia Rivers from the Office of the Deputy Undersecretary of Defense (Environmental Cleanup) spoke about

past and ongoing Department of Defense (DOD) efforts

to clean up contaminated sites and the challenges that

DOD faces at sites that involve chlorinated solvents.

She introduced several DOD efforts to foster innovative

approaches to remediate these sites and emphasized

that natural attenuation is an important tool at the disposal of remediators.

Herb Ward of Rice University followed with a talk entitled

“Where Are We Now? Moving to a Risk-Based Approach.” He discussed the history of regulation and

remediation at Superfund sites, stating that “pump and

treat” cannot always achieve stringent cleanup levels

and is not always the most cost-efficient approach to

remediating sites. Mr. Ward called for contaminated

sites to be managed based on risk assessment, rather

than remediated to meet absolute standards. Natural

attenuation is a viable risk management tool, the usefulness of which depends on site-specific characteristics.

Kenneth Lovelace of the Superfund Division of EPA’s

Office of Emergency and Remedial Response concluded the introductory talks with a presentation entitled

“Where Are We Now With Public and Regulatory Acceptance?” He emphasized EPA’s support for natural attenuation as a proactive remedial strategy well suited to

particular sites. Mr. Lovelace identified thorough site

characterization, careful monitoring of remedy progress,

and contingency measures to ensure long-term reliability and protection as important factors in convincing

regulators and the public to choose natural attenuation

at a given site.

The 29 papers and 32 posters presented at the symposium highlighted recent achievements in the use of natural attenuation at particular sites, new research on the

efficacy of natural attenuation, developments in site

characterization and modeling techniques, and the

place of natural attenuation in the current regulatory

framework. These presentations represent the state of

the art in the use of natural attenuation as a tool in

remediating hazardous waste sites.

ix
Introduction

On September 11-13, 1996, the U.S. Environmental

Protection Agency (EPA), the U.S. Air Force Armstrong

Laboratory’s Environics Directorate, and the U.S. Air

Force Center for Environmental Excellence hosted a

conference entitled, “Symposium on Natural Attenuation

of Chlorinated Organics in Groundwater.” Natural attenuation is the biodegradation and/or chemical destruction or stabilization of contaminants and can be an

important tool for stabilizing or remediating a contaminated site.

The symposium’s principal goals were to:

• Increase understanding of the natural attenuation

process.

• Review methods for screening sites.

• Help decision-makers determine the feasibility of using natural attenuation at sites contaminated with

chlorinated solvents.

• Solicit feedback from the regulatory and industrial

communities on the appropriate application of and

protocol for natural attenuation.

The intended audience for this symposium included

regulators, the regulated community, and research scientists. Representatives from these groups made presentations at the symposium that covered the following

topics:

• Laboratory studies and field demonstrations conducted in support of natural attenuation at government and industry sites.

• Methods for assessing the potential for natural attenuation at contaminated sites.

• Methods for measuring the effectiveness of natural

attenuation.

There are a large number of sites contaminated with

chlorinated organic solvents remaining in the United

States. The U.S. Department of Defense (DOD), for

example, estimates that more than 1,000 sites on 200

military installations contain chlorinated organic solvents. Remediation will begin at 400 of these sites in the

next two years. EPA, DOD, and other involved parties

have initiated a number of efforts to foster innovative

approaches to sites contaminated with chlorinated solvents. These efforts have facilitated the exchange of

information about the efficacy of natural attenuation as

an important tool for protecting human health and the

environment.

This document contains abstracts of paper and poster

presentations from the symposium, as well as summaries of a meeting and roundtable discussion on state and

federal regulatory issues affecting natural attenuation of

chlorinated solvents. The symposium agenda is included in the appendix.

x
Where Are We Now? Moving to a Risk-Based Approach

C.H. Ward

Rice University, Houston, Texas

Setting Cleanup Goals for Ground Water

When the Comprehensive Environmental Response,

Compensation, and Liability and Resource Conservation and Recovery Acts were implemented in the mid1980s, the cleanup goals for contaminants in ground

water often defaulted to concentration-based standards

for drinking water (maximum contaminants levels or

MCLs). These standards were designed for public water

supplies. Because water supply was seen as the important contribution of ground water, the application of

these standards seemed to be relevant and appropriate.

There was little awareness of the contribution of ground

water to the function of the landscape. The impact of

contaminants that discharged from ground water to sensitive receptor ecosystems received less attention.

Stringent drinking water standards were selected with

the expectation that they could be met with existing

pump-and-treat technology. Pump-and-treat was naively thought to be a quick, viable fix to ground-water

contamination. To budget for the first authorization of

Superfund, Congress estimated a unit cost for remediation that included application of pump-and-treat, then

multiplied this estimate by the number of sites (1).

The Failure To Meet Cleanup Goals for

Ground Water

In the mid-1990s, a National Research Council committee reviewed the performance of conventional pumpand-treat methods at 77 sites. At 69 of the sites, the

cleanup goal had not been reached. Based on a body

of science and empirical experience developed from the

mid-1980s to the mid-1990s, the committee identified

five reasons that pump-and-treat had failed to perform

as expected (2):

• The physical heterogeneity of the subsurface makes

contaminant migration pathways extremely difficult to

detect.

• Contaminants are often present as nonaqueousphase liquids (NAPLs) that are not efficiently removed by pumping ground water.

• Contaminants migrate to inaccessible regions so that

their recovery is controlled by the rate of diffusion

back out of the inaccessible regions, not by the rate

of ground-water extraction.

• Sorption of contaminants to subsurface materials results in an underestimate of the total contaminant

mass in the aquifer.

• Difficulties in characterizing the subsurface make it

difficult to extrapolate between sampling points and

produce uncertainty in engineering remedial designs.

The Ground-Water Remediation 

Treadmill

The default remedy selected to clean up ground water

contamination was not working at most sites. Concentrations of contaminants in pumped wells often reached

an asymptote that was above the cleanup goal. In the

instances in which major reductions in contaminant concentrations were achieved, the concentrations of contaminants would often rebound after the pumps were

turned off. As a result, major funds were being expended

to operate and maintain systems that were not meeting

cleanup objectives.

The NRC committee (2) evaluated alternative technologies and found that a substantial amount of performance

data existed for three alternative technologies: soil vapor extraction, R.L. Raymond’s process using hydrogen

peroxide for in situ bioremediation of hydrocarbons, and

bioventing. The Raymond process does not work for

most chlorinated solvents; in particular, it does not work

for tetrachloroethylene and trichloroethylene. Bioventing

and soil vapor extraction work only in the vadose zone,

not in aquifers.

The committee also evaluated developing technologies that still required more controlled field studies

1
and implementation at large-scale sites to generate reliable performance data. They considered pulsed or

variable pumping, in situ bioremediation designed for

chlorinated solvents, air sparging, steam-enhanced extraction, in situ thermal desorption, soil flushing, and in

situ chemical treatment.

It is difficult for technologies presently available or under

development to consistently clean aquifers contaminated with chlorinated solvents to drinking water MCLs.

Presently, we can be more effective preventing the

spread of contamination and reducing exposure.

Containment Instead of Cleanup

In the period from the early 1980s to mid-1990s, while

pump-and-treat was being implemented as a remedial

technology, microbiologists, hydrologists, engineers,

and chemists were working to develop a quantitative

understanding of the fate of chemical contaminants in

the subsurface. The pump-and-treat systems were being monitored, and many of the ground-water contaminants were recognized to be transformation products of

the chlorinated solvents that were originally spilled. For

example, cis-dichloroethylene and vinyl chloride were

often produced from reductive dechlorination of

tetrachloroethylene and trichloroethylene.

By the mid-1990s, 10 years of monitoring data existed

on many chlorinated solvent plumes. At many sites,

there was clear evidence that the plumes were not

expanding; some natural activity was preventing the

spread of contamination. At other plumes, containment

was not achieved, and contamination spread with the

flow of ground water. The effectiveness of pump-andtreat containment should thus be compared to the containment provided by the processes that naturally

attenuate contaminants in ground water. These processes include biodegradation, abiotic transformation,

sorption, and dilution.

Contribution of Natural Attenuation to

Containment

If natural attenuation can contain the spread of contamination, it is the philosophical equivalent of pump-andtreat, a cap on the source, a slurry wall, or an in situ

reactive barrier.

Some regulators have dismissed natural attenuation as

a “do nothing” approach. If site managers do nothing but

compile monitoring data on the contaminants of concern, the characterization is accurate. All they know is

the distribution of contaminants at their site. If site managers carry out careful and well-planned studies of the

hydrology, geochemistry, and microbiology at their site

and use this information to understand in detail the

behavior of contaminants, they in turn can use this

understanding to make rigorous and defensible predictions about the prospects for the spread of contaminants.

A good characterization study to predict containment by

natural attenuation is the equivalent of reliable performance data on a proactive technology for containment.

Because site characterizations often require sophisticated sampling techniques, new analytical approaches,

and state-of-the-art ground-water modeling, natural attenuation becomes very much a “high-tech” approach (3).

The emerging approach to risk management uses

ground-water science to predict the behavior of plumes,

then takes advantage of natural attenuation in a comprehensive risk management strategy. These comprehensive strategies usually have some element of source

removal or source control at the hot spots, with natural

attenuation reserved for the diffuse contamination some

distance from the source.

Impacts of Ground Water on Surface-Water

Ecosystems

Many plumes of chlorinated solvents discharge to surface water. Discharge from chlorinated solvent plumes

has been evaluated at the U.S. Army’s Picatinny Arsenal, at the St. Joseph, Michigan, national priority list site,

and at the fire training site at Plattsburgh Air Force Base

in New York. Case studies on these plumes appear

elsewhere in this volume.

When a plume discharges to surface water, the risk

management emphasis shifts. The concentration of contaminants is much less important than the mass flux of

contaminants to the receptor ecosystem. To manage

risk associated with ground-water discharge, the loading

of contaminants to the receptor ecosystem must be

compared with the loading that can be accepted without

damage to the receptor ecosystem. Chlorinated solvents do not bioaccumulate, and they rapidly volatilize

to the atmosphere. As a consequence, there is little

anecdotal evidence that discharge of chlorinated solvents from ground water has damaged surface-water

ecosystems; nonetheless, these issues deserve systematic evaluation.

The discipline of toxicological assessment of ecosystems has made extensive progress in the last decade.

No established and widely accepted protocol for making these assessments exists, however. As a result,

much of the science is not readily available to regulators. This makes it difficult for the regulators to participate as intellectual partners in the risk assessment

and risk management process. A protocol should be

developed to evaluate the transfer of contaminants

from ground-water to surface-water ecosystems. By

documenting appropriate sampling methods, analytical

procedures, procedures for interpreting the data, and

mathematical models to collate and integrate data, such

2
a protocol would greatly facilitate the task of determining

the loadings that surface-water ecosystems can receive

without being damaged.

References

1. National Research Council. 1994. Ranking hazardous waste sites

for remedial action. Washington, DC.

2. National Research Council. 1994. Alternatives for ground water

cleanup. Washington, DC.

3. National Research Council. 1993. In situ bioremediation: When

does it work? Washington, DC.

3
Where Are We Now With Public and Regulatory Acceptance? (Resource

Conservation and Recovery Act [RCRA] and Comprehensive Environmental

Response, Compensation, and Liability Act [CERCLA])

Kenneth Lovelace and Peter Feldman

U.S. Environmental Protection Agency,

Office of Emergency and Remedial Response (Superfund),

Washington, DC

Introduction

The U.S. Environmental Protection Agency (EPA) remains committed to the goal of restoring contaminated

ground waters to their beneficial uses. The Agency also

continues to support the use of natural attenuation as a

restoration method. EPA recognizes that, in certain circumstances, remedies using natural attenuation can be

more cost effective than “active” remediation approaches in achieving cleanup objectives equally protective of human health and the environment. The

Agency also recognizes that many technical questions

remain to be answered regarding the efficacy of this

approach, which underscores the importance of continued scientific research as well as the need to employ

remedies that use natural attenuation in a consistent

and responsible manner.

What Is Natural Attenuation?

Natural attenuation is discussed in the preamble of the

National Oil and Hazardous Substances Pollution Contingency Plan (NCP), which is the regulatory framework

for the Superfund program (1). In the NCP, natural attenuation is described as a process that “will effectively

reduce contaminants in the ground water” to concentrations “protective of human health and sensitive ecological environments in a reasonable timeframe.” The NCP

goes on to recognize that natural attenuation may include any or all of the following processes:

• Biodegradation

• Dilution

• Dispersion

• Adsorption

Thus, the NCP definition includes biodegradation, which

alters or destroys the contamination, as well as physical

processes that lower the concentrations and availability

of contaminants without necessarily altering the chemistry. Other processes not mentioned in the NCP are not

necessarily excluded from the definition (e.g., volatilization). Other EPA remediation programs also recognize

this definition, including the Corrective Action program

under the Resource Conservation and Recovery Act

(RCRA) and Underground Storage Tank (UST) programs.

Some terms, such as “intrinsic remediation” or “passive

remediation,” are essentially equivalent to the NCP’s

definition of natural attenuation. Other terms used in

recent literature, including “intrinsic bioremediation” or

“in situ bioremediation,” appear to be more restrictive in

scope than “natural attenuation.” In addition, natural

attenuation is the term used in existing EPA guidance

(e.g., U.S. EPA [2]).

Regulatory Framework

Natural attenuation is recognized as a legitimate remedial approach for ground-water cleanup under the Superfund, RCRA Corrective Action, and UST remediation

programs. A directive clarifying EPA’s policy regarding

the use of natural attenuation for remediation of sites

regulated under these programs is currently in development (3). Remedies selected for contaminated ground

water (and for other media) under these programs must

protect human health and the environment, regardless

of the particular remediation technology or approach

selected. Remedies may achieve protection through a

mix of treatments that reduce or destroy contaminants;

containment and other engineering controls, which limit

exposure; and other means identified as part of the

4
remedy selection process. Each EPA program offers

guidance suggesting when specific methods of protection may be more appropriate than others.

EPA recognizes that natural attenuation may be an appropriate remediation method for contaminated ground

water under the right circumstances. Natural attenuation

should continue to be carefully evaluated along with

other viable remedial approaches or technologies within

the existing remedy selection framework. Natural attenuation is not to be considered a default or presumptive remedy for a given site under any of these EPA

programs.

Cleanup policies for Superfund have addressed the

use of natural attenuation in some detail; most of

the relevant discussion can be found in the NCP preamble. The following NCP language specifies the definition

and cleanup expectation for remedies using natural

attenuation:

[S]election of natural attenuation by EPA does not

mean that the ground water has been written off and

not cleaned up but rather that biodegradation, dispersion, dilution, and adsorption will effectively reduce contaminants in the ground water to

concentrations protective of human health in a timeframe comparable to that which could be achieved

through active restoration. . . .(1).

Thus, the NCP expects that a remedy employing natural

attenuation will be fully protective and attain the required

cleanup levels for the aquifer in a timeframe that is not

unreasonably long. Since the other EPA remediation

programs present similar expectations, use of natural

attenuation as a remedy does not reduce EPA’s responsibility to protect human health and the environment and

to satisfy the cleanup levels and other remediation objectives selected for a given site. In short, use of natural

attenuation does not imply that EPA has agreed to a “no

action” remedy or that EPA or responsible parties may

“walk away” from their remedial obligations at a site.

When Is Natural Attenuation Appropriate?

Because of the longer timeframes needed for remediation strategies using natural attenuation, such an approach is best suited for sites where there is no demand

for the ground water in the near future. For example,

where adequate alternate water sources are available,

future demand for the contaminated ground water is

likely to be low. Also, the timeframe required for natural

attenuation should be reasonable compared with more

active alternatives. Other site conditions that favor the

use of natural attenuation as a remediation approach

are discussed below.

Large, Dilute Contaminant Plumes

The types of contaminants, their concentrations, and

hydrogeologic conditions should indicate that natural

attenuation is a viable remediation approach for a given

site. Natural attenuation is more likely to be an appropriate remediation approach at sites with large plumes and

relatively low contaminant concentrations. For these

types of sites, required cleanup le vels might be

attainable in a reasonable timeframe using natural attenuation and at a much lower cost than other alternatives.

Sources Controlled or Controllable

Natural attenuation will not be effectively used to reach

desired cleanup levels if the rate of contamination entering ground water exceeds the rate of the natural

attenuation processes. Therefore, the natural attenuation approach shall not be used unless contaminant

sources have been controlled or site characterization

data indicate that the sources are no longer present.

Measures for controlling contaminant sources, include

removal, treatment, or containment of source materials.

Sources of contaminants to ground water could include

surface facilities, landfill wastes, contaminated soils, or

nonaqueous-phase liquids (NAPLs) in the subsurface.

Protected Drinking Water or

Environmental Resources

Cross contamination of other aquifers or discharge of

contamination to surface waters or sensitive ecological

environments is more likely if contamination is left in the

subsurface for long periods. Site conditions should indicate a low potential for migration of contaminants into

uncontaminated media, or measures for controlling

plume migration should be included in remedies using

natural attenuation. In addition, the issue of whether

“daughter” products of natural attenuation will pose a

significant risk must be addressed.

Combining Natural Attenuation With

Other Methods

For sites where natural attenuation alone is not capable

of achieving desired cleanup levels in a reasonable

timeframe, natural attenuation combined with more active remediation methods may prove to be effective.

Some areas of the plume may require a much longer

time to attenuate naturally than others, such as areas

with relatively high contaminant levels (“hot spots”). In

this situation, natural attenuation of dilute plume areas

combined with extraction and treatment to control

source areas and remediate plume hot spots may be an

effective remediation approach, especially for sites

where dilute portions of the plume cover a relatively

large area.

5
In some cases, it may be appropriate for natural attenuation to be used as a followup to active remediation. In

this approach, active measures are used to reduce contaminant concentrations, followed by natural attenuation

as the final stage of remediation.

Promoting Regulatory and Public

Acceptance

In general, promoting acceptance of natural attenuation

will require detailed site characterization and analysis to

demonstrate that this approach will achieve remediation

goals, careful monitoring of remediation progress, and

identification of contingency measures. These provisions are necessary to convince regulatory agencies

and the public that natural attenuation is a valid remediation approach rather than a “walkaway” and that it will

be sufficiently protective.

Building confidence in the approach can also be promoted by involving the responsible regulatory agencies

as early in the process as possible. For example, upfront agreement on the type of characterization data

needed to demonstrate the efficacy of natural attenuation can save considerable effort later in the remedy

selection process.

Detailed Site Characterization

Convincing regulatory officials and local citizens that

natural attenuation will be effective starts with a detailed

site characterization and a clear conceptual model of

site conditions. A conceptual model of how natural attenuation will perform at a given site is essential to show

that natural attenuation will be effective and that potential adverse impacts to human health and the environment can be prevented over the long period required for

cleanup. The burden of proving the viability of natural

attenuation is on the proponent, not the regulator.

Site-specific data should be used to demonstrate that

the required cleanup levels can be attained in a reasonable timeframe compared with other remedial alternatives. Such a demonstration can be supported by the

following types of site data:

• Contaminant concentrations have decreased over

time.

• Geochemical or microbiological parameters are characterized to the extent needed to support predictive

models.

• Predictive models show required cleanup levels will

be attained in a timeframe that is reasonable for the

site.

Exposure Prevention Measures

Prevention of exposure to contaminated ground water

over the long period required for cleanup is critical to

ensure protectiveness. Remedies using natural attenuation should include effective measures for ensuring that

contaminated ground water does not reach public or

private wells, or for providing effective treatment prior to

use.

Performance Monitoring

A thorough monitoring network and plan are necessary

to evaluate the progress of natural attenuation. Remedies using natural attenuation should include a monitoring plan to ensure that the remedy’s performance

matches predictions, there are no adverse impacts, and

unanticipated events can be detected in time to develop

an appropriate response.

Contingency Measures

Contingencies for initiating active remediation measures

should be incorporated into strategies using natural attenuation. Such contingencies provide assurance that

the remedy’s protectiveness will be maintained should

natural attenuation not progress as expected. The trigger(s) for implementing such contingencies should be

clearly spelled out in site decision documents.

Summary

EPA believes that natural attenuation should continue to

play an important role in the cleanup of sites with contaminated ground water. Furthering the technical understanding of the underlying treatment processes and

promoting the responsible use of this remediation

method should serve to enhance the role that natural

attenuation plays in restoring the nation’s ground water.

Greater regulatory and public acceptance of natural attenuation will require demonstrating that such remedies

will be effective in meeting remediation goals and in

protecting human health and the environment over the

long period required for cleanup. Demonstrating the

effectiveness of remedies using natural attenuation will

involve thorough site characterization, careful monitoring of remedy progress, and contingency measures to

ensure long-term reliability and protectiveness.

References

1. U.S. EPA. 1990. National Oil and Hazardous Substances Pollution

Contingency Plan: Final rule (NCP). Fed. Reg. 55(46):8733-8734.

March 8.

2. U.S. EPA. 1988. Guidance on remedial actions for contaminated

ground water at Superfund sites. Office of Solid Waste and Emergency Response Directive 9283.1-2 EPA/540/G-88/003 (December).

3. U.S. EPA. 1996. Use of natural attenuation at Superfund, RCRA

Corrective Action, and UST remediation sites. Draft directive

available in Fall 1996 or Winter 1997. Office of Solid Waste and

Emergency Response.

6
Biotic and Abiotic Transformations of Chlorinated Solvents in Ground Water

Perry L. McCarty

Stanford University, Department of Civil Engineering, Stanford, California

Introduction

Chlorinated solvents and their natural transformation

products represent the most prevalent organic groundwater contaminants in the country. These solvents, consisting primarily of chlorinated aliphatic hydrocarbons

(CAHs), have been used widely for degreasing of aircraft engines, automobile parts, electronic components,

and clothing. Only during the past 15 years has it become recognized that CAHs can be transformed biologically (1). Such transformations sometimes occur under

the environmental conditions present in an aquifer in the

absence of planned human intervention, a process

called natural attenuation or intrinsic biotransformation (2).

The major chlorinated solvents are carbon tetrachloride

(CT), tetrachloroethene (PCE), trichloroethene (TCE),

and 1,1,1-trichloroethane (TCA). These compounds can

be transformed by chemical and biological processes in

soils to form a variety of other CAHs, including chloroform (CF), methylene chloride (MC), cis- and trans-1,2-

dichloroethene (cis-DCE, t-DCE), 1,1-dichloroethene

(1,1-DCE), vinyl chloride (VC), 1,1-dichloroethane

(DCA), and chloroethane (CA). Abiotic or chemical

transformations of some CAHs can occur within the time

frame of interest in ground water. CAHs can also be

transformed through the action of aerobic or anaerobic

microorganisms. In some cases, such transformations

may be co-metabolic, that is, fortuitous transformation

brought about by enzymes that microorganisms are using for other purposes. In such cases, the transforming

microorganisms must be actively growing, which requires the presence of primary substrates. In other

cases, the microorganisms may be using the CAHs in

energy metabolism, a condition now being commonly

found under anaerobic conditions. These are unique

reactions, because the microorganisms use CAHs as

electron acceptors just as aerobic organisms use oxygen. This in turn requires a suitable electron donor such

as hydrogen or organic compounds. Transformations

that are likely to occur in ground water and the environmental conditions required are discussed below.

Chemical Transformation

TCA is the only major chlorinated solvent that can be

transformed chemically in ground water under all likely

conditions within the one- to two-decade time span of

general interest, although chemical transformation of CT

through reductive processes is a possibility. TCA chemical transformation occurs by two different pathways, leading to the formation of 1,1-DCE and acetic acid (HAc):

The rate of each chemical transformation is given by the

first-order reaction:

C = Coe-kt (Eq. 3)

where C is the concentration of TCA at any time t, Co

represents the initial concentration at t = 0, and k is a

transformation rate constant. The overall rate constant

for TCA transformation (kTCA) is equal to the sum of the

individual rate constants (kDCE + kHAc). The transformation rate constants are functions of temperature:

k = Ae-E/0.008314K (Eq. 4)

where A and E are constants and K is the temperature

in degrees Kelvin. Table 1 lists A and E values for TCA

abiotic transformation reported by various investigators,

as well as calculated values for the TCA transformation

rate constant for 10°C, 15°C, and 20°C using Equation

4. Also given is the average calculated TCA half-life

based upon t1/2 = 0.69/k. The temperature effect on

TCA half-life is quite significant.

7
Cline and Delfino (4) found that kDCE equaled about 21

percent of KTCA, and Haag and Mill (3) found it to be 22

percent. This means that almost 80 percent of the TCA

is transformed into acetic acid. The 20-plus percent that

is converted to 1,1-DCE is of great significance, however, because 1,1-DCE is considered more toxic than

TCA, with an MCL of 7 micrograms per liter (μg/L)

compared with TCA’s MCL of 200 μg/L. Whenever TCA

is present as a contaminant, 1,1-DCE can also be expected. In general, TCA is probably the main source of

1,1-DCE contamination found in aquifers.

CA, formed through biological transformation of TCA,

can also be chemically transformed with a half-life on

the order of months by hydrolysis to ethanol, which can

then be biologically converted to acetic acid and harmless products (6).

Biological Transformation

CAHs can be oxidized or reduced, generally through

co-metabolism, as noted in Table 2. In ground water,

reductive transformations are most often noted, perhaps

because the presence of intermediate products that are

formed provide strong evidence that reductive transformations are taking place. Co-metabolic aerobic transformation of TCE is also possible, although if it did occur

the intermediate products formed are unstable and more

difficult, analytically, to measure. Thus, convincing evidence for the latter is more difficult to obtain. Also,

aerobic co-metabolism of TCE would only occur if sufficient dissolved oxygen and a suitable electron donor,

such as methane, ammonia, or phenol, were present.

Since circumstances under which the proper environmental conditions for significant aerobic co-metabolism

are not likely to occur often, natural attenuation by aerobic

co-metabolism of TCE is probably of little significance.

Ample evidence suggests that anaerobic reductive

transformation of CAHs occurs frequently, however, and

this process is of importance to the transformation of all

chlorinated solvents and their transformation products.

The major environmental requirement is the presence of

sufficient concentrations of other organics that can serve

as electron donors for energy metabolism, which is often

the case in aquifers. Indeed, the extent to which reductive dehalogenation occurs may be limited by the

amount of these co-contaminants present. Theoretically,

it would require only a 0.4-gram chemical oxygen demand (COD) equivalent of primary substrate to convert

1 gram of PCE to ethene (7), but many times more than

this is actually required because of competition by other

microorganisms for the electron donors present.

Figure 1 illustrates the potential chemical and biological

transformation pathways for the four major chlorinated

solvents under anaerobic environmental conditions (6).

Freedman and Gossett (8) provided the first evidence

for conversion of PCE and TCE to ethene, and de Bruin

et al. (9) reported complete reduction to ethane. Table 3

Table 1. Reported First-Order TCA Abiotic Transformation Rates (kTCA)

kTCA(yr-1)

A yr-1 E kJ 10°C 15°C 20°C References

3.47 (10)20 118.0 0.058 0.137 0.32 3

6.31 (10)20 119.3 0.060 0.145 0.34 4

1.56 (10)20 116.1 0.058 0.137 0.31 5

Average half-life (yr) 12 4.9 0.95

Table 2. Conditions for Biotic and Abiotic Transformations of Chlorinated Solvents

Carbon Tetrachloride

(CTC)

Trichloroethene

(PCE)

Tetrachloroethene

(TCE)

1,1,1-Trichloroethane

(TCA)

Biotic—Aerobic

Primary substrate No No No No

Co-metabolism No No Yes Perhaps

Biotic—Anaerobic

Primary substratea Perhaps Yes Yes Perhaps

Co-metabolism Yes Yes Yes Yes

Hazardous intermediates Yes Yes Yes Yes

Abiotic Perhaps No No Yes 

a

 Can be used as electron acceptor in energy metabolism.

8
indicates that while some transformations, such as that

of CT to CF and carbon dioxide, may take place under

mildly reducing conditions such as those associated

with denitrification, complete reductive transformation to

inorganic end products and of PCE and TCE to ethene

generally requires conditions suitable for methane fermentation. Extensive reduction can also occur under

sulfate-reducing conditions. For methane fermentation

to occur in an aquifer, the presence of sufficient organic

co-contaminant is required to reduce all of the oxygen,

nitrate, nitrite, and sulfate present. Some organics will

be required to reduce the CAHs, and perhaps iron(II) as

well, if present in significant amounts. If the potential for

natural biological attenuation of CAHs is to be evaluated, then the concentrations of nitrate, nitrite, sulfate,

iron(II), and methane, as well as organics as indicated

by COD or total organic carbon (TOC), should be determined. Unfortunately, such analyses are not considered

essential in remedial investigations—they should be.

Several pure cultures of microorganisms are now available that can also reduce PCE to cis-DCE (10-14). Only

one has been reported that can convert PCE completely

to ethene (15). Most of the isolates are strict anaerobes

and use hydrogen as an electron donor, with CAHs

being used as electron acceptors in energy metabolism.

One isolate, however, is a facultative aerobe (14) that

can use many organics, such as acetate, as the electron

donor and oxygen, nitrate, PCE, or TCE as electron

acceptors, which it does in that order of preference. It is

now believed that the CAH reducers compete for the

hydrogen they use, which is formed as an intermediate

in anaerobic organic oxidation, with sulfate reducers,

methanogens, and holoacetogens (16). This may explain

the excessive donor requirements for CAH reduction.

Concerns are frequently expressed over the VC formed

as an intermediate in reductive dehalogenation of PCE,

TCE, and DCE in ground water, because VC is a known

human carcinogen. It is possible to oxidize VC aerobically, however, with oxygen as an electron acceptor or

even under anoxic conditions with iron(III) (17). In addition, VC is readily and very efficiently co-metabolized

aerobically by methane, phenol, or toluene oxidizers

(18, 19). Here, transformation yields of over 1 gram of

VC per gram of methane have been obtained. Thus, at

the aerobic fringes of plumes with methane and VC

present, or where sufficient iron(III) is present, natural

attenuation of VC through oxidation can occur.

Case Studies

Major et al. (20) reported field evidence for intrinsic

bioremediation of PCE to ethene and ethane at a chemical transfer facility in North Toronto. In addition to high

concentrations of PCE (4.4 milligrams per liter [mg/L]),

high concentrations of methanol (810 mg/L) and acetate

(430 mg/L) were found as co-contaminants in the

ground water and served as electron donors for the

transforming organisms. Where high concentrations of

PCE were found, TCE (1.7 mg/L), cis-DCE (5.8 mg/L),

and VC (0.22 mg/L) were also found, but little ethene

(0.01 mg/L). At one downgradient well, however, no PCE

or TCE were found, but cis-DCE (76 mg/L), VC (9.7

mg/L) and ethene (0.42 mg/L) were present, suggesting

that significant dehalogenation had occurred. Microcosm studies also suggested that biotransformation was

occurring at the site, with complete disappearance of

PCE, TCE, and cis-DCE and production of both VC and

ethene. The conversions were accompanied by significant

methane production, indicating the presence of suitable

redox conditions for the transformation.

Fiorenza et al. (21) reported on PCE, TCE, TCA, and

dichloromethane (DCM) contamination of ground water

at a carpet backing manufacturing plant in Hawkesbury,

Ontario. The ground water contained 492 mg/L of volatile

Figure 1. Anaerobic chemical and biological transformation

pathways for chlorinated solvents.

Table 3. Environmental Conditions for Reductive Transformations of Chlorinated Solvents

Redox Environment 

Chlorinated Solvent All Denitrification Sulfate Reduction Methanogenesis

Carbon tetrachloride CT → CF CT → CO2+Cl1,1,1-Trichloroethane TCA → 1,1-DCE

+ CH3COOH

TCA → 1,1-DCA TCA → CO2+ClTetrachloroethene PCE → 1,2-DCE PCE → ethene 

Trichloroethene TCE → 1,2-DCE TCE → ethene

9
fatty acids and 4.2 mg/L of methanol, organics that

appeared to serve as electron donors for dehalogenation. Sulfate was nondetected, but the concentration in

native ground water was about 15 to 18 mg/L. Total

dissolved iron was quite high (19.5 mg/L) and above the

upgradient concentration of 2.1 mg/L. Methane was present. This supports conditions suitable for natural biodegradation of the chlorinated solvents. While some

chemical transformation of TCA to 1,1-DCE was indicated (0.4 mg/L) biotransformation was extensive, as

indicated by a 1,1-DCA concentration of 7.2 mg/L, compared with the TCA concentration of 5.5 mg/L. Some CA

was also present (0.19 mg/L). Transformation was also

indicated for PCE and TCE because the cis-DCE, VC,

and ethene concentrations were 56, 4.2, and 0.076

mg/L, respectively. Only traces of ethane were found.

Downgradient from the lagoon, the dominant products

were cis-DCE (4.5 mg/L), VC (5.2 mg/L), and 1,1-DCA

(2.1 mg/L). While good evidence for natural attenuation

exists for this site, the ethene and ethane concentrations

were low compared with the VC concentration, suggesting that biotransformation was not eliminating the chlorinated solvent hazard at the site, although it was

producing compounds that may be more susceptible to

aerobic co-metabolism.

Evidence for intrinsic biotransformation of chlorinated

solvents has also been provided from analyses of gas

from municipal refuse landfills where active methane

fermentation exists. A summary by McCarty and Reinhard (22) of data from Charnley et al. (23) reported

average gaseous concentrations in parts per million by

volume from eight refuse landfills as PCE, 7.15; TCE,

5.09; cis-DCE, not measured; trans-DCE, 0.02; and VC,

5.6. While these averages indicate that, in general,

transformation was not complete, the high VC concentration indicates the transformation was significant. For

TCA, gaseous concentrations were TCA, 0.17; 1,1-

DCE, 0.10; 1,1-DCA, 2.5; and CA, 0.37. These data

indicate that TCA biotransformation was quite extensive,

with the transformation intermediate, 1,1-DCA, present

at quite significant levels, as is frequently found in

ground water.

Perhaps the most extensively studied and reported intrinsic chlorinated solvent biodegradation is that at the

St. Joseph, Michigan, Superfund site (7, 24-27).

Ground-water concentrations of TCE as high as 100

mg/L were found, with extensive transformation to cisDCE, VC, and ethene. A high but undefined COD (400

mg/L) in ground water, resulting from waste leaching

from a disposal lagoon, provided the energy source for

the co-metabolic reduction of TCE. Nearly complete

conversion of the COD to methane provided evidence

of the ideal conditions for intrinsic bioremediation (7).

Extensive analysis near the source of contamination

indicated that 8 to 25 percent of the TCE had been

converted to ethene and that up to 15 percent of the

reduction in COD in this zone was associated with reductive dehalogenation (25). Through more extensive

analysis of ground water further downgradient from the

contaminating source, Wilson et al. (26) found a 24-fold

reduction in CAHs across the site. The great extent of

aerobic co-metabolic VC transformation in the methane

present suggests that aerobic oxidation at the plume

fringes is likely to be occurring (18). A review of the data

at individual sampling points indicated that conversion

of TCE to ethene was most complete where methane

production was highest and removal of nitrate and sulfate by reduction was most complete.

Since the above early reports, many others have reported on the natural biological attenuation of CAHs in

ground water, all showing conversion of PCE, TCE, or

TCA to nonchlorinated end points (28-31). Whether

complete dehalogenation is likely to occur over time at

these sites is still not clear. Review of this literature by the

reader interested in these processes is recommended.

References

 1. McCarty, P.L., and L. Semprini. 1994. Ground-water treatment for

chlorinated solvents. In: Norris, R.D., ed. Handbook of bioremediation. Boca Raton, FL: Lewis Publishers. pp. 87-116.

 2. National Research Council. 1993. In situ bioremediation. When

does it work? Washington, DC: National Academy Press. p. 207.

 3. Haag, W.R., and T. Mill. 1988. Transformation kinetics of 1,1,1-

trichloroethane to the stable product 1,1-dichloroethene. Environ.

Sci. Technol. 22:658-663.

 4. Cline, P.V., and J.J. Delfino. 1989. Effect of subsurface sediment

on hydrolysis of haloalkanes and epoxides. In: Larson, R.A., ed.

Biohazards of drinking water treatment. Chelsea, MI: Lewis Publishers, Inc. pp. 47-56.

 5. Jeffers, P., L. Ward, L. Woytowitch, and L. Wolfe. 1989. Homogeneous hydrolysis rate constants for selected chlorinated methanes, ethanes, ethenes, and propanes. Environ. Sci. Technol.

23(8):965-969.

 6. Vogel, T.M., C.S. Criddle, and P.L. McCarty. 1987. Transformations of halogenated aliphatic compounds. Environ. Sci. Technol.

21:722-736.

 7. McCarty, P.L., and J.T. Wilson. 1992. Natural anerobic treatment

of a TCE plume, St. Joseph, Michigan, NPL site. In: U.S. EPA.

Bioremediation of hazardous wastes. EPA/600/R-92/126. Cincinnati, OH. pp. 47-50.

 8. Freedman, D.L., and J.M. Gossett. 1989. Biological reductive

dechlorination of tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions. Appl. Environ. Microbiol.

55:2144-2151.

 9. de Bruin, W.P., M.J.J. Kotterman, M.A. Posthumus, G. Schraa,

and A.J.B. Zehnder. 1992. Complete biological reductive transformation of tetrachloroethene to ethane. Appl. Environ. Microbiol. 58:1996-2000.

10. Holliger, C., G. Schraa, A.J.M. Stams, and A.J.B. Zehnder. 1993.

A highly purified enrichment culture couples the reductive

dechlorination of tetrachloroethene to growth. Appl. Environ. Microbiol. 59:2991-2997.

11. Neumann, A., H. Scholz-Muramatsu, and G. Dickert. 1994.

Tetrachloroethene metabolism of Dehalospirillum multivorans.

Arch. Microbiol. 162:295-301.

10
12. Scholz-Muramatsu, H., A. Neumann, M. MeBmer, E. Moore, and

G. Diekert. 1995. Isolation and characterization of Dehalospirillum multivorans gen. sp. nov., a tetrachloroethene-utilizing,

strictly anaerobic bacterium. Arch. Microbiol. 163:48-56.

13. Holliger, C., and W. Schumacher. 1994. Reductive dehalogenation as respiratory process. Antonie Van Leeuwenhoek 66:239-

246.

14. Sharma, P., and P.L. McCarty. 1996. Isolation and characterization of facultative aerobic bacterium that reductively dehalogenates tetrachlorethene to cis-1,2-dichloroethene. Appl.

Environ. Microbiol. 62:761-765.

15. Maymo-Gatell, X., Y.T. Chien, T. Anguish, J. Gossett, and S.

Zinder. 1996. Isolation and characterization of an anaerobic

eubacterium which reductively dechlorinates tetrachloroethene

(PCE) to ethene. In: Abstracts of the 96th General Meeting of the

American Society of Microbiology, New Orleans. pp. Q-126.

16. Fennel, D.E., M.A. Stover, S.H. Zinder, and J.M. Gossett. 1995.

Comparison of alternative electron donors to sustain PCE anaerobic reductive dechlorination. In: Hinchee, R.E., A. Leeson,

and L. Semprini, eds. Bioremediation of chlorinated solvents.

Columbus, OH: Battelle Press. pp. 9-16.

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

of vinyl chloride in Fe(III)-reducing aquifer sediments. Environ.

Sci. Technol. 30:2084-2086.

18. Dolan, M.E., and P.L. McCarty. 1995. Small-column microcosm

for assessing methane-stimulated vinyl chloride transformation in

aquifer samples. Environ. Sci. Technol. 29:1892-1897.

19. Hopkins, G.D., and P.L. McCarty. 1995. Field evaluation of in situ

aerobic cometabolism of trichloroethylene and three dichloroethylene isomers using phenol and toluene as the primary substrates. Environ. Sci. Technol. 29:1628-1637.

20. Major, D.W., W.W. Hodgins, and B.J. Butler. 1991. Field and

laboratory evidence of in situ biotransformation of tetrachloroethene to ethene and ethane at a chemical transfer facility in

North Toronto. In: Hinchee, R.E., and R.F. Olfenbuttel, eds. Onsite bioreclamation. Stoneham, MA: Butterworth- Heinemann. pp.

147-171.

21. Fiorenza, S., E.L. Hockman, Jr., S. Szojka, R.M. Woeller, and

J.W. Wigger. 1994. Natural anaerobic degradation of chlorinated

solvents at a Canadian manufacturing plant. In: Hinchee, R.E.,

A. Leeson, L. Semprini, and S. Kom, eds. Bioremediation of

chlorinated and polycyclic aromatic hydrocarbon compounds.

Boca Raton, FL: Lewis Publishers. pp. 277-286.

22. McCarty, P.L., and M. Reinhard. 1993. Biological and chemical

transformations of halogenated aliphatic compounds in aquatic

and terrestrial environments in the biochemistry of global change.

In: Oremland, R.S., ed. The biogeochemistry of global change:

Radiative trace gases. New York, NY: Chapman & Hall. pp. 839-

852.

23. Charnley, G., E.A.C. Crouch, L.C. Green, and T.L. Lash. 1988.

Municipal solid waste landfilling: A review of environmental effects. Prepared by Meta Systems, Inc., Cambridge, MA.

24. McCarty, P.L., L. Semprini, M.E. Dolan, T.C. Harmon, C. Tiedeman, and S.M. Gorelick. 1991. In situ methanotrophic bioremediation for contaminated groundwater at St. Joseph, Michigan. In:

Hinchee, R.E., and R.G. Olfenbuttel, eds. In: On-site bioreclamation processes for xenobiotic and hydrocarbon treatment. Boston,

MA: Butterworth-Heinemann. pp. 16-40.

25. Semprini, L., P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson. 1995.

Anaerobic transformation of chlorinated aliphatic hydrocarbons in

a sand aquifer based on spatial chemical distributions. Water

Resour. Res. 31(4):1051-1062.

26. Wilson, J.T., J.W. Weaver, and D.H. Kampbell. 1994. Intrinsic

bioremediation of TCE in ground water at an NPL site in St.

Joseph, Michigan. In: U.S. EPA Symposium on Intrinsic Bioremediation of Ground Water. EPA/540/R-94/515. Washington, DC.

27. Haston, Z.C., P.K. Sharma, J.N. Black, and P.L. McCarty. 1994.

Enhanced reductive dechlorination of chlorinated ethenes. In:

U.S. EPA Symposium on Bioremediation of Hazardous Wastes:

Research, Development, and Field Evaluation. EPA/600/R94/075. Washington, DC. pp. 11-14.

28. Major, D., E. Cox, E. Edwards, and P. Hare. 1995. Intrinsic

dechlorination of trichloroethene to ethene in a bedrock aquifer.

In: Hinchee, R.E., J.T. Wilson, and D.C. Downey, eds. Intrinsic

bioremediation. Columbus, OH: Battelle Press. pp. 197-203.

29. Lee, M.D., P.F. Mazierski, R.J. Buchanan, D.E. Ellis, and L.S.

Sehayek. 1995. Intrinsic in situ anaerobic biodegradation of chlorinated solvents at an industrial landfill. In: Hinchee, R.E., J.T.

Wilson, and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH: Battelle Press. pp. 205-222.

30. Cox, E., E. Edwards, L. Lehmicke, and D. Major. 1995. Intrinsic

biodegradation of trichloroethene and trichloroethane in a sequential anaerobic-aerobic aquifer. In: Hinchee, R.E., J.T. Wilson,

and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH:

Battelle Press. pp. 223-231.

31. Buchanan, J.R.J., D.E. Ellis, J.M. Odom, P.F. Mazierski, and M.D.

Lee. 1995. Intrinsic and accelerated anaerobic biodegradation of

perchloroethylene in groundwater. In: Hinchee, R.E., J.T. Wilson,

and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH:

Battelle Press. pp. 245-252.

11
Microbiological Aspects Relevant to Natural Attenuation of Chlorinated Ethenes

James M. Gossett

Cornell University, School of Civil and Environmental Engineering, Ithaca, New York

Stephen H. Zinder

Cornell University, Section of Microbiology, Ithaca, New York

Introduction

Chlorinated ethenes are widely employed as solvents in

civilian and military applications. They are excellent degreasing agents, nearly inflammable, and noncorrosive,

and in most applications they do not pose an acute

toxicological hazard. Not surprisingly, tetrachloroethene

(PCE) and the less-chlorinated ethenes produced from

it via reductive dehalogenation—trichloroethene (TCE),

dichloroethene (DCE) isomers, and vinyl chloride

(VC)—have become common ground-water pollutants,

often present as co-contaminants with fuel-derived pollutants such as benzene, toluene, ethylbenzene, and

xylenes (BTEX).

Results from many field and laboratory studies have

shown that chlorinated ethenes can be sequentially,

reductively dechlorinated under anaerobic conditions,

ultimately yielding ethene, which is environmentally acceptable (1, 2). The process requires some form of

electron donor (shown in Figure 1 as 2[H] per step), with

the chlorinated ethene serving as electron acceptor.

Since most significantly contaminated subsurface environments are indeed anaerobic, reductive dechlorination to ethene offers promise that natural attenuation

may be exploited in many instances of contamination by

chlorinated ethenes. The completeness of the conversion to ethene is highly variable from site to site, however, with the responsible factors for this variation not

well understood.

This paper presents some of the microbiological factors

that the authors believe influence the natural attenuation

of chlorinated ethenes.

Co-metabolic Versus Direct Dechlorination

Many of the early observations of reductive dechlorination of PCE and TCE were studies in which the mediating microorganisms were either obviously methanogens

(e.g., the pure-culture studies of Fathepure et al. [3-5])

or likely so. Many classes of anaerobic organisms (e.g.,

methanogens, acetogens, and sulfate reducers) have

been found to possess metal-porphyrin-containing cofactors that can mediate the slow, incomplete reductive

dechlorination of PCE and TCE to (usually) DCE isomers (6). This process is co-metabolic in that it happens

more or less accidentally or incidentally as the organisms carry out their normal metabolic functions; the

organisms apparently derive no growth-linked or energy-conserving benefit from the reductive dechlorination. Such co-metabolic dechlorinations undoubtedly are

responsible for the incomplete, relatively slow transformations of chloroethenes observed at many field sites.

The organisms that can mediate such processes are

ubiquitous, but the process is sufficiently slow and incomplete that a successful natural attenuation strategy

cannot completely rely upon it.

On the other hand, more recent studies have demonstrated the existence of direct dechlorinators—microorganisms derived from contaminated subsurface environments

and treatment systems—that utilize chlorinated ethenes

as electron acceptors in an energy-conserving, growthcoupled metabolism termed dehalorespiration (7). Several

species that carry out direct dechlorination of chlorinated

ethenes are described below.

Figure 1. Reductive dechlorination of chlorinated ethenes 

(under anaerobic conditions).

12
To a large extent, then, success or failure of natural

attenuation can be linked to the specific type of dechlorinator present (i.e., co-metabolic or direct), as well as to

the relative supply of H2 precursors compared with the

supply of chlorinated ethene that must be reduced.

Competitive Aspects of Dechlorination

Unfortunately, many users compete for H2 in anaerobic

microbial environments. For example, direct dechlorinators must compete for available H2 with hydrogenotrophic methanogens and sulfate reducers. Thus, in any

comprehensive, meaningful assessment of prospects

for natural attenuation, assessing only the nature of the

dechlorinators and the quantities of available donors

and chlorinated ethenes is insufficient; one must also

take into account competing demands for H2.

Because of the relatively high energy available from

reductive dechlorination, it is reasonable to suspect that

dechlorinators may out-compete methanogens for H2 at

very low H2 levels. Experimental evidence for this

comes from studies in which lactate was the administered electron donor, supplying H2 as it was rapidly

fermented to acetate. During the period of high H2 levels, methane production co-existed with dechlorination.

As lactate was depleted, H2 production waned, and H2

levels dropped to low levels; beyond this point, methane

production was negligible while dechlorination continued slowly. In fact, kinetic analysis of mixed cultures

of Dehalococcus ethenogenes and hydrogenotrophic

methanogens showed that this dechlorinator has an

affinity for H2 10 times greater than that of the methanogens in the culture (8). We do not know whether this high

affinity for H2 is typical of dechlorinators, but thermodynamic arguments would suggest it. We also do not yet

know the differences in relative affinity for H2 between

dechlorinators and sulfate-reducers, important competitors in many subsurface environments.

Competition for H2 is thus important, and the partitioning

of H2 flows among the various competitors is a function

of the H2 concentration, which itself depends on the rates

of H2 production and utilization. Compounds such as

lactate or ethanol that can be rapidly fermented to acetate, producing high, short-lived peaks of H2, do not

favor dechlorination as well as would more persistent,

slowly fermented substrates such as benzoate or propionate (and by extension, probably BTEX components).

The quality of the donor needs to be considered as

much as does its quantity. Comprehensive assessment

are best performed with microcosm studies, along with

microbiological analyses of in situ relative populations of

competing organisms and data on subsurface chemistry

(particularly of potentially competing electron acceptors).

Microbiology of Direct Dechlorinators

As summarized in Table 1, several organisms have

recently been isolated that can carry out direct respiratory reductive dechlorination of chloroethenes. All of these

organisms have been isolated since 1993, and several

more will likely be added to the list in the next few years.

A few tentative conclusions may be drawn from this

table. First, organisms that reduce PCE as far as cisDCE are relatively abundant and easier to culture. This

ability seems to have evolved in several different phylogenetic groups in the eubacteria, as determined by 16S

rRNA sequence analysis. Many direct dechlorinators

seem to be related to either the gram-negative sulfatereducing bacteria (epsilon proteobacteria) or the grampositive group, including Desulfotomaculum. Sulfate reducers

Table 1. Properties of Some Direct PCE Dechlorinators

Organism

Dechlorination

Reactions

Electron

Donors

Other Electron

Acceptors Morphology

Phylogenetic

Position References

Dehalobacter

restrictus

PCE, TCE → cis-DCE H2 None Rod Gram +

Desulfotomaculum

group

9-11

Dehalospirillum

multivorans

PCE, TCE → cis-DCE H2, formate,

pyruvate, etc.

Thiosulfate, nitrate,

fumarate, etc.

Spirillum Epsilon

proteobacteria

12

Strain TT4B PCE, TCE → cis-DCE Acetate None Rod ? 13

Enterobacter

agglomerans

PCE, TCE → cis-DCE Nonfermentable

substrates

O2, nitrate, etc. Rod Gamma

proteobacteria

14

Desulfitobacterium

sp. strain PCE1

PCE, TCE → (cis-DCE)

o-chlorophenols

Lactate,

pyruvate,

butyrate,

ethanol, etc.

Sulfite, thiosulfate,

fumarate

Curved rod Gram +

Desulfotomaculum

group

15

Dehalococcus

ethenogenes

strain 195

PCE, others → ethene H2 None Irregular

coccus

Novel eubacterium 16a

a

 Maymó-Gatell, X., Y.-T. Chien, J.M. Gossett, and S.H. Zinder. 1996. Isolation of a novel bacterium capable of reductively dechlorinating

tetrachloroethene to ethene. Unpublished data.

13
tend to be versatile at using electron acceptors for anaerobic respiration. We know much less about organisms

capable of reducing chloroethenes past DCE. These organisms play a crucial role in either producing VC, which

is degradable aerobically and under ferric iron-reducing

conditions, or ethene, which is nontoxic.

Some PCE-dechlorinating organisms appear versatile

at using electron donors and acceptors, while others,

most notably “Dehalobacter restrictus.” “D. ethenogenes,” and strain TT4B apparently can only use a

single electron donor and only chlorinated aliphatic hydrocarbons as electron acceptors. These findings raise

questions about what these organisms used as electron

acceptors before widespread chlorinated ethene contamination. The organisms possibly use electron acceptors not yet tested, or may once have been more

versatile but lost the ability to use other electron acceptors

in chlorinated ethene-contaminated environments or when

cultured on PCE as the sole electron acceptor.

Another important aspect of the PCE direct dechlorinators that Table 1 does not address is their nutrition.

Some PCE dechlorinators, such as Dehalospirillum

multivorans, require only acetate and carbon dioxide

as a carbon source (PCE and its daughter products

are not carbon sources), while others have a complex

nutrition, such as D. ethenogenes, which requires acetate, vitamin B12, unidentified factors in sewage sludge

(16), and perhaps other factors. Indeed, this organism’s requirement for vitamin B12 allowed a plausible

explanation for methanol’s being the best H2-source

for PCE dechlorination by the original mixed dechlorinating culture (17), since methanol-utilizing methanogens and acetogens are rich in vitamin B12 and related

corrinoid compounds. A butyrate-fed bioreactor faltered

until it was amended with vitamin B12 (18), which is

not present in yeast extract and apparently is in low

concentrations in the butyrate-oxidizing consortium

present in that bioreactor.

The Importance of Assessing the Big

Picture

This paper has attempted to address some of the microbial complexities of assessing natural attenuation potential. It is important to keep in mind the competitive

aspects of electron donor flow. In essence, dechlorination is in a “foot race” with competing donor uses. If too

little donor is initially present, the pattern of its conversion to H2 is too unfavorable, or there is too much

competition for it, dechlorination may not proceed adequately to completion. As other papers in this volume

suggest, relying on reductive dechlorination to achieve

complete conversion to ethene may not be necessary in

all cases; for example, some aerobic and iron-reducing

microbial processes can oxidize/mineralize VC. Therefore, conversion of PCE and TCE to VC by the time a

plume reaches an aerobic or iron-reducing zone may be

sufficient in many instances.

More problematic are situations in which degradation

proceeds only as far as DCEs. At some sites, there may

not be enough electron donor present. At other sites, a

sufficient amount of potential electron donors appear to

be present, and it is unclear whether further dechlorination is limited by physical/chemical factors, nutrients, or

lack of the appropriate dechlorinating organisms. Particularly troubling are sites in which PCE, TCE, and

DCEs reach aerobic zones in which they are essentially

nondegradable under natural conditions. Unfortunately,

our present understanding of the diversity and properties of organisms dechlorinating chlorinated ethenes

past DCEs is rudimentary.

In summary, the goal of assessment should be to evaluate the potential for sustained conversion to at least VC

in anaerobic zones. Comprehensive assessment thus

requires knowledge of both the quantity and quality of

the electron donor, of competing, alternative electron

acceptors (e.g., sulfates, ferric iron), and of relative

population levels of dechlorinating organisms and potentially competing microbial activities.

References

 1. deBruin, W.P., M.J.J. Kotterman, M.A. Posthumus, G. Schraa,

and A.J.B. Zehnder. 1992. Complete biological reductive transformation of tetrachloroethene to ethane. Appl. Environ. Microbiol. 58:1996-2000.

 2. Freedman, D.L., and J.M. Gossett. 1989. Biological reductive

dechlorination of tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions. Appl. Environ. Microbiol.

55:2144-2151.

 3. Fathepure, B.Z. and S.A. Boyd. 1988. Dependence of tetrachloroethylene dechlorination on methanogenic substrate consumption by Methanosarcina sp. Strain DCM. Appl. Environ. Microbiol.

54:2976-2980.

 4. Fathepure, B.Z., and S.A. Boyd. 1988. Reductive dechlorination

of perchloroethylene and the role of methanogens. FEMS Microbiol. Lett. 49:149-156.

 5. Fathepure, B.Z., J.P. Nengu, and S.A. Boyd. 1987. Anaerobic

bacteria that dechlorinate perchloroethene. Appl. Environ. Microbiol. 53:2671-2674.

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

catalyzed by bacterial transition-metal coenzymes. Environ. Sci.

Technol. 25:715-722.

 7. Holliger, C., and W. Schumacher. 1994. Reductive dehalogenation as a respiratory process. Antonie van Leeuwenhoek 66:239-

246.

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

kinetics of hydrogen utilization for reductive dechlorination of

tetrachloroethene and methanogenesis in an in press anaerobic

enrichment culture. Environ. Sci. Technol.

 9. Holliger, C. 1992. Reductive dehalogenation by anaerobic bacteria. Ph.D. dissertation. Agricultural University, Wageningen, the

Netherlands.

14
10. Holliger, C., G. Schraa, A.J.M. Stams, and A.J.B. Zehnder. 1992.

Enrichment and properties of an anaerobic mixed culture reductively dechlorinating 1,2,3-trichlorobenzene to 1,3-dichlorobenzene. Appl. Environ. Microbiol. 58:1636-1644.

11. Holliger, C., G. Schraa, A.J.M. Stams, and A.J.B. Zehnder. 1993.

A highly purified enrichment culture couples the reductive

dechlorination of tetrachloroethene to growth. Appl. Environ. Microbiol. 59:2991-2997.

12. Neumann, A., H. Scholz-Muramatsu, and G. Diekert. 1994.

Tetrachloroethene metabolism of Dehalospirillum multivorans.

Arch. Microbiol. 162:295-301.

13. Krumholz, L.R. 1995. A new anaerobe that grows with tetrachloroethylene as an electron acceptor. Abstract presented at the

95th General Meeting of the American Society for Microbiology.

14. Sharma, P.K., and P.L. McCarty. 1996. Isolation and characterization of a facultatively aerobic bacterium that reductively dehalogenates tetrachloroethene to cis-1,2-dichloroethene. Appl.

Environ. Microbiol. 62:761-765.

15. Gerritse, J., V. Renard, T.M. Pedro-Gomes, P.A. Lawson, M.D.

Collins, and J.C. Gottschal. 1996. Desulfitobacterium sp. strain

PCE1, an anaerobic bacterium that can grow by reductive

dechlorination of tetrachloroethene or ortho-chlorinated phenols.

Arch. Microbiol. 165:132-140.

16. Maymó-Gatell, X., V. Tandoi, J.M. Gossett, and S.H. Zinder. 1995.

Characterization of an H2-utilizing enrichment culture that reductively dechlorinates tetrachloroethene to vinyl chloride and ethene

in the absence of methanogenesis and acetogenesis. Appl. Environ. Microbiol. 61:3928-3933.

17. DiStefano, T.D., J.M. Gossett, and S.H. Zinder. 1991. Reductive

dechlorination of high concentrations of tetrachloroethene to

ethene by an anaerobic enrichment culture in the absence of

methanogenesis. Appl. Environ. Microbiol. 57:2287-2292.

18. Fennell, D.E., M.A. Stover, S.H. Zinder, and J.M. Gossett. 1995.

Comparison of alternative electron donors to sustain PCE anaerobic reductive dechlorination. In Hinchee, R.E., A. Leeson,

and L. Semprini, eds. Bioremediation of chlorinated solvents.

Columbus, OH: Battelle Press. pp. 9-16.

15
Microbial Ecology of Adaptation and Response in the Subsurface

Guy W. Sewell

U.S. Environmental Protection Agency, National Risk Management Research Laboratory,

Robert S. Kerr Environmental Research Center, Ada, Oklahoma

Susan A. Gibson

South Dakota State University, Department of Biology, Brookings, South Dakota

Introduction

The release of bio-oxidizable organic contaminants into

the subsurface and ground water quickly drives the local

environment anoxic and initiates a series of complex

and poorly understood responses by subsurface microorganisms. Field and laboratory research suggests that

multiple, physiologically defined communities develop

that are spatially and chronologically separate. These

communities are most likely ecologically defined by the

flux of biologically available electron donors and acceptors.

Under anaerobic conditions most organics continue to

degrade although the apparent rate may be slower.

Some contaminants may not be oxidatively catabolized,

however, due to thermodynamic limitations, lack of

genomic potential, or physical/chemical properties.

The parent chloroethenes—tetrachloroethene (PCE)

and trichloroethene (TCE)—are all too common examples of this type of ground-water contaminant. While

PCE and TCE do not seem to serve as carbon/energy

sources for subsurface bacteria, they can be reductively

biotransformed. These microbially mediated, naturally

occurring transformations (both oxidative and reductive)

of subsurface and ground-water contaminants have

been observed at many sites and hold significant potential for use as in situ remediation methods as the basis

for active or passive biotreatment technologies. While

these processes are observable and in some cases

have been demonstrated as remedial technologies,

however, our ability to predict the onset, extent, and

rates of transformation is limited. This lack of predictive

ability is more pronounced under anaerobic or intrinsic

conditions, and is extremely limited when reductive transformations are the target processes. Little is known about

the environmental parameters, microbial interactions,

and metabolic responses that control these degradation

processes in the subsurface.

A more complete understanding of the ecological and

physiological factors is needed for accurate and appropriate predictions and evaluations, particularly for in situ

transformation processes under intrinsic (native) conditions, where engineered approaches are not available

to influence or dominate in situ hydrogeochemical conditions. Under “native” conditions, the heterogeneity of

the site may also have a profound effect on the fate of

the contaminants. An understanding of the three-dimensional distribution of geochemical and hydraulic conditions is important for evaluating the contaminant

interactions with the subsurface microbial ecology. To

evaluate the likelihood of contaminant transformation, it

is necessary to have some understanding of the physiology of microorganisms in the subsurface and of the

ecological constraints that effect biological processes in

that environment.

Metabolic Principles

Heterotrophic organisms (like humans and most bacteria) oxidize organic compounds to obtain energy. In this

process electrons, or reducing equivalents, from the

oxidizable organic compound (substrate) are transferred

to and ultimately reduce an electron acceptor. The electron acceptor may be an organic or inorganic compound.

During this electron transfer process, usable energy is recovered through a complex series of oxidation-reduction

(redox) reactions by the formation of energy storage compounds or electrochemical gradients. The oxidation of organic compounds coupled with the reduction of molecular

oxygen is termed aerobic heterotrophic respiration and has

been the basis of most applications of bioremediation.

16
When oxygen is unavailable, biotransformations can still

occur. In anaerobic respiration, the oxidation of organic

matter can be coupled with a number of other organic

or inorganic electron acceptors. Some microorganisms

carry out a process known as fermentation. Fermenting

microorganisms utilize substrates as both an electron

donor and an electron acceptor. In this process, an

organic compound is metabolized, with a portion of

molecule becoming a reduced end product(s) and another becoming an oxidized end product(s). A common

example of this process is the alcoholic fermentation of

starch to carbon dioxide (CO2) (oxidized product) and

ethanol (reduced product). Fermentative organisms play

a critical role in anaerobic consortia by transforming

organic substrates into simple products which can then

be used by other members of the community, such as

dehalgenators, for further oxidation.

The potential energy available from the oxidation of a

particular substrate when coupled with the reduction of

different electron acceptors varies considerably. A higher

energy-yielding process will tend to predominate if the

required electron acceptor is available at biologically

significant concentrations (i.e., oxygen utilized before

nitrate). Under anaerobic conditions, microorganisms

may enter into very tightly linked metabolic consortia.

That is, the catalytic entity responsible for the destruction of a contaminant is often not a single type of microorganism. Such consortia can develop regardless of the

nature of the terminal electron acceptor.

As a class, the chloroethenes offer a diverse array of

metabolic fates. The parent compounds PCE and TCE

have been shown to undergo reductive transformations in

subsurface systems under the appropriate environmental

conditions. This reductive transformation process, referred to as reductive dechlorination or biodehalogenation, is a sequential removal of chlorine moieties from

the ethene core during a biologically mediated twoelectron transfer. Microorganisms in the subsurface and

other environments use the chloroethenes as terminal

electron acceptors and gain useable metabolic energy

by linking the oxidation of electron donors such as molecular hydrogen or organic compounds to the reduction

of chloroethenes. The exact mechanism of this type of

anaerobic respiration and the enzymes and co-factors

involved have yet to be identified. It is important to note

that the reductive dechlorination process only supplies

useable metabolic energy if coupled to the oxidation of

an appropriate electron donor.

TCE, dichloroethenes (DCEs), and vinyl chloride have

been shown to undergo co-metabolic oxidative transformations. By definition, co-metabolic process do not directly benefit the organisms buy supplying energy or

material for cellular synthesis. The mono-oxygenase

systems that transform chloroethenes may be inactivated during the process (competitive inhibitor). For the

co-metabolic process to occur, the true parent substrate

for the mono-oxygenase system and chloroethenes

must be present, as well as molecular oxygen. This

activity has been demonstrated as an active biotreatment process, but it is of limited significant under native

or intrinsic conditions because of the anticompetitive

effects on the microorganisms involved and the environmental conditions needed for significant transformation

to occur.

The lesser-chlorinated DCEs can be reductively transformed, and a growing body of evidence suggests that

they may be oxidatively catabolized with oxygen or other

electron acceptors. Vinyl chloride (monochloroethene)

is regarded as the most hazardous of the chloroethene

series. A known carcinogen, vinyl chloride is more mobile than the parent compounds and is extremely volatile. Due to its toxicity, when vinyl chloride is detected in

the subsurface environment with the other chloroethenes, it is usually the focus of risk-based evaluations and drives the cleanup process. Vinyl chloride

can be reductively modified to the nonchlorinated and

environmentally acceptable end product ethene. It can

also be oxidatively catabolized to CO2 and Cl-

 under

aerobic and iron-reducing conditions.

Ecological Principles

The subsurface environment is a unique and underappreciated ecosystem. The appropriate application of intrinsic remediation at a site requires an understanding

of the ecological processes under site-specific conditions. Subsurface microorganisms will respond to take

advantage of all available resources (i.e., energy, nutrients, space) that allow them to survive and reproduce.

This response is bounded by evolutionary, physical, and

thermodynamic constraints, but in general we see that

microorganisms (and life in general) rapidly take advantage of resources and conditions. This concept is referred to as filling all available ecological niches. The

converse statement is also true, however: subsurface

microorganism do not respond in ways that are counterproductive to this competition for resources and survival.

This beguilingly simple concept is often overlooked in

the design and implementation of bioremediation.

The nature of the subsurface environment (i.e., its lack

of primary production) makes it useful to view subsurface (microbial) ecology in terms of energy transformation and transfer. Under intrinsic conditions, bioavailable

energy is the ecological resource that induces the observed biodegradation process. As noted earlier, the

transfer or harvesting of this resource by subsurface

microorganisms involves oxidation/reduction couples

available to subsurface microorganisms. Under pristine conditions, energy transduction in ground-water environments is

limited by the availability of carbon/energy sources

(electron donors). Energy transduction in contaminated

17
subsurface systems is usually limited by the availability

of the electron acceptor. When readily degradable organic matter enters the subsurface in sufficient quantities, it produces a series of zones defined by the

terminal electron accepting process (TEAP). These

zones are not necessarily mutually exclusive and depend on the availability of electron acceptors (O2, NO3=,

SO4=, Fe3+, CO2). There is no reason to assume a

similarity between the biodegradation potential in different metabolic zones. This potential will be based on the

energetics associated with the dominant redox processes, the metabolic diversity of the microbial communities, the immediate geochemical conditions, and the

chemical nature of the contaminant of concern.

The reductive dehalogenation process may be thought

of as another TEAP, and the microorganisms involved

compete for the available flow of energy (reducing

equivalents). As noted above, however, PCE or TCE, in

the absence of sufficient electron donors such as an

oxidizable co-contaminant or native organic matter,

does not represent a resource to the indigenous microorganisms. This is why PCE and TCE plumes with detectable levels of dissolved oxygen do not show

evidence of active biodegradation. The presence of dissolved oxygen indicates no significant quantities of oxidizable electron donor are present. Vinyl chloride (and

perhaps DCEs) under the same conditions may undergo

further transformation, however, if appropriate electron

acceptors such as O2 or Fe3+ are present. Under these

conditions, the oxidation of vinyl chloride represents a

resource (energy) to the subsurface microbial populations.

Mechanisms of Adaptation

While an understanding of the ecological processes is

useful in predicting whether a transformation is likely to

occur, an understanding of the adaptation processes’

mechanisms is needed to predict the onset of the degradation activity. Possible mechanisms of adaptation

include expression of catabolic potential (induction), selection of novel capabilities (mutation), growth of degradative populations, formation of degradative consortia,

and formation of metabolic intermediate pools. Laboratory research results indicate that the formation of catabolically competent consortia could be a limiting step

in the observed lag before the onset of degradation.

Historical exposure and total microbial mass did not

significantly affect the observed lag but did affect transformation rates. Environmental parameters that support

anaerobic microbial transformation processes (both oxidative and reductive) positively affected the observed

adaptation response. While more work is needed, these

preliminary observations offer some explanations for

varying field observations and suggest that clearer understanding of the mechanisms involved may lead to

greater predictive capabilities.

Conclusion

Biotransformations that serve as major mass removal

mechanisms for the intrinsic remediation of chloroethenes and other ground-water contaminants have

been demonstrated in the laboratory and the field. The

use of these processes as remedial technologies, however, is difficult to evaluate at field scale, which limits our

ability to predict the rate and extent of the degradation

of contaminants in complex, heterogeneous subsurface

environments. An understanding of the physiology and

ecology of subsurface microorganisms is one of the few

tools regulators and scientists have to evaluate the appropriate implementation of intrinsic remediation for

chloroethene sites. A greater understanding of the

mechanisms of adaptation in subsurface microbial communities could also prove to be useful in the appropriate

application of in situ bioremediation under active or

intrinsic conditions.

Additional Reading

Chapelle, F.H. 1993. Ground water microbiology and geochemistry.

New York, NY: John Wiley and Sons.

U.S. EPA. 1991. Environmental Research Brief: Anaerobic biotransformation of contaminants in the subsurface. EPA/600/M-90/024. 

February.

Vogel, T.M., C.S. Criddle, and P.L. McCarty. 1987. Transformations of

halogenated aliphatic compounds. Environ. Sci. Technol. 22:722-736.

18
Identifying Redox Conditions That Favor the Natural Attenuation of Chlorinated

Ethenes in Contaminated Ground-Water Systems

Francis H. Chapelle

U.S. Geological Survey, Columbia, South Carolina

Introduction

Over the last several years, it has been demonstrated

that petroleum hydrocarbons biodegrade in virtually all

ground-water systems (1), and that natural attenuation

can greatly reduce the transport of contaminants away

from particular hydrocarbon spills (2, 3). These results

have raised the prospect that chlorinated ethenes—perchloroethene (PCE), trichloroethene (TCE), dichloroethenes (DCEs), and vinyl chloride (VC)—will prove

similarly amenable to natural attenuation processes.

The microbial processes leading to biodegradation of

chlorinated ethenes, however, can be much different

from those that degrade petroleum hydrocarbons. Petroleum hydrocarbons universally serve as electron donors (i.e., as an energy source) in microbial metabolism.

In contrast, chlorinated ethenes, in addition to serving

as electron donors, can function as electron acceptors

(i.e., they are reduced via reductive dechlorination) or

can be fortuitously degraded by various co-metabolic

processes. Because of this diversity, it is not surprising

that the efficiency with which chlorinated ethenes are

naturally attenuated varies widely among ground-water

systems.

Under anoxic conditions, chlorinated ethenes are subject to reductive dechlorination according to the sequence PCE → TCE + Cl → DCE + 2Cl → VC + 3Cl →

ethylene + 4Cl (1). The efficiency of dechlorination,

however, appears to differ under methanogenic, sulfate-reducing, iron(III)-reducing, and nitrate-reducing

con- ditions. Dechlorination of PCE and TCE to DCE is

favored under mildly reducing conditions such as nitrate

or iron(III) reduction (4), whereas the transformations of

DCE to VC or of VC to ethylene seems to require the

more strongly reducing conditions of methanogenesis

(5-7). Further complicating this picture, lightly chlorinated ethenes such as VC can be oxidized under oxic

(8) or iron(III)-reducing conditions (9), and by various

co-metabolic degradation processes (10).

Clearly, an accurate delineation of redox conditions is

central to evaluating the potential for the natural attenuation of chlorinated ethenes in ground-water systems.

This paper summarizes a methodology for identifying

the zonation of redox conditions in the field. This methodology can serve as an a priori screening tool for

identifying ground-water systems in which redox conditions will favor natural attenuation of chlorinated

ethenes. Conversely, this methodology can identify systems for which natural attenuation of chlorinated

ethenes is not favored and other remediation technologies should be considered.

Methodology for Determining Redox

Processes in Ground-Water Systems

Platinum electrode redox potential measurement historically has been the most widely used method for determining redox conditions in ground-water systems. While

redox potential measurements can accurately distinguish oxic from anoxic ground water, they cannot distinguish between different anoxic processes such as

nitrate reduction, iron(III) reduction, sulfate reduction, or

methanogenesis. One reason is that many redox species, such as hydrogen sulfide (H2S) or methane (CH4),

are not electroactive on platinum electrode surfaces

(11). Because distinguishing between these processes

is critical to evaluate the natural attenuation of chlorinated ethenes, redox potential measurements alone cannot provide the needed information.

A different methodology, which is based on microbial

physiology, has recently been introduced for delineating

redox processes (12-14). This method relies on three

lines of evidence: the consumption of electron acceptors,

the production of metabolic end products, and the measurement of concentrations of transient intermediate

19
products. Molecular hydrogen (H2), the most ubiquitous

intermediate product of anaerobic microbial metabolism,

has proven to be especially useful in this context. Different electron-accepting processes have characteristic

H2-utilizing efficiencies. Nitrate reduction, the most energetically favorable anoxic process, maintains H2 concentrations below 0.1 nanomoles (nM) per liter. Iron(III)

reduction maintains H2 concentrations between 0.2 and

0.8 nM, whereas for sulfate reduction the characteristic

range is between 1 and 4 nM. Methanogenesis, the least

energetically favorable anoxic process, is characterized

by H2 in the 5 to 15 nM range.

Patterns of electron-acceptor consumption, final product

accumulation, and H2 concentrations can be combined to

logically identify redox processes (13). For example, if

sulfate concentrations are observed to decrease along an

aquifer flowpath, if sulfide concentrations are observed to

increase, and if H2 concentrations are in the 1 to 4 nM

range characteristic of sulfate reduction, it may be concluded with a high level of confidence that sulfate reduction

is the predominant redox process. If all three possible

indicators (electron acceptor consumption, end-product

production, and H2 concentrations) indicate the same redox process, a high degree of confidence in the delineation

is warranted. Conversely, if only one indicator is available,

or if lines of evidence conflict, proportionally less confidence in the redox delineation is warranted.

Measuring Hydrogen Concentrations in

Ground Water

With the exception of dissolved hydrogen (H2), all of the

redox-sensitive parameters (dissolved oxygen, nitrate,

nitrite, ferrous iron [Fe2+], H2S, sulfate, and methane)

needed to assess redox processes are routinely examined in ground-water chemistry investigations. Hydrogen concentrations in ground water can be made using

a gas-stripping procedure (13). A standard gas-sampling

bulb is attached to a stream of water produced from a

well and purged for several minutes (at approximately

500 milliliters/minute) to eliminate all gas bubbles. Next,

20 milliliters of nitrogen, made H2-free by passage

through a hopcalite column, is introduced to the bulb

through a septum. As water continues to purge the bulb,

H2 and other slightly soluble gases partition to the headspace and asymptotically approach equilibrium with the

dissolved phase. After 20 to 25 minutes, equilibrium is

achieved, and the gas bubble is sampled using a syringe. A duplicate sample is taken 5 minutes later. H2 is

then measured by gas chromatography with reduction

gas detection. Concentrations of aqueous H2 are then

calculated from H2 solubility data. For fresh water in

equilibrium with a gas phase at 1 atmosphere pressure,

1.0 parts per million H2 in the gas phase corresponds to

0.8 nM of dissolved H2.

An Example of Redox Zone Delineation

Related to the Natural Attenuation of

Chlorinated Ethenes—Cecil Field, Florida

An example of how redox processes can be delineated,

and how this delineation affects assessment of natural

attenuation of chlorinated ethenes, is a study performed

by the U.S. Geological Survey in cooperation with the

U.S. Navy at Site 8, Naval Air Station (NAS) Cecil Field.

Site 8 was a fire-training area used to train Navy personnel in firefighting procedures (Figure 1). Over the

operational life of Site 8, a variety of petroleum products

and chlorinated solvents seeped into the underlying

ground-water system.

Changes in the concentrations of redox-sensitive constituents along the flowpath of the shallow aquifer system are shown in Figure 2. Ground water at the site is

oxic upgradient of the fire pits but becomes anoxic

downgradient of the fire pits (Figure 2A). Once the water

becomes anoxic, concentrations of methane begin to

rise, peaking at about 7 milligrams per liter 200 feet

downgradient (Figure 2A) and indicating methanogenic

conditions. Between 170 and 400 feet downgradient,

concentrations of sulfate decrease and concentrations

of H2S increase (Figure 2B), indicating active sulfate

reduction. Concentrations of dissolved Fe2+ remain below 1 mg/L until about 400 feet along the flowpath, then

increase to about 2.5 mg/L, indicating active iron(III)

reduction. The H2 concentrations are consistent with the

redox zonation indicated by the other redox-sensitive

parameters (Figure 2C). H2 concentrations in the range

characteristic of methanogenesis are observed in

ground water near the fire-training pits where high methane concentrations are present.

Between 200 and 170 feet downgradient, where sulfate

concentrations decline and sulfide concentrations inFigure 1. Map showing location of fire-training pits and monitoring wells, Site 8, NAS Cecil Field, Florida.

20
crease, H2 concentrations are in the 1 to 4 nM range

characteristic of sulfate reduction. Finally, between 400

and 500 feet downgradient, where concentrations of

Fe2+ increase, H2 concentrations are in the 0.2 to 0.8 nM

range characteristic of iron(III) reduction.

A cross section showing the interpretation of these data

and including wells screened deeper in the flow system

is given in Figure 3. A methanogenic zone is present

near the contaminant source, surrounded by sulfate-reducing and iron(III)-reducing zones further downgradient. This redox zonation suggests that the natural

attenuation of chlorinated ethenes will be rapid and

efficient at this site. Near the contaminant source,

methanogenic and sulfate-reducing zones favor

dechlorination of PCE, TCE, and DCE. In the downgradient iron(III)-reducing zone, anoxic oxidation of VC

to carbon dioxide (CO2) can occur (Figure 3).

These biodegradation processes, which can be postulated solely on the basis of the observed redox zonation,

are consistent with the observed behavior of chlorinated

ethenes at this site (Figure 4A). PCE, TCE, and VC are

present in ground water near the fire-training pits but

drop below detectable levels along the flowpath. In fact,

natural attenuation of chlorinated ethenes at this site has

been so efficient that the best water-chemistry record of

the original contamination is probably the elevated concentrations of dissolved inorganic carbon (Figure 4B)

and dissolved chloride (Figure 4C) observed in downgradient ground water that currently lacks measurable

chlorinated ethene contamination. These patterns suggest

that most of the chlorinated ethenes have been completely transformed to CO2 and chloride by the cumulaFigure 2. Concentration changes of redox-sensitive parameters

along ground-water flowpaths in the shallow aquifer,

Site 8, NAS Cecil Field, Florida.

Figure 3. Concentrations of dissolved hydrogen (nM) and the

zonation of predominant redox processes, Site 8,

NAS Cecil Field, Florida.

Figure 4. Concentration changes of chloride, dissolved inorganic carbon, and chlorinated ethenes along groundwater flowpaths.

21
tive effects of reductive dehalogenation in the methanogenic and sulfate-reducing zones and oxidative processes in the downgradient iron(III)-reducing and oxic

zones.

Conclusion

An understanding of ambient redox conditions is a powerful tool for assessing the efficiency of natural attenuation

of chlorinated ethenes. The methodology for assessing

redox conditions involves tracking the disappearance of

electron acceptors, the appearance of end products, and

concentrations of H2. Using this information, it is possible

to logically deduce redox zonation at particular sites, and

assess the confidence that is appropriate for the delineation. This methodology was demonstrated at a site at

NAS Cecil Field, Florida. The progression from methanogenic → sulfate reduction → Iron(III) reduction → oxygen

reduction has efficiently decreased concentrations of chlorinated ethenes, indicating that natural attenuation is a

viable remedial option at this site.

References

 1. Hinchee, R.E., J.A. Kittel, and H.J. Reisinger, eds. 1995. Applied

bioremediation of petroleum hydrocarbons. Columbus, OH:

Batelle Press.

 2. Weidemeyer, T.H., D.C. Downey, J.T. Wilson, D.H. Kampbell,

R.N. Miller, and J.E. Hansen. 1995. Technical protocol for implementing the intrinsic remediation with long-term monitoring option

for natural attenuation of dissolved-phase fuel contamination in

ground water. U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, San Antonio, TX. p. 129.

 3. Chapelle, F.H., J.M. Landmeyer, and P.M. Bradley. 1996. Assessment of intrinsic bioremediation of jet fuel contamination in a

shallow aquifer, Beaufort, South Carolina. U.S. Geological Survey

Water Resources Investigations Report 95-4262.

 4. Vogel, T.M., C.S. Criddle, and P.L. McCarty. 1987. Transformations of halogenated aliphatic compounds. Environ. Sci. Technol.

21:721-736.

 5. Friedman, D.L., and J.M. Gossett. 1989. Biological reductive

dechlorination of tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions. Appl. Environ. Microbiol.

55:2144-2151.

 6. DeBrunin, W.P., M.J.J. Kotterman, M.A. Posthumus, G. Schraa,

and A.J.B. Zehnder. 1992. Complete biological reductive transformation of tetrachloroethene to ethene. Appl. Environ. Microbiol. 58:1996-2000.

 7. DiStefano, T.D., J.M. Gossett, and S.H. Zinder. 1991. Reductive

dechlorination of high concentrations of tetrachloroethene to

ethene by an anaerobic enrichment culture in the absence of

methanogenesis. Appl. Environ. Microbiol. 57:2287-2292.

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

of vinyl chloride in groundwater samples. Appl. Environ. Microbiol.

56:3878-3880.

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

of vinyl chloride in Fe(III) reducing aquifer sediments. Environ.

Sci. Technol. 40:2084-2086.

10. McCarty, P.L., and L. Semprini. 1994. Groundwater treatment for

chlorinated solvents. In: Handbook of bioremediation. Boca Raton, FL: Lewis Publishers. pp. 87-116.

11. Stumm, W. and J.J. Morgan. 1981. Aquatic chemistry. New York,

NY: John Wiley & Sons. p. 780.

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

as an indicator of the predominant terminal electron-accepting

reaction in aquatic sediments. Geochim. Cosmochim. Acta

52:2993-3003.

13. Chapelle, F.H., P.B. McMahon, N.M. Dubrovsky, R.F. Fujii, E.T.

Oaksford, and D.A. Vroblesky. 1995. Deducing the distribution of

terminal electron-accepting processes in hydrologically diverse

groundwater systems. Water Resour. Res. 31:359-371.

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

dissolved H2 concentrations to determine distribution of microbially catalyzed redox reactions in anoxic groundwater. Environ.

Sci. Technol. 28:1255-1210.

22
Design and Interpretation of Microcosm Studies for Chlorinated Compounds

Barbara H. Wilson and John T. Wilson

U.S. Environmental Protection Agency, National Risk Management Research Laboratory, 

Ada, Oklahoma

Darryl Luce

U.S. Environmental Protection Agency, Region 1, Boston, Massachusetts

Introduction

Three lines of evidence are used to support natural

attenuation as a remedy for chlorinated solvent contamination in ground water: documented loss of contaminant

at field scale, geochemical analytical data, and direct

microbiological evidence. The first line of evidence

(documented loss) involves using statistically significant

historical trends in contaminant concentration in conjunction with aquifer hydrogeological parameters (such

as seepage velocity and dilution) to show that a reduction in the total mass of contaminants is occurring at the

site. The second line of evidence (geochemical data)

involves the use of chemical analytical data in mass

balance calculations to show that decreases in contaminant concentrations can be directly correlated to increases in metabolic byproduct concentrations. This

evidence can be used to show that concentrations of

electron donors or acceptors in ground water are sufficient to facilitate degradation of the dissolved contaminants (i.e., there is sufficient capacity). Solute fate and

transport models can be used to aid the mass balance

calculations and to collate information on degradation.

Microcosm studies are often used to provide a third line

of evidence. The potential for biodegradation of the contaminants of interest can be confirmed using of microcosms

through comparison of removals in the living treatments

with removals in the controls. Microcosm studies also permit

an absolute mass balance determination based on biodegradation of the contaminants of interest. Further, the appearance of daughter products in the microcosms can be

used to confirm biodegradation of the parent compound.

When To Use Microcosms

Microcosms have two fundamentally different applications. First, they are frequently used in a qualitative way to

illustrate the important processes that control the fate of

organic contaminants. Second, they are used to estimate

rate constants for biotransformation of contaminants that

can be used in a site-specific transport-and-fate model of

a contaminated ground-water plume. This paper discusses

the second application.

Microcosms should be used when there is no other way

to obtain a rate constant for attenuation of contaminants,

particularly when estimating the rate of attenuation from

monitoring well data in the plume of concern is impossible. In some situations, there are legal or physical impediments to the comparison of concentrations in

monitoring wells along a flow path. In many landscapes,

the direction of ground-water flow (and water-table elevations in monitoring wells) can vary over short periods

due to tidal influences or changes in barometric pressure. Changes in the stage of a nearby river or pumping

wells in the vicinity can also affect the direction of

ground-water flow. These changes in ground-water flow

direction do not allow simple “snapshot” comparisons of

concentrations in monitoring wells because of uncertainties in identifying the flow path. Rate constants from

microcosms can be used with average flow conditions

to estimate attenuation at some point of discharge or

point of compliance.

Application of Microcosms

The primary objective of microcosm studies is to obtain

rate constants applicable to average flow conditions.

These average conditions can be determined by continuous monitoring of water-table elevations in the aquifer being evaluated. The product of the microcosm

study, and the continuous monitoring of water-table elevations, will be a yearly or seasonal estimate of the

extent of attenuation along average flow paths. Removals seen at field scale can be attributed to biological

23
activity. If removals in the microcosms duplicate removal

at field scale, the rate constant can be used for risk

assessment purposes.

Selecting Material for Study

Prior to choosing material for microcosm studies, the

location of major conduits of ground-water flow should

be identified, and the geochemical regions along the

flow path should be determined. The important geochemical regions for natural attenuation of chlorinated

aliphatic hydrocarbons are regions that are actively

methanogenic, exhibit sulfate reduction and iron reduction concomitantly, or exhibit iron reduction alone. The

pattern of chlorinated solvent biodegradation varies in

different regions. Vinyl chloride tends to accumulate

during reductive dechlorination of trichloroethylene

(TCE) or tetrachloroethylene (PCE) in methanogenic

regions (1, 2); it does not accumulate to the same extent

in regions exhibiting iron reduction and sulfate reduction

(3). In regions showing iron reduction alone, vinyl chloride is consumed but dechlorination of PCE, TCE, or

dichloroethylene (DCE) may not occur (4). Core material

must be acquired from each geochemical region in major flow paths represented by the plume, and the hydraulic conductivity of each depth at which core material is

acquired must be measured. If possible, the microcosms should be constructed with the most transmissive

material in the flow path.

Several characteristics of ground water from the same

interval used to collect the core material should be

determined, including temperature, redox potential, pH,

and concentrations of oxygen, sulfate, sulfide, nitrate,

ferrous iron, chloride, methane, ethane, ethene, total

organic carbon, and alkalinity. The concentrations of

compounds of regulatory concern and any breakdown

products for each site must be determined. The ground

water should be analyzed for methane to determine

whether methanogenic conditions exist and for daughter

products ethane and ethene. A comparison of the

ground-water chemistry from the interval in which the

cores were acquired with that in neighboring monitoring

wells will demonstrate whether the collected cores are

representative of that section of the contaminant plume.

Reductive dechlorination of chlorinated solvents requires an electron donor for the process to proceed. The

electron donor could be soil organic matter, low molecular weight organic compounds (e.g., lactate, acetate, methanol, glucose), H2, or a co-contaminant such as landfill

leachate or petroleum compounds (5-7). In many instances,

the actual electron donor(s) may not be identified.

Several characteristics of the core material should also

be evaluated. The initial concentration of the contaminated material (in micrograms per kilogram) should be

identified before constructing the microcosms. It is also

necessary to determine whether the contamination is

present as a nonaqueous-phase liquid (NAPL) or in

solution. A total petroleum hydrocarbon (TPH) analysis

will reveal the presence of any hydrocarbon-based oily

materials. The water-filled porosity, a parameter generally used to extrapolate rates to the field, can be calculated by comparing wet and dry weights of the aquifer

material.

To ensure sample integrity and stability during acquisition, it is important to quickly transfer the aquifer material

into a jar, exclude air by adding ground water, and seal

the jar without headspace. The material should be

cooled during transportation to the laboratory, then incubated at the ambient ground-water temperature in the

dark before the construction of microcosms.

At least one microcosm study per geochemical region

should be completed. If the plume is greater than 1

kilometer in length, several microcosm studies per geochemical region may need to be constructed.

Geochemical Characterization of the Site

The geochemistry of the subsurface affects the behavior

of organic and inorganic contaminants, inorganic minerals, and microbial populations. Major geochemical parameters that characterize the subsurface include

alkalinity, pH, redox potential, dissolved constituents (including electron acceptors), temperature, the physical

and chemical characterization of the solids, and microbial processes. The most important of these in relation

to biological processes are alkalinity, redox potential, the

concentration of electron acceptors, and the chemical

nature of the solids.

Alkalinity

Biologically active portions of a plume may be identified

in the field by their increased alkalinity (compared with

background wells), caused by the carbon dioxide resulting from biodegradation of the pollutants. Increases in

both alkalinity and pH have been measured in portions

of an aquifer contaminated by gasoline undergoing active utilization of the gasoline components (8). Alkalinity

can be one of the parameters used to identify where to

collect biologically active core material.

pH

Bacteria generally prefer a neutral or slightly alkaline pH

level, with an optimum pH range for most microorganisms between 6.0 and 8.0; many microorganisms, however, can tolerate a pH range of 5.0 to 9.0. Most ground

waters in uncontaminated aquifers are within these

ranges. Natural pH values may be as low as 4.0 or 5.0

in aquifers with active oxidation of sulfides, and pH

values as high as 9.0 may be found in carbonate-buffered systems (9). pH values as low as 3.0 have been

measured for ground waters contaminated with municipal

24
waste leachates, however, which often contain elevated

concentrations of organic acids (10). In ground waters contaminated with sludges from cement manufacturing, pH

values as high as 11.0 have been measured (9).

Redox Potential

The oxidation/reduction (redox) potential of ground

water is a measure of electron activity that indicates the

relative ability of a solution to accept or transfer electrons. Most redox reactions in the subsurface are microbially catalyzed during metabolism of native organic

matter or contaminants. The only elements that are

predominant participants in aquatic redox processes are

carbon, nitrogen, oxygen, sulfur, iron, and manganese

(11). The principal oxidizing agents in ground water are

oxygen, nitrate, sulfate, manganese(IV), and iron(III).

Biological reactions in the subsurface both influence and

are affected by the redox potential and the available

electron acceptors. The redox potential changes with

the predominant electron acceptor, with reducing conditions increasing through the sequence oxygen, nitrate,

iron, sulfate, and carbonate. The redox potential decreases in each sequence, with methanogenic (carbonate as the electron acceptor) conditions being most

reducing. The interpretation of redox potentials in

ground water is difficult (12). The potential obtained in

ground water is a mixed potential that reflects the potential of many reactions and cannot be used for quantitative interpretation (11). The approximate location of the

contaminant plume can be identified in the field by

measuring the redox potential of the ground water.

To overcome the limitations imposed by traditional redox

measurements, recent work has focused on measuring

molecular hydrogen to accurately describe the predominant in situ redox reactions (13-15). The evidence suggests that concentrations of H2 in ground water can be

correlated with specific microbial processes, and these

concentrations can be used to identify zones of

methanogenesis, sulfate reduction, and iron reduction in

the subsurface (3).

Electron Acceptors

Measuring the available electron acceptors is a critical

step in identifying the predominant microbial and geochemical processes occurring in situ at the time of sample collection. Nitrate and sulfate are found naturally in

most ground waters and will subsequently be used as

electron acceptors once oxygen is consumed. Oxidized

forms of iron and manganese can be used as electron

acceptors before sulfate reduction commences. Iron

and manganese minerals solubilize coincidently with

sulfate reduction, and their reduced forms scavenge

oxygen to the extent that strict anaerobes (some sulfate

reducers and all methanogens) can develop. Sulfate is

found in many depositional environments, and sulfate

reduction may be very common in many contaminated

ground waters. In environments where sulfate is depleted, carbonate becomes the electron acceptor, with

methane gas produced as an end product.

Temperature

The temperature at all monitoring wells should be measured to determine when the pumped water has stabilized and is ready for collection. Below approximately 30

feet, the temperature in the subsurface is fairly consistent on an annual basis. Microcosms should be stored

at the average in situ temperature. Biological growth can

occur over a wide range of temperatures, although most

microorganisms are active primarily between 10°C and

35°C (50°F to 95°F).

Chloride

Reductive dechlorination results in the accumulation of

inorganic chloride. In aquifers with a low background of

inorganic chloride, the concentration of inorganic chloride should increase as the chlorinated solvents degrade. The sum of the inorganic chloride plus the

contaminant being degraded should remain relatively

consistent along the ground-water flow path.

Tables 1 and 2 list the geochemical parameters, contaminants, and daughter products that should be measured during site characterization for natural attenuation.

The tables include the analyses that should be performed, the optimum range for natural attenuation of

chlorinated solvents, and the interpretation of the value

in relation to biological processes.

Table 1. Geochemical Parameters

Analysis Range Interpretation

Redox potential < 50 mV

against Ag/AgCl

Reductive pathway possible

Sulfate < 20 mg/L Competes at higher

concentrations with reductive

pathway

Nitrate < 1 mg/L Competes at higher

concentrations with reductive

pathway

Oxygen < 0.5 mg/L Tolerated; toxic to reductive

pathway at higher

concentrations

Oxygen > 1 mg/L Vinyl chloride oxidized

Iron(II) > 1 mg/L Reductive pathway possible

Sulfide > 1 mg/L Reductive pathway possible

Hydrogen > 1 nM Reductive pathway possible;

vinyl chloride may

accumulate 

Hydrogen < 1 nM Vinyl chloride oxidized

pH 5 < pH < 9 Tolerated range

25
Microcosm Construction

During construction of the microcosms, manipulations

should take place in an anaerobic glovebox. These

gloveboxes exclude oxygen and provide an environment in which the integrity of the core material may be

maintained, since many strict anaerobic bacteria are sensitive to oxygen. Stringent aseptic precautions are not

necessary for microcosm construction; maintaining the

anaerobic conditions of the aquifer material and solutions added to the microcosm bottles is more important.

The microcosms should have approximately the same

ratio of solids to water as the in situ aquifer material, with

minimal or negligible headspace. Most bacteria in the

subsurface are attached to the aquifer solids. If a microcosm has too much water and the contaminant is primarily in the dissolved phase, the bacteria must

consume or transform a great deal more contaminant to

produce the same relative change in the contaminant

concentration. As a result, the kinetics of removal at field

scale will be underestimated in the microcosms.

A minimum of three replicate microcosms for both living

and control treatments should be constructed for each

sampling event. Microcosms sacrificed at each sampling interval are preferable to microcosms that are repetitively sampled. The compounds of regulatory interest

should be added at concentrations representative of the

higher concentrations found in the geochemical region

of the plume being evaluated, and should be added as

concentrated aqueous solutions. If an aqueous solution

is not feasible, dioxane or acetonitrile may be used as

solvents. Carriers that can be metabolized anaerobically

should be avoided, particularly alcohols. If possible,

ground water from the site should be used to prepare

dosing solutions and to restore water lost from the core

barrel during sample collection.

Although no method is perfect, autoclaving is the preferred sterilization method for long-term microcosm

studies, and mercuric chloride is excellent for short-term

studies (weeks or months). Mercuric chloride complexes

to clays, however, and control may be lost as it is sorbed

over time. Sodium azide is effective in repressing metabolism of bacteria that have cytochromes but is not

effective on strict anaerobes.

The microcosms should be incubated in the dark at the

ambient temperature of the aquifer. Preferably, the microcosms should be inverted in an anaerobic glovebox

as they incubate; anaerobic jars are also available that

maintain an oxygen-free environment. Dry redox indicator strips can be placed in the jars to ensure that anoxic

conditions are maintained. If no anaerobic storage is

available, the inverted microcosms can be immersed in

approximately 2 inches of water during incubation. Teflon-lined butyl rubber septa are excellent for excluding

oxygen and should be used if the microcosms must be

stored outside an anaerobic environment.

The studies should last from 12 to 18 months. The

residence time of a plume may be several years to tens

of years at field scale. Rates of transformation that are

slow in terms of laboratory experimentation may have a

considerable environmental significance, and a microcosm study lasting only a few weeks to months may not

have the resolution to detect slow changes that are of

environmental significance. Additionally, microcosm

studies often distinguish a pattern of sequential biodegradation of the contaminants of interest and their

daughter products.

Microcosm Interpretation

As a practical matter, batch microcosms with an optimal

solids/water ratio that are sampled every 2 months in

triplicate for up to 18 months, can resolve biodegradation from abiotic losses with a detection limit of 0.001 to

0.0005 per day. Rates determined from replicated batch

microcosms are found to more accurately duplicate field

rates of natural attenuation than column studies. Many

plumes show significant attenuation of contamination at

field-calibrated rates that are slower than the detection

limit of microcosms constructed with that aquifer material. Although rate constants for modeling purposes are

more appropriately acquired from field-scale studies,

agreement between rates in the field and rates in the

laboratory is reassuring.

The rates measured in the microcosm study may be

faster than the estimated field rate. This may not be due

to an error in the laboratory study, particularly if estimation of the field-scale rate of attenuation did not account

for regions of preferential flow in the aquifer. The regions

Table 2. Contaminants and Daughter Products

Analysis Interpretation

PCE Material spilled

TCE Material spilled or daughter product of

perchloroethylene

1,1,1-Trichloroethane Material spilled

cis-DCE Daughter product of trichloroethylene

trans-DCE Daughter product of trichloroethylene

Vinyl chloride Daughter product of dichloroethylene

Ethene Daughter product of vinyl chloride

Ethane Daughter product of ethene

Methane Ultimate reductive daughter product

Chloride Daughter product of organic chlorine

Carbon dioxide Ultimate oxidative daughter product

Alkalinity Results from interaction of carbon

dioxide with aquifer minerals

26
of preferential flow may be determined using a down-hole

flow meter or a geoprobe method for determining hydraulic

conductivity in 1- to 2-foot sections of the aquifer.

Statistical comparisons can determine whether removals of contaminants of concern in the living treatments

are significantly different from zero or significantly different from any sorption that is occurring. Comparisons are

made on the first-order rate of removal, that is, the slope

of a linear regression of the natural logarithm of the

concentration remaining against time of incubation for

both the living and control microcosm. These slopes

(removal rates) are compared to determine whether

they are different and, if so, the extent of the difference

that can be detected at a given level of confidence

The Tibbetts Road Case Study

The Tibbetts Road Superfund site in Barrington, New

Hampshire, a former private home, was used to store

drums of various chemicals from 1944 to 1984. The

primary ground-water contaminants in the overburden

and bedrock aquifers were TCE and benzene, with respective concentrations of 7,800 μg/L and 1,100 μg/L.

High concentrations of arsenic, chromium, nickel, and

lead were also found.

Material collected at the site was used to construct a

microcosm study evaluating the removal of benzene,

toluene, and TCE. This material was acquired from the

waste pile near the origin of Segment A (Figure 1), the

most contaminated source at the site. Microcosms were

incubated for 9 months. The aquifer material was added

to 20-milliliter headspace vials; dosed with 1 milliliter of

spiking solution; capped with a Teflon-lined, gray butyl

rubber septa; and sealed with an aluminum crimp cap.

Controls were prepared by autoclaving the material

used to construct the microcosms overnight. Initial concentrations for benzene, toluene, and TCE were 380

μg/L, 450 μg/L, and 330 μg/L, respectively. The microcosms were thoroughly mixed by vortexing, then stored

inverted in the dark at the ambient temperature of 10°C.

The results (Figures 2 through 4 and Table 3) show that

significant biodegradation of both petroleum aromatic

hydrocarbons and the chlorinated solvent had occurred.

Significant removal in the control microcosms also occurred

for all compounds. The data exhibited more variability

Figure 1. TCE concentrations in the Tibbetts Road microcosm

study.

Figure 2. Benzene concentrations in the Tibbetts Road microcosm study.

Figure 3. Toluene concentrations in the Tibbetts Road microcosm study.

Figure 4. Location of waste piles and flow path segments at

the Tibbetts Road Superfund site.

27
in the living microcosms than in the control treatment, a

pattern that has been observed in other microcosm

studies. The removals observed in the controls are probably due to sorption; however, this study exhibited more

sorption than typically seen.

The rate constants determined from the microcosm

study for the three compounds are shown in Table 4. The

appropriate rate constant to be used in a model or a risk

assessment would be the first-order removal in the living

treatment minus the first-order removal in the control, in

other words, the removal that is in excess of the removal

in the controls.

The first-order removal in the living and control microcosms was estimated as the linear regression of the

natural logarithm of concentration remaining in each

microcosm in each treatment against time of incubation.

Student’s t distribution with n - 2 degrees of freedom was

used to estimate the 95 percent confidence interval. The

standard error of the difference of the rates of removal

in living and control microcosms was estimated as the

square root of the sum of the squares of the standard

errors of the living and control microcosms, with n - 4

degrees of freedom (16).

Table 5 presents the concentrations of organic compounds and their metabolic products in monitoring wells

used to define line segments in the aquifer for estimation

of field-scale rate constants. Wells in this aquifer

showed little accumulation of trans-DCE, 1,1-DCE, vinyl

chloride, or ethene, although removals of TCE and cisDCE were extensive. This can be explained by the

observation that iron-reducing bacteria can rapidly oxidize vinyl chloride to carbon dioxide (4). Filterable iron

accumulated in ground water in this aquifer.

The extent of attenuation from well to well (Table 5) and

the travel time between wells in a segment (Figure 4)

were used to calculate first-order rate constants for each

segment (Table 6). Travel time between monitoring wells

was calculated from site-specific estimates of hydraulic

conductivity and from the hydraulic gradient. In the area

sampled for the microcosm study, the estimated Darcy

Table 3. Concentrations of TCE, Benzene, and Toluene in the Tibbetts Road Microcosms

Compound

Time Zero

Microcosms

Time Zero

Controls

Week 23

Microcosms

Week 23

Controls

Week 42

Microcosms

Week 42

Controls

TCE 328 337 1 180 2 36.3

261 394 12.5 116 2 54.5

309 367 8.46 99.9 2 42.3

Mean ± standard deviation 299 ± 34.5 366 ± 28.5 7.32 ± 5.83 132 ± 42.4 2.0 ± 0.0 44.4 ± 9.27

Benzene 366 396 201 236 11.1 146

280 462 276 180 20.5 105

340 433 22.8 152 11.6 139

Mean ± standard deviation 329 ± 44.1 430 ± 33.1 167 ± 130 189 ± 42.8 14.4 ± 5.29 130 ± 21.9

Toluene 443 460 228 254 2 136

342 557 304 185 2.5 92

411 502 19.9 157 16.6 115

Mean ± standard deviation 399 ± 51.6 506 ± 48.6 184 ± 147 199 ± 49.9 7.03 ± 8.29 114 ± 22.0 

Table 4. First-Order Rate Constants for Removal of TCE, Benzene, and Toluene in the Tibbetts Road Microcosms

Parameter

Living

Microcosms

Autoclaved

Controls

Removal Above

Controls

First-Order Rate of Removal (per year)

TCE 6.31 2.62 3.69

95% confidence interval ± 2.50 ± 0.50 ± 2.31

Minimum rate significant at 95% confidence 1.38

Benzene 3.87 1.51 2.36

95% confidence interval ± 1.96 ± 0.44 ± 1.83

Minimum rate significant at 95% confidence 0.53

Toluene 5.49 1.86 3.63

95% confidence interval ± 2.87 ± 0.45 ± 2.64

Minimum rate significant at 95% confidence 0.99

28
flow was 2.0 feet per year. With an estimated porosity in

this particular glacial till of 0.1, this corresponds to a

plume velocity of 20 feet per year.

Summary

Table 7 compares the first-order rate constants estimated

from the microcosm studies with the rate constants estimated at field scale. The agreement between the independent estimates of rate is good, indicating that the rates

can appropriately be used in a risk assessment. The rates

of biodegradation documented in the microcosm study

could easily account for the disappearance of TCE,

trichloroethylene, benzene, and toluene observed at field

scale. The rates estimated from the microcosm study are

several-fold higher than the rates estimated at field scale,

which may reflect an underestimation of the true rate in

the field. The estimates of plume velocity assumed that

the aquifer was homogeneous. No attempt was made in

this study to correct the estimate of plume velocity for

Table 5. Concentration of Contaminants and Metabolic Byproducts in Monitoring Wells Along Segments in the Plume Used To

Estimate Field-Scale Rate Constants 

Parameter Segment A Segment B Segment C

Monitoring well 80S 79S 70S 52S 70S 53S

Upgradient g/L Downgradient g/L Upgradient g/L Downgradient g/L Upgradient g/L Downgradient g/L

TCE 200 13.7 710 67 710 3.1

cis-DCE 740 10.9 220 270 220 2.9

trans-DCE 0.41 < 1 0.8 0.3 0.8 < 1

1, 1-DCE 0.99 < 1 < 1 1.6 < 1 < 1

Vinyl Chloride < 1 < 1 < 1 < 1 < 1 < 1

Ethene < 4 < 4 7 < 4 7 < 4

Benzene 510 2.5 493 420 493 < 1

Toluene 10,000 < 1 3,850 900 3,850 < 1

o-Xylene 1,400 8.4 240 71 240 < 1

m-Xylene 2,500 < 1 360 59 360 < 1

p-Xylene 1,400 22 1,100 320 1,100 < 1

Ethylbenzene 1,300 0.7 760 310 760 < 1

Methane 353 77 8 3 8 < 2

Iron 27,000 

Table 7. Comparison of First-Order Rate Constants in a Microcosm Study and in the Field at the Tibbetts Road NPL Site

Parameter

Microcosms Corrected for Controls Field Scale

Average Rate

Minimum Rate Significant

at 95% Confidence

Segment 

A

Segment 

B

Segment 

C

First-Order Rate (per year)

Trichloroethylene 3.69 1.38 0.41 0.59 0.54

Benzene 2.36 0.53 0.82 0.04 > 0.62 

Toluene 3.63 0.99 > 1.42 0.36 > 0.83 

Table 6. First-Order Rate Constants in Segments of the

Tibbetts Road Plume

Flow Path Segments in Length and Time of

Ground-Water Travel

Compound

Segment A

130 feet =

6.4 years

Segment B

80 feet =

2.4 years

Segment C

200 feet =

10 years

First-Order Rate Constants in Segments (per year)

TCE 0.41 0.59 0.54

cis-DCE 0.65 Produced 0.43

Benzene 0.82 0.04 > 0.62

Toluene > 1.42 0.36 > 0.83

o-Xylene 0.79 0.30 > 0.55

m-Xylene > 1.20 0.45 > 0.59

p-Xylene 0.64 0.31 > 0.70

Ethylbenzene 1.16 0.22 > 0.66

29
the influence of preferential flow paths. Preferential flow

paths with a higher hydraulic conductivity than average

would result in a faster velocity of the plume, thus a lower

residence time and faster rate of removal at field scale.

References

 1. U.S. EPA. 1995. EPA project summary. EPA/600/SV-95/001. U.S.

EPA. Washington, DC.

 2. Wilson, J.T., D. Kampbell, J. Weaver, B. Wilson, T. Imbrigiotta,

and T. Ehlke. 1995. A review of intrinsic bioremediation of trichloroethylene in ground water at Picatinny Arsenal, New Jersey, and

St. Joseph, Michigan. In: U.S. EPA. Symposium on Bioremediation of Hazardous Wastes: Research, Development, and Field

Evaluations, Rye Brook, N.Y. EPA/600/R-95/076.

 3. Chapelle, F.H. 1996. Identifying redox conditions that favor the

natural attenuation of chlorinated ethenes in contaminated

ground-water systems. In: Proceedings of the Symposium on

Natural Attenuation of Chlorinated Organics in Ground Water,

September 11-13, Dallas, TX.

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

of vinyl chloride in Fe(III)-reducing aquifer sediments. Environ.

Sci. Technol. In press.

 5. Bouwer, E.J. 1994. Bioremediation of chlorinated solvents using

alternate electron acceptors. In: Handbook of bioremediation.

Boca Raton, FL: Lewis Publishers.

 6. Sewell, G.W., and S.A. Gibson. 1991. Stimulation of the reductive

dechlorination of tetrachloroethylene in anaerobic aquifer microcosms by the addition of toluene. Environ. Sci. Technol.

25(5):982-984.

 7. Klecka, G.M., J.T. Wilson, E. Lutz, N. Klier, R. West, J. Davis, J.

Weaver, D. Kampbell, and B. Wilson. 1996. Intrinsic remediation

of chlorinated solvents in ground water. In: Proceedings of the

IBC/CELTIC Conference on Intrinsic Bioremediation, March 18-

19, London, UK.

 8. Cozzarelli, I.M., J.S. Herman, and M.J. Baedecker. 1995. Fate of

microbial metabolites of hydrocarbons in a coastal plain aquifer:

The role of electron acceptors. Environ. Sci. Technol. 29(2):458-469.

 9. Chapelle, F.H. Ground-water microbiology and geochemistry.

New York, NY: John Wiley & Sons.

10. Baedecker, M.J., and W. Back 1979. Hydrogeological processes

and chemical reactions at a landfill. Ground Water 17(5):429-437.

11. Stumm, W., and J.J. Morgan. 1970. Aquatic chemistry. New York,

NY: Wiley Interscience.

12. Snoeyink, V.L., and D. Jenkins. 1980. Water chemistry. New York,

NY: John Wiley & Sons.

13. Chapelle, F.H., P.B. McMahon, N.M. Dubrovsky, R.F. Fugii, E.T.

Oaksford, and D.A. Vroblesky. 1995. Deducing the distribution of

terminal electron-accepting processes in hydrologically diverse

groundwater systems. Water Resour. Res. 31:359-371.

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

dissolved H2 concentrations to determine distribution of microbially catalyzed redox reactions in anoxic groundwater. Environ.

Sci. Technol. 28:1255-1210.

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

as an indicator of the predominant terminal electron-accepting

reactions in aquatic sediments. Geochim. Cosmochim. Acta

52:2993-3003.

16. Glantz, S.A. 1992. Primer of biostatistics. New York, NY:

McGraw-Hill.

30
Conceptual Models for Chlorinated Solvent Plumes and Their Relevance to

Intrinsic Remediation

John A. Cherry

University of Waterloo, Department of Earth Sciences, Waterloo, Ontario

Introduction

Plumes in which chlorinated solvents are the primary

contaminants of concern are common in aquifers in

North America and Europe. Most of these plumes have

existed for two decades or longer, but only a few were

delineated before the mid-1980s. In general, solvent

plumes are deeper and more extensive than other types

of plumes. Many unremediated solvent plumes have

shown little change in peak concentrations or shape

since monitoring began more than a decade ago.

Plumes subjected to pump-and-treat often have shown

an initial decline in solvent concentrations but thereafter

have nearly constant concentrations in the source

zones. Permanent restoration of these ground-water

systems has not yet been accomplished at significant

solvent contamination sites.

Conceptual Models for Dense

Nonaqueous-Phase Liquid Sites

The conceptual models that best explain chlorinated

solvent plumes have considerable immobile immisciblephase solvent mass (dense nonaqueous-phase liquid

[DNAPL]) situated below the water table that continually

contributes dissolved solvents to the plume. Within the

subsurface zone causing plume development in fractured porous media, the original DNAPL mass may have

undergone phase transfer so that the mass now resides

totally or partly as dissolved and sorbed mass in the

low-permeability matrix blocks between fractures. The

subsurface zone of plume origin is referred to as the subsurface source zone or simply the source zone, whether it

has DNAPL residual or free product or has phased-transferred DNAPL in low-permeability zones. Significant solvent mass may also reside above the water table, but

this mass is typically not a major contributor to the

ground-water plume relative to the deeper solvent mass.

Although many indirect lines of evidence indicate that

the solvent mass in the subsurface source zone is the

long-term cause of the plumes, reliable estimates of the

mass in this zone are very rare. The monitoring data

necessary for such estimates are usually not achievable

because of the excessive time and cost involved. At

nearly all solvent contamination sites, disposals, leakages, or spills have ceased; therefore, the solvent mass

in the source zone is now slowly diminishing and eventually the source zone will be depleted. This depletion,

however, is expected to take many decades or even

centuries.

Many solvent plumes have traveled sufficiently far to

encounter natural hydrologic boundaries such as

streams, lakes, or wetlands or induced boundaries such

as water wells. The fronts of some solvent plumes have

not yet encountered boundaries, and questions arise as

to how much farther these fronts will travel while maintaining hazardous concentration levels. These questions are linked to possibilities for the plume front to

achieve an effective steady-state position. If the frontal

zone of a plume achieves this steady-state or near

steady-state position, then in the context of downgradient receptors the plume can be viewed as having

achieved intrinsic remediation. It is unlikely that depletion of the source zones contributes to intrinsic remediation of chlorinated solvent plumes; therefore, intrinsic

remediation must depend on attenuation processes operating within the plume.

Intrinsic Remediation

Intrinsic remediation occurrences are well known at petroleum contamination sites, but little is known about the

actual applicability of this concept to chlorinated solvent

plumes. Use of the term “intrinsic remediation” implies

nothing about the specific subsurface processes that

cause the remediation other than that the various processes somehow combine to cause the plume front to

achieve steady state or near steady state or perhaps

cause plume shrinkage.

31
The three processes that can drive the plume front

towards the steady-state condition are mechanical dispersion, molecular diffusion, and degradation. Sorption

can contribute to the appearance of a quasi-steady state

for some interval of time, but it does not cause permanent mass removal from or dilution of the plume. In the

context of chlorinated solvent plumes, biotic or abiotic

processes commonly cause transformations of the parent contaminant, such as trichloroethene (TCE) or

tetrachloroethene (PCE), to hazardous transformation

products such as trans- or cis-dichloroethene (DCE) and

vinyl chloride. Unfortunately, this often renders the

plume more hazardous. The term “intrinsic remediation”

for chlorinated solvent sites should be reserved for

plumes in which the degree of hazard has diminished

sufficiently within the plume front to achieve a drinkingwater standard or some other state of acceptably low risk.

Thus, intrinsic remediation requires the degradation to

be sufficiently complete to attenuate the hazard of the

plume front or the dispersion to cause sufficient dilution

to reduce concentrations to acceptable levels.

It is not feasible using laboratory studies to draw conclusions on the propensity for chlorinated solvent plumes

to achieve intrinsic remediation. Conclusions about propensity must come from comprehensive observations of

the nature and fate of actual plumes. Whether or not a

set of observations can be regarded as comprehensive

depends on the conceptual model or models deemed to

be most applicable.

The Fringe and Core Hypothesis

This paper presents a conceptual model for the anatomy

of chlorinated solvent plumes. Emphasis is on plumes

in sandy or gravelly aquifers. Based primarily on field

observations, it argues the merits of a conceptual model

in which solvent plumes typically have two components:

a low-concentration fringe that surrounds a high-concentration core. Multiple cores can exist in some plumes

due to the complexity of the source zone. The fringe,

which has concentrations in the range of one to a thousand micrograms per liter, is commonly large relative to

the volume of the core, which commonly has concentrations between one and a few tens of milligrams per liter.

Although concentrations in most of the core are orders

of magnitude larger than those in most of the fringe, the

peak concentrations in the core are much less than

DNAPL solubility, except close to the source zone. To

achieve intrinsic remediation, plume concentrations in

the core must decline orders of magnitude to attain

maximum contaminant levels (MCLs) for drinking water.

Thus, the attenuation processes must act much more

strongly on the core than the fringe to reach MCLs. Such

strong attenuation is unlikely to occur in many plumes.

Delineation of chlorinated solvent plumes in the United

States began in the early to mid-1980s as a result of

Superfund and the Resource Conservation and Recovery Act. During the past 15 years, millions of conventional monitoring wells have been used at many

thousands of solvent sites in the United States and

several other countries. Solvent plumes present an exceptionally difficult monitoring challenge because the

spatial distribution of contamination is often complex

due to the variability of the subsurface source zones and

to geologic heterogeneity within the plumes. Conventional monitoring networks using monitoring wells usually

indicate the presence of the fringe, which is commonly taken

to represent the plume as a whole. Due to the sparseness of data points, conventional networks only rarely

establish the existence of cores, except perhaps close

to the source zones. Thus, such plumes with no observed cores are perceived to have relatively low concentrations and therefore small total contaminant mass.

Detailed monitoring of experimental solvent plumes produced at the Borden field site (an unconfined sand

aquifer located 60 kilometers northwest of Toronto, Canada [1]) using unconventional techniques, as well as

similar monitoring of several plumes at actual industrial

sites, provides exceptional spatial resolution of the distribution of contaminants and confirms the presence of

cores as well as fringes. Many if not most plumes in

which cores have not been identified based on conventional monitoring may actually have cores that have

gone undetected because of the sparseness of the

monitoring networks.

Conclusion

Information on the concentration distribution in solvent

plumes is limited, particularly at and near the plume

fronts. Conventional approaches to monitoring result in

data that are too sparse to identify cores. Cores extending far from the subsurface source zones are likely a

common feature of solvent plumes in sand or gravel

aquifers. Although thousands of solvent plumes have

been monitored for many years, the sparseness of data

severely limits possibilities for determining the number

of occurrences of intrinsic remediation. More detailed

data sets that can be obtained using new methods of

monitoring, primarily direct push methods for spatial

rather than temporal resolution, offer the best possibilities for examining the fringe-and-core conceptual model

and intrinsic remediation of solvent plumes.

Reference

1. Cherry, J.A., J.F. Barker, S. Feenstra, R.W. Gillham, D.M. Mackay,

and D.J.A. Smyth. The Borden site for groundwater contamination

experiments: 1978-1995. In: Kobus, H., B. Barczewski, and H.-P.

Koschitzky, eds. Groundwater and subsurface remediation: Research strategies for in-situ remediation. Berlin/New York: Springer-Verlag. pp.102-127.

32
Site Characterization Tools: Using a Borehole Flowmeter To Locate and

Characterize the Transmissive Zones of an Aquifer

Fred Molz

Clemson University, Environmental Systems Engineering Department,

Clemson, South Carolina

Gerald Boman

Auburn University, Civil Engineering Department, Auburn, Alabama

Introduction

A study in which both direct and indirect techniques for

developing hydraulic conductivity (K) logs of screened

wells and/or boreholes were examined concluded that

techniques relying on direct hydraulic measurements,

such as transient pressure changes or flow rates, offer

the most promising methodology for determining accurate logs of horizontal K versus elevation in aquifers (1).

The borehole flowmeter, which can be used to measure

the vertical flow distribution in pumped wells, offers one

of the most direct techniques available for measuring a

K log. These conclusions have been supported by more

recent studies (2-7).

Inadequate performance of many pump-and-treat systems has been attributed to improper design (8, 9). In

far too many cases, underestimation of aquifer heterogeneity plays a significant role in these design failures. Bioremediation design is also sensitive to

aquifer heterogeneity. For example, if rate constants

for attenuation of chlorinated contaminants are to be

used for exposure assessments, it is necessary to

estimate the residence time of the contaminant in the

aquifer as accurately as possible. Conventional estimates of plume velocity use the average hydraulic

conductivity as determined by an aquifer test. These

average hydraulic conductivities can underestimate

the local hydraulic conductivity of the geological interval carrying a plume of contamination by a factor of

ten or more. Proper characterization of aquifer hydraulic properties, especially the spatial variations, is

currently limited by the methods for measuring those

properties. The borehole flowmeter enables one to

determine two basic things: the natural (ambient) vertical flow that exists in most wells, and, through a

small pumping test, the flow distribution entering the well

from the surrounding formation. If certain conditions are

met, the distributions provide sufficient information to

determine the hydraulic conductivity of the aquifer zones

selected as measurement intervals (4, 10).

Interest in borehole flowmeters as a means to directly

measure the variation of hydraulic conductivity became

apparent in the 1980s through the publication of a number of papers (11-13). By the late 1980s, the electromagnetic (EM) flowmeter had been designed, developed,

and tested by the Tennessee Valley Authority. This

unique flowmeter has several practical advantages. This

paper presents the results of EM flowmeter studies and

explains the capabilities of the instrument. It is now

recognized that the application of such instruments to

the characterization of aquifer properties will greatly

enhance the understanding of heterogeneity and its effect on contaminant migration (3, 6).

Conducting a Flowmeter Test

A flowmeter test may be viewed as a natural generalization of a standard, fully penetrating pumping test. In

the latter application, only the steady pumping rate, QP,

is measured, whereas during a flowmeter test the vertical flow rate distribution within the borehole or well

screen, Q(z), is recorded as well as QP (Figure 1).

Shown in Figure 2 are the discharge rates that are

provided directly by the instrument. “Ambient flow” refers to the natural flow in a test well due to small

hydraulic head differences in the vertical direction that

are detectable in most aquifers. “Pumping induced

flow” represents the flow distribution in a test well

caused by a small pump, which is also illustrated in

Figure 1. The flow data that ultimately go into a hydraulic

33
conductivity computation are represented by the “net

pumping flow,” which is the difference between pumping

induced flow and ambient flow (4).

To convert flow distributions in a well into flow to or from

aquifer layers, the discharge data are differenced (i.e.,

value at lower elevation subtracted from value at upper

elevation) to produce the “differential ambient flow” and

the “differential net flow.” The result of doing this to the

hypothetical data in Figure 2 is shown in Figure 3. Once

the flow to the well has stabilized, the differential net flow

(DNF) is proportional to the horizontal hydraulic conductivity distribution, K(z). The process of converting a DNF

curve into K(z) involves only algebra and a small amount

of additional data (4).

Flowmeter technology is very cost effective. It may be

viewed as an extension of a standard pumping test,

since flow distribution in the pumping well is measured

in addition to pumping rate and drawdown, but the cost

is less. Only a few hundred gallons of water are produced per test versus thousands or tens of thousands

for conventional pumping tests. Thus, potential treatment and disposal costs are minimal. A typical flowmeter

test can be completed in an hour or two. Cost per data

point is less than standard pumping tests by a factor of

100 or more, and the quantity of hydraulic conductivity

information produced is increased dramatically.

Measured K(z) Distributions

A commercial version of the EM borehole flowmeter is now

available, and it has been applied recently at several sites,

including the Savannah River site, the Louisiana Army

Ammunition Plant, and George Air Force Base (AFB),

California. The Savannah River application was in a 14-

meter thick confined aquifer in an alluvial basin composed

of sand, silt, and clay strata of variable composition. The

K distribution obtained in well P26-M1 is shown in Figure

4. Heterogeneity is evident, with K varying by an order of

magnitude at various locations in the aquifer (3).

A particularly illuminating EM flowmeter application at

George AFB was reported by Wilson et al. (Figure 5) (6).

The concentration data were obtained from core samples, while the K data are based on EM flowmeter

measurements in Well MW-27, which was located nearby.

The transmissive layer identified by the flowmeter is

likely the main stratum where the benzene is migrating.

This inference is supported by additional flowmeter tests

in neighboring wells (6).

Figure 1. Apparatus and geometry associated with a borehole

flowmeter test. The flowmeter measures the vertical

discharge distribution in the well caused by the

pump. A hypothetical data set (Q(z) versus z) is

shown at the bottom of the figure.

Figure 2. The data actually recorded by a flowmeter are ambient flow and pumping-induced flow. In both cases,

positive values indicate upward flow. Net pumping

flow is pumping-induced flow minus ambient flow.

34
Conclusion

Flowmeter tests have now been conducted at sites in

many regions of the United States. Results document

that the EM flowmeter is capable of supplying a new

level of detail concerning K distributions in granular

aquifers (2-4, 6, 7, 11, 13) and flowpath delineation in

fractured-rock aquifers (5, 12, 14). The resulting information concerning hydraulic heterogeneity is unprecedented and promises to serve as valuable input to

monitoring well screen location and remediation design.

Because basic data input has been the “weak link”

in the chain of activities constituting subsurface remediation, the potential impact of flowmeters on site

Figure 3. Plots of the differences of neighboring values of ambient flow (differential ambient flow) and net flow (differential net flow). These values represent the flow

entering (positive) or leaving (negative) the well

from/to the various layers of the aquifer.

Figure 4. Hydraulic conductivity as a function of depth in Well

P26-M1 at the Savannah River site. The measurement

interval (layer thickness) was 1 foot.

Figure 5. Benzene concentration and hydraulic conductivity as

functions of elevation at George Air Force Base. The

benzene appears to be migrating in the high transmissivity layer defined by a flowmeter analysis (6).

35
characterization and modeling is dramatic. Simultaneously, the effort required to perform flowmeter tests is

practical and economical.

While we view the technology represented by the EM

flowmeter as a definite step forward, the instrument in

its present prototype form is rather awkward to use on

a routine basis (3). The flowmeter probe hangs from stiff

electrical (not logging) cable and requires a packer inflation gas line to be attached. One must raise and lower

the instrument by hand, usually using the cable, the

inflation line, and a measuring tape bound together with

ties. The cable is difficult to clean, and stretching leads

to depth placement errors with increasing cable length.

These shortcomings may be removed by a redesign

effort that we are attempting to initiate. None of the

existing shortcomings, however, prevent effective use of

the EM borehole flowmeter, and the resulting data provide hydraulic conductivity information far superior to

that derived from standard pumping tests.

References

 1. Taylor, K., S.W. Wheatcraft, J. Hess, J.S. Hayworth, and F.J.

Molz. 1990. Evaluation of methods for determining the vertical

distribution of hydraulic conductivity. Ground Water 27: 88-98.

 2. Boggs, J.M., S.C. Young, L.M. Beard, L.W. Gelhar, K.R. Rehfeldt,

and E.E. Adams. 1992. Field study of dispersion in a heterogeneous aquifer, 1. Overview and site description. Water Resour.

Res. 28(I2):3281-3292.

 3. Boman, G.K., F.J. Molz, and K.D. Boone. 1996. Borehole flowmeter application in fluvial sediments: methodology, results and

assessment. Ground Water. Submitted.

 4. Molz, F.J., and S.C. Young. 1993. Development and application

of borehole flowmeters for environmental assessment. The Log

Analyst 3:13-23.

 5. Paillet, F.L., K. Novakowski, and P. Lapcevic. 1992. Analysis of

transient flows in boreholes during pumping in fractured formations. In: 33rd Annual Logging Symposium Transactions. Society

of Professional Well Log Analysts, S1-S22.

 6. Wilson, J.T., G. Sewell, D. Caron, G. Doyle, and R. Miller. 1995.

Intrinsic bioremediation of jet fuel contamination at George Air

Force Base. In: Hinchee, R.E., J.T. Wilson, and D.C. Downey,

eds. Intrinsic bioremediation. Richland, WA: Battelle Press. pp.

91-100.

 7. Young, S.C., and H.S. Pearson. 1995. The electromagnetic borehole flowmeter: Description and application. Ground Water Monitoring and Remediation XV(4):138-146.

 8. Haley, J.L., B. Hanson, C. Enfield, and J. Glass. 1991. Evaluating

the effectiveness of groundwater extraction systems. Ground

Water Monitoring and Remediation XI:119-124.

 9. U.S. EPA. 1990. Basics of pump-and-treat remediation technology. EPA/600/8-90/003. Report prepared by Geo Trans Inc.,

Herndon, VA.

10. U.S. EPA. 1990. A new approach and methodologies for characterizing the hydrogeologic properties of aquifers. EPA/600/2-

90/002 (NTIS90-167063). Ada, OK.

11. Molz, F.J., R.H. Morin, A.E. Hess, J.G. Melville, and O. Güven.

1989. The impeller meter for measuring aquifer permeability variations: evaluations and comparison with other tests. Water Resour. Res. 25:1677-1683.

12. Morin, R.H., A.E. Hess, and F.L. Paillet. 1988. Determining the

distribution of hydraulic conductivity in a fractured limestone aquifer by simultaneous injection and geophysical logging. Ground

Water 26:587-595.

13. Rehfeldt, K.R., P. Huschmeid, L.W. Gelhar, and M.E. Schaefer.

1989. The borehole flowmeter technique for measuring hydraulic

conductivity variability. Report EM-6511. Electric Power Research

Institute, Palo Alto, CA.

14. Hess, A.E., and F.L. Paillet. 1990. Applications of the thermalpulse flowmeter in the hydraulic characterization of fractured

rock. ASTM STP 1101. American Society for Testing and Materials, Philadelphia, PA. pp. 99-112.

36
Overview of the Technical Protocol for Natural Attenuation of Chlorinated

Aliphatic Hydrocarbons in Ground Water Under Development for the 

U.S. Air Force Center for Environmental Excellence

Todd H. Wiedemeier, Matthew A. Swanson, and David E. Moutoux

Parsons Engineering Science, Inc., Denver, Colorado

John T. Wilson and Donald H. Kampbell

U.S. Environmental Protection Agency, National Risk Management Research Laboratory, 

Ada, Oklahoma

Jerry E. Hansen and Patrick Haas

U.S. Air Force Center for Environmental Excellence, Technology Transfer Division, 

Brooks Air Force Base, Texas

Introduction

Over the past several years, natural attenuation has

become increasingly accepted as a remedial alternative

for organic compounds dissolved in ground water. The

U.S. Environmental Protection Agency’s (EPA) Office of

Research and Development and Office of Solid Waste and

Emergency Response define natural attenuation as:

The biodegradation, dispersion, dilution, sorption,

volatilization, and/or chemical and biochemical stabilization of contaminants to effectively reduce contaminant toxicity, mobility, or volume to levels that

are protective of human health and the ecosystem.

In practice, natural attenuation has several other names,

such as intrinsic remediation, intrinsic bioremediation, or

passive bioremediation. The goal of any site characterization effort is to understand the fate and transport

of the contaminants of concern over time in order to

assess any current or potential threat to human health

or the environment. Natural attenuation processes, such

as biodegradation, can often be dominant factors in the

fate and transport of contaminants. Thus, consideration

and quantification of natural attenuation is essential to

more thoroughly understand contaminant fate and

transport.

This paper presents a technical protocol for data collection and analysis in support of remediation by natural

attenuation to restore ground water contaminated with

chlorinated aliphatic hydrocarbons and ground water

contaminated with mixtures of fuels and chlorinated aliphatic hydrocarbons. In some cases, the information

collected using this protocol will show that natural attenuation processes, with or without source removal, will

reduce the concentrations of these contaminants to below risk-based corrective action criteria or regulatory

standards before potential receptor exposure pathways

are completed. The evaluation should include consideration of existing exposure pathways as well as exposure pathways arising from potential future use of the

ground water.

This protocol is intended to be used within the established regulatory framework. It is not the intent of this

document to replace existing EPA or state-specific guidance on conducting remedial investigations.

Overview of the Technical Protocol

Natural attenuation in ground-water systems results

from the integration of several subsurface attenuation

mechanisms that are classified as either destructive or

nondestructive. Biodegradation is the most important

destructive attenuation mechanism. Nondestructive attenuation mechanisms include sorption, dispersion, dilution from recharge, and volatilization. The natural

attenuation of fuel hydrocarbons is described in the

Technical Protocol for Implementing Intrinsic Remediation With Long-Term Monitoring for Natural Attenuation

of Fuel Contamination Dissolved in Groundwater, recently

published by the U.S. Air Force Center for Environmental

37
Excellence (AFCEE) (1). This document differs from the

technical protocol for intrinsic remediation of fuel hydrocarbons because the individual processes of chlorinated

aliphatic hydrocarbon biodegradation are fundamentally

different from the processes involved in the biodegradation of fuel hydrocarbons.

For example, biodegradation of fuel hydrocarbons, especially benzene, toluene, ethylbenzene, and xylenes

(BTEX), is mainly limited by electron acceptor availability, and biodegradation of these compounds generally

will proceed until all of the contaminants are destroyed.

In the experience of the authors, there appears to be an

inexhaustible supply of electron acceptors in most, if not

all, hydrogeologic environments. On the other hand, the

more highly chlorinated solvents (e.g., perchloroethene

and trichloroethene) typically are biodegraded under

natural conditions via reductive dechlorination, a process that requires both electron acceptors (the chlorinated aliphatic hydrocarbons) and an adequate supply of

electron donors. Electron donors include fuel hydrocarbons or other types of anthropogenic carbon (e.g., landfill leachate, BTEX, or natural organic carbon). If the

subsurface environment is depleted of electron donors

before the chlorinated aliphatic hydrocarbons are removed, reductive dechlorination will cease, and natural

attenuation may no longer be protective of human health

and the environment. This is the most significant difference between the processes of fuel hydrocarbon and

chlorinated aliphatic hydrocarbon biodegradation.

For this reason, it is more difficult to predict the long-term

behavior of chlorinated aliphatic hydrocarbon plumes

than fuel hydrocarbon plumes. Thus, it is important to

have a thorough understanding of the operant natural

attenuation mechanisms. In addition to having a better

understanding of the processes of advection, dispersion, dilution from recharge, and sorption, it is necessary

to better quantify biodegradation. This requires a thorough understanding of the interactions between chlorinated aliphatic hydrocarbons, anthropogenic/natural

carbon, and inorganic electron acceptors at the site.

Detailed site characterization is required to adequately

understand these processes.

Chlorinated solvents are released into the subsurface

under two possible scenarios: 1) as relatively pure solvent mixtures that are more dense than water, or 2) as

mixtures of fuel hydrocarbons and chlorinated aliphatic

hydrocarbons which, depending on the relative proportion of each, may be more or less dense than water.

These products commonly are referred to as

“nonaqueous-phase liquids,” or NAPLs. If the NAPL is

more dense than water, the material is referred to as a

“dense nonaqueous-phase liquid,” or DNAPL. If the

NAPL is less dense than water, the material is referred

to as a “light nonaqueous-phase liquid,” or LNAPL. In

general, the greatest mass of contaminant is associated

with these NAPL source areas, not with the aqueous

phase.

As ground water moves through or past the NAPL

source areas, soluble constituents partition into the

moving ground water to generate a plume of dissolved

contamination. After further releases have been

stopped, these NAPL source areas tend to slowly

weather away as the soluble components, such as

BTEX or trichloroethene, are depleted. In cases where

source removal or reduction is feasible, it is desirable to

remove product and decrease the time required for complete remediation of the site. At many sites, however,

mobile NAPL removal is not feasible with available technology. In fact, the quantity of NAPL recovered by commonly used recovery techniques is a trivial fraction of

the total NAPL available to contaminate ground water.

Mobile NAPL recovery typically recovers less than 10

percent of the total NAPL mass in a spill.

Compared with conventional engineered remediation

technologies, natural attenuation has the following

advantages:

• During natural attenuation, contaminants are ultimately

transformed to innocuous byproducts (e.g., carbon dioxide, ethene, and water), not just transferred to another phase or location in the environment.

• Natural attenuation is nonintrusive and allows continuing use of infrastructure during remediation.

• Engineered remedial technologies can pose greater

risk to potential receptors than natural attenuation

because contaminants may be transferred into the

atmosphere during remediation activities.

• Natural attenuation is less costly than currently available remedial technologies, such as pump-and-treat.

• Natural attenuation is not subject to the limitations of

mechanized remediation equipment (e.g., no equipment downtime).

• Those compounds that are the most mobile and toxic

are generally the most susceptible to biodegradation.

Natural attenuation has the following limitations:

• Natural attenuation is subject to natural and anthropogenic changes in local hydrogeologic conditions,

including changes in ground-water gradients and velocity, pH, electron acceptor concentrations, electron

donor concentrations, and/or potential future contaminant releases.

• Aquifer heterogeneity may complicate site characterization and quantification of natural attenuation.

• Time frames for complete remediation may be relatively long.

38
• Intermediate products of biodegradation (e.g., vinyl

chloride) can be more toxic than the original contaminant.

This document describes those processes that bring

about natural attenuation, the site characterization activities that may be performed to support a feasibility

study to include an evaluation of natural attenuation,

natural attenuation modeling using analytical or numerical solute fate-and-transport models, and the postmodeling activities that should be completed to ensure

successful support and verification of natural attenuation. The objective of the work described herein is to

quantify and provide defensible data in support of natural attenuation at sites where naturally occurring subsurface attenuation processes are capable of reducing

dissolved chlorinated aliphatic hydrocarbon and/or fuel

hydrocarbon concentrations to acceptable levels. A

comment made by a member of the regulatory community (2) summarizes what is required to successfully

implement natural attenuation:

A regulator looks for the data necessary to determine that a proposed treatment technology, if properly installed and operated, will reduce the

contaminant concentrations in the soil and water to

legally mandated limits. In this sense the use of

biological treatment systems calls for the same level

of investigation, demonstration of effectiveness, and

monitoring as any conventional [remediation] system.

To support remediation by natural attenuation, the proponent must scientifically demonstrate that degradation

of site contaminants is occurring at rates sufficient to be

protective of human health and the environment. Three

lines of evidence can be used to support natural attenuation of chlorinated aliphatic hydrocarbons, including:

• Observed reduction in contaminant concentrations

along the flow path downgradient from the source of

contamination.

• Documented loss of contaminant mass at the field

scale using:

– Chemical and geochemical analytical data (e.g.,

decreasing parent compound concentrations, increasing daughter compound concentrations, depletion of electron acceptors and donors, and

increasing metabolic byproduct concentrations).

– A conservative tracer and a rigorous estimate of

residence time along the flow path to document

contaminant mass reduction and to calculate biological decay rates at the field scale.

• Microbiological laboratory data that support the occurrence of biodegradation and give rates of biodegradation.

At a minimum, the investigator must obtain the first two

lines of evidence or the first and third lines of evidence.

The second and third lines of evidence are crucial to the

natural attenuation demonstration because they provide

biodegradation rate constants. These rate constants are

used in conjunction with the other fate-and-transport

parameters to predict contaminant concentrations and

to assess risk at downgradient points of compliance.

The first line of evidence is simply an observed reduction

in the concentration of released contaminants downgradient from the NAPL source area along the groundwater flow path. This line of evidence does not prove

that contaminants are being destroyed because the reduction in contaminant concentration could be the result

of advection, dispersion, dilution from recharge, sorption, and volatilization with no loss of contaminant mass

(i.e., the majority of apparent contaminant loss could be

due to dilution). Conversely, an increase in the concentrations of some contaminants, most notably degradation products such as vinyl chloride, could be indicative

of natural attenuation.

To support remediation by natural attenuation at most

sites, the investigator will have to show that contaminant

mass is being destroyed via biodegradation. This is

done using either or both of the second or third lines of

evidence. The second line of evidence relies on chemical and physical data to show that contaminant mass is

being destroyed via biodegradation, not just diluted. The

second line of evidence is divided into two components:

• Using chemical analytical data in mass balance calculations to show that decreases in contaminant and

electron acceptor and donor concentrations can be

directly correlated to increases in metabolic end

products and daughter compounds. This evidence

can be used to show that electron acceptor and donor concentrations in ground water are sufficient to

facilitate degradation of dissolved contaminants. Solute fate-and-transport models can be used to aid

mass balance calculations and to collate information

on degradation.

• Using measured concentrations of contaminants

and/or biologically recalcitrant tracers in conjunction

with aquifer hydrogeologic parameters, such as

seepage velocity and dilution, to show that a reduction in contaminant mass is occurring at the site and

to calculate biodegradation rate constants.

The third line of evidence, microbiological laboratory

data, can be used to provide additional evidence that

indigenous biota are capable of degrading site contaminants at a particular rate. Because it is necessary to

show that biodegradation is occurring and to obtain

biodegradation rate constants, the most useful type of

microbiological laboratory data is the microcosm study.

This paper presents a technical course of action that

allows converging lines of evidence to be used to scientifically document the occurrence and quantify the rates

of natural attenuation. Ideally, the first two lines of evidence

39
should be used in the natural attenuation demonstration.

To further document natural attenuation, or at sites with

complex hydrogeology, obtaining a field-scale biodegradation rate may not be possible; in this case, microbiological laboratory data can be used. Such a

“weight-of-evidence” approach will greatly increase the

likelihood of successfully implementing natural attenuation at sites where natural processes are restoring the

environmental quality of ground water.

Collection of an adequate database during the iterative

site characterization process is an important step in the

documentation of natural attenuation. Site characterization should provide data on the location, nature,

and extent of contaminant sources. Contaminant sources generally consist of hydrocarbons present as mobile

NAPL (i.e., NAPL occurring at sufficiently high saturations to drain under the influence of gravity into a well)

and residual NAPL (i.e., NAPL occurring at immobile,

residual saturation that is unable to drain into a well by

gravity). Site characterization also should provide information on the location, extent, and concentrations of

dissolved contamination; ground-water geochemical

data; geologic information on the type and distribution

of subsurface materials; and hydrogeologic parameters

such as hydraulic conductivity, hydraulic gradients, and

potential contaminant migration pathways to human or

ecological receptor exposure points.

The data collected during site characterization can be

used to simulate the fate and transport of contaminants

in the subsurface. Such simulation allows prediction of

the future extent and concentrations of the dissolved

contaminant plume. Several models can be used to

simulate dissolved contaminant transport and attenuation. The natural attenuation modeling effort has three

primary objectives: 1) to predict the future extent and

concentration of a dissolved contaminant plume by

simulating the combined effects of advection, dispersion, sorption, and biodegradation; 2) to assess the potential for downgradient receptors to be exposed to

contaminant concentrations that exceed regulatory or

risk-based levels intended to be protective of human

health and the environment; and 3) to provide technical

support for the natural attenuation remedial option at

postmodeling regulatory negotiations to help design a

more accurate verification and monitoring strategy and

to help identify early source removal strategies.

Upon completion of the fate-and-transport modeling effort, model predictions can be used in an exposure

pathways analysis. If natural attenuation is sufficient to

mitigate risks to potential receptors, the proponent of

natural attenuation has a reasonable basis for negotiating this option with regulators. The exposure pathways

analysis allows the proponent to show that potential

exposure pathways to receptors will not be completed.

The material presented herein was prepared through

the joint effort of the AFCEE Technology Transfer Division; the Bioremediation Research Team at EPA’s National Risk Management Research Laboratory in Ada,

Oklahoma (NRMRL), Subsurface Protection and Remediation Division; and Parsons Engineering Science, Inc.

(Parsons ES). This compilation is designed to facilitate

implementation of natural attenuation at chlorinated aliphatic hydrocarbon-contaminated sites owned by the

U.S. Air Force and other U.S. Department of Defense

agencies, the U.S. Department of Energy, and public

interests.

Overview of Chlorinated Aliphatic

Hydrocarbon Biodegradation

Because biodegradation is the most important process

acting to remove contaminants from ground water, an

accurate estimate of the potential for natural biodegradation is important to obtain when determining whether

ground-water contamination presents a substantial

threat to human health and the environment. This information also will be useful when selecting the remedial

alternative that will be most cost-effective in eliminating

or abating these threats should natural attenuation

alone not prove to be sufficient.

Over the past two decades, numerous laboratory and

field studies have demonstrated that subsurface microorganisms can degrade a variety of hydrocarbons and

chlorinated solvents (3-23). Whereas fuel hydrocarbons

are biodegraded through use as a primary substrate

(electron donor), chlorinated aliphatic hydrocarbons

may undergo biodegradation through three different

pathways: through use as an electron acceptor, through

use as an electron donor, or through co-metabolism,

where degradation of the chlorinated organic is fortuitous and there is no benefit to the microorganism. At a

given site, one or all of these processes may be operating, although at many sites the use of chlorinated aliphatic hydrocarbons as electron acceptors appears to

be most important under natural conditions. In general,

but in this case especially, biodegradation of chlorinated

aliphatic hydrocarbons will be an electron-donor-limited

process. Conversely, biodegradation of fuel hydrocarbons is an electron-acceptor-limited process.

In a pristine aquifer, native organic carbon is used as an

electron donor, and dissolved oxygen (DO) is used first

as the prime electron acceptor. Where anthropogenic

carbon (e.g., fuel hydrocarbon) is present, it also will be

used as an electron donor. After the DO is consumed,

anaerobic microorganisms typically use additional electron acceptors (as available) in the following order of

preference: nitrate, ferric iron oxyhydroxide, sulfate, and

finally carbon dioxide. Evaluation of the distribution of

these electron acceptors can provide evidence of where

and how chlorinated aliphatic hydrocarbon biodegradation

40
is occurring. In addition, because chlorinated aliphatic

hydrocarbons may be used as electron acceptors or

electron donors (in competition with other acceptors or

donors), isopleth maps showing the distribution of these

compounds can provide evidence of the mechanisms of

biodegradation working at a site. As with BTEX, the driving

force behind oxidation-reduction reactions resulting in

chlorinated aliphatic hydrocarbon degradation is electron transfer. Although thermodynamically favorable,

most of the reactions involved in chlorinated aliphatic

hydrocarbon reduction and oxidation do not proceed

abiotically. Microorganisms are capable of carrying out

the reactions, but they will facilitate only those oxidationreduction reactions that have a net yield of energy.

Mechanisms of Chlorinated Aliphatic

Hydrocarbon Biodegradation

Electron Acceptor Reactions (Reductive

Dechlorination)

The most important process for the natural biodegradation of the more highly chlorinated solvents is reductive

dechlorination. During this process, the chlorinated hydrocarbon is used as an electron acceptor, not as a

source of carbon, and a chlorine atom is removed and

replaced with a hydrogen atom. In general, reductive

dechlorination occurs by sequential dechlorination from

perchloroethene to trichloroethene to dichloroethene to

vinyl chloride to ethene. Depending on environmental

conditions, this sequence may be interrupted, with other

processes then acting on the products. During reductive

dechlorination, all three isomers of dichloroethene can

theoretically be produced; however, Bouwer (24) reports

that under the influence of biodegradation, cis-1,2-dichloroethene is a more common intermediate than

trans-1,2-dichloroethene, and that 1,1-dichloroethene is

the least prevalent intermediate of the three dichloroethene isomers. Reductive dechlorination of chlorinated solvent compounds is associated with all

accumulation of daughter products and an increase in

the concentration of chloride ions.

Reductive dechlorination affects each of the chlorinated

ethenes differently. Of these compounds, perchloroethene is the most susceptible to reductive dechlorination because it is the most oxidized. Conversely, vinyl

chloride is the least susceptible to reductive dechlorination because it is the least oxidized of these compounds.

The rate of reductive dechlorination also has been observed to decrease as the degree of chlorination decreases (24, 25). Murray and Richardson (26) have

postulated that this rate decrease may explain the accumulation of vinyl chloride in perchloroethene and

trichloroethene plumes that are undergoing reductive

dechlorination.

Reductive dechlorination has been demonstrated under

nitrate- and sulfate-reducing conditions, but the most

rapid biodegradation rates, affecting the widest range of

chlorinated aliphatic hydrocarbons, occur under methanogenic conditions (24). Because chlorinated aliphatic hydrocarbon compounds are used as electron acceptors

during reductive dechlorination, there must be an appropriate source of carbon in order for microbial growth to

occur (24). Potential carbon sources include natural

organic matter, fuel hydrocarbons, or other organic compounds such as those found in landfill leachate.

Electron Donor Reactions

Murray and Richardson (26) write that microorganisms

are generally believed to be incapable of growth using

trichloroethene and perchloroethene as a primary substrate (i.e., electron donor). Under aerobic and some

anaerobic conditions, the less-oxidized chlorinated aliphatic hydrocarbons (e.g., vinyl chloride) can be used as

the primary substrate in biologically mediated redox reactions (22). In this type of reaction, the facilitating microorganism obtains energy and organic carbon from the

degraded chlorinated aliphatic hydrocarbon. This is the

process by which fuel hydrocarbons are biodegraded.

In contrast to reactions in which the chlorinated aliphatic

hydrocarbon is used as an electron acceptor, only the

least oxidized chlorinated aliphatic hydrocarbons can be

used as electron donors in biologically mediated redox

reactions. McCarty and Semprini (22) describe investigations in which vinyl chloride and 1,2-dichloroethane

were shown to serve as primary substrates under aerobic conditions. These authors also document that dichloromethane has the potential to function as a primary

substrate under either aerobic or anaerobic environments. In addition, Bradley and Chapelle (27) show

evidence of mineralization of vinyl chloride under ironreducing conditions so long as there is sufficient

bioavailable iron(III). Aerobic metabolism of vinyl chloride may be characterized by a loss of vinyl chloride

mass and a decreasing molar ratio of vinyl chloride to

other chlorinated aliphatic hydrocarbon compounds.

Co-metabolism

When a chlorinated aliphatic hydrocarbon is biodegraded via co-metabolism, the degradation is catalyzed

by an enzyme or cofactor that is fortuitously produced

by the organisms for other purposes. The organism

receives no known benefit from the degradation of the

chlorinated aliphatic hydrocarbon; in fact, the co-metabolic

degradation of the chlorinated aliphatic hydrocarbon

may be harmful to the microorganism responsible for the

production of the enzyme or cofactor (22).

Co-metabolism is best documented in aerobic environments, although it could occur under anaerobic conditions. It has been reported that under aerobic conditions

41
chlorinated ethenes, with the exception of perchloroethene, are susceptible to co-metabolic degradation

(22, 23, 26). Vogel (23) further elaborates that the cometabolism rate increases as the degree of dechlorination decreases. During co-metabolism, trichloroethene

is indirectly transformed by bacteria as they use BTEX

or another substrate to meet their energy requirements.

Therefore, trichloroethene does not enhance the degradation of BTEX or other carbon sources, nor will its co-metabolism interfere with the use of electron acceptors

involved in the oxidation of those carbon sources.

Behavior of Chlorinated Solvent Plumes

Chlorinated solvent plumes can exhibit three types of

behavior depending on the amount of solvent, the

amount of biologically available organic carbon in the

aquifer, the distribution and concentration of natural

electron acceptors, and the types of electron acceptors

being used. Individual plumes may exhibit all three types

of behavior in different portions of the plume. The different types of plume behavior are summarized below.

Type 1 Behavior

Type 1 behavior occurs where the primary substrate is

anthropogenic carbon (e.g., BTEX or landfill leachate),

and this anthropogenic carbon drives reductive dechlorination. When evaluating natural attenuation of a plume

exhibiting Type 1, behavior the following questions must

be answered:

1. Is the electron donor supply adequate to allow

microbial reduction of the chlorinated organic

compounds? In other words, will the microorganisms

“strangle” before they “starve”—will they run out of

chlorinated aliphatic hydrocarbons (electron

acceptors) before they run out of electron donors?

2. What is the role of competing electron acceptors

(e.g., DO, nitrate, iron(III), and sulfate)?

3. Is vinyl chloride oxidized, or is it reduced?

Type 1 behavior results in the rapid and extensive degradation of the highly chlorinated solvents such as perchloroethene, trichloroethene, and dichloroethene.

Type 2 Behavior

Type 2 behavior dominates in areas that are characterized by relatively high concentrations of biologically

available native organic carbon. This natural carbon

source drives reductive dechlorination (i.e., is the primary substrate for microorganism growth). When evaluating natural attenuation of a Type 2 chlorinated solvent

plume, the same questions as those posed for Type 1

behavior must be answered. Type 2 behavior generally

results in slower biodegradation of the highly chlorinated solvents than Type 1 behavior, but under the right

conditions (e.g., areas with high natural organic carbon

contents) this type of behavior also can result in rapid

degradation of these compounds.

Type 3 Behavior

Type 3 behavior dominates in areas that are characterized by low concentrations of native and/or anthropogenic carbon and by DO concentrations greater than

1.0 milligrams per liter. Under these aerobic conditions,

reductive dechlorination will not occur; thus, there is no

removal of perchloroethene, trichloroethene, and dichloroethene. The most significant natural attenuation

mechanisms for these compounds is advection, dispersion, and sorption. However, vinyl chloride can be rapidly oxidized under these conditions.

Mixed Behavior

A single chlorinated solvent plume can exhibit all three

types of behavior in different portions of the plume. This

can be beneficial for natural biodegradation of chlorinated aliphatic hydrocarbon plumes. For example,

Wiedemeier et al. (28) describe a plume at Plattsburgh

Air Force Base, New York, that exhibits Type 1 behavior

in the source area and Type 3 behavior downgradient

from the source. The most fortuitous scenario involves

a plume in which perchloroethene, trichloroethene, and

dichloroethene are reductively dechlorinated (Type 1 or

2 behavior), then vinyl chloride is oxidized (Type 3 behavior) either aerobically or via iron reduction. Vinyl

chloride is oxidized to carbon dioxide in this type of

plume and does not accumulate. The following sequence of reactions occurs in a plume that exhibits this

type of mixed behavior:

Perchloroethene → Trichloroethene →

Dichloroethene → Vinyl chloride → Carbon dioxide

The trichloroethene, dichloroethene, and vinyl chloride

may attenuate at approximately the same rate, and thus

these reactions may be confused with simple dilution.

Note that no ethene is produced during this reaction.

Vinyl chloride is removed from the system much faster

under these conditions than it is under vinyl chloride-reducing conditions.

A less desirable scenario—but one in which all contaminants may be entirely biodegraded— involves a plume

in which all chlorinated aliphatic hydrocarbons are reductively dechlorinated via Type 1 or Type 2 behavior.

Vinyl chloride is reduced to ethene, which may be further

reduced to ethane or methane. The following sequence

of reactions occurs in this type of plume:

Perchloroethene → Trichloroethene →

Dichloroethene → Vinyl chloride → Ethene → Ethane

42
This sequence has been investigated by Freedman and

Gossett (13). In this type of plume, vinyl chloride degrades more slowly than trichloroethene and thus tends

to accumulate.

Protocol for Quantifying Natural

Attenuation During the Remedial

Investigation Process

The primary objective of the natural attenuation investigation is to show that natural processes of contaminant

degradation will reduce contaminant concentrations in

ground water to below risk-based corrective action or regulatory levels before potential receptor exposure pathways

are completed. This requires a projection of the potential

extent and concentration of the contaminant plume in time

and space. The projection should be based on historic

variations in, and the current extent and concentrations

of, the contaminant plume, as well as the measured

rates of contaminant attenuation. Because of the inherent uncertainty associated with such predictions, the

investigator must provide sufficient evidence to demonstrate that the mechanisms of natural attenuation will

reduce contaminant concentrations to acceptable levels

before potential receptors are reached. This requires the

use of conservative solute fate-and-transport model input parameters and numerous sensitivity analyses so

that consideration is given to all plausible contaminant

migration scenarios. When possible, both historical data

and modeling should be used to provide information that

collectively and consistently supports the natural reduction and removal of the dissolved contaminant plume.

Figure 1 outlines the steps involved in the natural attenuation demonstration. This figure also shows the

important regulatory decision points in the process of

implementing natural attenuation. Predicting the fate of

a contaminant plume requires the quantification of solute transport and transformation processes. Quantification of contaminant migration and attenuation rates and

successful implementation of the natural attenuation remedial option requires completion of the following steps:

1. Review available site data, and develop a preliminary

conceptual model.

2. Screen the site, and assess the potential for natural

attenuation.

3. Collect additional site characterization data to support

natural attenuation, as required.

4. Refine the conceptual model, complete premodeling

calculations, and document indicators of natural

attenuation.

5. Simulate natural attenuation using analytical or

numerical solute fate-and-transport models that allow

incorporation of a biodegradation term, as necessary.

6. Identify potential receptors, and conduct an

exposure-pathway analysis.

7. Evaluate the practicability and potential efficiency of

supplemental source removal options.

8. If natural attenuation with or without source removal

is acceptable, prepare a long-term monitoring plan.

9. Present findings to regulatory agencies, and obtain

approval for remediation by natural attenuation.

Review Available Site Data, and Develop a

Preliminary Conceptual Model

Existing site characterization data should be reviewed

and used to develop a conceptual model for the site. The

preliminary conceptual model will help identify any

shortcomings in the data and will allow placement of

additional data collection points in the most scientifically

advantageous and cost-effective manner. A conceptual

model is a three-dimensional representation of the

ground-water flow and solute transport system based on

available geological, biological, geochemical, hydrological, climatological, and analytical data for the site. This

type of conceptual model differs from the conceptual site

models that risk assessors commonly use that qualitatively consider the location of contaminant sources, release mechanisms, transport pathways, exposure

points, and receptors. The ground-water system conceptual model, however, facilitates identification of these

risk-assessment elements for the exposure pathways

analysis. After development, the conceptual model can

be used to help determine optimal placement of additional data collection points (as necessary) to aid in the

natural attenuation investigation and to develop the solute fate-and-transport model.

Contracting and management controls must be flexible

enough to allow for the potential for revisions to the

conceptual model and thus the data collection effort. In

cases where few or no site-specific data are available,

all future site characterization activities should be designed to collect the data necessary to screen the site

to determine the potential for remediation by natural

attenuation. The additional costs incurred by such data

collection are greatly outweighed by the cost savings

that will be realized if natural attenuation is selected.

Moreover, most of the data collected in support of natural attenuation can be used to design and support other

remedial measures.

Table 1 contains the soil and ground-water analytical

protocol for natural attenuation of chlorinated aliphatic

hydrocarbons and/or fuel hydrocarbons. Table 1A lists a

standard set of methods, while Table 1B lists methods

that are under development and/or consideration. Any

plan to collect additional ground-water and soil quality

data should include targeting the analytes listed in Table

1A, and possibly Table 1B.

43
Figure 1. Natural attenuation of chlorinated solvents flow chart.

44
Table 1A. Soil and Ground-Water Analytical Protocola

Matrix Analysis Method/Referenceb-e Commentsf,g Data Use

Recommended 

Frequency of

Analysis

Sample Volume,

Sample Container,

Sample Preservation

Field or

Fixed-Base

Laboratory

Soil Volatile

organic

compounds 

SW8260A Handbook

method

modified for

field extraction

of soil using

methanol

Useful for determining

the extent of soil

contamination, the

contaminant mass

present, and the need

for source removal

Each soil

sampling round

Collect 100 g of soil

in a glass container

with Teflon-lined cap;

cool to 4°C

Fixed-base

Soil Total

organic

carbon

(TOC)

SW9060, modified 

for soil samples

Procedure

must be

accurate over

the range of

0.5 to 15%

TOC

The amount of TOC

in the aquifer matrix

influences

contaminant migration

and biodegradation

At initial

sampling

Collect 100 g of soil

in a glass container

with Teflon-lined cap;

cool to 4°C

Fixed-base

Soil

gas

O2, CO2 Field soil gas 

analyzer

Useful for determining

bioactivity in the

vadose zone

At initial

sampling and

respiration

testing

Reuseable 3-L 

Tedlar bags

Field

Soil

gas

Fuel and

chlorinated

volatile

organic

compounds

EPA Method 

TO-14

Useful for determining

the distribution of

chlorinated and BTEX

compounds in soil

At initial

sampling

1-L Summa canister Fixed-base

Water Volatile

organic

compounds

SW8260A Handbook

method;

analysis may

be extended to

higher

molecularweight alkyl

benzenes

Method of analysis for

BTEX and chlorinated

solvents/byproducts

Each sampling

round

Collect water

samples in a 40-mL

volatile organic

analysis vial; cool to

4°C; add hydrochloric

acid to pH 2

Fixed-base

Water Polycyclic

aromatic

hydrocarbons

(PAHs)

(optional;

intended

for diesel

and other

heavy oils)

Gas chromatography/

mass spectroscopy

Method SW8270B;

high-performance

liquid chromatography

Method SW8310

Analysis

needed only

when required

for regulatory

compliance

PAHs are components

of fuel and are

typically analyzed for

regulatory compliance

As required by

regulations

Collect 1 L of water 

in a glass container;

cool to 4°C

Fixed-base

Water Oxygen DO meter Refer to

Method A4500

for a

comparable

laboratory

procedure

Concentrations less

than 1 mg/L generally

indicate an anaerobic

pathway

Each sampling

round

Measure DO on site

using a flow-through

cell

Field

Water Nitrate Iron chromatography

Method E300; anion

method

Method E300

is a handbook

method; also

provides

chloride data

Substrate for microbial

respiration if oxygen

is depleted

Each sampling

round

Collect up to 40 mL

of water in a glass or

plastic container; add

H2SO4 to pH less

than 2; cool to 4°C

Fixed-base

Water Iron(II)

(Fe+2)

Colorimetric HACH

Method 8146

Filter if turbid May indicate an

anaerobic degradation

process due to

depletion of oxygen,

nitrate, and

manganese

Each sampling

round

Collect 100 mL of

water in a glass

container

Field

45
Matrix Analysis Method/Referenceb-e Commentsf,g Data Use

Recommended 

Frequency of

Analysis

Sample Volume,

Sample Container,

Sample Preservation

Field or

Fixed-Base

Laboratory

Water Sulfate

(SO4-2)

Iron chromatography

Method E300 or

HACH Method 8051

Method E300

is a handbook

method, HACH

Method 8051

is a

colorimetric

method; use

one or the

other

Substrate for

anaerobic microbial

respiration

Each sampling

round

Collect up to 40 mL

of water in a glass or

plastic container; cool

to 4°C

E300 =

Fixed-base

HACH

Method

8051 = Field

Water Methane,

ethane,

and ethene

Kampbell et al. (35)

or SW3810, modified

Method

published by

EPA

researchers

The presence of CH4

suggests

biodegradation of

organic carbon via

methanogensis;

ethane and ethane

are produced during

reductive

dechlorination

Each sampling

round

Collect water

samples in 50 mL

glass serum bottles

with butyl

gray/Teflon-lined

caps; add H2SO4 to

pH less than 2; cool

to 4°C

Fixed-base

Water Alkalinity HACH alkalinity test

kit Model AL AP MG-L

Phenolphtalein

method

Water quality

parameter used to

measure the buffering

capacity of ground

water; can be used to

estimate the amount

of CO2 produced

during biodegradation

Each sampling

round

Collect 100 mL of

water in glass

container

Field

Water Oxidationreduction

potential

A2580B Measurements

made with

electrodes,

results are

displayed on a

meter, protect

samples from

exposure to

oxygen; report

results against

a silver/silver

chloride

reference

electrode

The oxidationreduction potential 

of ground water

influences and is

influenced by the

nature of the

biologically mediated

degradation of

contaminants; the

oxidation-reduction

potential of ground

water may range from

more than 800 mV to

less than -400 mV

Each sampling

round

Collect 100 to 

250 mL of water 

in a glass container

Field

Water pH Field probe with

direct reading meter

Field Aerobic and

anaerobic processes

are pH-sensitive

Each sampling

round

Collect 100 to 

250 mL of water 

in a glass or plastic

container; analyze

immediately

Field

Water Temperature Field probe with

direct reading meter

Field only Well development Each sampling

round

Not applicable Field

Water Conductivity E120.1/SW9050,

direct reading meter

Protocols/

Handbook

methods

Water quality

parameter used as a

marker to verify that

site samples are

obtained from the

same ground-water

system

Each sampling

round

Collect 100 to 250

mL of water in a

glass or plastic

container

Field

Water Chloride Mercuric nitrate

titration A4500-Cl- C

Ion

chromatography 

Method E300;

Method

SW9050 may

also be used

Final product of

chlorinated solvent

reduction; can be

used to estimate

dilution in calculation

of rate constant

Each sampling

round

Collect 250 mL of

water in a glass

container

Fixed-base

Table 1A. Soil and Ground-Water Analytical Protocola (Continued)

46
Matrix Analysis Method/Referenceb-e Commentsf,g Data Use

Recommended 

Frequency of

Analysis

Sample Volume,

Sample Container,

Sample Preservation

Field or

Fixed-Base

Laboratory

Water Chloride

(optional;

see data

use)

HACH chloride test

kit Model 8-P

Silver nitrate

titration

As above, and to

guide selection of

additional data points

in real time while in

the field

Each sampling

round

Collect 100 mL of

water in a glass

container

Field

Water Total

organic

carbon 

SW9060 Laboratory Used to classify

plumes and to

determine whether

anaerobic metabolism

of chlorinated solvents

is possible in the

absence of

anthropogenic carbon

Each sampling

round

Collect 100 mL of

water in a glass

container; cool

Laboratory

a

 Analyses other than those listed in this table may be required for regulatory compliance.

b

 “SW” refers to the Test Methods for Evaluating Solid Waste, Physical, and Chemical Methods (29). c

 “E” refers to Methods for Chemical Analysis of Water and Wastes (30). d

 “HACH” refers to the Hach Company catalog (31).

e “A” refers to Standard Methods for the Examination of Water and Wastewater (32). f

 “Handbook” refers to the AFCEE Handbook to Support the Installation Restoration Program (IRP) Remedial Investigations and Feasibility

Studies (RI/FS) (33).

g

 “Protocols” refers to the AFCEE Environmental Chemistry Function Installation Restoration Program Analytical Protocols (34).

Table 1A. Soil and Ground-Water Analytical Protocola (Continued)

Table 1B. Soil and Ground-Water Analytical Protocol: Special Analyses Under Development and/or Considerationa,b

Matrix Analysis Method/Reference Comments Data Use

Recommended

Frequency

of Analysis

Sample Volume,

Container,

Preservation

Field or

Fixed-Base

Laboratory

Soil Biologically

available iron(III)

Under development HCl

extraction

followed by

quantification

of released

iron(III)

To predict the

possible extent of

iron reduction in

an aquifer

One round of

sampling in

five borings,

five cores

from each

boring

Collect minimum

1-inch diameter

core samples into

a plastic liner; cap

and prevent

aeration

Laboratory

Water Nutritional

quality of native

organic matter

Under development Spectrophotometric

method

To determine the

extent of reductive

dechlorination

allowed by the

supply of electron

donor

One round of

sampling in

two to five

wells

Collect 1,000 mL

in an amber glass

container

Laboratory

Water Hydrogen (H2) Equilibration with

gas in the field;

determined with a

reducing gas

detector

Specialized

analysis

To determine the

terminal electron

accepting process;

predicts the

possibility for

reductive

dechlorination

One round of

sampling 

Sampling at well

head requires the

production of 100

mL per minute of

water for 30

minutes

Field

Water Oxygenates

(including

methyl-tert-butyl

ether, ethers,

acetic acid,

methanol, and

acetone)

SW8260/8015c Laboratory Contaminant or

electron donors

for dechlorination

of solvents

At least one

sampling

round or as

determined

by regulators

Collect 1 L of

water in a glass

container;

preserve with HCl

Laboratory

a

 Analyses other than those listed in this table may be required for regulatory compliance.

b

 Site characterization should not be delayed if these methods are unavailable.

c

 “SW” refers to Test Methods for Evaluating Solid Waste, Physical and Chemical Methods (29).

47
Screen the Site, and Assess the Potential for

Natural Attenuation

After reviewing available site data and developing a

preliminary conceptual model, an assessment of the

potential for natural attenuation must be made. As stated

previously, existing data can be useful in determining

whether natural attenuation will be sufficient to prevent

a dissolved contaminant plume from completing exposure pathways, or from reaching a predetermined point

of compliance, in concentrations above applicable regulatory or risk-based corrective action standards. Determining the likelihood of exposure pathway completion is

an important component of the natural attenuation investigation. This is achieved by estimating the migration

and future extent of the plume based on contaminant

properties, including volatility, sorptive properties, and

biodegradability; aquifer properties, including hydraulic

gradient, hydraulic conductivity, porosity, and total organic carbon (TOC) content; and the location of the

plume and contaminant source relative to potential receptors (i.e., the distance between the leading edge of

the plume and the potential receptor exposure points).

These parameters (estimated or actual) are used in this

section to make a preliminary assessment of the effectiveness of natural attenuation in reducing contaminant

concentrations.

If, after completing the steps outlined in this section, it

appears that natural attenuation will be a significant

factor in contaminant removal, detailed site characterization activities in support of this remedial option

should be performed. If exposure pathways have already been completed and contaminant concentrations

exceed regulatory levels, or if such completion is likely,

other remedial measures should be considered, possibly in conjunction with natural attenuation. Even so, the

collection of data in support of the natural attenuation

option can be integrated into a comprehensive remedial

plan and may help reduce the cost and duration of other

remedial measures, such as intensive source removal

operations or pump-and-treat technologies. For example, dissolved iron concentrations can have a profound

influence on the design of pump-and-treat systems.

Based on the experience of the authors, in an estimated

80 percent of fuel hydrocarbon spills at federal facilities,

natural attenuation alone will be protective of human

health and the environment. For spills of chlorinated

aliphatic hydrocarbons at federal facilities, however,

natural attenuation alone will be protective of human

health and the environment in an estimated 20 percent

of the cases. With this in mind, it is easy to understand

why an accurate assessment of the potential for natural

biodegradation of chlorinated compounds should be

made before investing in a detailed study of natural

attenuation. The screening process presented in this

section is outlined in Figure 2. This approach should

allow the investigator to determine whether natural attenuation is likely to be a viable remedial alternative before

additional time and money are expended. The data required to make the preliminary assessment of natural

attenuation can also be used to aid the design of an

engineered remedial solution, should the screening process suggest that natural attenuation alone is not feasible.

The following information is required for the screening

process:

• The chemical and geochemical data presented in

Table 2 for a minimum of six samples. Figure 3

shows the approximate location of these data collection points. If other contaminants are suspected, then

data on the concentration and distribution of these

compounds also should be obtained.

• Locations of source(s) and receptor(s).

• An estimate of the contaminant transport velocity and

direction of ground-water flow.

Once these data have been collected, the screening

process can be undertaken. The following steps summarize the screening process:

1. Determine whether biodegradation is occurring using

geochemical data. If biodegradation is occurring,

proceed to Step 2. If it is not, assess the amount and

types of data available. If data are insufficient to

determine whether biodegradation is occurring,

collect supplemental data.

2. Determine ground-water flow and solute transport

parameters. Hydraulic conductivity and porosity may

be estimated, but the ground-water gradient and flow

direction may not. The investigator should use the

highest hydraulic conductivity measured at the site

during the preliminary screening because solute

plumes tend to follow the path of least resistance

(i.e., highest hydraulic conductivity). This will give the

“worst case” estimate of solute migration over a

given period.

3. Locate sources and receptor exposure points.

4. Estimate the biodegradation rate constant. Biodegradation rate constants can be estimated using

a conservative tracer found commingled with the

contaminant plume, as described by Wiedemeier et

al. (36). When dealing with a plume that contains

only chlorinated solvents, this procedure will have to

be modified to use chloride as a tracer. Rate

constants derived from microcosm studies can also

be used. If it is not possible to estimate the

biodegradation rate using these procedures, then

use a range of accepted literature values for

biodegradation of the contaminants of concern.

48
Figure 2. Initial screening process flow chart.

49
Table 2. Analytical Parameters and Weighting for Preliminary Screening

Analyte

Concentration in Most

Contaminated Zone Interpretation

Points

Awarded

Oxygena < 0.5 mg/L Tolerated; suppresses reductive dechlorination at higher

concentrations

3

Oxygena > 1 mg/L Vinyl chloride may be oxidized aerobically, but reductive

dechlorination will not occur

-3

Nitratea < 1 mg/L May compete with reductive pathway at higher

concentrations

2

Iron (II)a > 1 mg/L Reductive pathway possible 3

Sulfatea < 20 mg/L May compete with reductive pathway at higher

concentrations

2

Sulfidea > 1 mg/L Reductive pathway possible 3

Methanea > 0.1 mg/L Ultimate reductive daughter product 2

> 1 Vinyl chloride accumulates 3

< 1 Vinyl chloride oxidizes

Oxidation reduction

potentiala < 50 mV against Ag/AgCl Reductive pathway possible < 50 mV = 1

< -100 mV = 2 

pHa 5 < pH < 9 Tolerated range for reductive pathway

DOC > 20 mg/L Carbon and energy source; drives dechlorination; can be

natural or anthropogenic

2

Temperaturea > 20°C At T > 20ÉC, biochemical process is accelerated 1

Carbon dioxide > 2x background Ultimate oxidative daughter product 1

Alkalinity > 2x background Results from interaction of carbon dioxide with aquifer

minerals

1

Chloridea > 2x background Daughter product of organic chlorine; compare chloride

in plume to background conditions

2

Hydrogen > 1 nM Reductive pathway possible; vinyl chloride may

accumulate

3

Hydrogen < 1 nM Vinyl chloride oxidized

Volatile fatty acids > 0.1 mg/L Intermediates resulting from biodegradation of aromatic

compounds; carbon and energy source

2

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

Perchloroethenea Material released

Trichloroethenea Material released or daughter product of perchloroethene 2b

Dichloroethenea Material released or daughter product of trichloroethene;

if amount of cis-1,2-dichloroethene is greater than 80%

of total dichloroethene, it is likely a daughter product of

trichloroethene

2b

Vinyl chloridea Material released or daughter product of dichloroethenes 2b

Ethene/Ethane < 0.1 mg/L Daughter product of vinyl chloride/ethene > 0.01 mg/L= 2

> 0.1 = 3

Chloroethanea Daughter product of vinyl chloride under reducing

conditions

2

1,1,1-Trichloroethanea Material released

1,1-dichloroethenea Daughter product of trichloroethene or chemical reaction

of 1,1,1-trichloroethane

a

 Required analysis.

b

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

50
5. Compare the rate of transport to the rate of attenuation,

using analytical solutions or a screening model such

as BIOSCREEN.

6. Determine whether the screening criteria are met.

Each of these steps is described in detail below.

Step 1: Determine Whether Biodegradation Is

Occurring

The first step in the screening process is to sample at

least six wells that are representative of the contaminant

flow system and to analyze the samples for the parameters listed in Table 2. Samples should be taken 1) from

the most contaminated portion of the aquifer (generally

in the area where NAPL currently is present or was

present in the past); 2) downgradient from the NAPL

source area but still in the dissolved contaminant plume;

3) downgradient from the dissolved contaminant plume;

and 4) from upgradient and lateral locations that are not

affected by the plume.

Samples collected in the NAPL source area allow determination of the dominant terminal electron-accepting

processes at the site. In conjunction with samples collected in the NAPL source zone, samples collected in

the dissolved plume downgradient from the NAPL

source zone allow the investigator to determine whether

the plume is degrading with distance along the flow path

and what the distribution of electron acceptors and donors and metabolic byproducts might be along the flow

path. The sample collected downgradient from the dissolved plume aids in plume delineation and allows the

investigator to determine whether metabolic byproducts

are present in an area of ground water that has been

remediated. The upgradient and lateral samples allow

delineation of the plume and indicate background concentrations of the electron acceptors and donors.

After these samples have been analyzed for the parameters listed in Table 2, the investigator should analyze the data to determine whether biodegradation is

occurring. The right-hand column of Table 2 contains

scoring values that can be used for this task. For example, if the DO concentration in the area of the plume with

the highest contaminant concentration is less than 0.5

milligrams per liter, this parameter is awarded 3 points.

Table 3 summarizes the range of possible scores and

gives an interpretation for each score. If the site scores

a total of 15 or more points, biodegradation is probably

occurring, and the investigator can proceed to Step 2.

This method relies on the fact that biodegradation will

cause predictable changes in ground-water chemistry.

Consider the following two examples. Example 1 contains data for a site with strong evidence that reductive

dechlorination is occurring. Example 2 contains data for

a site with strong evidence that reductive dechlorination

is not occurring.

Figure 3. Data collection points required for screening.

Table 3. Interpretation of Points Awarded During Screening Step 1

Score Interpretation

0 to 5 Inadequate evidence for biodegradation

of chlorinated organics

6 to 14 Limited evidence for biodegradation of

chlorinated organics

15 to 20 Adequate evidence for biodegradation of

chlorinated organics

> 20 Strong evidence for biodegradation of

chlorinated organics

Example 1. Strong Evidence for Biodegradation of

Chlorinated Organics

Analyte

Concentration in Most 

Contaminated Zone

Points

Awarded

DO 0.1 mg/L 3

Nitrate 0.3 mg/L 2

Iron(II) 10 mg/L 3

Sulfate 2 mg/L 2

Methane 5 mg/L 3

Oxidation-reduction

potential

-190 mV 2

Chloride 3x background 2

Perchloroethene

(released)

1,000 μg/L 0

Trichloroethene 

(none released)

1,200 μg/L 2

cis-1,2-Dichloroethene

(none released)

500 μg/L 2

Vinyl chloride 

(none released)

50 μg/L 2

Total points awarded 23 

51
In this example, the investigator can infer that biodegradation is occurring and may proceed to Step 2.

In this example, the investigator can infer that biodegradation is probably not occurring or is occurring too slowly

to be a viable remedial option. In this case, the investigator cannot proceed to Step 2 and will likely have to

implement an engineered remediation system.

Step 2: Determine Ground-Water Flow and Solute

Transport Parameters

After biodegradation has been shown to be occurring, it

is important to quantify ground-water flow and solute

transport parameters. This will make it possible to use

a solute transport model to quantitatively estimate the

concentration of the plume and its direction and rate of

travel. To use an analytical model, it is necessary to

know the hydraulic gradient and hydraulic conductivity

for the site and to have estimates of the porosity and

dispersivity. The coefficient of retardation also is helpful

to know. Quantification of these parameters is discussed

by Wiedemeier et al. (1).

To make modeling as accurate as possible, the investigator must have site-specific hydraulic gradient and hydraulic conductivity data. To determine the ground-water

flow and solute transport direction, the site must have at

least three accurately surveyed wells. The porosity and

dispersivity are generally estimated using accepted literature values for the types of sediments found at the

site. If the investigator does not have TOC data for soil,

the coefficient of retardation can be estimated; however,

assuming that the solute transport and ground-water

velocities are the same may be more conservative.

Step 3: Locate Sources and Receptor Exposure

Points

To determine the length of flow for the predictive modeling conducted in Step 5, it is important to know the

distance between the source of contamination, the

downgradient end of the dissolved plume, and any potential downgradient or cross-gradient receptors.

Step 4: Estimate the Biodegradation Rate

Constant

Biodegradation is the most important process that degrades contaminants in the subsurface; therefore, the

biodegradation rate is one of the most important model

input parameters. Biodegradation of chlorinated aliphatic hydrocarbons can commonly be represented as

a first-order rate constant. Site-specific biodegradation

rates generally are best to use. Calculation of site-specific biodegradation rates is discussed by Wiedemeier

et al. (1, 36, 37). If determining site-specific biodegradation rates is impossible, then literature values for the

biodegradation rate of the contaminant of interest must

be used. It is generally best to start with the average

value and then to vary the model input to predict “best

case” and “worst case” scenarios. Estimated biodegradation rates can be used only after biodegradation has

been shown to be occurring (see Step 1).

Step 5: Compare the Rate of Transport to the

Rate of Attenuation

At this early stage in the natural attenuation demonstration, comparison of the rate of solute transport to the rate

of attenuation is best accomplished using an analytical

model. Several analytical models are available, but the

BIOSCREEN model is probably the simplest to use.

This model is nonproprietary and is available from the

Robert S. Kerr Laboratory’s home page on the Internet

(www.epa.gov/ada/kerrlab.html). The BIOSCREEN

model is based on Domenico’s solution to the advectiondispersion equation (38), and allows use of either a

first-order biodegradation rate or an instantaneous reaction between contaminants and electron acceptors to

simulate the effects of biodegradation. To model transport of chlorinated aliphatic hydrocarbons using

BIOSCREEN, only the first-order decay rate option

should be used. BIOCHLOR, a similar model, is under

development by the Technology Transfer Division of

AFCEE. This model will likely use the same analytical

solution as BIOSCREEN but will be geared towards

evaluating transport of chlorinated compounds under

the influence of biodegradation.

The primary purpose of comparing the rate of transport

with the rate of attenuation is to determine whether the

Example 2. Biodegradation of Chlorinated Organics Unlikely

Analyte

Concentration in Most

Contaminated Zone

Points

Awarded

DO 3 mg/L -3 

Nitrate 0.3 mg/L 2

Iron(II) Not detected 0

Sulfate 10 mg/L 2

Methane ND 0

Oxidation-reduction

potential

100 mV 0

Chloride Background 0

Trichloroethene

(released)

1,200 μg/L 0

cis-1,2-Dichloroethene Not detected 0

Vinyl chloride ND 0

Total points awarded 1

52
residence time along the flow path is adequate to be

protective of human health and the environment (i.e., to

qualitatively estimate whether the contaminant is attenuating at a rate fast enough to allow degradation of the

contaminant to acceptable concentrations before receptors are reached). It is important to perform a sensitivity

analysis to help evaluate the confidence in the preliminary screening modeling effort. If modeling shows that

receptors may not be exposed to contaminants at concentrations above risk-based corrective action criteria,

then the screening criteria are met, and the investigator

can proceed with the natural attenuation feasibility study.

Step 6: Determine Whether the Screening Criteria

Are Met

Before proceeding with the full-scale natural attenuation

feasibility study, the investigator should ensure that the

answers to all of the following criteria are “yes”:

• Has the plume moved a distance less than expected,

based on the known (or estimated) time since the

contaminant release and the contaminant velocity, as

calculated from site-specific measurements of hydraulic conductivity and hydraulic gradient, as well as estimates of effective porosity and contaminant

retardation?

• Is it likely that the contaminant mass is attenuating

at rates sufficient to be protective of human health

and the environment at a point of discharge to a

sensitive environmental receptor?

• Is the plume going to attenuate to concentrations

less than risk-based corrective action guidelines before reaching potential receptors?

Collect Additional Site Characterization Data

To Support Natural Attenuation, As Required

Detailed site characterization is necessary to document

the potential for natural attenuation. Review of existing

site characterization data is particularly useful before

initiating site characterization activities. Such review

should allow identification of data gaps and guide the most

effective placement of additional data collection points.

There are two goals during the site characterization

phase of a natural attenuation investigation. The first is

to collect the data needed to determine whether natural

mechanisms of contaminant attenuation are occurring

at rates sufficient to protect human health and the environment. The second is to provide sufficient site-specific

data to allow prediction of the future extent and concentration of a contaminant plume through solute fate-andtransport modeling. Because the burden of proof for

natural attenuation is on the proponent, detailed site

characterization is required to achieve these goals and

to support this remedial option. Adequate site characterization in support of natural attenuation requires that

the following site-specific parameters be determined:

• The extent and type of soil and ground-water

contamination.

• The location and extent of contaminant source area(s)

(i.e., areas containing mobile or residual NAPL).

• The potential for a continuing source due to leaking

tanks or pipelines.

• Aquifer geochemical parameters.

• Regional hydrogeology, including drinking water

aquifers and regional confining units.

• Local and site-specific hydrogeology, including local

drinking water aquifers; location of industrial, agricultural, and domestic water wells; patterns of aquifer

use (current and future); lithology; site stratigraphy,

including identification of transmissive and nontransmissive units; grain-size distribution (sand versus silt

versus clay); aquifer hydraulic conductivity; groundwater hydraulic information; preferential flow paths;

locations and types of surface water bodies; and

areas of local ground-water recharge and discharge.

• Identification of potential exposure pathways and

receptors.

The following sections describe the methodologies that

should be implemented to allow successful site characterization in support of natural attenuation. Additional information can be obtained from Wiedemeier et al. (1, 37).

Soil Characterization

To adequately define the subsurface hydrogeologic system and to determine the amount and three-dimensional

distribution of mobile and residual NAPL that can act as

a continuing source of ground-water contamination, extensive soil characterization must be completed. Depending on the status of the site, this work may have

been completed during previous remedial investigation

activities. The results of soils characterization will be

used as input into a solute fate-and-transport model to

help define a contaminant source term and to support

the natural attenuation investigation.

The purpose of soil sampling is to determine the subsurface distribution of hydrostratigraphic units and the distribution of mobile and residual NAPL. These objectives

can be achieved through the use of conventional soil

borings or direct-push methods (e.g., Geoprobe or cone

penetrometer testing). All soil samples should be collected, described, analyzed, and disposed of in accordance with local, state, and federal guidance. Wiedemeier

et al. (1) present suggested procedures for soil sample

collection. These procedures may require modification

to comply with local, state, and federal regulations or to

accommodate site-specific conditions.

53
The analytical protocol to be used for soil sample analysis is presented in Table 1. This analytical protocol

includes all of the parameters necessary to document

natural attenuation, including the effects of sorption and

biodegradation. Knowledge of the location, distribution,

concentration, and total mass of contaminants of regulatory concern sorbed to soils or present as residual

and/or mobile NAPL is required to calculate contaminant

partitioning from NAPL into ground water. Knowledge of

the TOC content of the aquifer matrix is important for

sorption and solute-retardation calculations. TOC samples should be collected from a background location in

the stratigraphic horizon(s) where most contaminant

transport is expected to occur. Oxygen and carbon dioxide measurements of soil gas can be used to find

areas in the unsaturated zone where biodegradation is

occurring. Knowledge of the distribution of contaminants

in soil gas can be used as a cost-effective way to

estimate the extent of soil contamination.

Ground-Water Characterization

To adequately determine the amount and three-dimensional distribution of dissolved contamination and to

document the occurrence of natural attenuation,

ground-water samples must be collected and analyzed.

Biodegradation of organic compounds, whether natural

or anthropogenic, brings about measurable changes in

the chemistry of ground water in the affected area. By

measuring these changes, documentation and quantitative evaluation of natural attenuation’s importance at a

site are possible.

Ground-water sampling is conducted to determine the

concentrations and distribution of contaminants, daughter products, and ground-water geochemical parameters. Ground-water samples may be obtained from

monitoring wells or with point-source sampling devices

such as a Geoprobe, Hydropunch, or cone penetrometer. All ground-water samples should be collected in

accordance with local, state, and federal guidelines.

Wiedemeier et al. (1) suggest procedures for groundwater sample collection. These procedures may need to

be modified to comply with local, state, and federal

regulations or to accommodate site-specific conditions.

The analytical protocol for ground-water sample analysis is presented in Table 1. This analytical protocol includes all of the parameters necessary to document

natural attenuation, including the effects of sorption and

biodegradation. Data obtained from the analysis of

ground water for these analytes is used to scientifically

document natural attenuation and can be used as input

into a solute fate-and-transport model. The following

paragraphs describe each ground-water analytical parameter and the use of each analyte in the natural

attenuation demonstration.

Volatile organic compound analysis (by Method

SW8260a) is used to determine the types, concentrations, and distributions of contaminants and daughter

products in the aquifer. DO is the electron acceptor most

thermodynamically favored by microbes for the biodegradation of organic carbon, whether natural or anthropogenic. Reductive dechlorination will not occur,

however, if DO concentrations are above approximately

0.5 milligrams per liter. During aerobic biodegradation of

a substrate, DO concentrations decrease because of

the microbial oxygen demand. After DO depletion, anaerobic microbes will use nitrate as an electron acceptor, followed by iron(III), then sulfate, and finally

carbon dioxide (methanogenesis). Each sequential reaction drives the oxidation-reduction potential of the

ground water further into the realm where reductive

dechlorination can occur. The oxidation-reduction potential range of sulfate reduction and methanogenesis is

optimal, but reductive dechlorination may occur under

nitrate- and iron(III)-reducing conditions as well. Because reductive dechlorination works best in the sulfatereduction and methanogenesis oxidation-reduction

potential range, competitive exclusion between microbial sulfate reducers, methanogens, and reductive

dechlorinators can occur.

After DO has been depleted in the microbiological treatment zone, nitrate may be used as an electron acceptor

for anaerobic biodegradation via denitrification. In some

cases iron(III) is used as an electron acceptor during

anaerobic biodegradation of electron donors. During this

process, iron(III) is reduced to iron(II), which may be

soluble in water. Iron(II) concentrations can thus be used

as an indicator of anaerobic degradation of fuel compounds. After DO, nitrate, and bioavailable iron(III) have

been depleted in the microbiological treatment zone,

sulfate may be used as an electron acceptor for anaerobic biodegradation. This process is termed sulfate reduction and results in the production of sulfide. During

methanogenesis (an anaerobic biodegradation process), carbon dioxide (or acetate) is used as an electron

acceptor, and methane is produced. Methanogenesis

generally occurs after oxygen, nitrate, bioavailable

iron(III), and sulfate have been depleted in the treatment

zone. The presence of methane in ground water is

indicative of strongly reducing conditions. Because

methane is not present in fuel, the presence of methane

in ground water above background concentrations in

contact with fuels is indicative of microbial degradation

of fuel hydrocarbons.

The total alkalinity of a ground-water system is indicative

of a water’s capacity to neutralize acid. Alkalinity is

defined as “the net concentration of strong base in

excess of strong acid with a pure CO2-water system as

the point of reference” (39). Alkalinity results from the

presence of hydroxides, carbonates, and bicarbonates

of elements such as calcium, magnesium, sodium, po54
tassium, or ammonia. These species result from the

dissolution of rock (especially carbonate rocks), the

transfer of carbon dioxide from the atmosphere, and the

respiration of microorganisms. Alkalinity is important in

the maintenance of ground-water pH because it buffers

the ground-water system against acids generated during both aerobic and anaerobic biodegradation.

In general, areas contaminated by fuel hydrocarbons

exhibit a total alkalinity that is higher than that seen in

background areas. This is expected because the microbially mediated reactions causing biodegradation of fuel

hydrocarbons cause an increase in the total alkalinity in

the system. Changes in alkalinity are most pronounced

during aerobic respiration, denitrification, iron reduction,

and sulfate reduction, and are less pronounced during

methanogenesis (40). In addition, Willey et al. (41) show

that short-chain aliphatic acid ions produced during

biodegradation of fuel hydrocarbons can contribute to

alkalinity in ground water.

The oxidation-reduction potential of ground water is a

measure of electron activity and an indicator of the

relative tendency of a solution to accept or transfer

electrons. Redox reactions in ground water containing

organic compounds (natural or anthropogenic) are usually

biologically mediated; therefore, the oxidation-reduction

potential of a ground-water system depends on and

influences rates of biodegradation. Knowledge of the

oxidation-reduction potential of ground water also is

important because some biological processes operate

only within a prescribed range of redox conditions. The

oxidation-reduction potential of ground water generally

ranges from -400 to 800 millivolts (mV). Figure 4 shows

the typical redox conditions for ground water when different electron acceptors are used.

Oxidation-reduction potential can be used to provide

real-time data on the location of the contaminant plume,

especially in areas undergoing anaerobic biodegradation. Mapping the oxidation-reduction potential of the

ground water while in the field helps the field scientist to

determine the approximate location of the contaminant

plume. To perform this task, it is important to have at

least one redox measurement (preferably more) from a

well located upgradient from the plume. Oxidation-reduction potential measurements should be taken during

well purging and immediately before and after sample

Figure 4. Redox potentials for various electron acceptors.

55
acquisition using a direct-reading meter. Because most

well purging techniques can allow aeration of collected

ground-water samples (which can affect oxidation-reduction

potential measurements), it is important to minimize

potential aeration.

Dissolved hydrogen concentrations can be used to determine the dominant terminal electron-accepting process in an aquifer. Because of the difficulty in obtaining

hydrogen analyses commercially, this parameter should

be considered optional at this time. Table 4 presents the

range of hydrogen concentrations for a given terminal

electron-accepting process. Much research has been

done on the topic of using hydrogen measurements to

delineate terminal electron-accepting processes (42-

44). Because the efficiency of reductive dechlorination

differs for methanogenic, sulfate-reducing, iron(III)-reducing, or denitrifying conditions, it is helpful to have

hydrogen concentrations to help delineate redox conditions when evaluating the potential for natural attenuation of chlorinated ethenes in ground-water systems.

Collection and analysis of ground-water samples for

dissolved hydrogen content is not yet commonplace or

standardized, however, and requires a relatively expensive field laboratory setup.

Because the pH, temperature, and conductivity of a

ground-water sample can change significantly shortly

following sample acquisition, these parameters must be

measured in the field in unfiltered, unpreserved, “fresh”

water collected by the same technique as the samples

taken for DO and redox analyses. The measurements

should be made in a clean glass container separate from

those intended for laboratory analysis, and the measured values should be recorded in the ground-water

sampling record.

The pH of ground water has an effect on the presence

and activity of microbial populations in the ground water.

This is especially true for methanogens. Microbes capable of degrading chlorinated aliphatic hydrocarbons and

petroleum hydrocarbon compounds generally prefer pH

values varying from 6 to 8 standard units. Ground-water

temperature directly affects the solubility of oxygen and

other geochemical species. The solubility of DO is temperature dependent, being more soluble in cold water

than in warm water. Ground-water temperature also affects

the metabolic activity of bacteria. Rates of hydrocarbon

biodegradation roughly double for every 10°C increase

in temperature (“Q”10 rule) over the temperature range

between 5°C and 25°C. Ground-water temperatures

less than about 5°C tend to inhibit biodegradation, and

slow rates of biodegradation are generally observed in

such waters.

Conductivity is a measure of the ability of a solution to

conduct electricity. The conductivity of ground water is

directly related to the concentration of ions in solution;

conductivity increases as ion concentration increases.

Conductivity measurements are used to ensure that

ground water samples collected at a site are representative of the water in the saturated zone containing

the dissolved contamination. If the conductivities of

samples taken from different sampling points are radically different, the waters may be from different hydrogeologic zones.

Elemental chlorine is the most abundant of the halogens. Although chlorine can occur in oxidation states

ranging from Cl-

 to Cl+7, the chloride form (Cl-) is the only

form of major significance in natural waters (45). Chloride forms ion pairs or complex ions with some of the

cations present in natural waters, but these complexes

are not strong enough to be of significance in the chemistry of fresh water (45). The chemical behavior of chloride is neutral. Chloride ions generally do not enter into

oxidation-reduction reactions, form no important solute

complexes with other ions unless the chloride concentration is extremely high, do not form salts of low solubility, are not significantly adsorbed on mineral surfaces,

and play few vital biochemical roles (45). Thus, physical

processes control the migration of chloride ions in the

subsurface.

Kaufman and Orlob (46) conducted tracer experiments

in ground water and found that chloride moved through

most of the soils tested more conservatively (i.e., with

less retardation and loss) than any of the other tracers

tested. During biodegradation of chlorinated hydrocarbons dissolved in ground water, chloride is released into

the ground water. This results in chloride concentrations

in the ground water of the contaminant plume that are

elevated relative to background concentrations. Because of the neutral chemical behavior of chloride, it can

be used as a conservative tracer to estimate biodegradation rates using methods similar to those discussed

by Wiedemeier et al. (36).

Field Measurement of Aquifer Hydraulic

Parameters

The properties of an aquifer that have the greatest impact on contaminant fate and transport include hydraulic

conductivity, hydraulic gradient, porosity, and dispersivTable 4. Range of Hydrogen Concentrations for a Given

Terminal Electron-Accepting Process

Terminal

Electron-Accepting Process

Hydrogen Concentration

(nanomoles per liter)

Denitrification < 0.1

Iron(III) reduction 0.2 to 0.8

Sulfate reduction 1 to 4

Methanogenesis > 5

56
ity. Estimating hydraulic conductivity and gradient in the

field is fairly straightforward, but obtaining field-scale

information on porosity and dispersivity can be difficult.

Therefore, most investigators rely on field data for hydraulic conductivity and hydraulic gradient and on literature values for porosity and dispersivity for the types of

sediments present at the site. Methods for field measurement of aquifer hydraulic parameters are described

by Wiedemeier et al. (1, 37).

Microbiological Laboratory Data

Microcosm studies are used to show that the microorganisms necessary for biodegradation are present and

to help quantify rates of biodegradation. If properly designed, implemented, and interpreted, microcosm studies can provide very convincing documentation of the

occurrence of biodegradation. Such studies are the only

“line of evidence” that allows an unequivocal mass balance determination based on the biodegradation of environmental contaminants. The results of a well-designed

microcosm study will be easy for decision-makers with

nontechnical backgrounds to interpret. Results of such

studies are strongly influenced by the nature of the

geological material submitted for study, the physical

properties of the microcosm, the sampling strategy, and

the duration of the study. Because microcosm studies

are time-consuming and expensive, they should be undertaken only at sites where there is considerable skepticism concerning the biodegradation of contaminants.

Biodegradation rate constants determined by microcosm studies often are much greater than rates

achieved in the field. Microcosms are most appropriate

as indicators of the potential for natural bioremediation

and to prove that losses are biological, but it may be

inappropriate to use them to generate rate constants.

The preferable method of contaminant biodegradation

rate-constant determination is in situ field measurement.

The collection of material for the microcosm study, the

procedures used to set up and analyze the microcosm,

and the interpretation of the results of the microcosm

study are presented by Wiedemeier et al. (1).

Refine the Conceptual Model, Complete

Premodeling Calculations, and Document

Indicators of Natural Attenuation

Site investigation data should first be used to refine the

conceptual model and quantify ground-water flow, sorption, dilution, and biodegradation. The results of these

calculations are used to scientifically document the occurrence and rates of natural attenuation and to help simulate

natural attenuation over time. Because the burden of

proof is on the proponent, all available data must be

integrated in such a way that the evidence is sufficient to

support the conclusion that natural attenuation is occurring.

Conceptual Model Refinement

Conceptual model refinement involves integrating newly

gathered site characterization data to refine the preliminary conceptual model that was developed based on

previously existing site-specific data. During conceptual

model refinement, all available site-specific data should

be integrated to develop an accurate three-dimensional

representation of the hydrogeologic and contaminant

transport system. This conceptual model can then be

used for contaminant fate-and-transport modeling. Conceptual model refinement consists of several steps, including preparation of geologic logs, hydrogeologic

sections, potentiometric surface/water table maps, contaminant contour (isopleth) maps, and electron acceptor

and metabolic byproduct contour (isopleth) maps. Refinement of the conceptual model is described by

Wiedemeier et al. (1).

Premodeling Calculations

Several calculations must be made prior to implementation of the solute fate-and-transport model. These calculations include sorption and retardation calculations,

NAPL/water-partitioning calculations, ground-water flow

velocity calculations, and biodegradation rate-constant

calculations. Each of these calculations is discussed in

the following sections. Most of the specifics of each

calculation are presented in the fuel hydrocarbon natural

attenuation technical protocol by Wiedemeier et al. (1),

and all will be presented in the protocol incorporating

chlorinated aliphatic hydrocarbon attenuation (37).

Biodegradation Rate Constant Calculations

Biodegradation rate constants are necessary to simulate accurately the fate and transport of contaminants

dissolved in ground water. In many cases, biodegradation of contaminants can be approximated using first-order kinetics. To calculate first-order biodegradation rate

constants, the apparent degradation rate must be normalized for the effects of dilution and volatilization. Two

methods for determining first-order rate constants are

described by Wiedemeier et al. (36). One method involves the use of a biologically recalcitrant compound

found in the dissolved contaminant plume that can be

used as a conservative tracer. The other method, proposed by Buscheck and Alcantar (47) involves interpretation of a steady-state contaminant plume and is based

on the one-dimensional steady-state analytical solution

to the advection-dispersion equation presented by Bear

(48). The first-order biodegradation rate constants for

chlorinated aliphatic hydrocarbons are also presented

(J. Wilson et al., this volume).

57
Simulate Natural Attenuation Using Solute

Fate-and-Transport Models

Simulating natural attenuation using a solute fate-andtransport model allows prediction of the migration and

attenuation of the contaminant plume through time. Natural attenuation modeling is a tool that allows site-specific

data to be used to predict the fate and transport of

solutes under governing physical, chemical, and biological processes. Hence, the results of the modeling effort

are not in themselves sufficient proof that natural attenuation is occurring at a given site. The results of the

modeling effort are only as good as the original data

input into the model; therefore, an investment in thorough site characterization will improve the validity of the

modeling results. In some cases, straightforward analytical models of contaminant attenuation are adequate

to simulate natural attenuation.

Several well-documented and widely accepted solute

fate-and-transport models are available for simulating

the fate-and-transport of contaminants under the influence of advection, dispersion, sorption, and biodegradation. The use of solute fate-and-transport modeling in

the natural attenuation investigation is described by

Wiedemeier et al. (1).

Identify Potential Receptors, and Conduct an

Exposure-Pathway Analysis

After the rates of natural attenuation have been documented and predictions of the future extent and concentrations of the contaminant plume have been made

using the appropriate solute fate-and-transport model,

the proponent of natural attenuation should combine all

available data and information to negotiate for implementation of this remedial option. Supporting the natural

attenuation option generally will involve performing a

receptor exposure-pathway analysis. This analysis includes identifying potential human and ecological receptors and points of exposure under current and future

land and ground-water use scenarios. The results of

solute fate-and-transport modeling are central to the

exposure pathways analysis. If conservative model input parameters are used, the solute fate-and-transport

model should give conservative estimates of contaminant plume migration. From this information, the potential for impacts on human health and the environment

from contamination present at the site can be estimated.

Evaluate Supplemental Source Removal

Options

Source removal or reduction may be necessary to reduce plume expansion if the exposure-pathway analysis

suggests that one or more exposure pathways may be

completed before natural attenuation can reduce chemical concentrations below risk-based levels of concern.

Further, some regulators may require source removal in

conjunction with natural attenuation. Several technologies suitable for source reduction or removal are listed

in Figure 1. Other technologies may also be used as

dictated by site conditions and local regulatory requirements. The authors’ experience indicates that source

removal can be very effective at limiting plume migration

and decreasing the remediation time frame, especially

at sites where biodegradation is contributing to natural

attenuation of a dissolved contaminant plume. The impact of source removal can readily be evaluated by

modifying the contaminant source term if a solute fateand-transport model has been prepared for a site; this

will allow for a reevaluation of the exposure-pathway

analysis.

Prepare a Long-Term Monitoring Plan

Ground-water flow rates at many Air Force sites studied

to date are such that many years will be required before

contaminated ground water could potentially reach Base

property boundaries. Thus, there frequently is time and

space for natural attenuation alone to reduce contaminant concentrations in ground water to acceptable levels. Experience at 40 Air Force sites contaminated with

fuel hydrocarbons using the protocol presented by

Wiedemeier et al. (1) suggests that many fuel hydrocarbon plumes are relatively stable or are moving very

slowly with respect to ground-water flow. This information is complemented by data collected by Lawrence

Livermore National Laboratories in a study of over 1,100

leaking underground fuel tank sites performed for the

California State Water Resources Control Board (49).

These examples demonstrate the efficacy of long-term

monitoring to track plume migration and to validate or

refine modeling results. There is not a large enough

database available at this time to assess the stability of

chlorinated solvent plumes, but in the authors’ experience chlorinated solvent plumes are likely to migrate

further downgradient than fuel hydrocarbon plumes before reaching steady-state equilibrium or before receding.

The long-term monitoring plan consists of locating

ground-water monitoring wells and developing a

ground-water sampling and analysis strategy. This plan

is used to monitor plume migration over time and to

verify that natural attenuation is occurring at rates sufficient to protect potential downgradient receptors. The

long-term monitoring plan should be developed based

on site characterization data, the results of solute fateand-transport modeling, and the results of the exposurepathway analysis.

The long-term monitoring plan includes two types of

monitoring wells: long-term monitoring wells are intended to determine whether the behavior of the plume

is changing; point-of-compliance wells are intended to

detect movements of the plume outside the negotiated

perimeter of containment, and to trigger an action to

58
manage the risk associated with such expansion. Figure

5 depicts 1) an upgradient well in unaffected ground

water, 2) a well in the NAPL source area, 3) a well

downgradient of the NAPL source area in a zone of

anaerobic treatment, 4) a well in the zone of aerobic

treatment, along the periphery of the plume, 5) a well

located downgradient from the plume where contaminant concentrations are below regulatory acceptance

levels and soluble electron acceptors are depleted with

respect to unaffected ground water, and 6) three pointof-compliance wells.

Although the final number and placement of long-term

monitoring and point-of-compliance wells is determined

through regulatory negotiation, the following guidance is

recommended. Locations of long-term monitoring wells

are based on the behavior of the plume as revealed

during the initial site characterization and on regulatory

considerations. Point-of-compliance wells are placed

500 feet downgradient from the leading edge of the

plume or the distance traveled by the ground water in

2 years, whichever is greater. If the property line is less

than 500 feet downgradient, the point-of-compliance

wells are placed near and upgradient from the property line. The final number and location of point-ofcompliance monitoring wells also depends on regulatory

considerations.

The results of a solute fate-and-transport model can be

used to help site the long-term monitoring and point-ofcompliance wells. To provide a valid monitoring system,

all monitoring wells must be screened in the same hydrogeologic unit as the contaminant plume. This generally requires detailed stratigraphic correlation. To

facilitate accurate stratigraphic correlation, detailed visual descriptions of all subsurface materials encountered

during borehole drilling should be prepared prior to

monitoring-well installation.

A ground-water sampling and analysis plan should be

prepared in conjunction with point-of-compliance and

long-term monitoring well placement. For long-term

monitoring wells, ground-water analyses should include

volatile organic compounds, DO, nitrate, iron(II), sulfate,

and methane. For point-of-compliance wells, groundwater analyses should be limited to determining volatile

organic compound and DO concentrations. Any statespecific analytical requirements also should be addressed in the sampling and analysis plan to ensure that

all data required for regulatory decision-making are collected. Water level and LNAPL thickness measurements

must be made during each sampling event. Except at

sites with very low hydraulic conductivity and gradients,

quarterly sampling of long-term monitoring wells is recommended during the first year to help determine the

direction of plume migration and to determine baseline

data. Based on the results of the first year’s sampling,

the sampling frequency may be reduced to annual sampling in the quarter showing the greatest extent of the

plume. Sampling frequency depends on the final placement of the point-of-compliance monitoring wells and

ground-water flow velocity. The final sampling frequency

should be determined in collaboration with regulators.

Present Findings to Regulatory Agencies, and

Obtain Approval for Remediation by Natural

Attenuation

The purpose of regulatory negotiations is to provide

scientific documentation that supports natural attenuation as the most appropriate remedial option for a given

site. All available site-specific data and information developed during the site characterization, conceptual

model development, premodeling calculations, biodegradation rate calculation, ground-water modeling,

model documentation, and long-term monitoring plan

preparation phases of the natural attenuation investigation should be presented in a consistent and complementary manner at the regulatory negotiations. Of

particular interest to the regulators will be proof that

natural attenuation is occurring at rates sufficient to

meet risk-based corrective action criteria at the point of

compliance and to protect human health and the environment. The regulators must be presented with a

“weight-of-evidence” argument in support of this remedial option. For this reason, all model assumptions

should be conservative, and all available evidence in

support of natural attenuation must be presented at the

regulatory negotiations.

A comprehensive long-term monitoring and contingency

plan also should be presented to demonstrate a commitment to proving the effectiveness of natural attenuation as a remedial option. Because long-term

monitoring and contingency plans are very site specific,

they should be addressed in the individual reports generated using this protocol.

Figure 5. Hypothetical long-term monitoring strategy.

59
References

 1. Wiedemeier, T.H., J.T. Wilson, D.H. Kampbell, R.N. Miller, and

J.E. Hansen. 1995. Technical protocol for implementing intrinsic

remediation with long-term monitoring for natural attenuation of

fuel contamination dissolved in groundwater. San Antonio, TX:

U.S. Air Force Center for Environmental Excellence.

 2. National Research Council. 1993. In-situ bioremediation: When

does it work? Washington, DC: National Academy Press.

 3. Bouwer, E.J., B.E. Rittman, and P.L. McCarty. 1981. Anaerobic

degradation of halogenated 1- and 2-carbon organic compounds.

Environ. Sci. Technol. 15 (5):596-599.

 4. Wilson, J.T., and B.H. Wilson. 1985. Biotransformation of trichloroethylene in soil. Appl. Environ. Microbiol. 49(1):242-243.

 5. Miller, R.E., and F.P. Guengerich. 1982. Oxidation of trichloroethylene by liver microsomal cytochrome P-450: Evidence for

chlorine migration in a transition state not involving trichloroethylene oxide. Biochemistry 21:1090-1097.

 6. Nelson, M.J.K., S.O. Montgomery, E.J. O’Neille, and P.H.

Pritchard. 1986. Aerobic metabolism of trichloroethylene by a

bacterial isolate. Appl. Environ. Microbiol. 52 (2):949-954.

 7. Bouwer, E.J., and J.P. Wright. 1988. Transformations of trace

halogenated aliphatics in anoxic biofilm columns. J. Contam. Hydrol. 2:155-169.

 8. Lee, M.D. 1988. Biorestoration of aquifers contaminated with

organic compounds. CRC Crit. Rev. Environ. Control 18:29-89.

 9. Little, C.D., A.V. Palumbo, S.E. Herbes, M.E. Lidstrom, R.L. Tyndall, and P.J. Gilmer. 1988. Trichloroethylene biodegradation by

a methane-oxidizing bacterium. Appl. Environ. Microbiol.

54(4):951-956.

10. Mayer, K.P., D. Grbi-Gali, L. Semprini, and P.L. McCarty. 1988.

Degradation of trichloroethylene by methanotrophic bacteria in a

laboratory column of saturated aquifer material. Water Sci. Tech.

20(11/12):175-178.

11. Arciero, D., T. Vannelli, M. Logan, and A.B. Hooper. 1989. Degradation of trichloroethylene by the ammonia-oxidizing bacterium

Nitrosomonas europaea. Biochem. Biophys. Res. Commun.

159:640-643.

12. Cline, P.V., and J.J. Delfino. 1989. Transformation kinetics of

1,1,1-trichloroethane to the stable product 1,1-dichloroethene. In:

Biohazards of drinking water treatment. Chelsea, MI: Lewis Publishers.

13. Freedman, D.L., and J.M. Gossett. 1989. Biological reductive

dechlorination of tetrachloethylene and trichloroethylene to ethylene under methanogenic conditions. Appl. Environ. Microbiol.

55:2144-2151.

14. Folsom, B.R., P.J. Chapman, and P.H. Pritchard. 1990. Phenol

and trichloroethylene degradation by Pseudomonas cepacia G4:

Kinetics and interactions between substrates. Appl. Environ. Microbiol. 56(5):1279-1285.

15. Harker, A.R., and Y. Kim 1990. Trichloroethylene degradation by

two independent aromatic-degrading pathways in Alcaligenes eutrophus JMP134. Appl. Environ. Microbiol. 56(4):1179-1181.

16. Alvarez-Cohen, L.M., and P.L. McCarty. 1991. Effects of toxicity,

aeration, and reductant supply on trichloroethylene transformation by a mixed methanotrophic culture. Appl. Environ. Microbiol.

57(1):228-235.

17. Alvarez-Cohen, L.M., and P.L. McCarty. 1991. Product toxicity

and cometabolic competitive inhibition modeling of chloroform

and trichloroethylene transformation by methanotrophic resting

cells. Appl. Environ. Microbiol. 57(4):1031-1037.

18. DeStefano, T.D., J.M. Gossett, and S.H. Zinder. 1991. Reductive

dehalogenation of high concentrations of tetrachloroethene to

ethene by an anaerobic enrichment culture in the absence of

methanogenesis. Appl. Environ. Microbiol. 57(8):2287-2292.

19. Henry, S.M. 1991. Transformation of trichloroethylene by

methanotrophs from a groundwater aquifer. Ph.D. thesis. Stanford University, Palo Alto, CA.

20. McCarty, P.L., P.V. Roberts, M. Reinhard, and G. Hopkins. 1992.

Movement and transformations of halogenated aliphatic compounds in natural systems. In: Schnoor, J.L., ed. Fate of pesticides and chemicals in the environment. New York, NY: John

Wiley & Sons.

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

metabolism in Mycobacterium aurum Li. Appl. Environ. Microbiol.

58(4):1220-1226.

22. McCarty, P.L., and L. Semprini. 1994. Ground-water treatment for

chlorinated solvents, In: Norris, R.D., R.E. Hinchee, R. Brown,

P.L. McCarty, L. Semprini, J.T. Wilson, D.H. Kampbell, M. Reinhard, E.J. Bouwer, R.C. Borden, T.M. Vogel, J.M. Thomas, and

C.H. Ward, eds. Handbook of bioremediation. Boca Raton, FL:

Lewis Publishers.

23. Vogel, T.M. 1994. Natural bioremediation of chlorinated solvents.

In: Norris, R.D., R.E. Hinchee, R. Brown, P.L. McCarty, L. Semprini, J.T. Wilson, D.H. Kampbell, M. Reinhard, E.J. Bouwer, R.C.

Borden, T.M. Vogel, J.M. Thomas, and C.H. Ward, eds. Handbook of bioremediation. Boca Raton, FL: Lewis Publishers.

24. Bouwer, E.J. 1994. Bioremediation of chlorinated solvents using

alternate electron acceptors. In: Norris, R.D., R.E. Hinchee, R.

Brown, P.L. McCarty, L. Semprini, J.T. Wilson, D.H. Kampbell, M.

Reinhard, E.J. Bouwer, R.C. Borden, T.M. Vogel, J.M. Thomas,

and C.H. Ward, eds. Handbook of bioremediation. Boca Raton,

FL: Lewis Publishers.

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

tetrachloroethylene to trichloroethylene, dichloroethylene, vinyl

chloride, and carbon dioxide under methanogenic conditions.

Appl. Environ. Microbiol. 49(5):1080-1083.

26. Murray, W.D., and M. Richardson. 1993. Progress toward the

biological treatment of C1 and C2 halogenated hydrocarbons.

Crit. Rev. Environ. Sci. Technol. 23(3):195-217.

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

of vinyl chloride in Fe(III)-reducing aquifer sediments. Environ.

Sci. Technol. 40:2084-2086.

28. Wiedemeier, T.H., L.A. Benson, J.T. Wilson, D.H. Kampbell, J.E.

Hansen, and R. Miknis. 1996. Patterns of natural attenuation of

chlorinated aliphatic hydrocarbons at Plattsburgh Air Force Base,

New York. Platform abstracts presented at the Conference on

Intrinsic Remediation of Chlorinated Solvents, Salt Lake City, UT,

April 2.

29. U.S. EPA. 1986. Test methods for evaluating solid waste, physical

and chemical methods, 3rd ed. SW-846. Washington, DC.

30. U.S. EPA. 1983. Methods for chemical analysis of water and

wastes. EPA/16020-07-71. Cincinnati, OH.

31. Hach Co. 1990. Hach Company Catalog: Products for Analysis.

Ames, IA.

32. American Public Health Association. 1992. Standard methods for

the examination of water and wastewater, 18th ed. Washington,

DC.

33. AFCEE. 1993. Handbook to support the Installation Restoration

Program (IRP) remedial investigations and feasibility studies

(RI/FS). U.S. Air Force Center for Environmental Excellence.

September. Brooks Air Force Base, TX.

34. AFCEE. 1992. Environmental chemistry function Installation Restoration Program analytical protocols. June.

60
35. Kampbell, D.H., J.T. Wilson, and S.A. Vandegrift. 1989. Dissolved

oxygen and methane in water by a GC headspace equilibrium

technique. Int. J. Environ. Anal. Chem. 36:249-257.

36. Wiedemeier, T.H., M.A. Swanson, J.T. Wilson, D.H. Kampbell,

R.N. Miller, and J.E. Hansen. 1996. Approximation of biodegradation rate constants for monoaromatic hydrocarbons (BTEX) in

groundwater. Ground Water Monitoring and Remediation. In

press.

37. Wiedemeier, T.H., M.A. Swanson, D.E. Moutoux, J.T. Wilson,

D.H. Kampbell, J.E. Hansen, P. Haas, and F.H. Chapelle. 1996.

Technical protocol for natural attenuation of chlorinated solvents

in groundwater. San Antonio, TX: U.S. Air Force Center for Environmental Excellence. In preparation.

38. Domenico, P.A. 1987. An analytical model for multidimensional

transport of a decaying contaminant species. J. Hydrol. 91:49-58.

39. Domenico, P.A., and F.W. Schwartz. 1990. Physical and chemical

hydrogeology. New York, NY: John Wiley and Sons.

40. Morel, F.M.M., and J.G. Hering. 1993. Principles and applications

of aquatic chemistry. New York, NY: John Wiley & Sons.

41. Willey, L.M., Y.K. Kharaka, T.S. Presser, J.B. Rapp, and I. Barnes.

1975. Short chain aliphatic acid anions in oil field waters and their

contribution to the measured alkalinity. Geochim. Cosmochim.

Acta 39:1707-1711.

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

as an indicator of the predominant terminal electron-accepting

reaction in aquatic sediments. Geochim. Cosmochim. Acta

52:2993-3003.

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

dissolved H2 concentrations to determine distribution of microbially catalyzed redox reactions in anoxic groundwater. Environ.

Sci. Technol. 28(7):1205-1210.

44. Chapelle, F.H., P.B. McMahon, N.M. Dubrovsky, R.F. Fujii, E.T.

Oaksford, and D.A. Vroblesky. 1995. Deducing the distribution of

terminal electron-accepting processes in hydrologically diverse

groundwater systems. Water Resour. Res. 31:359-371.

45. Hem, J.D. 1985. Study and interpretation of the chemical characteristics of natural water. U.S. Geological Survey Water Supply

Paper 2254.

46. Kaufman, W.J., and G.T. Orlob. 1956. Measuring ground water

movement with radioactive and chemical tracers. Am. Water

Works Assn. J. 48:559-572.

47. Buscheck, T.E., and C.M. Alcantar. 1995. Regression techniques

and analytical solutions to demonstrate intrinsic bioremediation.

In: Proceedings of the 1995 Battelle International Conference on

In-Situ and On Site Bioreclamation. April.

48. Bear, J. 1979. Hydraulics of groundwater. New York, NY:

McGraw-Hill.

49. Rice, D.W., R.D. Grose, J.C. Michaelsen, B.P. Dooher, D.H. MacQueen, S.J. Cullen, W.E. Kastenberg, L.G. Everett, and M.A.

Marino. 1995. California leaking underground fuel tank (LUFT)

historical case analyses. California State Water Resources Control Board.

61
The BIOSCREEN Computer Tool

Charles J. Newell and R. Kevin McLeod

Groundwater Services, Inc., Houston, Texas

James R. Gonzales

U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, Texas

Introduction

BIOSCREEN is an easy-to-use screening tool for simulating the natural attenuation of dissolved hydrocarbons

at petroleum fuel release sites. The software, programmed in the Microsoft Excel spreadsheet environment and based on the Domenico analytical solute

transport model (1), has the ability to simulate advection,

dispersion, adsorption, and aerobic decay, as well as

anaerobic reactions that have been shown to be the

dominant biodegradation processes at many petroleum

release sites. BIOSCREEN includes three different

model types: solute transport without decay, solute

transport with biodegradation modeled as a first-order

decay process (simple, lumped-parameter approach),

and solute transport with biodegradation modeled as an

“instantaneous” biodegradation reaction (the approach

used by BIOPLUME models) (2).

Intended Uses for BIOSCREEN

BIOSCREEN attempts to answer two fundamental

questions regarding intrinsic remediation (3):

• How far will the plume extend if no engineered control

or source zone reduction is implemented?

BIOSCREEN uses an analytical solute transport model

with two options for simulating in situ biodegradation:

first order decay and instantaneous reaction. The

model predicts the maximum extent of plume

migration, which may then be compared with the

distance to potential points of exposure (e.g., drinking

water wells, ground-water discharge areas, or

property boundaries).

• How long will the plume persist until natural attenuation processes cause it to dissipate?

BIOSCREEN uses a simple mass balance approach,

based on the mass of dissolvable hydrocarbons in the

source zone and the rate of hydrocarbons leaving the

source zone, to estimate the source zone concentration

versus time. Because an exponential decay in source

zone concentration is assumed, the predicted plume

lifetimes can be large, usually ranging from 5 to 500

years. Note that this is an unverified relationship (there

are little data showing source concentrations versus

long periods), and the results should be considered

order-of-magnitude estimates of the time to dissipate

the plume.

BIOSCREEN is intended to be used in two ways:

• As a screening model to determine whether intrinsic

remediation is feasible at a given site. In this case,

BIOSCREEN is used early in the remediation process

and before site characterization activities are completed. Some data, such as electron acceptor concentrations, may not be available, so typical values are

used. The BIOSCREEN results are used to determine

whether an intrinsic remediation field program should

be implemented to quantify the natural attenuation occurring at a site. In addition, BIOSCREEN is an excellent communication and teaching tool that can be used

to present information in a graphical manner and help

explain the concepts behind natural attenuation.

• As the primary intrinsic remediation ground-water

model at smaller sites. The U.S. Air Force Intrinsic

Remediation Protocol describes how intrinsic remediation models may be used to help verify that natural

attenuation is occurring and to help predict how far

plumes might extend under an intrinsic remediation

scenario. At large, high-effort sites, such as Superfund and Resource Conservation and Recovery Act

sites, a more sophisticated intrinsic remediation

model is probably more appropriate. At smaller,

lower-effort sites, such as service stations, BIOSCREEN

62
may be sufficient to complete the intrinsic remediation study.

BIOSCREEN Input and Output

To run BIOSCREEN, the user enters site data in the

following categories: hydrogeologic, dispersion, adsorption, biodegradation, general information, source characteristics, and observed data. For several parameters

(e.g., seepage velocity), the user can either enter the

value directly or use supporting data (hydraulic conductivity, hydraulic gradient, and effective porosity) to calculate the value. Figure 1 shows the actual input screen.

BIOSCREEN output includes plume centerline graphs,

three-dimensional color plots of plume concentrations,

and mass balance data showing the contaminant mass

removal by each electron acceptor (instantaneous reaction option). Figures 2 and 3 show the two output

screens. The input and output screens have on-line help

built into the software. A detailed user’s manual is also

available (4).

BIOCHLOR: A BIOSCREEN for

Chlorinated Solvents

While BIOSCREEN was originally designed to simulate intrinsic remediation at petroleum release sites,

the system can be modified to simulate intrinsic remediation of chlorinated hydrocarbons. Current plans call

for converting the BIOSCREEN model to BIOCHLOR.

Key changes are:

• Biodegradation using first-order decay only: Microbial constraints on kinetics are much more important

for chlorinated solvents than for petroleum compounds. Therefore, the first-order decay approach

will be emphasized in both the BIOCHLOR software

and manual. A detailed survey of solute decay data

and source decay data from existing sites and the

literature will be provided.

• More detailed information on source terms: Chlorinated solvents are associated with the presence of

free-phase and residual dense nonaqueous phase

liquids (DNAPLs) rather than residual light

nonaqueous phase liquids (LNAPLs) such as gasoline and JP-4. The source terms will be discussed in

more detail to ensure that model input data and preliminary calculations are representative of DNAPL

sites.

• Evaluation of biodegradation products: The generation of products of chlorinated solvent biodegradation

will be discussed. Simple analytical tools may be

developed and incorporated into BIOCHLOR.

BIOSCREEN is available by contacting EPA’s Center for

Subsurface Modeling Support (CSMoS), NRMRL/SPRD,

P.O. Box 1198, Ada, OK 74821-1198, telephone 405-436-

8594, fax 405-436-8718, bulletin board 405-436-8506

Figure 1. BIOSCREEN input screen.

63
Figure 2. BIOSCREEN centerline output screen.

Figure 3. BIOSCREEN concentration array output screen.

64
(14,400 baud, 8 bits, 1 stopbit available, no parity), and

Internet http://www.epa.gov/ada/kerrlab.html. Electronic

manuals will be available in .pdf format; the Adobe

Acrobat Reader is necessary to read and print .pdf files.)

References

1. Domenico, P.A. 1987. An analytical model for multidimensional

transport of a decaying contaminant species. J. Hydro. 91:49-58.

2. Rifai, H.S., P.B. Bedient, R.C. Borden, and J.F. Haasbeek. 1987.

BIOPLUME II—computer model of two-dimensional transport under the influence of oxygen limited biodegradation in ground water.

User’s manual, Ver. 1.0. Rice University, Houston, TX.

3. Newell, C.J., J.W. Winters, H.S. Rifai, R.N. Miller, J. Gonzales, and

T.H. Wiedemeier. 1995. Modeling intrinsic remediation with multiple electron acceptors: Results from seven sites. In: Proceedings

of the Petroleum Hydrocarbons and Organic Chemicals in Ground

Water Conference, Houston, TX, November. National Ground

Water Association. pp. 33-48.

4. Newell, C.J., R.K. McLeod, and J.R. Gonzales. 1996.

BIOSCREEN Natural Attenuation Decision Support System, Version 1.3, U.S. Air Force Center for Environmental Excellence,

Brooks AFB, San Antonio, TX.

65
Case Study: Naval Air Station Cecil Field, Florida

Francis H. Chapelle and Paul M. Bradley

U.S. Geological Survey, Columbia, South Carolina

Redox processes at a fire-training area at Naval Air

Station Cecil Field in Florida are segregated into distinct

and clearly definable zones. Near the source of contamination, methanogenesis predominates. As ground water

flows downgradient, distinct sulfate-reducing, iron(III)-reducing, and oxygen-reducing zones are encountered. This

naturally occurring sequence favors the reductive dehalogenation of chlorinated ethenes near the contamination

source, followed by oxidative degradation of vinyl chloride

to carbon dioxide and chloride downgradient of the

source. This sequence of redox processes has created a

natural bioreactor that effectively treats contaminated

ground water without human intervention. These results

show that mapping the zonation of redox processes at

individual sites is an important step in evaluating the potential for natural attenuation of chlorinated ethenes.

66
Case Study of Natural Attenuation of Trichloroethene at St. Joseph, Michigan

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

U.S. Environmental Protection Agency,

National Risk Management Research Laboratory, Ada, Oklahoma 

Introduction

Trichloroethene (TCE) was found in ground water at the

St. Joseph, Michigan, Superfund site in 1982. The site,

located 4 miles south of St. Joseph and 0.5 mile east of

Lake Michigan, has been used for auto parts manufacturing since 1942. The aquifer is primarily composed of

medium, fine, and very fine glacial sands. The base of

the aquifer is defined by a clay layer that lies between

21 and 29 meters below the ground surface, with elevation of the clay layer increasing toward Lake Michigan.

Investigation at the site included an exhaustive study of

41 possible contaminant sources but did not definitively

identify the source.

The source was apparently situated over a groundwater divide, however, as the contamination was divided

into eastern and western plumes. Both plumes were

found to contain TCE, cis- and trans-1,2-dichloroethene

(cis-DCE and t-DCE), 1,1-dichloroethene (1,1-DCE),

and vinyl chloride (VC). Initial investigation indicated

that natural anaerobic degradation of the TCE was occurring in the western plume, because of the presence

of transformation products and significant levels of

ethene and methane (1, 2).

This paper describes the investigation at the site and

presents the field evidence for natural attenuation of

TCE. Since degradation of TCE is known to occur anaerobically under differing redox conditions and to produce specific daughter products, the relationships

between measured concentrations of chlorinated

ethenes and various redox indicators are emphasized.

Sampling Strategy

Water samples were taken in October 1991 and March

1992 from a 5-foot long slotted auger (3). Seventeen

boreholes were completed near the source of the western plume (1), which formed three transect crossing the

contaminant plume. Data from these first three transect

were analyzed by Semprini et al. (4). 

In 1992, two additional transect (4 and 5 on Figure 1)

consisting of nine additional slotted auger borings

were completed. These two transect were chosen to

sample the plume in the vicinity of Lake Michigan. In

each boring, water samples were taken in 5-foot intervals from the water table to the base of the aquifer.

Onsite gas chromatography was used to determine

the width of the plume and the point of highest

concentration in each transect. The onsite gas chromatography ensured that the entire width of the contaminant plume was captured within each transect. In

August 1994, data were collected from a transect

located about 100 meters offshore that was roughly

parallel to the shore line and contained four borings.

Water samples were taken with a barge-mounted geoprobe (3). Data from the lake transect showed the

location of the plume by the observed reduction in

dissolved oxygen concentrations and the measured

redox potentials. 

Results

Figures 2 through 5 show the data from all the boreholes

separated by transect, which in effect also separates

them by sample date. By compositing the data set,

sitewide trends can be seen. These figures are supplemented by Figures 6 through 9, which show contaminant

distribution with depth in single boreholes from representative locations. Significant methane concentrations

occurred where dissolved oxygen concentrations were

low (Figure 2). Variation in concentration occurring on a

scale smaller than the length of the auger is not accurately represented, as waters of differing chemistry may

mix upon sampling. This may explain why a few data

points simultaneously have high methane and high oxygen concentrations. Most importantly, the figure indicates that a large number of sample locations at the site

had the necessary strong reducing conditions for reductive dechlorination to occur.

67
Figure 3 shows the distribution of sulfate and oxygen.

Many of the points cluster near the sulfate axis, which

generally corresponds to the regions of high methane

concentration. Generally the sulfate concentrations

around the site are high, and completely clear patterns

of sulfate depletion are not found in the composited data

set. The trends that exist in the data are best illustrated

by the contaminant distribution in individual boreholes.

These data suggest that transformation of TCE to DCE

occurred where the sulfate concentration showed some

sign of decline. Clustering of the sulfate data occurred

at sulfate concentrations of 300 micromolar (mM) or less

and oxygen concentrations of 50 mM or less.

Many of these points occurred in the transition between

the aerobic and methanogenic region. Compared with

higher sulfate concentrations at other locations, these

points tentatively indicate sulfate reduction zones, which

may be concurrent with methanogenesis (5). In particular, in Transect 4 and 5 the uppermost sample locations

Figure 1. St. Joseph Superfund site plan.

Figure 2. Scatter plot of methane and oxygen data. Figure 3. Scatter plot of sulfate and oxygen data.

68
were devoid of contaminants and were oxygenated.

Sulfate concentrations in the range of 300 to 500 μM at

these points indicate background sulfate levels.

The entire chlorinated ethene (TCE, DCEs, and VC) and

ethene data set is plotted in Figures 4 and 5 as a

chlorine number, NCl, that is defined by 

NCl = Σ wi Ci

Σ Ci

where wi is the number of chlorine atoms in molecule i

and Ci

 is the molar concentration of each ethene species. The chlorine number composites the ethene concentrations and scales them from 0 to 3. At 0 no

chlorinated species are present, and at 3 all of the

ethene is in the form of TCE. Generally, the integer

chlorine numbers (0, 1, 2, 3) are obtained with non-0

concentration only of the ethene with that number of

chlorine atoms. There are fortuitous combinations, however, of positive non-0 concentrations that give integer

chloride numbers. None of these combinations occurred

in the St. Joseph data set. 

High chlorine numbers were associated with many of

the high dissolved oxygen concentrations (Figure 4),

indicating that most of the chlorine was contained in

TCE molecules at these sampling points. Some of these

had chlorine numbers of 3, indicating that TCE was the only

species present. The majority of locations with chlorine

numbers below 3 were anaerobic, which also corresponded to methanogenic locations. The latter condition, in conjunction with the presence of the TCE

degradation products (indicated by the low chlorine

numbers), indicates degradation of the TCE. When the

data set is plotted against the methane concentration

(Figure 5), the data appeared scattered over most of the

graph. Some of the lowest chloride numbers were associated with the high methane concentrations.

Generally, many of the downgradient locations (squares

on Figures 4 and 5) showed chlorine numbers above 2

and lower methane concentrations. These data suggest

that in the downgradient transect, TCE degraded to DCE

under other than methanogenic conditions.

Data from selected borings represent the general trends

with depth in each of the transect (Figures 6 and 7). In

Transect 2, located near the presumed source of contamination, dissolved oxygen was depleted below the

60-foot depth (Figure 6). Between 45 and 60 feet, the 45-

and 55-foot depths showed significant dissolved oxygen

as well as significant methane concentrations. Sulfate

showed a weak declining trend with depth to about 70

Figure 4. Composited chlorine number plotted against oxygen

concentration.

Figure 5. Composited chlorine number plotted against methane

concentration.

Figure 6. Distribution of chloride, sulfate, dissolved oxygen,

and methane with depth in Borehole T23. Figure 7. Distribution of ethenes with depth at Borehole T23.

69
feet. Significant TCE and cis-DCE concentrations

were found only from 75 to 85 feet below the surface

(Figure 7). VC was found at concentrations of 40 μM or

less over most of the borehole. Ethene was found at

highest concentrations at the bottom of the borehole,

where methane concentrations also were highest.

Borehole T42 had the highest chlorinated ethene concentrations recorded for Transect 4, and it also represents the general chemical distribution for the

downgradient transect (Figures 8 and 9). From the water

table to the depth of 60 feet, oxygen concentrations

were high but decreasing (Figure 8). This contrasts with

the upgradient transects, which showed less consistent

depletion of oxygen near the water table. Sulfate concentrations decreased from 60 to 70 feet, roughly the

same zone in which oxygen was declining. From 70 to

85 feet, sulfate concentrations remained low but increased from 80 feet to the bottom of the borehole.

Methane was not present in the aerobic zone above 65

feet, but it increased sharply in concentration from 70 to

80 feet before decreasing. 

Figure 9 shows the distribution of the chlorinated

ethenes and ethene in T42. TCE was found from 60 feet

downward, with its maximum concentration occurring at

the 70-foot depth. The region above the 60-foot depth

was free from chlorinated ethenes, so the high sulfate

and oxygen concentrations found there correspond with

no activity due to TCE degradation. 

The cis-DCE concentration was also highest at the 70-

foot depth. Methane first appeared at 65 feet, and the

peak cis-DCE concentration occurred where sulfate

concentrations declined to the minimum. VC was found

from 65 feet to the bottom of the borehole. Ethene was

found from 70 feet downward, corresponding closely to

the most methanogenic part of the borehole.

Conclusion

Because of a variety of evidence, the data set from St.

Joseph suggest the occurrence of natural attenuation.

The composited data set indicate that, with the exception of a few points, the oxygenated and methanogenic

zones of the aquifer are clearly separated. The presence

of many methanogenic locations in the aquifer show that

the strongly reducing conditions required for production

of VC existed in the aquifer. The distribution of the

chloride number indicate that the majority of sample

locations where daughter products were present were

also anaerobic. Data from individual boreholes indicate

that high cis-DCE concentrations were commonly associated with declines in oxygen and sulfate concentrations and appeared on the upper edge of the

methanogenic zone. Generally, ethene was found in the

most methanogenic portions of the aquifer and was also

associated with relatively high VC concentrations, suggesting that the ethene production was limited to those

sample locations.

References

1. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T. Wilson. 1993.

Natural anaerobic bioremediation of TCE at the St. Joseph, Michigan, Superfund site. In: U.S. EPA. Symposium on Bioremediation

of Hazardous Wastes: Research, Development, and Field Evaluations. EPA/600/R-93/054. pp. 57-60.

2. McCarty, P.L., and J.T. Wilson. 1992. Natural anaerobic treatment

of a TCE plume St. Joseph, Michigan, NPL sites. In: U.S. EPA.

Bioremediation of Hazardous Wastes. EPA/600/R-92/126. pp. 47-50.

3. U.S. EPA. 1995. Natural bioattenuation of trichloroethene at the

St. Joseph, Michigan, Superfund site. EPA/600/V-95/001.

4. Semprini, L., P.K. Kitanidis, D. Kampbell, and J.T. Wilson. 1995.

Anaerobic transformation of chlorinated aliphatic hydrocarbons in

a sand aquifer based on spatial chemical distributions. Water Resource. Res. 31(4):1051-1062.

5. Lovley, D.R., D.F. Dwyer, and M.J. Klug. 1982. Kinetic analysis of

competition between sulfate reducers and methanogens for hydrogen in sediments. Appl. Environ. Microbiol. 43(6):1373-1379.

Figure 8. Distribution of chloride, sulfate, dissolved oxygen,

and methane with depth in Borehole T42.

Figure 9. Distribution of ethenes with depth in Borehole T42.

70
Extraction of Degradation Rate Constants From the St. Joseph, Michigan,

Trichloroethene Site

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

U.S. Environmental Protection Agency, 

National Risk Management Research Laboratory, Ada, Oklahoma

Background

Anaerobic biodegradation of trichloroethene (TCE) occurs through successive dechlorination from TCE to

dichloroethene (DCE), vinyl chloride (VC), and ethene

(1). The process produces three isomers of DCE: 1,1-

DCE, cis-1,2-DCE, and trans-1,2-DCE. Although TCE

was commonly used in industry, the DCEs were not, and

ethene would not be expected in most ground waters.

Thus, the presence of these compounds is indicative of

degradation when found in anaerobic ground waters.

Implicit in the work of Kitanidis et al. (2) and McCarty

and Wilson (3) is the fact that degradation of TCE at the

St. Joseph, Michigan, site was not predicted from theoretical considerations; rather, degradation of TCE was

established from the field data as described in these

proceedings (Weaver et al. a, this volume). The purpose

of this paper is to present estimates of averaged concentrations, mass flux, and degradation rate constants.

Ground-Water Flow

Ground water flows at the St. Joseph site from the

contaminant source toward Lake Michigan. The average

hydraulic conductivity at the site was estimated at 7.5

meters per day from a calibrated ground-water flow

model (4). The estimated travel time for TCE between

the source and the lake is approximately 18 years (Table

1). If the contamination was released only in the aqueous phase, one would expect that contaminants released 18 years or longer ago would by now have

discharged into the lake. The observed contaminant distribution suggests a continuing source, most likely a DNAPL.

Averaged Concentrations

Data were collected from the site from sets of borings

that formed four on-shore and one off-shore transects

that crossed the plume (Weaver et al. a, this volume).

These range from 130 to 855 meters from the suspected source of contamination. From the borings, a

three-dimensional view of the contamination was developed. A field gas chromatograph was used to determine

the boundaries of the plume. Sampling continued until

the entire width of the plume was crossed at each

transect. By following this procedure, the transects are

known to have contained the entire plume. This approach allows calculation of total mass that crosses

each transect and thus gives an estimate of flux of each

contaminant as a function of distance from the lake.

Transect-averaged concentration estimates were developed by using the SITE-3D graphics package (5). The

data were represented as sets of blocks that are centered around each boring. The blocks were each 5 feet

high, corresponding to the length of the slotted auger. At

each transect, the average concentration was calculated

Table 1. Attenuation of the Chlorinated Ethenes Along the

Length of the Plume

Average Concentration (μg/L)

Highest Concentration (μg/L)

Distance

From

Source

(m)

Transport

Time

(y)

Transect

Width

(m) TCE cis-DCE

Vinyl

Chloride

130 3.2 108 6,500

68,000 

8,100

128,000 

 930

4,400 

390 9.7 150 520

8,700

 830

9,800

 450

1,660 

550 12.5 192 15

 56

 18

 870

 106

 205

855 17.9 395 <1

 0.4

 <1

 0.8

 <1

 0.5

71
by summing over the blocks and dividing by the area of

the transects.

In Table 2, concentration estimates are presented for the

perpendicular transects ordered from furthest upgradient

(Transect 2) to furthest downgradient (Transect 5). The

concentration estimates are based only on blocks from

the anaerobic portion of the aquifer (and thus differ from

the averages in Table 1). All of the chlorinated ethenes

show decreasing concentration with distance downgradient; thus, all of the rate coefficients developed below reflect

a net loss of the species. The chloride concentrations

increase downgradient as expected from the dechlorination of the ethenes. On a molar basis, however, the

increase in average chloride concentration is greater

than that which would result from dechlorination alone.

Mass Flux

The concentration results (Table 2) show that by the time

the contaminants reach the lake, their concentrations

are reduced to very low levels. It is equally important to

determine the mass of chemicals released to the lake

per year. Given the approximate ground-water velocities

and the contaminant concentrations in the transects, an

estimate of the mass flux of chemicals can also be

estimated. Advective mass fluxes of each chemical were

estimated per transect by multiplying the seepage velocity by concentration in each block formed, using

SITE-3D. The results are given in Table 3, which shows

a decline in the mass flux of each chlorinated ethene.

The flux reduction ranged from a factor of 10 to 123. The

flux of methane showed no consistent pattern. Chloride

flux increased beyond Transect 1.

Degradation Rates

The transport of each chemical is parametrized by the

ground-water flow velocity, the retardation coefficient,

the dispersivities, and the decay constant. Specifically,

two-dimensional solute transport with first-order decay

obeys

R ∂c

∂t

= Dxx

∂2

c

∂x2 + Dyy

∂2

c

∂y2 − v

∂c

∂x

− λ∗ c

(Eq. 1)

where R is the retardation coefficient; c is the concentration; t is time; Dxx and Dyy are the longitudinal and

transverse dispersion coefficients, respectively; x is longitudinal distance; y is the distance transverse to the

plume centerline in the horizontal plane; v is the seepage velocity; and λ* is the first-order decay constant.

First-order decay is assumed for this analysis because

it is the usual way to report degradation rates of chlorinated hydrocarbons (6). This form of the transport equation assumes that ground-water flow is uniform and

aligned with the axis of the plume, as observed for the

plume. This assumption also allows application of analytic solutions as described in the appendix.

The concentration of dissolved chemicals can change

because of the effect of the terms on the right-hand side

of Equation 1. Dispersion is used to characterize apparent physical dilution in aquifers. Dispersion is currently

Table 2. Transect-Averaged Concentrations (μg/L) From the

Anaerobic Zone

Chemical Transect 2 Transect 4 Transect 5

Lake

Transect

TCE 7,411 864 30.1 (1.4)

cis-DCE 9,117 1,453 281 (0.80)

t-DCE 716 34.4 5.39 (1.1)

1,1-DCE 339 24.3 2.99 blq

VC 998 473 97.7 (0.16)

Ethene 480 297 24.2 No data

Sum of the

ethenes

19,100 3,150 442 (3.5)

Chloride 65,073 78,505 92,023 44,418

Note: Values in parentheses were based on one or more estimated

values; blq indicates no detection above the limit of quantitation.

Table 3. Flux Estimates for Transects 1, 2, 4, and 5

Mass Flux (kg/y)

Transect TCE cis-DCE VC Ethene

Total

Ethenes Methane Chloride

1 (August-September 1991) 50.0 45.2 16.8 7.95 125 49.2 545

2 (August-September 1991) 117 133 16.8 7.60 283 65.7 1,456

4 (March 1992) 30.9 41.7 3.87 10.8 88.4 101 4,610

5 (April 1992) 0.95 10.0 1.68 0.164 13.1 46.7 5,290

Reduction ratio 123 13 10 46 22

Note: The reduction ratio is the ratio of mass flux at Transect 2 to that at Transect 5. 

72
understood to result primarily from ground-water flow

through heterogeneous materials. In multidimensional

flow, advection can cause concentrations to decrease

because of the divergence of flow lines. Advection does

not directly change concentrations in one-dimensional

flow but influences the contribution of dispersion. Decay

changes concentration through removal of mass from

the aquifer.

The significance of these observations is that when

presented with a set of contaminant concentrations, the

distribution of contamination may depend on physiochemical and biological processes. Observed concentrations in themselves do not indicate the contribution of

each process to the plume shape. Extraction of apparent

rates from the field data needs to account for the multiple processes. In Table 4, estimated rate constants are

given for St. Joseph. These constants were determined

from the solution of the transport equation presented in

the appendix. The solution included advection, retardation, longitudinal and transverse dispersion, and firstorder loss. Inclusion of transverse dispersion is

important because this characterizes downgradient

spreading of the plume. The observed widths of the

plume at St. Joseph are given in Table 1 and were used

to estimate the transverse dispersivity according to the

procedure given in the appendix. The effect of transverse dispersivity on the estimated rate constants, however, decreases as the plume widens and the centerline

concentrations decrease. Longitudinal dispersivity has

been shown to have a minor impact on the estimated

rate constants at distances between transects on the

order of 100 meters (7).

The rates given in Table 5 are called net rates because,

for the daughter products, the observed concentrations

are a result of production of the daughter from both

decay of the parent and decay of the daughter itself. The

gross rate of decay of the daughter (Table 4) does not

include its production and was determined by the procedure given in the appendix. The two rates are the same for

TCE, since no production of TCE occurred. The gross

rates are, as expected, higher than the net rates, because production of a compound must be balanced by

high gross rates to attain the observed net rate.

Conclusion

The western TCE plume at St. Joseph, Michigan,

showed a decrease of maximum TCE concentration by

a factor of 50,000 from the furthest upgradient transect

to the lake transect. Concentrations of each contaminant declined to values below the respective maximum

contaminant levels when sampled from the lake sediments. Mass fluxes decrease by factors of 10 to 123

from the source to the last on-shore transect (Transect

5); thus, not only do the concentrations decline, but

so does the loading in the ground water. The reduction

in loading is attributed to degradation, because of the

geochemical evidence presented (Weaver et al. a, this

volume).

Further, when site-specific estimates of the transport

parameters are used in solutions of the transport equations, the apparent reduction in concentration is only

accounted for by loss of mass. These apparent degradation rate constants were calculated from the St.

Joseph data set through application of a two-dimensional analytical solution of the transport equation. Since

transverse spreading of the plume reduces the contaminant concentrations, the effect of transverse dispersivity

was included in the analysis.

Appendix

Extraction of Rate Constants via

Two-Dimensional, Steady-State Transport

Analysis

The two-dimensional transport equation, subject to the

boundary conditions

c (x,y,0) = 0

c (0,y,t) = co exp (

−y2

2 σ2)

c (∞,y,t) = c (x,−∞,t) = c (x,∞,t) = 0

(Eq. 2)

Table 4. Apparent Degradation Rate Constants (One Per

Year) From the Two-Dimensional Model (Equation 3)

and the Gross Rate Correction Given by Equation 7

Chemical

Transect

2 to 4

Transect

4 to 5

Transect

5 to Lake

TCE 0.30 1.7 1.7

cis-DCE 0.54 1.1 4.0

Vinyl chloride 2.6 3.1 20 

Table 5. Net Apparent Degradation Rate Constants 

(One Per Year) From the Two-Dimensional Model 

(Equation 3)

Chemical

Transect

2 to 4

Transect

4 to 5

Transect

5 to Lake

TCE 0.30 1.7 1.7

cis-DCE 0.26 0.58 3.3

Vinyl chloride 0.15 0.78 2.6

73
has the approximate steady state solution (9)

(Eq. 3)

Vertically averaged concentrations and the distances

between each borehole were used to develop the

boundary condition (c(0,y,t) in Equation 2) for application

of Equation 3. The unknown parameters are the upgradient peak concentration, co, and the standard deviation, σ, of the distribution. Since the width of the plume,

W, was established via the field sampling program, the

standard deviation of the distribution can be estimated

as W = 6σ. A mass balance can then be solved for the

peak concentration of the Gaussian distribution, co, from

∫nc^dy = ∫ nco exp(

−y2

2σ2) dy = nco σ √2π

(Eq. 4)

where n is the porosity, c

^ is the vertically averaged concentration, and the y coordinate runs parallel to the transect.

The transverse dispersivity can also be estimated from the

measured widths of the transects. The width of a contaminant

distribution is related to the transverse dispersivity through

Dyy = ayy v = 1 dσ2

2 dt

(Eq. 5)

where ayy is the transverse dispersivity. By applying Equation 5 in a discrete form and substituting ∆ t = ∆ xR/v, an

expression for ayy is obtained in terms of the seepage

velocity, retardation coefficient, distance between transects (∆ x), and change in variance of the Gaussian

distributions for the transect concentrations (∆ σ2

):

ayy = 1 ∆σ2

2R ∆x

(Eq. 6)

The only remaining unknown in Equation 3 is the decay

constant λ*, which is determined through a bisection

search. Table 5 gives the rate constants from the twodimensional model.

Net and Gross Decay Rates

The rate constants derived from the solution (Equation

3 and Table 5) are net rates that include the production

and decay of a given daughter product. It is necessary

to separate production of the compound from its decay

to estimate the gross apparent decay rates for cis-DCE,

t-DCE, 1,1-DCE, and VC. Previous work (7) used a

reaction rate model that simultaneously solved ordinary

differential equations for this purpose. Here, simplified

expressions for the rates were used to estimate the

apparent decay rates:

λj(n) = fj λj+1(n) S + λj(g) (Eq. 7)

where λj(n) is the net decay rate determined by Equation

3, fj

 is the fraction of an isomer (j) produced from the

degradation of the parent (j+1), λj+1(n) is the apparent

decay rate of the parent defined from Equation 3, S is

the ratio of molar concentration of parent j+1 to daughter

j, and λj(g) is the gross apparent decay rate of daughter

j. For the DCE isomers, fj

 is approximated by the average ratio of an isomer j to the sum of the DCEs over the

pairs of transects. For VC, fj

 is equal to 1.0. The gross

apparent decay rates for cis-DCE, t-DCE, 1,1-DCE, and

VC appear in Table 4. Although Equation 7 is concentration dependent because S was assumed to be the

average of the up- and downgradient ratios, the results

presented in Table 4 are essentially the same as determined from the reaction rate model (8).

References

1. McCarty, P.L., and L. Semprini. 1994. Ground-water treatment for

chlorinated solvents. In: Norris R.D., R.E. Hinchee, R. Brown, P.L.

McCarty, L. Semprini, J.T. Wilson, D.H. Kampbell, M. Reinhard,

E.J. Bouwer, R.C. Borden, T.M. Vogel, J.M. Thomas, and C.H.

Ward, eds. Handbook of bioremediation. Chelsea, MI: Lewis Publishers. pp. 87-116.

2. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T. Wilson. 1993.

Natural anaerobic bioremediation of TCE at the St. Joseph, Michigan, Superfund site. In: U.S. EPA. Symposium on Bioremediation

of Hazardous Wastes: Research, Development, and Field Evaluations. EPA/600/R-93/054. pp. 57-60.

3. McCarty, P.L., and J.T. Wilson. 1992. Natural anaerobic treatment of

a TCE plume at the St. Joseph, Michigan, NPL site. In: U.S. EPA.

Bioremediation of hazardous wastes. EPA/600/R-92/126. pp. 47-50.

4. Tiedeman, C., and S. Gorelick. 1993. Analysis of uncertainty in

optimal groundwater contaminant capture design. Water Resour.

Res. 29:2139-2153.

5. U.S. EPA. 1996. Animated three-dimensional display of field data

with SITE-3D: User’s guide for version 1.00. Technical report.

EPA/600/R-96/004.

6. Rifai, H.S., R.C. Borden, J.T. Wilson, and C.H. Ward. 1995. Intrinsic bioattenuation for subsurface restoration. In: Hinchee, R.E., J.T.

Wilson, and D.C. Downey, eds. Intrinsic bioremedation, Vol. 3.

Columbus, OH: Battelle Press. pp. 1-29.

74
7. Weaver, J.W., J.T. Wilson, D.H. Kampbell, and M.E. Randolph.

1995. Field-derived transformation rates for modeling natural

bioattenuation of trichloroethene and its degradation products.

Presented at the Next Generation of Computational Models Computational Methods, August 17-19, Bay City, MI. Society of Industrial and Applied Mathematics.

8. Smith, V.J., and R.J. Charbeneau. 1990. Probabilistic soil contamination exposure assessment procedures. J. Environ. Engineer.

116(6):1143-1163.

75
Natural Attenuation of Chlorinated Aliphatic Hydrocarbons at Plattsburgh 

Air Force Base, New York

Todd H. Wiedemeier

Parsons Engineering Science, Inc., Denver, Colorado

John T. Wilson and Donald H. Kampbell

U.S. Environmental Protection Agency, National Risk Management Research Laboratory,

Subsurface Protection and Remediation Division, Ada, Oklahoma

Introduction

Activities at a former fire training area (Site FT-002) at

Plattsburgh Air Force Base (AFB) in New York resulted

in contamination of shallow soils and ground water with

a mixture of chlorinated solvents and fuel hydrocarbons.

Ground water contaminants include trichloroethene

(TCE), cis-1,2-dichloroethene (cis-1,2-DCE), vinyl chloride, and benzene, toluene, ethylbenzene, and xylenes

(BTEX). Table 1 contains contaminant data for selected

wells at the site.

Contaminant plumes formed by chlorinated aliphatic hydrocarbons (CAHs) dissolved in ground water can exhibit three types of behavior based on the amount and

type of primary substrate present in the aquifer. Type 1

behavior occurs where anthropogenic carbon such as

BTEX or landfill leachate is being utilized as the primary

substrate for microbial degradation. Such plumes typically are anaerobic, and the reductive dechlorination of

highly chlorinated CAHs introduced into such a system

can be quite rapid. Type 2 behavior occurs in areas that

are characterized by high natural organic carbon concentrations and anaerobic conditions. Under these conditions, microorganisms utilize the natural organic carbon

as a primary substrate; if redox conditions are favorable,

highly chlorinated CAHs introduced into this type of

system will be reductively dechlorinated. Type 3 behavior occurs in areas characterized by low natural organic

carbon concentrations, low anthropogenic carbon concentrations, and aerobic or weakly reducing conditions.

Biodegradation of CAHs via reductive dechlorination will

not occur under these conditions. Biodegradation of the

less chlorinated compounds such as vinyl chloride, however, can occur via oxidation.

Plattsburgh AFB is located in northeastern New York

State, approximately 26 miles south of the Canadian

border and 167 miles north of Albany. Site FT-002 (Figure 1) is located in the northwest corner of the base on

a land surface that slopes gently eastward toward the

confluence of the Saranac and the Salmon Rivers, approximately 2 miles east of the site. The site, which is

approximately 700 feet wide and 800 feet long, was

used to train base and municipal fire-fighting personnel

from the mid-1950s until it was permanently closed to

fire-training activities in May 1989.

Four distinct stratigraphic units underlie the site: sand,

clay, till, and carbonate bedrock. Figure 2 shows three

of the four stratigraphic units at the site. The sand unit

consists of well-sorted, fine- to medium-grained sand

with a trace of silt, and generally extends from ground

surface to as much as 90 feet below ground surface

(bgs) in the vicinity of the site. A 7-foot thick clay unit has

been identified on the eastern side of the site. The

thickness of the clay on the western side of the site has

not been determined. A 30- to 40-foot thick clay till unit

is also present from 80 to 105 feet bgs in the vicinity of

the site. Bedrock is located approximately 105 feet bgs.

Ground-Water Hydraulics

The depth to ground water in the sand aquifer ranges

from 45 feet bgs on the west side of the site to zero on

the east side of the runway, where ground water discharges to a swamp (Figure 2). Ground-water flow at the

site is to the southeast, with the average gradient approximately 0.010 foot per foot (ft/ft). Hydraulic conductivity of the upper sand aquifer was measured using

constant drawdown tests and rising head tests. Hydraulic conductivity values for the unconfined sand aquifer

76
underlying the site range from 0.059 to 90.7 feet per day

(ft/day). The average hydraulic conductivity for the site

is 11.6 ft/day. Freeze and Cherry (1) give a range of

effective porosity for sand of 0.25 to 0.50. Effective

porosity was assumed to be 0.30. The horizontal gradient of 0.010 ft/ft, the average hydraulic conductivity

value of 11.6 ft/day, and an effective porosity of 0.30

yields an average advective ground-water velocity for

the unconfined sand aquifer of 0.39 ft/day, or approximately 142 feet per year. Because of low background

total organic carbon (TOC) concentrations at the site,

retardation is not considered to be an important transport parameter.

Ground Water and Light

Nonaqueous-Phase Liquid Chemistry

Contaminants

Figure 1 shows the approximate distribution of light

nonaqueous-phase liquid (LNAPL) at the site. This LNAPL

is a mixture of jet fuel and waste solvents that partitions

BTEX and TCE to ground water. Analysis of the LNAPL

shows that the predominant chlorinated solvents are

tetrachloroethane (PCE) and TCE; DCE and vinyl chloride are not present in measurable concentrations. For

the most part, ground water beneath and downgradient

from the LNAPL is contaminated with dissolved fuel-related compounds and solvents consistent with those

identified in the LNAPL. The most notable exceptions

are the presence of cis-1,2-DCE and vinyl chloride,

which, because of their absence in the LNAPL, probably

were formed by reductive dechlorination of TCE.

The dissolved BTEX plume currently extends approximately 2,000 feet downgradient from the site, and has

a maximum width of about 500 feet. Total dissolved

BTEX concentrations as high as 17 milligrams per liter

(mg/L) have been observed in the source area. Figure

3 shows the extent of BTEX dissolved in ground water.

As indicated on this map, dissolved BTEX contamination is migrating to the southeast in the direction of

ground-water flow. Five years of historical data for the

site show that the dissolved BTEX plume is at steadystate equilibrium and is no longer expanding.

Detectable concentrations of dissolved TCE, DCE, and

vinyl chloride currently extend approximately 4,000 feet

Table 1. Analytical Data, Plattsburgh Air Force Base

Point Date

Distance

From

Source

(feet)

TMB

(μg/L)

BTEX

(μg/L)

TCE

(μg/L)

Total

DCEa

(μg/L)

Vinyl

Chloride 

(μg/L)

Methane

(μg/L)

Ethene

(μg/L)

Chloride 

(mg/L)

Dissolved

Oxygen

(mg/L)

Nitrate

(mg/L)

Iron(II) 

(mg/L)

Sulfate 

(mg/L)

Hydrogen

(nM)

Total

Organic 

Carbon 

(mg/L)

A Aug.

95

 0 1,757 16,790 25,280 51,412 0 1,420 < 0.001 63 0.1 0.2 4.0 5.5 6.70 80

May

96

 828 6,598 580 12,626 0 1,600 < 0.001 82 0.5 0.0 45.6 1.0 2.00 94

B Aug.

95

 970 491 3,060 2 14,968 897 305 35.00 48 0.5 0.2 15.3 0.0 1.66 30

May

96

 463 4,198 1 9,376 1,520 339 13.00 43 0.1 0.0 16.0 0.0 1.40 31

C Aug.

95

1,240 488 3,543 3 10,035 1,430 1,010 182.00 46 0.4 0.2 13.8 0.0 NA 21

May

96

 509 3,898 1 10,326 1,050 714 170.00 57 0.2 0.0 19.3 0.0 11.13 24

D Aug.

95

2,050 NA NA NA NA NA NA NA NA NA NA NA NA NA NA 

May

96

 9 89 0 1,423 524 617 4.00 14 0.2 0.1 2.5 1.5 NA 14

E Aug.

95

2,560 0 40 24 2,218 8 3,530 < 0.001 20 0.9 0.3 0.7 0.5 NA 8

May

96

 0 40 17 1,051 12 1,800 < 0.001 18 0.1 0.0 0.0 1.0 0.81 8

F Aug.

95

3,103 0 2 1 226 5 115 < 0.001 3 0.4 10.4 0.0 14.7 0.22 NA 

May

96

 0 2 0 177 4 44 < 0.001 3 0.2 9.5 0.1 14.4 0.25 NA 

a

 Greater than 99% of DCE is cis-1,2-DCE.

NA = Not analyzed.

Point A = MW-02-108, B = MW-02-310, C = 84DD, D = 84DF, E = 34PLTW12, F = 35PLTW13.

77
downgradient from FT-002. Concentrations of TCE,

DCE, and vinyl chloride as high as 25 mg/L, 51 mg/L,

and 1.5 mg/L, respectively, have been observed at the

site. As stated previously, no DCE was detected in the

LNAPL plume at the site, and greater than 99 percent of

the DCE found in ground water is the cis-1,2-DCE isomer. Figure 3 shows the extents of CAH compounds

dissolved in ground water at the site. As indicated on

this map, contamination is migrating to the southeast in

the direction of ground-water flow. Five years of historical data for the site show that the dissolved CAH plume

is at steady-state equilibrium and is no longer expanding.

Indicators of Biodegradation

The distribution of electron acceptors used in microbially

mediated oxidation-reduction reactions is shown in Figure 4. Electron acceptors displayed in this figure include

dissolved oxygen, nitrate, and sulfate. There is a strong

correlation between areas with elevated BTEX concentrations and areas with depleted dissolved oxygen, nitrate, and sulfate. The absence of these compounds in

contaminated ground water suggests that aerobic respiration, denitrification, and sulfate reduction are working

to biodegrade fuel hydrocarbons at the site. Background

dissolved oxygen, nitrate, and sulfate concentrations

are on the order of 10 mg/L, 10 mg/L, and 25 mg/L,

respectively.

Figure 5 shows the distribution of metabolic byproducts

produced by microbially mediated oxidation-reduction

reactions that biodegrade fuel hydrocarbons. Metabolic

byproducts displayed in this figure include iron(II) and

methane (Figure 5). There is a strong correlation between areas with elevated BTEX concentrations and

areas with elevated iron(II) and methane. The presence

of these compounds in concentrations above background in contaminated ground water suggests that

iron(III) reduction and methanogenesis are working to

biodegrade fuel hydrocarbons at the site. Background

iron(II) and methane concentrations are less than 0.05

and 0.001 mg/L, respectively.

The pE of ground water is also shown in Figure 5. Areas

of low pE correspond to areas with contamination, indicating that biologically mediated oxidation-reduction reactions are occurring in the area with ground-water

contamination.

Figure 1. Site map.

78
Figure 3 illustrates the distribution of chloride in ground

water and compares measured concentrations of total

BTEX and CAHs in the ground water with chloride and

ethene. There is a strong correlation between areas with

contamination and areas with elevated chloride and

ethene concentrations relative to measured background

concentrations. The presence of elevated concentrations of chloride and ethene in contaminated ground

water suggests that TCE, DCE, and vinyl chloride are

being biodegraded. Background chloride concentrations

at the site are approximately 2 mg/L; background ethene

concentrations at the site are less than 0.001 mg/L.

Dissolved hydrogen concentrations can be used to determine the dominant terminal electron-accepting process in an aquifer. Table 2 presents the range of hydrogen

concentrations for a given terminal electron-accepting

process. Much research has been done on the topic of

using hydrogen measurements to delineate terminal

electron-accepting processes (2-4). Table 1 presents

hydrogen data for the site.

Biodegradation Rate Constant Calculations

Apparent biodegradation rate constants were calculated

using the method presented in Wiedemeier et al. (5, 6)

for trimethylbenzene (TMB). A modified version of this

method that takes into account the production of chloride during biodegradation also was used to calculate

approximate biodegradation rates. Table 3 presents the

results of these rate-constant calculations.

Primary Substrate Demand for Reductive

Dechlorination

For reductive dechlorination to occur, a carbon source

that can be used as a primary substrate must be present

in the aquifer. This carbon substrate can be in the form

of anthropogenic carbon (e.g., fuel hydrocarbons) or

native organic material.

Figure 2. Hydrogeologic section.

Table 2. Range of Hydrogen Concentrations for a Given

Terminal Electron-Accepting Process

Terminal Electron

Accepting Process

Hydrogen Concentration 

(nM/L)

Denitrification < 0.1

Iron(III) reduction 0.2 to 0.8

Sulfate reduction 1 to 4

Methanogenesis > 5

79
Figure 3. Chlorinated solvents and byproducts (1995).

Figure 4. BTEX and electron acceptors (1995).

80
Reductive Dechlorination Supported by Fuel

Hydrocarbons (Type 1 Behavior)

Fuel hydrocarbons are known to support reductive

dechlorination in aquifer material (7). Equation 1 below

describes the oxidation of BTEX compounds (approximated as CH) to carbon dioxide during reduction of

carbon to chlorine bonds (represented as C-Cl) to carbon to hydrogen bonds (represented as C-H).

CH + 2H2O + 2.5C-Cl → CO2 + 2.5H+

 + 

2.5Cl-

 + 2.5C-H (Eq. 1)

Based on Equation 1, each 1.0 milligram (mg) of BTEX

that is oxidized via reductive dechlorination requires the

consumption of 6.8 mg of organic chloride and the liberation of 6.8 mg of biogenic chloride. PCE loses two

C-Cl bonds while being reduced to vinyl chloride. Based

on Equation 1, 1⁄2 x 2.5 = 1.25 moles of TCE that would

have to be reduced to vinyl chloride to oxidize 1 mole of

BTEX to carbon dioxide. Therefore, each 1.0 mg of

BTEX oxidized would consume 12.6 mg of TCE. If DCE

were reduced to vinyl chloride, each 1.0 mg of BTEX

oxidized would consume 18.6 mg of DCE. To be more

conservative, these calculations should be completed

assuming that TCE and DCE are reduced to ethene.

Because the amount of ethene produced is trivial compared with the amount of TCE and DCE destroyed,

however, we have omitted this step here.

Reductive Dechlorination Supported by

Natural Organic Carbon (Type 2 Behavior)

Wershaw et al. (8) analyzed dissolved organic material

in ground water underneath a dry well that had received

TCE discharged from the overflow pipe of a degreasing

unit. The dissolved organic material in ground water

exposed to the TCE was 50.57 percent carbon, 4.43

percent hydrogen, and 41.73 percent oxygen. The elemental composition of this material was used to calculate an empirical formula for the dissolved organic

matter, and to estimate the number of moles of C-Cl

bonds required to reduce one mole of dissolved organic

carbon in this material:

C1.0H1.051O0.619 + 1.38H2O + 1.91C-Cl → CO2 + 

1.91Cl-

 + 1.91C-H + 1.91H+ (Eq. 2)

Based on Equation 2, each 1.0 mg of dissolved organic

carbon that is oxidized via reductive dechlorination requires the consumption of 5.65 mg of organic chloride

and the liberation of 5.65 mg of biogenic chloride. Using

Equation 2, 1⁄2 x 1.91 = 0.955 moles of TCE that would

have to be reduced to vinyl chloride to oxidize 1 mole of

organic carbon to carbon dioxide. Therefore, 1.0 mg of

organic carbon oxidized would consume 10.5 mg of

TCE. If DCE were reduced to vinyl chloride, each 1.0 mg of

organic carbon oxidized would consume 15.4 mg of DCE.

Table 4 compares the electron donor demand required

to dechlorinate the alkenes remaining in the plume with

the supply of potential electron donors. Table 3 reveals

that removal of TCE and cis-1,2-DCE slows or ceases

between points C and E. This correlates with the exhaustion of BTEX in the plume. Over this interval, the

supply of BTEX is a small fraction of the theoretical

demand required for dechlorination. There are adequate

supplies of native organic matter, suggesting that native

organic matter may not be of sufficient nutritional quality

to support reductive dechlorination in this aquifer.

Discussion and Conclusions

Available geochemical data indicate that the geochemistry of ground water in the source area and about 1,500

feet downgradient is significantly different than the ground

water found between 1,500 and 4,000 feet downgradient from the source. Near the source the plume exhibits

Type 1 behavior. At about 1,500 feet downgradient from

Table 3. Approximate First-Order Biodegradation Rate Constants

Compound

Correction 

Method

A - B

0 to

970 feet

(1/year)

B - C

970 to

1,240 feet

(1/year)

C - E

1,240 to

2,560 feet

(1/year)

TCE Chloride 1.27 0.23 -0.30

TMB 1.20 0.52 NA 

Average 1.24 0.38 -0.30

DCE Chloride 0.06 0.60 0.07

TMB 0.00 0.90 NA 

Average 0.03 0.75 0.07

Vinyl chloride Chloride 0.00 0.14 0.47

TMB 0.00 0.43 NA 

Average 0.00 0.29 0.47

BTEX Chloride 0.13 0.30 0.39

TMB 0.06 0.60 NA 

Average 0.10 0.45 0.39

Table 4. Comparison of the Estimated Electron Donor

Demand To Support Reductive Dechlorination to the

Supply of BTEX and Native Organic Carbon

Point

Chloride

(mg/L)

Organic 

Chloride

(mg/L)

BTEX

Available

(mg/L)

BTEX

Demand

(mg/L)

TOC

Supply

(mg/L)

Organic 

Carbon

Demand 

(mg/L)

A 63 58.1 16.8 8.5 80.4 10.3 

B 43 7.72 4.2 1.13 31.1 1.37

C 57 8.26 3.9 1.21 24.3 1.46

D 13.6 1.34 0.09 0.20 13.8 0.24

E 18.4 0.78 0.04 0.114 8.2 0.14

81
the source, the plume reverts to Type 3 behavior. Figure

6 shows the zones of differing behavior at the site.

Type 1 Behavior

In the area extending to approximately 1,500 feet downgradient from the former fire- training pit (source area),

the dissolved contaminant plume consists of commingled BTEX and TCE and is characterized by anaerobic

conditions that are strongly reducing (i.e., Type 1 behavior). Dissolved oxygen concentrations are on the order

of 0.1 mg/L (background = 10 mg/L), nitrate concentrations are on the order of 0.1 mg/L (background = 10

mg/L), iron(II) concentrations are on the order of 15 mg/L

(background = less than 0.05 mg/L), sulfate concentrations are less than 0.05 mg/L (background = 25 mg/L),

and methane concentrations are on the order of

3.5 mg/L (background = mg/L). Hydrogen concentrations in the source area range from 1.4 to 11 nanomoles

(nM). As shown by Table 2, these hydrogen concentrations are indicative of sulfate reduction and methanogenesis, even though there is no sulfate available and

relatively little methane is produced. Thus, reductive

dechlorination may be competitively excluding these

processes.

In this area BTEX is being used as a primary substrate,

and TCE is being reductively dechlorinated to cis-1,2-

DCE and vinyl chloride. This is supported by the fact that

no detectable DCE or vinyl chloride was found in the

LNAPL present at the site and is strong evidence that

the DCE and vinyl chloride found at the site are produced by the biogenic reductive dechlorination of TCE.

Furthermore, the dominant isomer of DCE found at the

site is cis-1,2-DCE, the isomer preferentially produced

during reductive dechlorination. Average calculated firstorder biodegradation rate constants in this zone are as

high as 1.24, 0.75, and 0.29 per year for TCE, cis-1,2-

DCE, and vinyl chloride, respectively. Figure 6 shows

the approximate extent of this type of behavior. Because

reductive dechlorination of vinyl chloride is slower than

direct oxidation, vinyl chloride and ethene are accumulating in this area (Figure 7).

Type 3 Behavior

Between 1,500 and 2,000 feet downgradient from the

source area, the majority of the BTEX has been biodegraded and the system begins to exhibit Type 3 behavior. Dissolved oxygen concentrations are on the order of

0.5 mg/L (background = 10 mg/L). Nitrate concentrations

start increasing downgradient of where Type 3 behavior

begins and are near background levels of 10 mg/L at the

downgradient extent of the CAH plume. Iron(II) concentrations have significantly decreased and are on the

Figure 5. BTEX and metabolic byproducts (1995).

82
order of 1 mg/L (background = less than 0.05 mg/L).

Sulfate concentrations start increasing to 15 mg/L at the

downgradient extent of the CAH plume. Methane concentrations are the highest in this area but could have

migrated from upgradient locations. The hydrogen concentrations at Points E and F are 0.8 nM and 0.25 nM,

respectively, suggesting that the dominant terminal electron-accepting process in this area is iron(III) reduction.

These conditions are not optimal for reductive dechlorination, and it is likely that vinyl chloride is being oxidized

via iron (III) reduction or aerobic respiration. Average

calculated rate constants in this zone are -0.3, 0.07, and

0.47 per year for TCE, cis-1,2-DCE, and vinyl chloride,

respectively. The biodegradation rates of TCE and DCE

slow because reductive dechlorination stops when the

plume runs out of primary substrate (i.e., BTEX). The

rate of vinyl chloride biodegradation in this area increases, probably because vinyl chloride is being oxidized. Because biodegradation of vinyl chloride is faster

under Type 3 geochemical conditions than the biodegradation of other CAH compounds, the accumulation of

vinyl chloride ceases and the accumulated vinyl chloride

rapidly degrades. Ethene concentrations also begin to

decrease because ethene is no longer being produced

from the reductive dechlorination of vinyl chloride (Figure 7).

Figure 6. Zonation of CAH plume.

Figure 7. Plot of TCE, DCE, and ethene versus distance downgradient.

83
References

1. Freeze, R.A., and J.A. Cherry. 1979. Groundwater. Englewood

Cliffs, NJ: Prentice-Hall, Inc.

2. Lovley, D.R., and S. Goodwin. 1988. Hydrogen concentrations as

an indicator of the predominant terminal electron-accepting reaction in aquatic sediments. Geochim. Cosmochim. Acta 52:2993-

3003.

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

dissolved H2 concentrations to determine distribution of microbially

catalyzed redox reactions in anoxic ground water. Environ. Sci.

Technol. 28(7):1205-1210.

4. Chapelle, F.H., P.B. McMahon, N.M. Dubrovsky, R.F. Fujii, E.T.

Oaksford, and D.A. Vroblesky. 1995. Deducing the distribution of

terminal electron-accepting processes in hydrologically diverse

groundwater systems. Water Resour. Res. 31:359-371.

5. Wiedemeier, T.H., J.T. Wilson, D.H. Kampbell, R.N. Miller, and J.E.

Hansen. 1995. Technical protocol for implementing intrinsic remediation with long-term monitoring for natural attenuation of fuel

contamination dissolved in groundwater. San Antonio, TX: U.S. Air

Force Center for Environmental Excellence.

6. Wiedemeier, T.H., M.A. Swanson, J.T. Wilson, D.H. Kampbell, R.N.

Miller, and J.E. Hansen. 1996. Approximation of biodegradation

rate constants for monoaromatic hydrocarbons (BTEX) in groundwater. Ground Water Monitoring and Remediation. Summer.

7. Sewell, G.W., and S.A. Gibson. 1991. Stimulation of the reductive

dechlorination of tetrachloroethene in anaerobic aquifer microcosms by the addition of toluene. Environ. Sci. Technol. 25:982-

984.

8. Wershaw, R.L., G.R. Aiken, T.E. Imbrigiotta, and M.C. Goldberg.

1994. Displacement of soil pore water by trichloroethylene. J. Environ. Quality 23:792-798.

84
Case Study: Natural Attenuation of a Trichloroethene Plume 

at Picatinny Arsenal, New Jersey

Thomas E. Imbrigiotta and Theodore A. Ehlke

U.S. Geological Survey, West Trenton, New Jersey

Barbara H. Wilson and John T. Wilson

U.S. Environmental Protection Agency, Ada, Oklahoma

Introduction

Past efforts to clean up aquifers contaminated with chlorinated solvents typically have relied on engineered remediation systems that were costly to build and operate.

Recently, environmental regulatory agencies have begun to give serious consideration to the use of natural

attenuation as a more cost-effective remediation option.

The successful use of natural attenuation to remediate

chlorinated-solvent contaminated sites depends on understanding the processes that control the transport and

fate of these compounds in the ground-water system. 

To this end, the U.S. Geological Survey, as part of its

Toxic Substances Hydrology Program, has been conducting an interdisciplinary research study of groundwater contamination by chlorinated solvents at Picatinny

Arsenal, New Jersey. The objectives of the study are to

identify and quantify the physical, chemical, and biological processes that affect the transport and fate of chlorinated solvents, particularly trichloroethene (TCE), in

the subsurface; determine the relative importance of

these processes at the site; and develop predictive models of chlorinated-solvent transport that may have transfer value to other solvent-contaminated sites in similar

hydrogeologic environments. 

This paper reports on the results of efforts to identify and

quantify the natural processes that introduce and remove TCE to and from the plume at Picatinny Arsenal,

and to determine which natural TCE-attenuation mechanisms are the most important on a plume-wide basis.

Geohydrology

Picatinny Arsenal is a weapons research and development facility located in a narrow glaciated valley in north

central New Jersey (Figure 1). The site is underlain by

a 15- to 20-meter thick unconfined aquifer consisting

primarily of fine to coarse sand with some gravel and

discontinuous silt and clay layers. Ground-water flows

from the sides of the valley toward the center, where it

discharges to Green Pond Brook. Within the unconfined

aquifer, flow is generally horizontal, with some downward flow near the valley walls and upward flow near

Green Pond Brook. Estimated ground-water flow velocities range from 0.3 to 1.0 meters per day (m/d) at the

site on the basis of hydraulic conductivities that range

from 15 to 90 m/d, gradients that range from 1.5 to 3.0

m per 500 m, and an average porosity of 0.3 (1-4).

Ground-Water Contamination

Ground water at Picatinny Arsenal was contaminated

over a period of 30 years as a result of activities associated with metal plating and degreasing operations in

Building 24 (5, 6). The areal and vertical extent of TCE

contamination at the site, determined using data from

October and November 1991, is shown in Figure 1.

Areally, the plume, as defined by the 10 micrograms per

liter (μg/L) line, extends about 500 m from Building 24

to Green Pond Brook and is approximately 250 m wide

where it enters the brook. Vertically, TCE contamination

is found at shallow depths near the source, over the

entire 15- to 20-m thickness of the unconfined aquifer in

the plume center, and at shallow depths as it discharges

upward to the brook (Figure 1B). Whereas TCE concentrations greater than 1,000 μg/L are found in the source

area, the TCE concentrations are highest (greater than

10,000 μg/L) near the base of the aquifer midway between the source and discharge.

Geochemistry of the Plume

Determination of the pH and redox conditions present in

a plume is essential to predicting the types of natural

biological interactions that may take place in the aquifer.

85
Results of water-quality analyses indicate that the pH of

ground water in the plume is near neutral (6.5 to 7.5),

and concentrations of both dissolved oxygen (less than

0.5 milligrams per liter [mg/l]) and nitrate (less than 1

mg/L) are very low. Concentrations of iron(II) are greater

than 1 mg/L in some areas of the plume, whereas sulfate

and carbon dioxide are consistently plentiful (greater than

40 mg/L and 100 mg/L as bicarbonate, respectively) as

Figure 1. Location of Building 24 study area at Picatinny Arsenal, New Jersey: (A) areal extent of ground-water trichloroethene

plume and (B) vertical distribution of ground-water trichloroethene concentrations, October to November 1991. (Location

of section A-A’ is shown in Figure 1A.)

86
potential terminal electron acceptors. In addition, sulfide

odor was noted in water from many wells within the

plume, and methane was present at concentrations

ranging from 1 to 85 μg/L. 

These findings indicate the plume is primarily anaerobic

and contains a variety of reducing redox environments

controlled in different areas by iron(III) reduction, sulfate

reduction, and methanogenesis. Under these conditions, reductive dechlorination of TCE can take place if

sufficient electron donors are available. Dissolved organic carbon (DOC), consisting primarily of humic and

fulvic acids, may fulfill the electron donor requirement in

this system. Concentrations of DOC are highest immediately downgradient from the source area (5 to 14

mg/L) and also are elevated near the discharge point (1

to 2 mg/L).

The presence of cis-1,2-dichloroethene (cis-DCE) and

vinyl chloride (VC)—TCE breakdown products—in 75

percent of the wells sampled in and around the plume

indicates that reductive dechlorination of TCE is taking

place in the aquifer. Because neither of these compounds was used in Building 24, they are believed to

originate from the biologically mediated breakdown of

TCE. Further evidence for reductive dechlorination of

TCE is the similarity among the distributions of TCE,

cis-DCE, and VC in the aquifer, although the concentrations of cis-DCE and VC are highest in the downgradient

portion of the plume near the discharge point.

Trichloroethene Mass Distribution

The mass of TCE dissolved in the ground water in the

plume was estimated on the basis of results of six

synoptic sampling taken from 1987 to 1991. By using a

plume volume of 2.3 x 106

 cubic meters (m3) and a

porosity of 0.3, and by assuming that each well represents a finite volume of the aquifer, the average mass

of TCE dissolved in the plume was determined to be

1,000 ± 200 kilograms (kg) (7). This estimate did not

show a consistent increasing or decreasing trend over

the six sampling, which implies that the plume was

essentially at steady state. Most of the dissolved TCE

mass (57 percent) is present in the ground water near

the base of the unconfined aquifer, where TCE concentrations are greater than 10,000 μg/L.

The mass of sorbed TCE within the plume was estimated from methanol-extraction analyses of sediments

from six sites along the centerline of the plume (8). The

ratio of the masses of sorbed TCE to dissolved TCE per

unit volume of aquifer ranged from 3:1 to 4:1 at these

six sites. Therefore, 3,000 to 4,000 kg of TCE is calculated to be sorbed to aquifer sediments within the plume.

A sorbed mass of 3,500 kg of TCE was used in all

calculations.

Trichloroethene Mass-Flux Estimates

The major naturally occurring processes that affect the

input or removal of TCE to or from the plume were

identified and studied independently as part of the Toxic

Substances Hydrology Program project at Picatinny Arsenal (9, 10). The TCE removal processes that were

considered include advective transport, lateral dispersion, anaerobic biotransformation, diffusion-driven volatilization, advection-driven volatilization, and sorption.

The TCE input processes evaluated include desorption,

infiltration, and dissolution. Each of these processes is

described briefly below, and a TCE mass-flux estimate

is made for each on the basis of the results of research

conducted in the Picatinny Arsenal plume.

Removal-Process Flux Estimates

Advective transport is the process by which dissolved

TCE is removed from the plume in ground water that is

discharging to Green Pond Brook. The mass flux of TCE

was calculated by using an advective flux rate of 800

liters per meter squared per week (based on modeling

analyses [4, 11]), a median ground-water TCE concentration of 1,200 μg/L, and a cross-sectional area of 980

square meters where the aquifer discharges to the

brook. On the basis of these values, approximately 50

kilograms per year (kg/yr) of TCE are removed from the

plume by discharge to Green Pond Brook.

Lateral dispersion is the process that causes plume

spreading by transport of TCE out of the side boundaries

of the plume where the concentration is 10 μg/L (Figure

1). Using Fick’s Law, the lateral TCE-concentration gradient, and the estimated area of the sides of the plume,

researchers calculated that less than 1 kg/yr of TCE is

lost from the plume by this mechanism. 

Anaerobic biotransformation is the biologically mediated

process of reductive dechlorination whereby TCE undergoes the sequential replacement of the chlorine atoms on the molecule with hydrogen atoms to form

cis-DCE, VC, and ethene as breakdown products (12,

13). Biotransformation rate constants were determined

in laboratory batch microcosm studies of core samples

from five sites along the centerline of the plume (14, 15).

The first-order TCE-degradation rate constants obtained

in these studies range from -0.004 to -0.035 per week,

with a median of -0.007 per week. If this latter rate

constant is applied to the 1,000 kg of TCE dissolved in the

plume, about 360 kg/yr of TCE are removed from the

plume by naturally occurring anaerobic biotransformation.

Volatilization is the loss of TCE from ground water into

the soil gas of the unsaturated zone across the water

table. Volatilization is driven by diffusive and advective

mechanisms. The rate of loss of TCE in diffusion-driven

volatilization is determined by the TCE gradient in the

soil gas of the unsaturated zone. Diffusion-driven vola87
tilization was estimated using Fick’s Law, field-measured unsaturated-zone soil-gas TCE gradients, bulk diffusion coefficients from the literature for sites with similar

soils, and the area of the plume. Removal of TCE from

the plume by diffusion-driven volatilization is calculated

to be less than 1 kg/yr over the area of the plume (7,

16). In advection-driven volatilization, the rate of loss of

TCE is controlled by pressure and temperature changes

in the unsaturated-zone soil gas. Advection-driven volatilization was investigated using a prototype vertical-flux

measuring device at Picatinny Arsenal (16). On the

basis of flux measurements made with the device at

eight sites and the area of the plume, the TCE removed

from the plume by advection-driven volatilization is calculated to be approximately 50 kg/yr.

Sorption is the partitioning of TCE from the ground water

into the organic-carbon fraction of the aquifer sediments. Field partition coefficients measured at several

locations within the plume (8) indicate that more TCE

was sorbed to aquifer organic materials at all sites than

would be predicted if the sorbed TCE concentrations

were in equilibrium with the ground-water TCE concentrations. Therefore, desorption processes rather than

sorption processes most likely predominate. Removal of

TCE by sorption is estimated to be less than 1 kg/yr.

Input-Process Flux Estimates

Desorption is the process by which TCE partitions out

of the organic phase on the contaminated sediments

back into the ground water in response to concentration

gradients. This process at Picatinny Arsenal was characterized as having two parts: an initial rapid phase of

desorption, in which 0 to 10 percent of the TCE releases,

and a second, slower phase of desorption, in which most

of the TCE releases over a longer period (8). First-order

desorption rate constants ranging from -0.003 to -0.015

per week were measured in flow-through column experiments. Because these experiments were conducted

with clean water, the desorption rates obtained probably

are higher than in situ desorption rates. For this reason,

the smaller of the desorption rate constants (-0.003 per

week) and the total amount of TCE estimated to be

sorbed to the plume sediments (3,500 kg) were used to

calculate that 550 kg/yr of TCE is being input to the

plume by means of desorption.

Infiltration, the process by which TCE in the soil gas or

on the unsaturated-zone soil is dissolved by percolating

recharge to the ground water, was studied with laboratory soil columns, field infiltration experiments, and multiphase solute-transport modeling (17). Because

concentrations of TCE in the soil gas generally are low

over most of the plume, and because infiltration occurs

only during recharge events rather than continuously

throughout the year, it was estimated that the input of

TCE to the plume by this process is less than 1 kg/yr.

Dissolution is the process by which dense nonaqueousphase liquid (DNAPL) TCE dissolves into the ground

water. The presence of DNAPL TCE at the base of the

unconfined aquifer midway between the source and the

brook has been suspected because concentrations of

TCE in ground water at this location are much higher

than those immediately upgradient. Concentrations of

TCE in deep wells in this area consistently exceed 2

percent of saturation, which is one indication of DNAPL

presence (18). DNAPL TCE has not been confirmed by

measurement or observation of free-phase TCE in any

water or soil sample from the arsenal. Consequently, the

mass of DNAPL TCE that is input by dissolution cannot

be calculated directly but can only be estimated by the

difference between the sum of the mass removed by all

removal processes and the sum of the mass introduced

by all other input processes.

Mass-Balance Analysis

The estimated mass balance for the TCE plume at

Picatinny Arsenal is shown in Figure 2. All inputs are

represented with open arrows; all outputs are represented with solid arrows.

Approximately 460 kg/yr of dissolved TCE is estimated

to be removed from the plume by natural processes. Of

this, 360 kg/yr, or 78 percent of the TCE removed annually, is removed as a result of anaerobic biotransformation. This is by far the most important TCE removal

process operating in the Picatinny Arsenal plume. Removal by advective transport to Green Pond Brook and

advection-driven volatilization are each estimated at 50

kg/yr. Therefore, each of these processes is responsible

for the removal of about 11 percent of the total TCE

removed annually from the plume. Lateral dispersion,

diffusion-driven volatilization, and sorption are all of minor importance compared with these major processes.

The finding that natural anaerobic biotransformation is

the principal mechanism for removal of TCE from the

plume at Picatinny Arsenal is significant. Anaerobic

biotransformation has been reported to be a major natural removal process for TCE at only a few sites (19), and

this conclusion has not previously been reached by

quantifying and comparing the magnitude of all other

removal processes occurring at a site. This result is

likely to have great transfer value to other sites with

similar geochemistry, hydrology, and geology.

The process of desorption is the most important input

mechanism evaluated at Picatinny Arsenal; it accounts

for the introduction of an estimated 550 kg/yr of TCE.

Input by infiltration is very small in comparison (less than

1 kg/yr). Because the sum of the inputs is larger than

the sum of the outputs, dissolution of DNAPL TCE in the

system cannot be estimated.

88
The fact that long-term desorption is a significant continuing source of TCE to the aquifer may explain why

the TCE concentrations are still relatively high in the

source area (greater than 1,000 μg/L) 13 years after

TCE use was discontinued at the site. This finding is

significant because it shows that desorption can be an

important input mechanism even at sites where the

sediment organic content is low (less than 0.5 percent).

Because the mass of TCE in the plume was at steady

state during these studies, the sources of TCE ideally

should equal the sinks of TCE. Although the estimated

inputs do not equal the estimated outputs in the mass

balance, they are of the same order of magnitude. Additional study of the individual processes would be necessary to refine the mass balance further. Because

confidence in the output-process mass-flux estimates is

Figure 2. Mass-balance estimates of fluxes of naturally occurring processes that affect the fate and transport of trichloroethene in

the ground-water system at Picatinny Arsenal, New Jersey.

89
high and the TCE desorption rate constants used probably were on the high side, the desorption mass-flux

estimate may be higher than the actual value.

Field-Scale Estimate of Natural

Attenuation Rate 

The natural attenuation rate of TCE at Picatinny Arsenal

was calculated from field data and compared with the

anaerobic biotransformation rates calculated in the laboratory microcosm studies. Assuming first-order kinetics

and considering the decrease in TCE concentrations

from the source area to the discharge area (1,900 μg/L

to 760 μg/L), the time of travel for TCE between these

two points in the plume (3.1 years), and the distance

between these two sites (470 m), then the field-scale

natural attenuation rate constant is calculated to be

-0.006 per week. This field-calculated rate constant is

nearly identical to the median rate constant of -0.007 per

week determined in the laboratory microcosm experiments. That both methods yield rate constants of similar

magnitude confirms that most of the natural attenuation

that occurs in the Picatinny Arsenal plume is due to

anaerobic biotransformation. In addition, it indicates that

the methods used to make these estimates and measurements are valid.

Comparison of Natural Attenutation

Processes to Pump-and-Treat

Remediation

A pump-and-treat system was installed in the Picatinny

Arsenal TCE plume as an interim remediation measure

in September 1992. It consists of a set of five withdrawal

wells from which an average of 440,000 liters per day

are pumped to a treatment system equipped with stripping towers and granulated activated carbon filters. On

the basis of average pumpage values and ground-water

TCE concentrations in each withdrawal well during

1995, the pump-and-treat system is currently removing

about 70 kg/yr at a cost of $700,000 per year. This is

about one-fifth the amount of TCE being removed from

the plume each year by anaerobic biotransformation,

and just slightly more than the mass of TCE being

removed by each of the processes of advective transport and advection-driven volatilization.

Conclusion

The relative importance of all naturally occurring processes that introduce or remove TCE to or from a contamination plume at Picatinny Arsenal, New Jersey, was

determined. Anaerobic biotransformation is the most

important process for TCE removal from the plume by

almost an order of magnitude over advective transport

and advection-driven volatilization. Anaerobic biotransformation accounts for an estimated 78 percent of the

total mass of TCE removed from the plume annually.

Other removal processes—lateral dispersion, diffusiondriven volatilization, and sorption—are minor in comparison. Desorption is the most significant TCE input

process evaluated. A mass-balance analysis shows that

the removal of TCE from the plume by natural attenuation processes is of the same order of magnitude as

the input of TCE to the plume. The natural attenuation

rate constant calculated from field TCE concentrations

and time-of-travel data is in close agreement with anaerobic biotransformation rate constants measured in

laboratory microcosm studies.

Anaerobic biotransformation removes approximately

five times the mass of TCE removed by an interim

pump-and-treat remediation system operating at the Picatinny Arsenal site. The pump-and-treat system removes just slightly more mass per year than each of the

processes of advective transport to Green Pond Brook

and advection-driven volatilization.

References

 1. Martin, M. 1989. Preliminary results of a study to simulate trichloroethylene movement in ground water at Picatinny Arsenal, New

Jersey. In: Mallard, G.E., and S.E. Ragnone, eds. U.S. Geological

Survey Toxic Substances Hydrology Program—proceedings of

the technical meeting, Phoenix, AZ, September 26-30, 1988. U.S.

Geological Survey Water-Resources Investigations Report 88-

4220. pp. 377-383.

 2. Martin, M. 1991. Simulation of reactive multispecies transport in

two dimensional ground-water-flow systems. In: Mallard, G.E.,

and D.A. Aronson, eds. U.S. Geological Survey Toxic Substances

Hydrology Program—proceedings of the technical meeting, Monterey, CA, March 11-15. U.S. Geological Survey Water-Resources Investigations Report 91-4034. pp. 698-703.

 3. Martin, M. 1996. Simulation of transport, desorption, volatilization,

and microbial degradation of trichloroethylene in ground water at

Picatinny Arsenal, New Jersey. In: Morganwalp, D.W., and D.A.

Aronson, eds. U.S. Geological Survey Toxic Substances Hydrology Program—proceedings of the technical meeting, Colorado

Springs, CO, September 20-24, 1993. U.S. Geological Survey

Water-Resources Investigations Report 94-4015.

 4. Voronin, L.M. 1991. Simulation of ground-water flow at Picatinny

Arsenal, New Jersey. In: Mallard, G.E., and D.A. Aronson, eds.

U.S. Geological Survey Toxic Substances Hydrology Program—

proceedings of the technical meeting, Monterey, CA, March 11-

15. U.S. Geological Survey Water-Resources Investigations

Report 91-4034. pp. 713-720.

 5. Sargent, B.P., T.V. Fusillo, D.A. Storck, and J.A. Smith. 1990.

Ground-water contamination in the area of Building 24, Picatinny

Arsenal, New Jersey. U.S. Geological Survey Water-Resources

Investigations Report 90-4057. p. 94.

 6. Benioff, P.A., M.H. Bhattacharyya, C. Biang, S.Y. Chiu, S. Miller,

T. Patton, D. Pearl, A. Yonk, and C.R. Yuen. 1990. Remedial

investigation concept plan for Picatinny Arsenal, Vol. 2: Descriptions of and sampling plans for remedial investigation sites. Argonne National Laboratory, Environmental Assessment and

Information Sciences Division, Argonne, IL. pp. 22-1 - 22-24.

 7. Imbrigiotta, T.E., T.A. Ehlke, M. Martin, D. Koller, and J.A. Smith.

1995. Chemical and biological processes affecting the fate and

transport of trichloroethylene in the subsurface at Picatinny Arsenal, New Jersey. Hydrological Sci. Technol. 11(1-4):26-50.

90
 8. Koller, D., T.E. Imbrigiotta, A.L. Baehr, and J.A. Smith. 1996.

Desorption of trichloroethylene from aquifer sediments at Picatinny Arsenal, New Jersey. In: Morganwalp, D.W., and D.A. Aronson, eds. U.S. Geological Survey Toxic Substances Hydrology

Program—proceedings of the technical meeting, Colorado

Springs, CO, September 20-24, 1993. U.S. Geological Survey

Water-Resources Investigations Report 94-4015.

 9. Imbrigiotta, T.E., and M. Martin. 1991. Overview of research activities on the movement and fate of chlorinated solvents in

ground water at Picatinny Arsenal, New Jersey. In: Morganwalp,

D.W., and D.A. Aronson, eds. U.S. Geological Survey Toxic Substances Hydrology Program—proceedings of the technical meeting, Monterey, CA, March 11-15. U.S. Geological Survey

Water-Resources Investigations Report 91-4034. pp. 673-680.

10. Imbrigiotta, T.E., and M. Martin. 1996. Overview of research activities on the transport and fate of chlorinated solvents in ground

water at Picatinny Arsenal, New Jersey, 1991-93. In: Morganwalp, D.W., and D.A. Aronson, eds. U.S. Geological Survey Toxic

Substances Hydrology Program—proceedings of the technical

meeting, Colorado Springs, CO, September 20-24, 1993. U.S.

Geological Survey Water-Resources Investigations Report 94-

4015. 

11. Martin, M., and T.E. Imbrigiotta. 1994. Contamination of ground

water with trichloroethylene at the Building 24 site at Picatinny

Arsenal, New Jersey. In: U.S. EPA Symposium on Intrinsic Bioremediation of Ground Water, Denver, CO, August 30-September

1, 1994. EPA/540/R-94/515. pp. 143-153.

12. Parsons, F.Z., P.R. Wood, and J. DeMarco. 1984. Transformations of tetrachloroethene and trichloroethene in microcosms and

ground water. J. Am. Waterworks Assoc. 76(2):56-59.

13. Vogel, T.M., C.S. Criddle, and P.L. McCarty. 1987. Transformations of halogenated aliphatic compounds. Environ. Sci. Technol.

21(8):722-736.

14. Wilson, B.H., T.A. Ehlke, T.E. Imbrigiotta, and J.T. Wilson. 1991.

Reductive dechlorination of trichloroethylene in anoxic aquifer

material from Picatinny Arsenal, New Jersey. In: Morganwalp,

D.W., and D.A. Aronson, eds. U.S. Geological Survey Toxic Substances Hydrology Program—proceedings of the technical meeting, Monterey, CA, March 11-15. U.S. Geological Survey

Water-Resources Investigations Report 91-4034. pp. 704-707.

15. Ehlke, T.A., T.E. Imbrigiotta, B.H. Wilson, and J.T. Wilson. 1991.

Biotransformation of cis-1,2-dichloroethylene in aquifer material

from Picatinny Arsenal, Morris County, New Jersey. In: Morganwalp, D.W., and D.A. Aronson, eds. U.S. Geological Survey Toxic

Substances Hydrology Program—proceedings of the technical

meeting, Monterey, CA, March 11-15. U.S. Geological Survey

Water-Resources Investigations Report 91-4034. pp. 689-697.

16. Smith, J.A., A.K. Tisdale, and H.J. Cho. In press. Quantification

of natural vapor fluxes of trichloroethene in the unsaturated zone

at Picatinny Arsenal, New Jersey. Environ. Sci. Technol.

17. Cho, H.J., P.R. Jaffe, and J.A. Smith. 1993. Simulating the volatilization of solvents in unsaturated soils during laboratory and

field infiltration experiments. Water Resour. Res. 29(10):3329-

3342.

18. Cohen, R.M., and J.W. Mercer. 1993. DNAPL site evaluation.

Boca Raton, FL: C.K. Smoley.

19. Wilson, J.T., J.W. Weaver, and D.H. Kampbell. 1994. Intrinsic

Bioremediation of TCE in ground water at an NPL site in St.

Joseph, Michigan. In: U.S. EPA Symposium on Intrinsic Bioremediation of Ground Water, Denver, CO, August 30-September 1.

EPA/540/R-94/515. pp. 154-160.

91
Case Study: Plant 44, Tucson, Arizona

Hanadi S. Rifai and Philip B. Bedient

Rice University, Houston, Texas

Kristine S. Burgess

Montgomery Watson, Salt Lake City, Utah

Introduction

A pump-and-treat remediation system operating for the

past 10 years at the Plant 44 site in Tucson, Arizona,

allowed hydraulic control of the dissolved chlorinated

solvents contaminant plume. Additionally, the pumpand-treat network removed a total of approximately

6,000 kilograms (kg) of trichloroethene (TCE) in its first

5 years of operation. Recent observations using site

data, however, include resurgence of TCE concentrations after pump turnoff and the emergence of the “tailing” phenomenon at a number of the pumping wells.

A detailed analysis of the site’s historical information as

well as extensive data collected before and after system

startup suggested the presence of dense nonaqueous

phase liquids (DNAPLs) at the site and revealed that the

pump-and-treat system would not achieve the desired

site cleanup within a reasonable time frame.

Plant 44 Site Description

The site hydrogeology consists of four stratigraphic units

(1): a relatively thick unsaturated zone extending between 110 to 130 feet below the surface; an upper zone

extending to a depth of 180 to 220 feet; an aquitard

consisting of 100 to 150 feet of low-permeability clay; and

a lower zone. Pump tests indicate that hydraulic conductivity ranges from 2 x 10-4 to 3 x 10-3 feet per second for

the upper zone. The background hydraulic gradient is

0.006 feet per foot toward the northwest, and the groundwater velocity ranges from 250 to 800 feet per year (2).

Activities at Plant 44 include development, manufacturing, testing, and maintenance of missile systems from

1952 until the present. Historical data indicate that

greater than 50 drums per year of TCE, 1,1-dichloroethene (1,1-DCE), and 1,1,1-trichloroethane (1,1,1-

TCA) were used at the site. The resulting area of TCE

contamination was approximately 5 miles long by 1.6

miles wide in 1986, before remediation startup (Figure

1). A maximum TCE concentration of 2.7 parts per million (ppm) was measured in 1986, although concentrations of up to 15.9 ppm have been observed in the

ground water (2). Potential sources of contamination

include pits, ponds, trenches, and drainage ditches in

which disposal of solvents and waste water was reported from 1952 through 1977 (Figure 2).

Ground-Water Extraction System

The pump-and-treat system began operation in April

1987. The system consists of 17 extraction wells and 13

recharge wells, shown in Figure 1. Water-level elevation

and contaminant concentration data for the pumping

wells and 40 monitoring wells are collected monthly. The

total dissolved mass removed by the system in its first

5 years of operation (approximately 6,000 kg) exceeds

the 3,800 kg dissolved mass present in the plume in

1986. This would suggest the presence of a continuing

source of contamination in the aquifer.

The concentration of dissolved TCE in the extracted

ground water decreases during remediation, particularly

in those wells with initially high TCE concentrations. In

the majority of cases, the TCE concentration appears to

level out between 3 and 5 years, usually to a value that

exceeds the TCE drinking water standard of 5 parts per

billion (ppb). Numerous spikes of high TCE concentration are observed in a number of the pumping wells,

possibly due to continuing sources.

Fate-and-Transport Modeling

Modeling of the TCE plume at the site was completed

to evaluate the time required for cleanup. Aqueousphase flow in the upper zone was simulated along with

the pump-and-treat remediation system. Source locations used in the modeling were based on the location

92
of “hot spots” in the plume, areas where formation of

DNAPL pools is likely and areas where the confining

clay layer is thin. An overall mass transfer rate due to a

continuing source of contamination was estimated

based on the difference between the mass pumped in

the first 5 years of operation of the system and the mass

present in the aquifer.

Additionally, source dissolution mechanisms were analyzed assuming the following four potential configurations of DNAPL in the subsurface: unsaturated zone

residual, a DNAPL pool, saturated zone residual, and

DNAPL located in a nonadvective zone. Dissolution

times, for example, due to unsaturated zone source

areas ranged from 1,100 to 13,000 years, while those

for DNAPL pools ranged from 1 to 60,000 years depending on the source assumptions that were made.

A comparison between the estimated mass transfer rate

and the dissolution data indicated that the two most

likely dissolution mechanisms present at the site include

unsaturated zone residuals and DNAPL pools. The associated dissolution times ranged from 100 to 1,000

years. The fate-and-transport modeling results, assuming no continuing sources of TCE into the aquifer, indicate that 50 more years of the remediation system’s

operation are required. If the estimated mass transfer

rates are incorporated into the model, the required remediation time exceeds hundreds of years.

Conclusion

Data from the Plant 44 site indicate that DNAPL may be

present. Further contamination of the ground water

might occur because of sources present in the unsaturated zone and the potential dissolution from TCE plumes.

Figure 1. TCE plume prior to remediation, December 1986 (ppb).

Figure 2. Historical onsite disposal locations.

93
Complete dissolution of the DNAPL pools may take as

long as 100 years under pumped conditions, while dissolution of unsaturated residual by infiltrating ground

water may continue for thousands of years. The groundwater extraction system at the site has contained the

dissolved plume and removed significant amounts of

chlorinated compounds. If DNAPL is present at the site,

however, complete removal of TCE using pump-and-treat

will require a very lengthy and costly operation period.

References

1. Hargis and Montgomery, Inc. 1982. Phase II investigation of subsurface conditions in the vicinity of abandoned waste disposal

sites, Hughes Aircraft Company manufacturing facility, Tucson,

Arizona, Vol. I. Tucson, AZ.

2. Groundwater Resources Consultants, Inc. 1992. Quarterly ground

water monitoring report, well field reclamation system, July through

September 1991, U.S. Air Force Plant 44.

94
Remediation Technology Development Forum Intrinsic Remediation Project at

Dover Air Force Base, Delaware

David E. Ellis and Edward J. Lutz

DuPont Specialty Chemicals-CRG, Wilmington, Delaware

Gary M. Klecka

Dow Chemical Company, Midland, Michigan

Daniel L. Pardieck

Ciba-Geigy, Greensboro, North Carolina

Joseph J. Salvo

General Electric Corporate Research and Development, Schenectady, New York

Michael A. Heitkamp

Monsanto Company, St. Louis, Missouri

David J. Gannon

Zeneca Bioproducts, Mississauga, Ontario

Charles C. Mikula and Catherine M. Vogel

U.S. Air Force, Tyndall Air Force Base, Florida

Gregory D. Sayles, Donald H. Kampbell, and John T. Wilson

U.S. Environmental Protection Agency, 

National Risk Management Research Laboratory, Ada, Oklahoma

Donald T. Maiers

U.S. Department of Energy, Idaho National Engineering Laboratory, Idaho Falls, Idaho

Introduction

The Remediation Technology Development Forum

(RTDF) Bioremediation Consortium is conducting a

large, integrated field and laboratory study of intrinsic

remediation in a plume at the Dover Air Force Base

(AFB) in Delaware. The work group is a consortium of

industrial companies and government agencies working

on various aspects of bioremediation of chlorinated solvents, such as tetrachloroethylene (PCE) and trichloroethene (TCE). The intrinsic bioremediation program is

part of an integrated study that also includes cometabolic bioventing and accelerated anaerobic treatment. The combination of these three methods can treat

all parts of a solvent contamination area.

The goals of the 4-year intrinsic remediation study are

to evaluate whether the contaminants at the site are

being destroyed through intrinsic remediation, to identify

the degradation mechanisms, and to develop and validate protocols for implementing intrinsic remediation at

other sites.

A wide variety of geological, geochemical, and biological

research is being integrated into this study. This presentation emphasizes the geochemical aspects of the study

for the following reasons: the geochemical data were

available early in the study; it clearly shows that solvent

destruction is happening; and the primary author’s expertise lies in geochemistry. The participants who focused on the biological aspects of this study will undoubtedly

be presenting their conclusions at future meetings.

95
Background

The RTDF Bioremediation Consortium initiated this

study in February 1995. Dover AFB was chosen over the

many other sites evaluated for the study because:

• The plume is well-characterized.

• Analyses of ground-water chemistry provided clear

evidence that chlorinated solvent contaminants are

being biodegraded.

• The deep zone of the aquifer has relatively simple

geology and is underlain by a thick confining layer.

• Access for sampling and testing is good, and the site

is easily reached by offsite personnel and visitors.

• The base has a proactive environmental program.

The plume contains primarily TCE and dichloroethene

(DCE), with smaller amounts of vinyl chloride (VC). It

occupies an area north and south of U.S. Highway 113

approximately 9,000 feet long and 3,000 feet wide.

There are multiple sources of solvent contamination in

the area north of the highway, as well as several minor

sources of petroleum hydrocarbons. There appear to be

at least three sources of TCE.

The water-bearing unit in the study area is composed of

fine- to coarse-grained sands ranging in thickness from

30 to 60 feet. The ground-water elevation ranges from

approximately 13 feet mean sea level (MSL) at the north

end of the plume to less than 3 feet MSL near the

southern end. Ground water flows to the south. The

plume velocity ranges from about 150 feet per year in

the northern portion of the study area to over 200 feet

per year beneath the southern area. The consortium

believes that the aquifer contains aerobic and anaerobic

microzones. This simple sand aquifer exhibits complex

metabolic activity that might not be apparent from a

cursory examination of geochemical information.

This paper focuses on the lower third of the aquifer,

which has the highest permeability and contains the

majority of the contaminants.

Current Findings

Intrinsic remediation is clearly occurring in the ground

water, and results suggest that multiple biodegradation

pathways are operating. These findings are based on data

on the plume profile, contaminant concentrations and geochemical markers, and the presence of the soluble chloride ion produced by biodegradation of the solvents.

Plume Profile

Figure 1 shows the relationship of the constituents

within the plume. Note that the plumes of the different

solvent species are “stacked.” There is no chromatographic

separation, as would be expected based on the much

different mobilities of these compounds in ground water.

This suggests that the more mobile compounds, such

as VC, are degrading before they can move away from

the less mobile ones.

Figure 1. Plume configuration in the deep zone at Dover AFB.

96
Contaminants and Geochemical Markers

TCE

TCE concentrations in the ground water range up to

20 milligrams per liter (mg/L.) The TCE concentration

declines rapidly near Highway 113. TCE is degraded

before reaching the St. Jones River to the south of the

plume.

DCE

DCE concentrations are over 10 mg/L in two areas. The

DCE is primarily cis-1,2-DCE, the isomer produced by

biodegradation of TCE. Chemically manufactured DCE

can be distinguished from biogenic DCE because

chemically manufactured DCE contains a mixture of

isomers, of which cis-DCE is a minor component. The

DCE plume overlaps the TCE plume. DCE concentrations also decline rapidly south of Highway 113.

VC

There is a smaller VC plume with concentrations up to

1 mg/L. Since VC was never used on the base, the

consortium believes that it is present as a biodegradation product of DCE. If DCE were being lost primarily by

reduction to VC, we should be able to detect low, transient

concentrations of VC throughout the area containing

DCE, regardless of the relative degradation rates of the

two compounds. The area containing VC, however, is

considerably smaller than the DCE plume.

Ethylene

Ethylene is also present, showing that complete reductive dehalogenation of TCE does occur in the deep

zone. The amount of ethylene is small, however: 50

micrograms per liter (μg/L) or less. This is much too low

to account for the observed losses of TCE and DCE.

Soluble Chloride Ion

The best evidence that chlorinated solvents are being

destroyed is the simultaneous increase in soluble chloride ions and decrease in solvent concentrations. This

is clearly observable at Dover AFB, as shown in Figures

2 and 3. While the total chlorocarbon concentrations

decrease from 15 to around 1 mg/L in the area of

Highway 113 (Figure 2), the dissolved chloride concentration increases to over 40 mg/L. Background chloride

levels are approximately 10 mg/L. The dissolved chloride (Figure 3) increases in the deep zone of the aquifer

but not the shallow zone. This eliminates other, extraneous chloride sources such as road salt. This evidence

clearly supports the hypothesis that solvents are being

destroyed by an intrinsic process.

Figure 2. Total chlorinated compounds (mg/L) in the deep zone at Dover AFB.

97
Biodegradation Mechanisms

Processes other than reductive dehalogenation account

for the majority of the degradation of DCE because of

low levels of VC and ethylene. The consortium has

extensively measured the geochemistry of the ground

water to understand this environment. Clues to the

mechanisms are found in dissolved oxygen levels,

methane data, the redox state of the aquifer, and laboratory studies.

In the vicinity of the plume, the dissolved oxygen concentration is depleted to below 1 mg/L in the ground

water. Dissolved oxygen begins increasing in the vicinity

of Highway 113. Outside the contaminated zone, dissolved oxygen is greater than 4 mg/L.

The methane data show a pattern generally the inverse

of the dissolved oxygen data. Methane concentrations

ranging from 20 to greater than 500 μg/L are found

within the contaminated zone, while no methane is observed outside of the plume. This indicates that

methanogenesis appears to be an important microbial

process in the anaerobic portion of the aquifer. The

occurrence of both methane and oxygen south of Highway 113 suggests that cooxidation is likely occurring at

Dover AFB.

The redox state of the deep zone at Dover AFB is

relatively high. In most of the plume, the bulk phase

redox is above 200 millivolts. All redox potentials are

above 50 millivolts. Sharma and McCarty (1) showed

that bacterial reductive dehalogenation of PCE and TCE

to DCE can occur in relatively oxidizing conditions, requiring only the absence of oxygen or nitrate, similar to

conditions at Dover AFB. Reductive dehalogenation of

DCE to VC or ethylene, however, appears to require

sulfate-reducing or methanogenic conditions (2, 3), processes that occur at redox levels below -200 millivolts.

These low oxidation states are probably found in microenvironments but do not dominate the aquifer. Finally, ongoing RTDF microcosm studies of Dover AFB

samples are showing clear production of 14CO2 from

14C-labeled DCE under oxygenated conditions. Therefore, the consortium believes that at Dover AFB TCE is

transformed to DCE by reductive dehalogenation and that

DCE is then biodegraded by a combination of direct

oxidation and cooxidation, with a minor component of

reductive dehalogenation to VC and ethylene.

Biodegradation Rates

Table 1 gives estimated half-lives and goodness of fit

values (r2

) at Dover AFB as calculated by two different

Figure 3. Dissolved chloride in the deep zone at Dover AFB.

Table 1. Half-Life Calculations for Dover AFB

Method PCE → TCE TCE → DCE DCE → VC VC → ETH

Buscheck 2.4 2.8 1.4 2.2

r

2 0.99 0.94 0.94 0.93

Graphical

extraction

 2.80 4.19 2.81 1.84

98
methods. The method developed by Buscheck et al. (4)

gives the values shown in the first row of the table. The

values in the second row were calculated by a simple

graphical extrapolation method. The values in both rows

are fairly consistent, all on the order of 1 to 2 years.

These rate constants are consistent with other chlorinated solvent rate constants determined to date. This

consistency suggests that a similar set of degradation

mechanisms operates at other sites as well.

If the plume is assumed to be in a steady state, isoconcentration maps can be used to calculate that about

250 pounds of chlorinated solvents are being biodegraded each year. This is equivalent to destroying 25

gallons of dense nonaqueous-phase liquid every year.

Conclusion

The RTDF project at Dover AFB is in the second of 4

years. The evidence clearly demonstrates that active

intrinsic remediation of chlorinated solvents is occurring.

The key evidence supporting this conclusion is:

• The contaminant plumes are “stacked,” indicating

that the more mobile contaminants are being destroyed before they can move away from the less

mobile contaminants.

• The chloride ion concentration in solution increases

as the solvent concentration declines. The increase

is large enough to account for the entire observed

loss of solvents.

• There is clear field evidence of reductive dehalogenation and oxidation, and possible evidence for co-oxidation.

References

1. Sharma, P.K., and P.L. McCarty, 1996. Isolation and characterization of a facultative bacterium that reductively dechlorinates

tetrachloroethene to cis-1,2 dichloroethene. Appl. Environ. Microbiol. 62(3):761-765.

2. Kastner, M. 1991. Reductive dechlorination of tri- and tetrachloroethylenes depends on transition from aerobic to anaerobic conditions. Appl. Environ. Microbiol. 57(7):2039-2046.

3. Holliger, C., and G. Schraa. 1994. Physiological meaning and

potential for application of reductive dechlorination by anaerobic

bacteria. FEMS Microbiology Reviews 15:297-305.

4. Buscheck, T.E., K.T. OReilly, S.N. Nelson. 1993. Evaluation of

intrinsic bioremediation at field sites. Proceedings of the conference Petroleum Hydrocarbons and Organic Chemicals in Ground

Water: Prevention, Detection and Restoration, Houston, TX, pp. 367-

381.

99
Case Study: Wurtsmith Air Force Base, Michigan

Michael J. Barcelona

University of Michigan, The National Center for Integrated Bioremediation Research and

Development, Department of Civil and Environmental Engineering, Ann Arbor, Michigan

Introduction

Wurtsmith Air Force Base (WAFB) in Oscoda, Michigan,

was decommissioned in June of 1993. Shortly thereafter, the U.S. Environmental Protection Agency (EPA),

the Strategic Environmental Research and Development Program (SERDP) of the Department of Defense

(DoD), the University of Michigan, and the Michigan

Department of Environmental Quality contributed resources to develop the National Center for Integrated

Bioremediation Research and Development (NCIBRD).

NCIBRD is a DoD National Environmental Technology

Test Site (NETTS) whose mission is to provide a welldefined and controlled research and development platform for the evaluation of in situ site characterization and

remediation technologies. The emphasis is on bioremediation techniques applied to subsurface and sediment

contamination problems. In situ biological technologies

with the potential to remediate unsaturated- and saturatedzone fuel and organochlorine solvent contamination in

subsurface and sediment systems are of particular interest. NCIBRD focused its early activities on the development of an expanded database of contaminant,

hydrogeologic, and geochemical conditions at several

contamination sites. Spatial and temporal variability in

these conditions makes evaluating the progress of intrinsic

bioremediation technology applications difficult.

Physical Setting

WAFB is located in Iosco County in northeast Michigan,

in the coastal zone of Lake Huron north of Oscoda.

Oscoda is accessible by rail, highway, and commercial

air routes north of Saginaw-Bay City, Michigan. WAFB

is under the authority of the Oscoda-Wurtsmith Airport

Authority and the Wurtsmith Area Economic Development Commission. The U.S. Air Force Base Conversion

Authority (BCA) is charged with remediating contaminated sites to enable the transition of site facilities to

civilian use. At present, 10 private or public concerns

have leased sites on the base for operations, including

an aircraft maintenance facility, a plastics manufacturer,

engineering firms, and educational institutions. The

base occupies 7 square miles bounded by the AuSable

River/AuSable River wetlands complex to the south,

Lake Van Etten to the east, and bluffs fronting a 5-milewide plain extending onto the base to the west. Lake

Huron receives the discharge from the associated

ground-water flow system and the Au Sable River approximately 0.5 mile south of the base boundary. The

altitude of the land surface ranges from 580 to 750 feet

above mean sea level. Figure 1 shows the base detail,

with an emphasis on Installation Restoration Program

(IRP) sites.

Mean monthly temperatures range from 21°F (-6°C) in

January to 68°F (20°C) in July. The lowest recorded

temperature was -22°F (-30°C), the highest 102°F

(39°C). Average annual precipitation is 30 inches (76

centimeters), and average snowfall is 44 inches (112

centimeters). Surficial geologic materials are of quaternary glaciofluvial and aeolian origins, made up largely

of medium to fine sands and coarse sand and gravel

deposits to depths of 60 to 90 feet (18 to 27 meters).

Below the glacial deposits, a confining lacustrine clay

layer (125 to 250 feet thick) separates the upper aquifer

ground water from lower, more saline waters in bedrock

units. In the eastern regions of the area, intermittent

sand, sand/gravel, and clay layers of 1 to 3 feet (less

than 0.3 to approximately 1 meter) thickness have been

observed in the saturated zone. These features are site

specific. Depths to ground water in the upper aquifer

range from less than 10 to approximately 30 feet (less

than 3 to 9 meters) in areas remote from pumping.

Average ground-water recharge rates range from 8 to

18 inches per year (20 to 46 centimeters per year). The

aquifer solids are greater than 85 percent quartz minerals, with organic carbon and inorganic carbon contents

below 0.1 and approximately 6.0 percent, respectively.

Hydraulic conductivities at the base range from 75 to

310 feet per day (23 to 95 meters per day), with a weighted

100
average of approximately 100 feet per day (31 meters

per day) based on selected slug or pump tests and

estimations from particle size distributions. Flow in the

sand and gravel upper aquifer is generally eastward

towards Lake Van Etten and south-southeast to the

AuSable River discharge areas at average rates of 1.0

to 0.3 feet per day (0.3 to 0.1 meter per day). In general,

vertical flow gradients are negligible except in zones of

ground-water discharge to surface-water bodies or near

pumping centers.

Figure 1. Map of Installation Restoration Program sites at WAFB.

101
Contaminant Profile

Contaminant investigations at the base began in the late

1970s. The Air Force and the U.S. Geological Survey

had been involved in formal studies since 1979. More

than 50 known and potential contamination sites have

been identified at the base through their efforts and

those of other contractors. Principal contaminants of

concern at the base include components of petroleum

hydrocarbon fuels, oils, and lubricants (POLs); organochlorine solvents (e.g., trichloroethylene [TCE], dichloroethylene [DCE]); fire-fighting compounds; combustion

products (e.g., naphthalene and phenanthrene); and

chlorinated aromatic compounds (e.g., dichlorobenzenes). Soil, aquifer solids, sediments, and ground

water are the major environmental media involved. Of

the 58 high-priority sites at the base, 13 include chlorinated solvents or partial microbial degradation products

as primary contaminants. Twelve of these 13 sites identify perchloroethylene (PCE) and TCE as primary contaminants in soil, aquifer solids, and ground water, and

show evidence of reductive dechlorination processes

(i.e., the presence of cis-1,2-DCE, vinyl chloride monomer). These sites have abundant levels of nonchlorinated organic matter and exhibit reduction to suboxic

redox conditions, as evidenced by the results from Fire

Training Area 2. The only major site at which sparse

evidence for microbial dechlorination of TCE exists is

the Pierces Point Plume, where oxic to transitional redox

conditions exist in the dissolved plume. The extent of

contamination of aquifer materials remains unknown.

Facilities

EPA (Region 5), Michigan Department of Natural Resources, and the BCA actively cooperate in the ongoing

IRP activities as well as the efforts of NCIBRD. Currently, NCIBRD occupies seven buildings on the base in

addition to 10,000 square feet of office and laboratory

space in Ann Arbor. Facilities for offices, laboratories,

storage, field operation, staging, and decontamination

have been developed to support activities at three sites

of intensive investigations. Mobile laboratory and drilling

vehicles provide additional support for year-round infield sampling and analysis assisted by experienced

field and laboratory staff. A basewide ground-water flow

model has been developed and refined by estimates of

hydraulic conductivity and mass water level measurements at more than 500 wells. Site-wide water balance

and refined ground-water transport models exist for

sites of current or future technology demonstration activity as well as for a controlled in situ injection experimental facility. This facility, the Michigan Integrated

Remediation Technology Laboratory (MIRTL), will be the

site of a natural gradient reactive tracer test in the

summer of 1996 for aerobic fuel bioremediation. MIRTL

will eventually consist of instrumented parallel test lanes

for both natural and induced gradient in situ testing of

cleanup technologies.

Case Study

Fire Training Area 2, in the southwest portion of the

base, has been the site of the most intensive monitoring

attention in the past decade at the base. Forty years of

fire training exercises using waste solvents and fuels

have resulted in soil and subsurface contamination with

hydrocarbons, chlorinated alkenes, aromatics, and polycylic aromatics. Early detective monitoring results were

collected by the U.S. Geological Survey from a network

of shallow and deep wells developed in 1987 (1). Focusing on the dissolved volatile organic compounds (i.e.,

aromatics and chlorinated alkenes), the plume was delineated to be approximately 200 to 300 feet (30 to 90

meters) wide, approximately 1,800 to 2,000 feet (550 to

610 meters) long, and approximately 6 to 25 feet (2 to

8 meters) thick. Concentrations of benzene, toluene,

ethylbenzene, and xylene compounds ranged from

greater than 2,000 to less than 10 micrograms per liter.

In both cases, the contaminant concentrations were

highest near the pad at the site (Figure 2). Although not

the source, the pad was certainly the locus of recent fire

training activity. Figure 2 shows the rough outline of the

major chlorinated alkene (i.e., principally cis-1,2-dichloroethylene, trichloroethylene, and perchloroethylene)

plume, which was restricted to the upper 6 feet (approximately 2 meters) of this water table aquifer in 1993.

Here, the cis-1,2-DCE metabolite of PCE and TCE was

the major constituent, accounting for over 90 percent of

the dissolved contaminants.

In 1994, quarterly contaminant and geochemical monitoring in the ground water was undertaken as the initial

part of a demonstration of intrinsic bioremediation.

Quarterly monitoring results since that time have disclosed variable dissolved concentrations of the chlorinated parent compounds as well as the DCE major

metabolite. It should be noted that vinyl chloride has

been detected only once in mid-field shallow wells. Figure 3 shows representative dissolved concentration

variability of TCE and DCE in the major portion of the

plume from available data over the past 9 years. Farfield wells have generally shown diminished concentrations of DCE, while near-field (i.e., near-pad) wells have

shown some increase, particularly in the last year. The

major plume dimensions evidenced in 1993 (Figure 2)

have remained stable, and iron- and sulfate-reducing or

methanogenic conditions prevail in its interior.1

The question arises in this case whether significant mass

removal has occurred during the course of the investigation. Based on dissolved concentrations, distribution

1

 Chapelle, F.H., S.K. Haack, P. Adriaens, M.A. Henry, and P.M.

Bradley. 1996. Comparison of Eh and H2 measurements for delineating redox processes in a contaminated aquifer. In preparation.

102
variability and roughly ±20 percent precision of sampling

and analysis would have to conclude that no significant

reduction in dissolved mass has occurred in the main

body of the plume.

To approach the net loss of chlorinated alkene compound mass from the plume, 13 borings were made in

1994 along the axis of the plume coincident with dominant ground-water flow direction. A total of 300 core

subsamples were taken by Geoprobe techniques collecting field-preserved samples subsequently analyzed

for major contaminants by static headspace techniques

(2). Companion cores were collected adjacent to these

locations for determination of oily phase, porosity, and

water contents by the methods of Hess et al. (3).

Table 1 contains the average results of these determinations at the near- and mid-field locations of the monitoring wells. It should be noted that, in contrast to the

water samples, which were contaminated by reductive

dechlorination metabolites, the solid-associated chlorinated hydrocarbon distributions were dominated by parent compounds, principally PCE and TCE. It is clear

Figure 2. Plan view of Fire Training Area 2 showing dimensions of major chlorinated VOC plume in ground water in 1993.

Table 1. Comparison of Average Dissolved and Aquifer

Solid-Associated Masses of Total Volatile

Chlorinated Compounds in the Fire Training

Area 2 Plume (masses expressed in milligrams 

per liter aquifer material)a

Location in Major

Plume

Ground

Water

(mg/L)

Aquifer 

Solids

(mg/L)

% of Total

Associated 

With

Solids

Volume

% of

Oily

Phaseb

Near field

(Approximately 

200 feet

downgradient 

from Pad Boring 6;

Well 4S)

0.08 10.5 99% 2.9 

Mid-field

(Approximately 

450 feet

downgradient from

Pad Boring 12;

Wells 8S and 8M)

0.003 6.3 99% 0.006

a

 One liter of aquifer material was assumed to contain approximately

1.75 kilograms of aquifer solids and 300 milliliters of ground water

in average unit volume in major plume. b

 Oily phase determined on field preserved (dry ice freezing) of cores

B-10 and B-12 respectively by the method of Hess et al. (3).

103
from these data that aqueous concentration variability in

determinations of metabolite concentrations are a

negligible portion of the total mass of chlorinated hydrocarbon contaminants. The determinations must include

considerations of oily-phase, solid-associated, and

aqueous masses on a volume basis. It is therefore

necessary to determine the relative mass distributions

of both parent and metabolite compounds to evaluate

net mass losses due to intrinsic bioremediation via reductive dechlorination processes. The apparent trends

in aqueous contaminant concentrations represent

symptoms of the ensemble processes contributing to

net mass loss, particularly in the near field of the presumed contaminant source.

Acknowledgments

The author would like to thank his staff, all collaborators,

and students for their constructive inputs to this slowly

developing but important field. Special thanks to Mark

Henry, Chris Till, Ron Lacasse, Amir Salezedeh, Debbie

Patt, and Patty Laird. NCIBRD staff and collaborators

welcome the contributions and participation of interested

Figure 3. Concentration trends over time for TCE and DCE at Fire Training Area 2 wells.

104
groups in future investigations of promising site characterization and bioremediation cleanup technologies.

References

1. U.S. Geological Survey. 1993. Data submission via memo to U.S.

Air Force Base Conversion Agency, Wurtsmith Air Force Base,

U.S. Geological Survey Lansing Regional Office, Michigan.

2. Barcelona, M.J. 1995. Verification of active and passive groundwater contamination remediation efforts. In: Gambolati, G. and G.

Verri, eds. Advanced methods for ground water pollution central.

International Center for Mechanical Sciences, University of Udine,

University of Padua, May 5-6, 1994, Udine, Italy. Courses and

Series No. 364. Wien/New York: Springer-Verlag. pp. 161-175.

3. Hess, K.M., W.N. Herkelrath, and H.I. Essaid. 1992. Determination

of subsurface fluid contents at a crude-oil spill site. J. Contam.

Hydrol. 10:75-96.

105
Case Study: Eielson Air Force Base, Alaska

R. Ryan Dupont, K. Gorder, D.L. Sorensen, and M.W. Kemblowski

Utah Water Research Laboratory, Utah State University, Logan, Utah

Patrick Haas

U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, Texas

Introduction

One innovative plume management approach that has

been the subject of a great deal of recent interest is that

of intrinsic remediation or natural attenuation, the process of site assessment and data reduction and interpretation that focuses on the quantification of the capacity

of a given aquifer system to assimilate ground-water

contaminants through physical, chemical, and/or biological means. The intrinsic remediation approach is

appropriate for a given site if the plume has not affected

a downgradient receptor and if the rate of contaminant

release from the source area is equal to or less than the

contaminant degradation rate observed at the site.

While many field sampling protocols are available from

a variety of sources describing approaches for collecting

and analyzing data necessary to verify that intrinsic

remediation processes are taking place at a given site,

the connection between these data and decisions regarding source removal activities or estimates of source

lifetime has not generally been presented in the literature. An approach for implementing intrinsic remediation

concepts from data collection through source removal

and source lifetime considerations has been developed

for the U.S. Environmental Protection Agency and the

U.S. Air Force (1-3), and these concepts and procedures

are presented in this paper through a case study at a

mixed solvent/hydrocarbon contaminated site (Site

45/57) at Eielson Air Force Base (AFB), Alaska.

Intrinsic Remediation Protocol

The intrinsic remediation assessment carried out at the

field site at Eielson AFB involved the seven-step process outlined in Figure 1. This process provides a logical

approach to evaluating the feasibility and appropriateness of implementing intrinsic remediation at a given

site and includes: 1) determining whether steady-state

plume conditions exist; 2) estimating contaminant degradation rates; 3) estimating the source mass; 4) estimating the source lifetime; 5) predicting long-term plume

behavior with and without source removal; 6) making

decisions regarding the use of intrinsic remediation and

the impact and desirability of source removal at a given

site; and 7) developing a long-term monitoring strategy

if intrinsic remediation is selected for plume management. Elements of this methodology will be highlighted

through the following case study.

Site Description

Eielson AFB is located in the Tanana River Valley in

Central Alaska, approximately 200 kilometers south of

the Arctic Circle. Most of the base is constructed on fill

material underlain by an unconfined aquifer consisting

of 60 to 90 meters of alluvial sands and gravels overlying a low-permeability bedrock formation (4). The aquifer

Figure 1. Components of the intrinsic remediation assessment

approach.

106
system below the base is bounded to the northeast by

the Yuon-Tanana uplands and is approximately 70 to 80

kilometers wide in the area of the base (5). The direction

of ground-water flow throughout the base is generally to

the north, with ground-water encountered at 2.5 to 3.5

meters below ground surface at various times of the year.

Fire training and fueling operations are believed to be

the source of ground-water contamination at Site 45/57.

Dissolved trichloroethene (TCE) concentrations as high

as 90 milligrams per liter (mg/L) have been observed at

the site and are thought to have resulted from releases

occurring within the last 20 to 40 years. No evidence of

free-phase TCE exists from soil boring or ground-water

monitoring data collected from 1992 through 1995. Anaerobic dechlorination reactions, evident as dechlorination products (cis- and trans-dichlorethene [DCE], vinyl

chloride [VC], and ethylene), have been observed in the

ground water at the site (Figure 2).

Assessment of Intrinsic Remediation at

Site 45/57

Steady-State Conditions

Steady-state conditions were assessed by inspection of

plume centerline concentrations over time (Figure 3),

and through an analysis of integrated plume mass data

for the site. Center of mass (CoM) and total mass results

for Site 45/57 were generated from ground-water

concentration data collected in this field study using a

Thiessen area approach (1-3). Both TCE centerline

concentrations and dissolved plume mass estimates

using a consistent set of sampling locations over time

indicated a decreasing plume mass, with CoM locations

indicating no net plume migration over the sampling

interval. The data indicated a finite source producing a

stable TCE plume at Site 45/57 (2, 3, 6).

Estimation of Contaminant Degradation Rate

Estimation of contaminant degradation rates can be

carried out using dissolved contaminant mass data if a

declining mass of contaminant is observed over time in

the plume. With estimated dissolved TCE concentrations in May 1994 (Mo) and July 1995 (M) being 40.1 and

33.1 kilograms, respectively, and assuming first-order

degradation of TCE in the plume, the estimated TCE

degradation rate (k1) is found by:

k1 = -ln (M/Mo)/t = 

-ln(33.1/40.1)/420 = 0.0005/d (Eq. 1)

where t = the time between sampling events = 14 months

= 420 days.

In addition, degradation rates can be estimated through

the calibration of contaminant fate-and-transport models

to field ground-water data. These models provide

improved estimates of contaminant degradation and

mobility because they integrate transport, retardation, and

degradation processes using site-specific contaminant

and aquifer properties. An analytical, three-dimensional

model developed by Domenico (7), the subject of a previous

Figure 2. Overlay plot of TCE and its degradation products measured in July 1995 at Site 45/57, Eielson AFB, Alaska.

107
paper by Gorder et al. (8), has been incorporated into

the intrinsic remediation methodology described in this

paper and was used to develop an independent estimate of a TCE degradation rate at Site 45/57 based on

July 1995 ground-water data. Calibration of this model

is described elsewhere (7, 8) and involves matching

predicted and measured centerline and cross-plume

contaminant concentrations through the adjustment of

aquifer dispersion properties and contaminant degradation rates. Through this process, a mean TCE degradation rate of 0.0026 per day (0.0006 to 0.007 per day)

was determined.

Estimation of Source Mass and Source

Lifetime

The source of the TCE plume at Site 45/57 had not been

completely identified. Site investigations conducted in

the past by the Pacific Northwest Laboratory and Harding Lawson Associates, as well as soil and ground-water

sampling conducted in the source area by the Utah

Water Research Laboratory, have not identified residual

phase TCE in either the vadose zone or capillary fringe,

nor below the ground-water table. In addition, the finding

of a decreasing dissolved TCE plume mass over time 

strengthens the argument that a residual phase does not

exist at the site. If it is assumed that a distinct freeproduct phase does not exist in the source area, an

estimate of source mass can be made assuming contaminated soil in equilibrium with the measured source

area dissolved TCE concentration, Co. Using this approach, the source area mass was estimated using the

following equation:

Msource = Co (Y) (L) (b) (R) (θ) (Eq. 2)

where Y = transverse source dimension = 22.5 meters;

L = source length in direction of ground-water flow = 15

meters; b = source area thickness = 3 meters; R = TCE

retardation factor = 2.5; and θ = aquifer total porosity =

0.38. Source dimensions were estimated based on interpolation of ground-water data collected within and outside

the source area, while R and θ were based on aquifer-specific characteristics determined from cores collected from

the site. Using these values, a source mass of 37.5 kilograms was estimated to exist at the site.

If the assumption of a finite source is appropriate at Site

45/57, then Equation 1 applies. With maximum source

area TCE concentrations of 90 mg/L and a ground-water

Figure 3. Plume centerline TCE concentrations measured from fall 1992 through July 1995 at Site 45/57, Eielson AFB, Alaska.

108
impact concentration (maximum contaminant level [MCL])

of 5 micrograms per liter (μg/L) established for TCE, an

estimated source lifetime for TCE at this site is:

Source lifetime = ln(5/90,000)/k1 = 

ln(5.6 x 10-5)/(-0.0026) = 10.3 years (Eq. 3)

A worst-case scenario can be formulated for contamination at Site 45/57 using an assumption that a small mass

of residual phase material, which has been undetected

in site investigation activities, exists within the source

area. The dimensions of this residual-phase source area

are defined by the sampling grid within which it must

exist, making its aerial extent no more than 15 by 5

meters (Ys by Ls). The residual phase volume, Sr, contained within the sandy aquifer at Site 45/57 is approximated to be 25 percent of the pore volume (9), or 9.5

percent of the source area volume. Based on a measured source area TCE concentration of 90 mg/L and a

TCE solubility of 1,377 mg/L, the mole fraction of TCE

in residual-phase material is estimated from Raoult’s

law to be 90/1,377 = 0.065, making the estimated concentration of TCE in the residual phase, CTCE:

(MassTCE)/(MassResidual Phase) = 

0.065 (MWTCE/MWResidual Phase) = 

0.065 (131.4/120) = 0.071

Based on these calculations, the estimated mass of TCE

that could exist at Site 45/57 in an unidentified residual

phase is:

MassTCE Residual = 

Ys (Ls) (b) (θ) (Sr) (ρResidual Phase) (CTCE) (Eq. 4)

MassTCE Residual = 

(15 m) (5 m) (3 m) (0.38) (0.25) (1,200 kg/m3

) (0.071)

= 1,677 kg

With this estimate of residual-phase source mass, the

lifetime of the source can be predicted based on the mass

flux of TCE out of this source area, as indicated below:

Mass flux = Ys (b) (v) (θ) (Co) = 

(15 m) (3 m) (0.1 m/d) (0.38) (0.09 kg/m3

) = 

0.15 kg/d (Eq. 5)

where v = ground-water velocity. With this mass flux

value, an estimate can be made for the source lifetime

assuming a residual-phase TCE mass of 1,677 kilograms exists at the site:

Source lifetime = 

MassTCE Residual/Mass flux = 

(1,677 kg)/(0.15 kg/d) = 10,897 d = 29.9 yy (Eq. 6)

As this example illustrates, if residual mass does exist,

the lifetime of the plume is extended significantly, increasing the overall cost of plume management at the

site. More information regarding residual-phase distribution at the site is needed to narrow the range of source

lifetime predictions.

Prediction of Long-Term Plume Behavior

Consideration of long-term plume behavior involves an

evaluation of the plume footprint over time with and

without source removal implemented at a given site.

Following source depletion or removal, the dissolved

plume will begin to contract as the assimilation of contaminants in the aquifer exceeds their release rate from

the source area. The impact of source removal can be

modeled by superimposing a plume with a negative

source concentration, initiated at the time of source

removal or depletion, on top of the existing contaminant

plume (7). This allows the prediction of the time required

for the entire dissolved plume to degrade below a level

of regulatory concern. Based on this information, a decision can be made regarding the expected benefit from

source removal in terms of reducing the time required for

management of the site to ensure long-term risk reduction.

If it is assumed that no free-phase product exists within

the source area of Site 45/57, then the projected source

lifetime is relatively short: approximately 10 years. With

a residual phase existing at the site, the projected

source lifetime is increased to approximately 30 years.

Using the field-data-calibrated Domenico model (6, 7),

a rapidly shrinking plume is predicted to be assimilated to below MCL values within 8 years following

100 percent source removal, as shown in Figure 4.

While removal of the source reduces the projected lifetime of contamination at the site by a factor of two to five,

the cost of such a removal action is high, it is highly

disruptive of current site uses, and the efficiency of

contaminant removal is uncertain. The recommendation

made for this site was against an active source removal

effort because of the marginal and high-cost benefit

expected from such an action.

Long-Term Monitoring Plan for the Site

With implementation of intrinsic remediation recommended at Site 45/57, a long-term monitoring network

is required. To have this network serve multiple purposes,

a combination of upgradient, downgradient, and withinplume monitoring locations is desirable.

Two sets of wells would be installed at Site 45/57 as part

of the long-term monitoring strategy for the intrinsic

remediation plume management approach. The first set,

the long-term monitoring wells, consists of a transect of

plume centerline wells composed of a proposed well

located upgradient of the TCE source area at monitoring

point SP16, three existing wells (45MW01, 45MW03,

109
and 45MW08) located within the observed TCE plume,

and two additional monitoring wells located near the

TCE source area. These wells are used to verify the

functioning of the intrinsic remediation process and allow updating of the conceptual model for plume and

source area configuration over time. The second set of

monitoring wells consists of a transect of three wells

perpendicular to the direction of plume migration, approximately 250 feet (75 meters) downgradient from Monitoring Well 45MW04 to establish the point-of-compliance

(POC) for this site. The purpose of the POC wells is to

verify that no TCE exceeding the federal MCL (5 μg/L)

migrates beyond the area under institutional control.

A sampling frequency of 1 to 2 year intervals was recommended for this site. This interval provides sufficient

data over time to verify plume stability and source area

depletion, at a reasonable frequency based on cost

considerations without compromising human health or

environmental quality.

Conclusion

This paper highlights the application of an intrinsic remediation protocol to a hydrocarbon/solvent contaminated site, Site 45/57, at Eielson AFB, Alaska. This

process involves 1) assessment of steady-state plume

conditions; 2) determination of degradation rates; 3)

estimation of the source term; 4) estimation of the

source lifetime; 5) prediction of the long-term behavior

of the plume with and without source removal; 6) assessment of aquifer assimilative capacity and the desirability of source removal at the site; and 7) development

of a long-term monitoring strategy for verification of

intrinsic remediation process performance and regulatory compliance purposes.

Intrinsic remediation of solvent contaminated ground

water was demonstrated at Site 45/57 through the identification of TCE dechlorination products in the plume

(Figure 1), the recognition of decreasing TCE dissolved

plume mass over time, and calibration of field data to a

Figure 4. Projected TCE plume concentrations 0, 3, and 8 years following source removal or depletion and proposed long-term

monitoring network at Site 45/57, Eielson AFB, Alaska.

110
fate-and-transport model. No residual-phase product

was identified within the source area based on historical

and recent site investigation activities; however, a worstcase estimate was made of the potential residual TCE

mass that might exist within the source area. Source

lifetime estimates ranged from approximately 10 years

without residual-phase TCE to approximately 30 years

with residual-phase material remaining at the site. From

an analysis of source depletion and plume attenuation

rates, it was determined that source removal may reduce

the projected site management lifetime from approximately

20 to 40 years to less than 10 years. Due to the difficulty

and expense of source removal, and to the overall short

timeframe for complete site remediation by intrinsic processes without source removal, long-term monitoring without source removal was recommended and has become

the basis of the record of decision for this site.

References

1. Dupont, R.R., D.L. Sorensen, M. Kemblowski, M. Bertleson, D.

McGinnis, I. Kamil, and Y. Ma. 1996. Monitoring and assessment

of in situ biocontainment of petroleum contaminated ground-water

plumes. Final report submitted to the U.S. Environmental Protection Agency, Analytical Sciences Branch, Characterization Research Division, Las Vegas, NV.

2. Dupont, R.R., D.L. Sorensen, M. Kemblowski, K. Gorder, and G.

Ashby. 1996. Assessment and quantification of intrinsic remediation at a chlorinated solvent/hydrocarbon contaminated site,

Eielson AFB, Alaska. Paper presented at the Conference on Intrinsic Remediation of Chlorinated Solvents, Salt Lake City, UT.

April 2. Battelle Memorial Institute.

3. Dupont, R.R., D.L. Sorensen, M. Kemblowski, K. Gorder, and G.

Ashby. 1996. An intrinsic remediation assessment methodology

applied at two contaminated ground-water sites at Eielson AFB,

Alaska. Paper presented at the First International IBC Conference

on Intrinsic Remediation, IBC, London, UK. March 18-19.

4. U.S. Air Force. 1994. OUs 3, 4, 5 RI Report, Vol. 1. Eielson AFB, AK.

5. CH2M-Hill. 1982. Installation restoration program records search,

Eielson Air Force Base, AK.

6. Utah Water Research Laboratory. 1995. Intrinsic remediation engineering evaluation/cost analysis for Site 45/57, Eielson AFB,

Alaska. Final report. Submitted to the U.S. Air Force Center for

Environmental Excellence, San Antonio, TX, and Eielson AFB, AK.

December.

7. Domenico, P.A. 1987. An analytical model for multidimensional

transport of decaying contaminant species. J. Hydrol. 91:49-58.

8. Gorder, K., R.R. Dupont, D.L. Sorensen, M.W. Kemblowski, and

J.E. McLean. 1996. Application of a simple ground-water model to

assess the potential for intrinsic remediation of contaminated

ground-water. Presented to the First IBC International Conference

on Intrinsic Remediation, London, UK. March 18-19.

9. Parker, J.C., R.J. Lenhard, and T. Kuppusamy. 1987. A parametric

model for constitutive properties governing multiphase flow in porous media. Water Resour. Res. 23:618-624.

111
Considerations and Options for Regulatory Acceptance of Natural Attenuation

in Ground Water

Mary Jane Nearman

U.S. Environmental Protection Agency, Seattle, Washington

Introduction

When approaching areas of ground-water contamination, both technical and regulatory options must be identified and evaluated to ensure compliance with state and

federal regulatory requirements. A strong technical defense presented in the appropriate regulatory framework

is necessary for the selection of natural attenuation as

a component of the remedy. At Eielson Air Force Base

(AFB) near Fairbanks, Alaska, this approach was used

to select natural attenuation as a major component of

the remedy for all areas of ground-water contamination.

This paper summarizes the various options evaluated

for addressing both the technical and regulatory issues

associated with the selection of natural attenuation for

ground-water contamination.

Background

Ground-water contamination at Eielson AFB generally

consists of relatively limited areas of contamination that

have an adverse impact on the beneficial uses of the

aquifer but are not currently posing an immediate risk to

receptors. This type of situation is frequently encountered under the Superfund program and poses a difficult

dilemma from both technical and regulatory perspectives for compliance with the U.S. Environmental Protection Agency’s (EPA’s) Ground Water Protection

Strategy. This strategy, which is outlined in the preamble

to the National Contingency Plan (NCP), includes a goal

to return usable ground waters to their beneficial uses

within a timeframe that is reasonable given the particular

circumstances of the site. The preamble to the NCP

further states that ground-water remediation levels

should generally be attained throughout the contaminated plume, or at and beyond the edge of the waste

management area when waste is left in place.

To comply with the Ground Water Protection Strategy, it

was necessary to first gain an understanding of the

source of the contamination, its fate and transport, and

the feasibility of contaminant removal. Once it was clear

what the technical approach should be, the second task

was to identify the most appropriate regulatory approach to accommodate the proposed technical solution. Options and combinations considered and used to

address ground-water contamination at Eielson AFB are

outlined below.

Technical Options

The first task in the Superfund process is to gain a

thorough understanding of the type of contamination,

the location and extent of the remaining source in both

the unsaturated and saturated zones, and the anticipated fate and transport of the contamination. Once this

is accomplished, alternatives for addressing the contamination can be evaluated.

In the feasibility study, a range of alternatives are developed and evaluated to determine the appropriate level

of source reduction and/or ground-water treatment. The

alternatives considered at Eielson AFB included:

• No action.

• Limited action, including institutional controls and

ground-water monitoring.

• Source removal (either in situ or ex situ) in the subsurface soils and smear zone combined with institutional controls and ground-water monitoring.

• Ground-water extraction and physical/chemical treatment combined with institutional controls and

ground-water monitoring.

The limited action alternative differed from the no action

alternative by the inclusion of institutional controls to

prevent exposure to contaminated ground water. This

definition comes from the NCP (55 Federal Register

[FR] 8711), which states that institutional controls, while

not actively cleaning up the contamination at the site,

112
can control exposure and, therefore, are considered to

be limited-action alternatives.

The selected remedy could be one of the alternatives or

a combination, depending on the degree of active restoration required. Considerations for remedy selection

were the amount of contamination remaining in the

unsaturated and saturated zones and the availability of the contamination for removal and treatment.

The technical evaluation is largely an issue of a

balance between the need for and feasibility of contaminant removal in the unsaturated and/or saturated

zones and the efficiency of natural attenuation. The

NCP (55 FR 8734) addresses this balance by describing

ground-water extraction and treatment as generally the

most effective method of reducing concentrations of

highly contaminated ground water. It subsequently

notes, however, that pump-and-treat systems are less

effective in further reducing low levels of contamination

to achieve remediation goals and allows for the use of

natural attenuation to complete cleanup actions in

some circumstances. If ground-water extraction and

treatment is not warranted due to the low levels present, then the attention is directed at any residual

source of contamination.

At Eielson AFB, residual source contamination typically

fit into two categories. In one category, an equilibrium

existed in which the rate of contaminant migration from

the source was approximately the same as the rate of

natural attenuation in the aquifer. In the second category, the source was continuing to overwhelm the rate

of natural attenuation, resulting in an expanding contaminant plume.

Even in situations in which the system was in equilibrium

and the plume was not expanding, source removal was

evaluated to determine whether reduction of contaminant mass would return the aquifer to its beneficial

uses throughout the plume in a significantly greater

timeframe than natural attenuation alone. This evaluation was not a trivial task given the difficulties in estimating the source term, accurately evaluating the

contaminant fate and transport in the subsurface, and

assessing the effectiveness of source removal. In evaluations conducted at Eielson AFB, modeling was generally the mechanism chosen to evaluate the benefits of

source removal. Generally, these modeling efforts used

conservative assumptions for fate-and-transport analysis and potentially overly optimistic assumptions for

source removal. In combination, the modeling results

indicated a significant benefit of source removal. Results

from subsequent pilot studies, however, indicated low

removal rates for the subsurface contamination, and

contradicted the conclusions of the model. Source removal, therefore, was not expected to significantly reduce risks or remediation timeframes.

Regulatory Options

If, based on the technical evaluation, natural attenuation

was identified as a major component of the selected

remedy, regulatory options were reviewed to determine the most relevant approach for the specific situation. All of the regulatory options considered have

several common requirements or considerations, which

are outlined below.

• The contaminant plume must be contained by the

contaminant source leach rate being in equilibrium

with the rate of natural attenuation or by hydraulic

containment of the leading edge of the aqueous

plume.

• Institutional controls must be effective, reliable, and

enforceable in preventing exposure to the contaminated ground water.

• Further contaminant reduction in the subsurface is

not indicated either due to technical impracticability

or because contamination reduction would not result

in significant risk reduction.

• Ground-water monitoring is necessary to confirm the

conceptual site model developed during the investigation and to ensure that the remedy remains protective.

• Statutory 5-year reviews are required whenever the

selected remedy will leave contamination on site

above levels that allow for unlimited use and unrestricted exposure (NCP §300.430(f)(4)(ii)).

At Eielson AFB, the regulatory options considered are

described below.

Alternate Concentration Limits

Alternate concentration limits (ACLs, 55 FR 8732) are

considered when the ground water has a known or

projected point of entry to surface water with no statistically significant increases in contaminant concentration

in the surface water. Natural attenuation is the mechanism for cleanup in ground water between the contamination and the point of surface-water discharge. If ACLs

are used, the remedial action must include enforceable

measures (e.g., institutional controls) that will preclude

human exposure to the contaminated ground water.

ACLs should only be used when active restoration of the

ground water is not practicable (55 FR 8754).

For Eielson AFB, ACLs were not applicable because

contaminated ground water did not discharge into surface water on base.

Alternate Points of Compliance

As stated previously, remediation levels should generally be attained throughout the contaminated plume, or

at and beyond the edge of the waste management area.

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For situations in which the risk of exposure is very slight

(i.e., because of remoteness of the site), however, alternate points of compliance may be considered in combination with natural attenuation provided contamination

in the aquifer is controlled from further migration (55 FR

8735). When releases from several distinct sources in

close geographical proximity cause a plume, the most

effective cleanup strategy may address the problem as

a whole, with the point of compliance encompassing the

sources of release (55 FR 8753).

For Eielson AFB, an alternate point of compliance was

established for a previously used base landfill. Consistent with expectations outlined in the Ground Water

Protection Strategy, an alternate point of compliance

was established at the edge of the waste management

area (i.e., the landfill boundary).

Technical Impracticability Waiver

The Superfund regulations allow Applicable or Relevant

and Appropriate Standards, Limitations, Criteria, and

Requirements (ARARs) to be waived under certain circumstances if the remedy can be demonstrated to be

protective. One of the six ARAR waivers provided by the

Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA §121(d)(4)) is technical

impracticability (TI). The use of the TI waiver requires a

demonstration that compliance with ARARs, including

maximum contaminant levels (MCLs) or non-zero maximum contaminant level goals (MCLGs), is technically

impracticable from an engineering perspective. A demonstration that ground-water restoration is technically

impracticable generally should be accompanied by a

demonstration that contaminant sources have been or

will be identified and removed or treated to the extent

practicable.

In the event that the requirements outlined above are

demonstrated and a TI waiver is invoked, an alternative

remedial strategy must be established that includes 1)

exposure control using enforceable, reliable institutional

controls such as deed notifications and restrictions on

water supply well construction and use; 2) source control through treatment or containment where feasible

and where significant risk reduction will result; and 3)

aqueous plume remediation by preventing contaminant

migration (e.g., through hydraulic containment), establishing a less-stringent cleanup level, and/or using natural attenuation.

At Eielson AFB, TI waivers are being invoked for two

lead contamination plumes caused by leaded gasoline

releases. The lead has degraded from the organic lead

contained in the gasoline to a relative immobile inorganic lead. Ground-water contamination is confined to

areas approximately 600 feet in length. Ground-water

remediation is technically impracticable because the inorganic lead is so strongly adsorbed to the soils.

Reliability of institutional controls is very good; Eielson

AFB is not a target of base closure. These institutional

controls preventing use of the ground water will protect

human health.

Selection of Natural Attenuation With or

Without Institutional Controls

Natural attenuation is generally recommended only

when more active restoration is not practicable, cost-effective, or warranted because of site-specific conditions

(e.g., ground water that is unlikely to be used in the

foreseeable future and therefore can be remediated

over an extended timeframe), or in situations in which

the method is expected to reduce the concentration of

contaminants in the ground water to remediation goals

in a reasonable timeframe (i.e., in a period comparable

to that achievable using other restoration methods). Institutional controls may be necessary to ensure that

such ground waters are not used before levels protective of human health are reached (55 FR 8734).

The limited action alternative (natural attenuation with

institutional controls and ground-water monitoring) has

been selected for numerous areas at Eielson AFB contaminated with both petroleum compounds and chlorinated organics. For all of these areas, the plume is

believed to have reached equilibrium where the rate of

contaminant leaching from the source is balanced with

the rate of natural attenuation. The use of institutional

controls was also a critical component of the selected

remedy to prevent exposure to contaminated ground

water until ARARs are achieved throughout the aquifer

and beneficial uses are restored.

Building a “Safety Net”

As with any environmental decision, it is prudent to

develop a “safety net” of contingencies to alleviate apprehensions associated with the selection of natural

attenuation.

The uncertainty associated with environmental decisions, specifically the selection of natural attenuation,

was addressed at Eielson AFB through the use of the

observational method. Key components of the observational method are 1) a decision based on the most

probable site conditions; 2) identification of reasonable

deviations from those conditions; 3) identification of parameters to monitor to detect deviations; and 4) preparation of contingency plans for each potential deviation

(1). The conceptual site model developed through the

investigation will be tested and confirmed through continued ground-water monitoring. A phased approach

with contingencies for additional remediation was established in the event that the conceptual site model is not

confirmed.

114
In addition, statutory 5-year reviews require an evaluation for additional remediation if it becomes apparent

that the remedy is not protective of human health or the

environment.

For sites where natural attenuation is selected and

ground-water contamination remains, reliable institutional controls are a critical component for a protective

remedy. For federal facilities, institutional controls to

prevent exposure to contaminated ground water are

generally effective and reliable and are further enhanced

by the statutory requirements for property transfer under

Section 120(h) of CERCLA.

Summary

Existing regulations and guidance were used to support

a technically defensible selection of natural attenuation

as a component of the selected remedy for all groundwater contamination areas at Eielson AFB. The selected

remedies included a sound regulatory framework that is

consistent with EPA’s Ground Water Protection Strategy.

Continued monitoring, contingencies for implementing

additional remediation if necessary, statutory 5-year protectiveness reviews, and the base closure requirements

of CERCLA Section 120 provide additional checks and

reviews to ensure that the selected remedy remains

protective.

Reference

1. Brown, S.M., D.R. Lincoln, and W.A. Wallace. 1989. Application of

the observational method to remediation of hazardous waste sites.

CH2M Hill, Bellevue, WA. April.

115
Lessons Learned: Risk-Based Corrective Action

Matthew C. Small

U.S. Environmental Protection Agency, Office of Underground Storage Tanks,

Region 9, San Francisco, California

Introduction

With over 300,000 leaking underground storage tanks

(LUSTs) nationwide (1) that have contaminated soil,

ground-water, and surface-water resources, operation

of an underground storage tank (UST) clearly is no

longer a casual undertaking. U.S. Environmental Protection Agency (EPA) regulations (2) created guidelines

and requirements for safe and responsible operation of

USTs, with provisions for early leak detection, leak reporting, financial responsibility, and cleanup of leaks.

Using a franchise approach, these EPA regulations have

been adopted—and sometimes supplemented—by

state UST programs in an effort to clean up existing

leaks and prevent future leaks. Most state programs

also instituted petroleum cleanup funds to assist owners

and operators of USTs in complying with financial responsibility requirements and to provide money for

cleanup of existing releases.

Initially many state programs required cleanup of LUST

sites to very low levels of compounds of concern (petroleum products) or in some cases even to background or

nondetectable levels at all sites, regardless of the actual

hazard posed by the site. These levels often proved to

be unattainable both technologically and economically,

however, making site closure difficult to obtain, stalling

property transfers, driving cleanup costs higher, and

frustrating all parties concerned. Even though state UST

programs have made strong efforts to prioritize sites for

cleanup and streamline oversight, only about 45 percent

(1) of known LUST sites nationwide had cleanups completed by the end of 1995, and some state cleanup funds

were almost exhausted, bordering on insolvency.

The American Society for Testing and Materials’ document “Standard Guide for Risk-Based Corrective Action

(RBCA) Applied at Petroleum Release Sites” (3) was

introduced as a logical framework for determining the

extent and urgency of corrective action required at a LUST

site. The RBCA standard provides a tiered approach to

evaluating risk, progressing from generic, conservative

calculation of risk-based screening levels (RBSLs) to

more site-specific target levels (SSTLs) derived from

increasingly site-specific data. Only completed pathways from contaminant source to potential receptors are

evaluated. Risk levels are used to back-calculate acceptable concentrations (RBSLs or SSTLs) for each

compound of concern for each completed pathway. Site

conditions are then compared with the RBSLs or SSTLs

to determine the extent of cleanup required.

Currently 43 states have entered the RBCA training process. Of these, 6 have implemented RBCA, 12 are working

on the program design, and 25 are still training (4). This

paper presents some of the lessons learned during the

process of developing and implementing RBCA.

Lessons Learned

RBCA Program Development

The process of implementing an RBCA program at the

state level requires commitment on the part of the entire

organization. Training is usually required for all interested parties, including state regulators, environmental

consultants, UST owners and operators, and the general public. All of these interested parties or stakeholders

must be involved in the process up front to avoid misconceptions and misunderstandings. It is especially important that key decision-makers understand and “buy

into” the process early on.

All interested parties must be involved in making the risk

management decisions necessary for implementing

RBCA. This includes determination of risk levels, pathways to be considered, compounds of concern, and

other key parameters used to calculate Tier 1 RBSLs.

Once the RBSLs have been calculated, it is important

to avoid the temptation to adjust the parameters in an

effort to make the RBSLs fit some preconceived or

pre-existing level.

116
Creation of RBCA lookup tables and cleanup numbers

can be a contentious process. The general public and

many regulators often want to retain cleanup to background for Tier 1 regardless of the actual hazard posed

by the site. Education is the only way to solve this

problem. People must be made aware that background

levels are unrealistic and often unattainable goals at

most sites, given available technologies and resources.

The RBCA process provides a method for determining

cleanup goals to adequately protect human health and

the environment.

Because RBCA often involves some major philosophical

changes, regulatory and policy changes may also be

needed. Stakeholders may feel wary if a state changes

from fixed, numerical cleanup standards to risk-based

cleanup goals without legislative authority to do so.

These stakeholders may feel more comfortable if the law

says the change is appropriate. Legislative mandates,

however, may also impose limitations that impede or

compromise RBCA implementation. Therefore, a balanced approach is required to ensure that regulators

and other stakeholders feel that the implementation

process is legitimate but not that RBCA is being forced

upon them against their will.

RBCA Program Implementation

Implementation can be difficult initially. States should

have a clear and thorough strategy for implementing as

complete a RBCA program as possible before they start.

If not, the state may end up haphazardly creating pieces

of the program in response to problems and issues as

they arise. For example, requirements and definitions

relating to key issues such as alternate points of compliance, acceptable sampling methodologies, and extent

of site assessment for Tier 1 versus Tier 2 should be

available before program implementation.

Modeling data can sometimes be misleading. In particular, estimates of indoor air concentrations that result

from a given soil concentration are often overestimated.

Monitoring and sampling are important to confirm any

modeling estimates used in the RBCA process.

RBCA is not a cure-all—some difficult issues will remain.

For example, third party liability for compounds of concern left behind at LUST sites following property transfer

may still cause uncertainty and potential problems. A site

closed using cleanup levels determined through RBCA

or by previous standards will leave some level of compounds of concern in place. In most cases, however,

RBCA provides a more sound and defensible basis for

site closure and levels of compounds of concern left in

place should third-party issues arise. Another issue is

the fear of having sites reopened after a closure letter

has been granted. Again, RBCA provides a clear, logical

framework for making site closure decisions that can

be easily revisited should the closure be questioned in

the future.

Some consultants and regulators may view RBCA as a

threat to their livelihood. Long cleanup times and low site

closure rates ensure continued work for both consultants and regulators. Sites will have to be closed eventually, however, and the RBCA process is one of the best

ways to achieve this goal.

Considerations for the Future

The implementation and acceptance of RBCA involves

a shift in perspective from asking the question “How

much or what levels of the compounds of concern can

we cleanup?” to asking “How much of the compounds

of concern can we safely leave in place?” Again, this is

not a significant change in the way we manage sites

because some level of compounds of concern have

always been left in place. RBCA simply asks the question early in the cleanup process to better utilize resources to clean up sites posing the most threat. We

must, however, guard against allowing ourselves to ask

“How much contamination can we allow to happen?”

It is extremely important to supplement an RBCA program with a strong program of leak prevention and

early leak detection.

References

1. Lund, L. 1996. EPA fiscal year 1996 semi-annual (1st and 2nd

quarter) UST activity report. May 3.

2. U.S. EPA. 1995. 40 Code of Federal Regulations, Part 280. July 1.

3. American Society for Testing and Materials. 1995. Standard guide

for risk-based corrective action (RBCA) applied at petroleum release sites. E-1739-95. September 10.

4. Partnership in RBCA Implementation (PIRI). 1996. RBCA implementation summary graph from EPA/ASTM data. February 6.

117
Informal Dialog on Issues of Ground-Water and Core Sampling

Donald H. Kampbell

National Risk Management Research Laboratory, 

R.S. Kerr Research Center, Ada, Oklahoma

An assessment of natural attenuation can be no better

than the site characterization activities that collect the

data used in the assessment. The following issues

should be considered when planning for field sampling:

• When is a conventional well required, and when can

a Geoprobe or CPT push technology be used for well

installation?

• What are the advantages and disadvantages of conventional wells, mini wells, or water samples collected with a Hydropunch and a CPT rig?

• How much water should be purged to prepare a well

for sampling?

– What is the evidence that a well is ready for sampling?

– What should be measured: conductivity, temperature, pH, turbidity, oxygen, or redox potential?

• What flow rate should be used to purge a well?

• What flow rate should be used to sample a well?

• What is the best way to measure oxygen in ground water?

– What are the relative advantages of oxygensensing electrodes and indicator dye kits?

– What level of training is required to use the equipment intelligently?

– What problems may arise?

• What is the best way to measure sulfide in ground water?

– What are the relative advantages of lead acetate

indicator paper, colorimeter assays, or ion specific

electrodes?

– How accurate should the measurement be?

– What problems may arise?

• What is the best way to measure iron(II) in ground water?

– What field methods are available?

– How accurate should the measurement be?

– What problems may arise?

• How should samples for methane, ethylene, and ethane be collected?

– Where can the samples be analyzed, and how

much should analysis cost?

• What is the best preservative for ground-water samples?

• How is ground water sampled for hydrogen?

– What are the limitations of this technique?

– What problems may arise?

• Must alkalinity be analyzed in the field, or can samples be shipped back to the laboratory?

• When should ground-water samples be acquired for

volatile fatty acids (VFAs)?

– How are VFA samples stabilized and extracted?

• What is the best way to collect core samples?

– What are the advantages and disadvantages of

available equipment?

– How should the samples be stabilized for analysis

of contaminants?

– What is the best way to screen samples in the field?

– How should the samples be stabilized for analysis

of microbial indices?

• How is soil gas analysis used to locate and identify

nonaqueous-phase liquid source areas?

– What parameters should be measured?

– What equipment is available?

• What new analyses could be developed to improve

understanding of natural attenuation?

– What new tracers might be used?

118
• What are the cost tradeoffs of these analyses compared with the benefit of improved understanding of

plume behavior?

• What should be the relative investment in sample

acquisition, sample analysis, data reduction, mathematical modeling, and report preparation?

• How many wells or cores samples are needed?

– To examine plume flow velocity?

– To examine proximity to sensitive receptors?

119
Appropriate Opportunities for Application of Natural Attenuation in the Civilian

Sector

Fran Kremer

U.S. Environmental Protection Agency,

Office of Research and Development, Cincinnati, Ohio

Introduction

This paper examines the historical and current use of

natural attenuation, with an emphasis on Superfund

sites and underground storage tank sites. The paper

also presents the results of a survey of natural attenuation policies and guidelines for natural attenuation in

both ground water and soils. Finally, it briefly addresses

the factors that are important for promoting greater acceptance of natural attenuation as a remediation approach for contaminated sites in the civilian sector.

Analysis of Ground-Water Records of

Decision

Earlier this year, EPA’s Superfund office completed an

analysis of ground-water Records of Decision (RODs)

from 1982 to 1994. The overall trend in ground-water

RODs using natural attenuation is shown in Figure 1. In

addition, in 1995, 26 RODs were signed using natural

attenuation; 10 percent of RODs for ground water were

specifically related to natural attenuation. Of the 73

RODs shown in Figure 1, 41 were Potentially Responsible Party (PRP) lead proposals.

EPA’s Technology Innovation Office analyzed these data

based on the use of natural attenuation in the 10 EPA

Regions. These results are summarized in Figure 2.

Region 1 through 6 are the heaviest users of natural

attenuation, while Regions 7 through 10 were marginal

users. No definitive conclusions have been drawn from

this, but the arid climate in western regions may not be

as conducive to selection of natural attenuation in

ground water. Region 4 had the greatest number of

natural attenuation RODs (16 sites) followed by Region

5 (12 sites). Regions 5 and 9, in general, are the most

frequent users of both active and passive bioremediation processes.

Figure 3 presents additional analysis of the 73 RODs by

type of site. The figure shows that 35 of the RODs

originated from industrial and municipal landfills, with

fewer from chemical and industrial manufacturing and

other types of sites. It should be noted that one out of

five National Priority List (NPL) sites is a landfill, a fact

that will skew these results.

Figure 4 shows the contaminants present at the 73 sites

for which natural attenuation was selected. Solvents

(chlorinated and nonchlorinated) and inorganics were

present most often, with both types of contaminants

present at 35 sites. It should be noted that more than

one contaminant can be identified for a given site.

Natural attenuation is often used with a combination of

other active approaches. This was the case for all but 6

of the 73 sites. In many cases the contaminant concentrations were found to be very low. Contingencies exist

for other active measures in the event that the natural

attenuation approach does not prove to be effective for

the particular site.

0

5

10

15

20

85 86 87 88 89 90 91 92 93 94

Fiscal Year

Number of RODs

1 1

3 3

7 7

9

12

13

17

Figure 1. Ground-water RODs using natural attenuation.

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Figure 2. Number of Natural Attenuation RODs per state by EPA region (1982-1994).

0 5 10 15 20 

Number of Sites*

Coal Gasification

Fuel Storage

Junkyard

Wood Treatment

Oil Reclamation

Metal Plating/Mining

Chemical/Industrial Mfg.

Industrial Landfill

Municipal Landfill

*Some sites have more than one use

Figure 3. Types of sites at which natural attenuation was selected.

0 5 10 15 20 25 30 35 

Number of Sites*

PCBs, PestIcides

PAHs, Phenols

BTEX, MTBE

Solvents

Inorganics

*Some sites have more than one contaminant

Figure 4. Contaminants present at sites for which natural attenuation was specified.

121
The Use of Natural Attenuation at Leaking

Underground Storage Tank Sites

As recently as 1993, landfilling was the predominant

remediation approach for addressing contaminated soils

at underground storage tank sites, while pump-and-treat

was the most common option for ground-water treatment at such sites. As of 1995, natural attenuation of

soils at such sites constituted 28 percent of the treatment scenarios utilized, second to landfilling (34 percent), while natural attenuation of ground water

increased to 47 percent. Natural attenuation will play a

vital role in remediation at underground storage tanks

sites. Depending on the location, however, bioventing

may play a more predominant role, particularly in urban

areas. The use of other options for remediation at underground storage tank sites is shown in Figure 5.

Analysis of State Guidelines and Policies

for Natural Attenuation

EPA recently conducted a survey to collect information

from the states regarding their policies on natural attenuation, asking the following questions:

• Do they encourage or discourage the use of natural

attenuation?

• Are there any formal or informal policies or guidelines

for natural attenuation?

• Do they use any particular model when deciding on

natural attenuation?

• Would they consider compounds other than petroleum hydrocarbons for natural attenuation?

The survey found that most states neither encouraged

or discouraged the use of natural attenuation, but were

willing to entertain proposals. The survey found no

trends regarding the use of particular models for deciding on natural attenuation; generally the states have

requirements for peer-reviewed models and will consider data from such models.

The extent of state policies for natural attenuation in

ground water is summarized in Figure 6. As the figure

shows, the states of North Carolina and New Jersey

have policies on ground water. Both require:

• Full plume definition and receptor analyses

• Appropriate modeling to predict plume degradation

• Source removal or control

• Monitoring program to demonstrate natural attenuation

North Carolina developed a rule on natural attenuation

in 1993, and has approved approximately 150 sites for

natural attenuation. Most are petroleum sites, but some

have included solvents and even lead. Under the rule,

the state:

• Looks at sorption and source removal as part of

natural attenuation, allowing the consideration of

natural attenuation for lead.

• Requires that the potential for toxic by-products be

assessed.

• Might not require source removal if no further leaching to ground water is proven.

• Requires that future land use options in the vicinity

of the site be examined.

New Jersey’s natural attenuation rules require the party

seeking to use natural attenuation to:

• Assess potential impacts, ensure no impact to receptors, and remove/remediate sources.

• Identify current and potential ground-water use

based on a 25-year plan.

• Include long-term monitoring in the costs of the remedy.

• Determine the duration and frequency of sampling

through historical data.

The New Jersey rules also allow the use of natural

attenuation at sites deemed technically impractical for

active remediation.

With respect to soils, both Texas and Wisconsin have

developed formal guidance on using natural attenuation

for soils. Several other states are developing soils guidance for natural attenuation: Alaska, Arizona, Florida,

Michigan, South Dakota, Vermont, and West Virginia.

Factors That Promote Greater Acceptance

of Natural Attenuation

Experience has shown that promoting greater acceptance of natural attenuation involves several key components. First, good communication is critical in getting

natural attenuation implemented at a site. It should be

conveyed that natural attenuation is a responsible, manNatural 

Attenuation 

(48%)

Pump-and-Treat

(29%)

Dual Phase Extraction

(2%)

Biosparging

(4%)

In Situ Bioremediaton

(4%)

Air Sparging

(13%)

Adapted from Dana Tulls, 

EPA UST/LUST National Conference Talk, 

March 11, 1996

Figure 5. Use of ground-water cleanup technology at UST sites.

122
aged remediation approach, not a “walk-away” option.

All parties, including regulatory agencies, need to be

involved in the consideration of natural attenuation very

early in the process.

Second, it is critical to have the information to support

natural attenuation. This includes site-specific data and

analyses that demonstrate occurrence, a defensible

conceptual model, and defensible predictive models,

where appropriate.

Other factors that are important for acceptance of natural

attenuation include plans to control, treat, or remove

sources; to thoroughly monitor plume and downgradient

areas; and to implement contingencies for other measures

in case natural attenuation fails to meet the desired goals.

Finally, the burden of proof is on the proponent, not the

regulator. In implementing natural attenuation, as in any

active remediation process, the project is not complete

until the goals of the regulatory agency have been met.

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Do not have any formal policy

Figure 6. State policies regarding natural attenuation in ground water.

123
Appropriate Opportunities for Application: U.S. Air Force and Department of

Defense

Patrick E. Haas

U.S. Air Force Center for Environmental Excellence (AFCEE), Technology Transfer Division,

Brooks Air Force Base, Texas

Introduction

U.S. Air Force spill or release sites are most commonly

contaminated with petroleum hydrocarbons. Chlorinated solvents, such as trichloroethylene (TCE), are the

second most common group of Air Force site contaminants. Sites contaminated with heavy metals are a distant third. The objective of this paper is to outline the role

of natural attenuation in an overall risk-based strategy

for remediating chlorinated solvent sites. The possible

benefits and importance of considering natural attenuation throughout the site characterization and remedial

phases are discussed.

Site Characterization Process

The goal of site characterization is to define the nature

and extent of contamination as well as the fate and

transport of constituents of concern over time. Natural

attenuation is the combination of the nondestructive

processes of dilution, sorption, and volatilization and the

destructive processes of biotic and abiotic degradation.

Thus, although the objective of site characterization is

to have a reasonable understanding of fate and transport, specific quantification of potentially major processes, such as biodegradation, often is overlooked in

traditional site characterizations—an egregious oversight at sites contaminated with petroleum hydrocarbons.

The Air Force Center for Environmental Excellence (AFCEE), Technology Transfer Division, has completed 50

site characterizations, inclusive of natural attenuation

parameters. Each of these investigations has demonstrated the dominant role natural biodegradation plays

in the fate and transport of petroleum hydrocarbons. The

message is not simply “Natural attenuation happens,

don’t worry,” but that including natural attenuation parameters gives the owner and regulator the ability to

discern which sites will pose a potential future risk and

which sites do not or will not pose any unacceptable risk.

In other words, fate and transport cannot be accurately

predicted without the inclusion of natural biodegradation

effects.

Natural attenuation can be evaluated as a treatment

option alongside more intrusive source removal approaches. The final remedial action may include natural

attenuation alone or in concert with these more intrusive

source removal approaches.

Background and Lessons Learned From

Petroleum Sites

The California LUFT Historical Case Analyses (1) provides dramatic confirmation of the dominant effects of

natural attenuation on fuel-contaminated sites. The

“demographics” of 1,500 State Leaking Underground

Fuel Tank (LUFT) cases with confirmed ground-water

contamination were:

• At 50 percent of sites, mean ground-water depth was

less than 15 feet, 25 percent were less than 7.5 feet.

• Ground-water plumes were less than 200 feet at 85

percent of sites.

• If all fuel-contaminated ground water was remediated, statewide water basin capacity would increase

by 5/10,000ths of 1 percent.

The current body of evidence supports the above profile

as being primarily the result of natural biodegradation.

Thus, not considering the effects of natural attenuation

would be an oversight.

Natural Attenuation of Chlorinated

Organics

The current debate centers around whether natural

processes can effectively “treat,” contain, and remove a

significant mass of chlorinated organics from ground124
water plumes. A growing body of evidence supports the

contention that natural degradation processes are important in understanding the fate and transport of chlorinated organics. Thus, a practical, field-based protocol

for the estimation of these effects is needed and is

currently under development.

Almost invariably, any concept or approach goes

through the sequential stages of first being ridiculed and

considered preposterous, secondly being dismissed as

trivial, and thirdly being proclaimed as his or her own

idea. Overall, natural attenuation of fuels is currently

somewhere between the second and third stages of

acceptance. In contrast, attitudes toward natural attenuation of chlorinated organics appear to be closer to the

first stage or, more fairly, in the “prove it to me” stage.

The point of the above discussion is to underline that the

inclusion of natural attenuation in the site characterization process improves the quality and accuracy of

the site investigation and facilitates defensible, riskbased decision-making. Also, by proceeding to the field

and evaluating natural attenuation as part of the site

characterization process, data will be generated that will

answer many of the practical questions regarding the

applicability of natural attenuation as part of the cleanup

process. These data are also directly useful in the evaluation and design of other remedial systems.

Given that the contribution of natural attenuation processes to chlorinated organics cleanup is less established than that of petroleum hydrocarbons, the fact

remains that the cost and potential risk implications

of many chlorinated organic contaminated sites are

higher than those of the majority of fuel-contaminated

sites. Many may currently predict that natural attenuation may not effectively contain or abate chlorinated

organic plumes. It is important to point out that less than

5 years ago the bulk of the environmental community

believed that oxygen was the primary electron acceptor

in the natural biodegradation of petroleum-contaminated

ground water. The above data greatly dispels this theory.

In short, natural attenuation alone may only apply to 20

percent of all chlorinated solvent sites.1

AFCEE Technology Transfer Focus: Field

Protocol

Regarding the natural attenuation of chlorinated organics, the AFCEE Technology Transfer Division is convinced that the current focus should be on “How can we

measure natural attenuation processes, like biodegradation, during the site characterization process?” and

“How can we reduce the cost of sampling and analysis

procedures?” Current efforts center around the development and validation of cost-effective and diagnostic

natural attenuation sampling and analysis procedures.

Thus, similar to the Technical Protocol for Implementing

Intrinsic Remediation with Long-Term Monitoring for

Natural Attenuation of Fuel Contamination Dissolved in

Ground Water (2), a field-based natural attenuation protocol for chlorinated organics is under development and

is currently in draft form. Development and field validation of this protocol are the first and most necessary

steps toward the development of an integrated risk management strategy for this class of sites. Natural attenuation is not selected presumptively. In fact, the selection

of natural attenuation alone or in concert with supplemental remedial systems is based on the collection of a

reasonable weight of site-specific supporting evidence.

Lines of Evidence

Site-specific data in support of natural attenuation include:

• Documented loss of contaminants at the field level

(e.g., shrinking, stable, or retarded plumes).

• Geochemical indicators of degradation (e.g., degradation byproduct formation, electron acceptor usage,

or redox potential).

In some cases, it may be beneficial to supplement the

above evidence with laboratory microcosm data. Laboratory microcosms can be used to demonstrate that

under specific conditions the constituent of concern is

biodegraded, as well as to provide estimates of degradation rates. Microcosm data are particularly relevant if

the laboratory conditions closely approximate field conditions. Confirmed contaminant removal and containment strongly establish that conditions are adequate to

support attenuation; however, confirmatory microcosm

data can more clearly elucidate the role of biodegradation.

Controlling Factors of Natural Attenuation

The natural attenuation of petroleum hydrocarbons is

largely controlled by the availability or utilization of electron acceptors (i.e., sulfate, nitrate, carbon dioxide) (see

the BIOSCREEN user’s guide [3], for details on electron

acceptor utilization profiles at 28 fuel sites). Thus, natural attenuation of petroleum hydrocarbons is overall an

electron acceptor-limited process. In contrast, natural

attenuation of chlorinated organics appears to be an

electron donor-limited process for parent compounds

such as tetrachloroethylene (PCE) or trichloroethylene

(TCE). That is, co-metabolism or reductive dehalogenation of chlorinated organics requires the presence of a

substrate or electron donor. As a result of this dependence, the draft Technical Protocol for Natural Attenuation of Chlorinated Solvents in Ground Water (4)

outlines different types of plume behavior.

1

 Wilson, J.T. 1996. Personal communication between the author and

John T. Wilson, U.S. Environmental Protection Agency, National Risk

Management Research Laboratory, Ada, OK.

125
Types of Plume Behavior

Type 1 behavior is associated with areas within the

plume where anthropogenic carbon (e.g., petroleum hydrocarbons or landfill leachate) is present and being

utilized as a carbon source or electron donor. Type 2

behavior is associated with areas within the plume

where a source of naturally occurring carbon is present

and being utilized. Type 3 behavior is associated with

areas within a plume where an inadequate carbon

source(s) is present or being utilized. All of these behaviors can be exhibited within a single plume simultaneously or in various combinations. The behavior profile

could also change over time.

These plume behaviors or profiles are important from

the perspective of 1) determining whether site constituents will be degraded over time, thereby preventing

exposure to receptors; 2) determining which metabolic

byproducts (e.g., vinyl chloride) might be produced; and

3) determining whether metabolic byproducts, such as

vinyl chloride, will subsequently be destroyed.

Common Primary Remedial Goals

Plume containment is one of the primary remedial goals

as well as a traditional remedial action (i.e., pump and

treat). Thus, by understanding natural attenuation and

plume behaviors, one can evaluate whether plume containment can be achieved with natural attenuation alone

or in concert with supplemental remedial systems. Another primary remedial objective is to achieve reductions

in the concentrations of constituents of concern to riskbased levels. Considering natural attenuation effects will

assist in the assessment of when concentration-based

goals will be met.

By way of example, the presence of Type 1 or 2 areas

may promote the breakdown of parent compounds (e.g.,

PCE, TCE), resulting in significant mass reduction. In

other words, this area, typically near the source, may

represent an effective treatment zone or system. Yet,

vinyl chloride can also be produced as a metabolic

byproduct in these typically anaerobic zones. The presence of vinyl chloride might be viewed as a disastrous

end result; however, vinyl chloride has been shown

overall to be more biodegradable than the parent compounds and to be biodegradable under aerobic and/or

iron-reducing conditions. Thus, with respect to risk management, a point of critical importance would be to

determine whether the vinyl chloride will enter into a

zone where the geochemical conditions support biodegradation. In certain instances, the zone where the constituent(s) of concern is located or the area directly

downgradient can be viewed and managed as an effective treatment zone.

Natural Attenuation and Alternative Plume

Management

Understanding natural attenuation effects allows the

consideration and implementation of alternative remedial approaches. Pump and treat systems are traditionally used to achieve hydraulic containment; however,

more limited ground-water pumping could be used to

create aerobic conditions in downgradient zones (e.g.,

aerobic) that would effectively complete the biodegradation of vinyl chloride. Alternative methods of groundwater aeration are also possible. Pumping upgradient of

source zones could slow ground-water velocities sufficiently so that the kinetics of biodegradation are compatible with plume containment. These approaches may

be “pump and not have to treat” alternatives. Other

traditional technologies can be used in concert with

natural attenuation in conventional source removal approaches (e.g., soil vapor extraction [SVE]). This combination would involve reducing contribution from the

source term down to levels where natural attenuation

mechanisms can remediate constituents of concern.

When evaluated as part of the site-specific feasibility

study, these approaches may be far more favorable

because of their lower pumping rates, infrastructure requirements, and operation and maintenance requirements compared with trying to achieve total hydraulic

control or complete source removal. In many cases, total

hydraulic control (via pump and treat) and extensive

source removal, especially below the water table, are

technically impracticable. Natural attenuation with or

without supplement remedial systems, however, may be

technically practicable.

Conceptual Models

TCE proved to be a highly effective degreasing agent.

As a result, the waste stream released to the environment often contained a combination of degreasing agent

and solvated petroleum hydrocarbons. Chlorinated solvents are also typically found at Air Force fire training

areas because waste solvents were combined with

waste fuels during training exercises. Thus, chlorinated

solvents are commonly found to be codistributed with

anthropogenic carbon sources. This codistribution can

often result in zones of Type 1 behavior. Due to natural

attenuation effects, typical petroleum hydrocarbon

plumes are relatively short. And in fact, in the California

LUFT Historical Case Analyses (1), 85 percent of the

1,500 ground-water plumes surveyed were less than

200 feet in length. The volume of the Type 1 zone may

be roughly delineated by the volume of the petroleumcontaminated area. Given that petroleum hydrocarbons

appear to be excellent microbial substrates, however,

this area of codistributed solvents and fuels will likely

have significantly higher reaction kinetics than areas in

which the substrate is less favorable or absent. Even if

126
its extent is somewhat limited, this area may be an

effective treatment zone.

Regarding the remediation of fuel sites, AFCEE has

strongly advocated source removal at sites where the

natural attenuation assessment has demonstrated that

the contribution from the source term exceeds the assimilative capacity of the ground water. This concept of

supplemental source removal should be directly applicable to sites where chlorinated organics are the single

predominant species. Source removal processes that

preferentially remove petroleum hydrocarbons or in effect alter the ratio of fuels to chlorinated organics could,

however, theoretically reduce the rate and quantity of

natural attenuation in the Type 1/source zone.

Possible Effects of Source Removal

The Plattsburgh Air Force Base case study presented at

this symposium by Wiedemeier et al. illustrates a decommissioned fire training area where waste chlorinated solvents, primarily TCE, are codistributed with

petroleum hydrocarbons. Free-phase petroleum hydrocarbons are present and detectable in significant apparent thicknesses in site monitoring wells. All the dissolved

TCE at this site appears to be transformed within close

proximity to the fuel-contaminated region. Planned light

nonaqueous-phase liquid (LNAPL) free product recovery and SVE should effectively reduce the source term.

It will be interesting to determine, however, whether

source removal activities significantly alter the ratio of

petroleum to chlorinated ground-water constituents. Initial half-life estimates in the Type 1/source zone are

estimated to be 54 days, whereas half-lives in Type 3

zones without suitable organic substrates are estimated

to be on the order of years.2

 Theoretically, then, if

ground-water conditions change from a Type 1 status of

sufficient substrate or electron donor and high dehalogenation kinetics to a Type 3 status with insufficient

useful substrate and lower dehalogenation kinetics, the

fate and transport of TCE could possibly change from a

half-life on the order of days to a half-life of years.

Abstaining from petroleum source removal at mixed

petroleum and chlorinated sites may serve the important

function of maintaining proper substrate or electron donor supply. This concept is being extended for discussion purposes only and should not be interpreted as a

blanket argument against source removal. Deliberate

substrate addition (e.g., methane, butane, methanol,

benzoate) is being advanced by the research community; therefore, an approach that takes advantage of an

existing substrate should be considered and may be

appropriate. The main consideration should be whether

the source—petroleum or chlorinated—needs to be removed to prevent unacceptable exposure. The land use

controls and commonly distant property boundaries of

military installations make this alternative even more

plausible. Since natural attenuation of chlorinated organics depends on substrate or electron donor availability, any alteration could dramatically affect transport. The

main point of this discussion is to underline the importance of understanding plume behavior because it is

likely to change over time, naturally or in response to

remedial systems.

Summary

In summary, the consideration of natural attenuation is

an essential aspect of the site characterization process.

Needless to say, smaller, less complex sites with lower

risk status and/or remediation costs may not merit a

detailed natural attenuation study. All other sites within

the AFCEE Technology Transfer Division Program will

be characterized according to the draft Technical Protocol for Natural Attenuation of Chlorinated Solvents in

Ground Water (4). This protocol will be updated and

refined based on actual field experience and regulatory

input. The final version will contain only those parameters that are both cost effective and diagnostic with

respect to fate and transport/natural attenuation effects.

This document will serve as a chlorinated solvent site

characterization “how-to” manual aimed at providing

needed upfront information to achieve final remedial

alternative selection earlier in the process. The emphasis will be on the inclusion of natural attenuation as a

remedial alternative alongside and in concert with conventional remedial alternatives—to include the “no action alternative.” Thus, this approach is an inclusive

analysis, not an exclusive analysis. The AFCEE hopes

to promote the proper consideration of the natural attenuation alternative since this alternative, to date, has

probably been excluded more often than it has been

included.

The AFCEE Technology Transfer Division is also investigating the technical and economic feasibility of other

remedial technologies and their applicability. Also, the

compatibility of each technology with the natural attenuation alternative is being considered. For Example, AFCEE has installed and is currently evaluating a zero

valence iron passive treatment wall to determine

whether this technology can provide long-term site remediation. Source reduction approaches such as surfactant-enhanced aquifer remediation, in-well aeration,

dual-phase extraction, and co-metabolic bioventing are

being applied at a variety of field sites. In addition,

innovative off-gas treatment systems are being placed

in real-world SVE remediation applications to verify applicability, cost effectiveness, and reliability.

The above examples are not meant to be an exhaustive

list of all technologies to be considered in feasibility

2

 Wiedemeier, T.H. 1996. Personal communication between the

author and Todd H. Wiedemeier, Parsons Engineering Science, Inc.,

Denver, CO.

127
studies; this listing is provided to communicate that

several specific alternatives are being evaluated in the

field in addition to natural attenuation. The overall effort

is to collect sufficient real-world applicability and performance data to complete the “remediation tool box” for

chlorinated solvents.

As stated above, natural attenuation will be applied as

part of a risk management strategy alone or in concert

with supplemental remedial systems. This risk management strategy shall emphasize plume management

concepts within current regulatory frameworks to include prioritized focus on higher mobility, higher risk

compounds; emphasis on early pathways analysis to

identify potentially completed pathways; determination

of whether plumes are contained, thus excluding any

future completed pathways; implementation of source

removal or other supplemental remedial systems when

exposure pathways are complete or natural attenuation

processes alone will not satisfy remedial objectives; and

use of the feasibility study process to balance technical,

economic, public, and regulatory factors. Consideration

will be site specific with respect to data and the regulatory framework. The use of natural attenuation as a

treatment alternative, however, is considered scientifically supportable, and when properly applied it is considered compatible with federal and state regulatory

programs.

References

1. Rice, D.W., R.D. Grose, J.C. Miachaelson, B.P. Dooher, D.H.

McQueen, S.J. Cullen, W.E. Kastenburg, L.G. Everett, and M.A.

Marino. 1995. California Leaking Underground Fuel Tank (LUFT)

Historical Case Analysis. California State Water Resources Control

Board.

2. Weidemeier, T.H., J.T. Wilson, D.H. Kampbell, R.N. Miller, and J.E.

Hansen. 1995. Technical Protocol for Implementing Intrinsic Remediation with Long-Term Monitoring for Natural Attenuation of

Fuel Contamination Dissolved in Ground Water. San Antonio, TX:

U.S. Air Force Center for Environmental Excellence.

3. Newell, C.J., R.K. McLeod, and J.R. Gonzalez. 1996.

BIOSCREEN Natural Attenuation Decision Support System, Version 1.30. San Antonio, TX: U.S. Air Force Center for Environmental Excellence.

4. Weidemeier, T.H., M.A. Swanson, D.E. Moutoux, J.T. Wilson, D.H.

Kampbell, J.E. Hansen, P. Haas, and F.H. Chapelle. 1996. Technical Protocol for Natural Attenuation of Chlorinated Solvents in

Ground Water. San Antonio, TX: U.S. Air Force Center for Environmental Excellence. In preparation.

128
Intrinsic Remediation in the Industrial Marketplace

David E. Ellis

DuPont Specialty Chemicals-CRG, Wilmington, Delaware

Introduction

Intrinsic remediation of chlorinated solvents is a common phenomenon. Most sites contain bacteria that can

both dechlorinate and oxidize chlorinated solvents to

nontoxic compounds. The challenge for site owners and

for regulators is to determine whether intrinsic remediation is a safe and effective remedy at individual sites.

Intrinsic remediation is an important development for

industry because it protects human health and the environment yet is more cost-effective than the competing,

intrusive ground water remediation techniques.

When To Consider Intrinsic Remediation

Decision-makers should determine whether the following criteria are met when evaluating the appropriateness

of intrinsic remediation for a given site:

• Intrinsic remediation protects human health and the

environment.

• Geochemical and volatile organic compound (VOC)

analyses demonstrate that intrinsic degradation of

contaminants is occurring.

• The contaminant source is continuing or cannot be

removed (e.g., dense nonaqueous-phase liquids

[DNAPLs]), so ground water will need long-term

treatment.

• Ground water receptors are not affected or can be

protected.

• Minimal disruption of plant operations or property is

desired.

• Alternative remedial technologies pose additional

risks, such as transferring contaminants to other environmental media or disrupting adjacent ecosystems.

• The rate of degradation balances the rate of migration and the potential for exposure, considering the

likely nature and timing of potential exposures. For

example, if a plume will degrade within 10 years and

the ground water is not likely to be used for 20 years,

intrinsic bioremediation should be seriously considered.

The Data Needed for an Intrinsic

Remediation Determination

Determination of the appropriateness of an intrinsic remediation demonstration considers the extent of the

data-gathering effort and the cost of the resources required. DuPont has developed the following list of minimum data to be gathered at all potential intrinsic

remediation sites:

• VOCs, including isomers.

• Dissolved oxygen, redox potential, and conductivity.

• Methane, ethane, ethylene, and propane.

• Total organic carbon (TOC).

• Major anions and cations (sodium, potassium, calcium, chloride, iron, magnesium, manganese, nitrate,

sulfate, and alkalinity).

DuPont recommends a tiered approach to intrinsic site

assessment, based on the complexity of the site, to better

understand what will be needed for a credible intrinsic

remediation demonstration. Table 1 characterizes the

three tiers of sites. For further information on requirements

for demonstrations, consult the newly issued Remediation

Technology Development Forum (RTDF) guidelines (1).

The Economics of Intrinsic Remediation

Those who have been involved in selecting the “best”

remedy for a site know that this is a time-consuming

task, which typically requires expensive sampling and

analysis; the more parameters, the greater the analytical

cost. Therefore, there is often a reluctance to evaluate

a large number of remedial alternatives. DuPont has

found, however, that the incremental cost of evaluating

intrinsic bioremediation along with other options is relatively small. This incremental cost may be more than

offset if intrinsic remediation is chosen over a technology

129
that would be more expensive to implement. This conclusion is based on using a “template” site to perform an

engineering cost estimate. The template site has the

following characteristics:

• 10-acre site

• Contaminant: tetrachloroethene (PCE)

• Concentration: 10 mg/L

• 20 monitoring wells, sampled twice a year for 30 years

• Completed remedial investigation

• Long-term monitoring costs are brought to present

costs using an inflation rate of 3 percent and a discount rate of 12 percent, the corporate cost of capital

Much of the investigation cost is the same regardless of

the remedy chosen. Therefore, only incremental costs

are considered in this analysis. The incremental present

cost of an intrinsic remediation demonstration above

that of a standard investigation and long-term monitoring is approximately $100,000 over 30 years. The simplest pump-and-treat remedy (air stripping and vaporphase granulated activated carbon) has a present cost

of $2.1 million over 30 years. A comparable intrinsic

remediation remedy has a present cost of $900,000.

(See Table 2 for cost details.) If intrinsic remediation is

protective, the saving is $1.2 million.

The Average Plume

DuPont recently surveyed over 50 sites and plumes to

get a statistical picture of how and where intrinsic biodegradation is operating. The survey looked for evidence

of reductive dehalogenation at these sites, which were

primarily DuPont Resource Conservation and Recovery

Act and Comprehensive Environmental Response,

Compensation, and Liability Act sites. Some outside

sites were included where data were available, as well

as several sites clearly described in the scientific literature. To be included, the sites needed to have either

SW846 Method 8240 analyses for VOCs, good geological delineation, and credible isoconcentration maps, or

to be thoroughly described in the technical literature.

Biodegradation

The sites selected for analysis were ones at which the

original contaminants could be identified; thus, field data

could be examined for the biodegradation byproducts of

those contaminants. The presence of these byproducts

indicates activity by naturally occurring bacteria. For

example, if most of the dichloroethene (DCE) present in

ground water is the cis-1,2-DCE isomer, that is conclusive evidence of the biological degradation of trichloroethene (TCE). The biodegradation results are

presented in Table 3. The data showed that:

• 88 percent of the sites have bacteria that can biodegrade PCE and TCE to DCE.

Table 1. Tiered Site Characteristics

Tier 1 (Easy Sites) Tier 2 (Intermediate Sites) Tier 3 (Difficult Sites)

Simple hydrogeology Moderately complex hydrogeology Hydrogeology complex

Single parent compound Few contaminants Confusing mixture

Size known (areal extent) Plume size questionable Large plume

Source and mass known Source and mass not well defined Very large or inaccessible source

Highest ground-water concentration < 10 mg/L Highest ground-water concentration < 100 mg/L Mobil DNAPLs

Static or shrinking plume Plume-size trend not known Growing plume or trend not known

Bioindicators obvious Some bioindicators No bioindicators

Receptors very far away Receptors are “not too far” Receptors close or affected

Analytical model sufficient Flow-and-transport model needed Needs a detailed fate-and-transport

model

Table 2. Present Cost of Intrinsic Remediation Versus Investigation and Long-Term Monitoring

Cost Element

Investigation

and Long-Term

Monitoring Cost

Intrinsic

Remedy

Cost

Incremental 

Cost—Intrinsic

Versus Investigation 

and Monitoring

Simple

Pump-andTreat Cost

Incremental Cost—

Pump-and-Treat

Versus Intrinsic

Up front $95,000 $35,000 $40,000 $650,000 $515,000

Annual $62,000 $68,000 $6,000 $35,000 $67,000

Present cost (30 years) $800,000 $900,000 $100,000 $2,100,000 $1,200,000

Note: 12 percent discount rate, 3 percent annual inflation. 

130
• 75 percent of the sites have bacteria that can biodegrade DCE to vinyl chloride (VC) or ethylene.

Some sites may not have enough bioavailable substrate

to complete the degradation reactions. Insufficient substrate should always be suspected at sites where biodegradation stops at either TCE or VC. Sites without the

full bacterial population needed for complete degradation would be expected to show either no degradation

or degradation that stops at DCE.

Half-Lives

Two simple methods were used to estimate half-lives.

The first method, developed by Buscheck et al. (2), is a

semilog plot of individual well analyses versus time of

transport. The second method is a simple graphical

extrapolation. The graphical extraction method assumes

that the plume is at steady state so that dilution, dispersion, and sorption factors are constant; measures concentration declines along the centerline of plumes on

high-quality isoconcentration maps; and calculates the

time for the water package to move each of those distances. The results of the two methods show good

agreement, with the graphical extraction method giving

somewhat longer half-lives. These data suggest that the

key factor in evaluating intrinsic remediation should be

the time of residence of contaminants in a plume before

it reaches a potential receptor, if it ever does. The average solvent half-lives are shown in Table 4.

Intrinsic Remediation Capacity

As a further criteria, it may be advantageous to calculate

the assimilative capacity of the aquifer, which is defined

as its capacity to biodegrade a contaminant. At many sites,

there appears to be no synthetic source of substrate. This

implies that natural organic material in the aquifer is

supplying electrons to drive the biodegradation reactions. Based on this assumption, one can calculate the

amount of chlorinated solvent that an aquifer can biodegrade, although this estimate can only apply to sites at

which the soils contain bacteria that can degrade chlorinated solvents.

Here is an example calculation. Typical aquifers contain

between 0.3 percent and 1 percent natural organic carbon. This equals 8 to 28 pounds of organic carbon per

cubic yard of soil at 2,800 pounds of soil per cubic yard.

A conservative assumption is that the aquifer contains

only 0.1 percent TOC and only 10 percent of the natural

organic carbon is bioavailable. If bacteria can use only

10 percent of the bioavailable organic carbon as food for

biodegrading chlorinated solvents, 0.03 pounds of

organic carbon per cubic yard (1 percent of the total

carbon present) is used as food in chlorocarbon degradation. Electron balance indicates that bacteria use 0.25

to 0.50 pounds of organic carbon to degrade 1 pound of

solvents (3). Therefore, each cubic yard of this hypothetical aquifer has the capacity to biodegrade at least

0.06 pounds of chlorinated solvent.

The plume that the RTDF is studying at Dover Air Force

Base in Delaware involves approximately 7.5 million

cubic yards of aquifer. Using the previous estimate, bacteria

in this aquifer should be able to biodegrade a minimum

of 450,000 pounds of solvents—the equivalent of 820

drums of DNAPL. It is very unlikely that there are 820

drums of DNAPL at Dover. Therefore, the bacteria in this

aquifer have an adequate supply of organic carbon to

biodegrade all the contaminants that are currently in it.

What About Existing Pump-and-Treat

Systems?

Shutting down a pump-and-treat system to let intrinsic

processes complete the restoration is now regarded as

acceptable during hydrocarbon remediation. Benzene is

the main component of concern in most hydrocarbon

plumes and is regulated at levels similar to those required for VC. Why shouldn’t chlorinated solvent pumpand-treat systems be shut down at some logical point

and intrinsic remediation be allowed to finish their work

as well? Many chlorinated solvent pump-and-treat systems have already reached their useful lifetime for contaminant removal.

All of the following criteria should be met before intrinsic

remediation can replace an existing, operating pumpand-treat system that treats chlorinated solvents:

• It can be demonstrated that intrinsic activity is already occurring in the aquifer.

• It is possible to predict how far the plume might

extend if the pump-and-treat system was not operating, and it can be shown that no receptor will be

affected.

Table 3. Biodegradation Results at Survey Sites

Reaction

Number of 

Sites Present

Total

Sites Percentage

PCE to TCE 27 31 87

TCE to DCE 39 44 88

DCE to VC 28 37 75

VC to ethane 18 31 58a

a

 Ethane data often are unavailable.

Table 4. Half-Lives Calculated by Graphical Extraction

Reaction Average Half-Life (years) Number of Sites

PCE to TCE 1.20 7

TCE to DCE 1.19 15

DCE to VC 1.05 12

VC to ethane 1.22 9

131
• Intrinsic remediation is protective of human health

and the environment.

Conclusions

• Intrinsic remediation is real. It is protective when

properly employed.

• Intrinsic biodegradation occurs at many sites. Each

biodegradation step has an average half-life of 1 to 2

years.

• The most important factors in determining the effectiveness of intrinsic remediation are plume residence

time and the half-lives of the sequential biodegradation reactions.

• Most aquifers contain much more organic carbon

than necessary to support intrinsic bioremediation.

While anthropogenic carbon may help support intrinsic degradation, it is not essential.

• Intrinsic remediation is not a “do nothing” approach,

and there is a moderate cost associated with it. The

present cost of an intrinsic remediation remedy is

approximately $900,000, compared with $2.1 million

for the cheapest pump-and-treat system.

References

1. Remediation Technology Development Forum Consortium for

Bioremedaiation of Chlorinated Solvents. 1996. Guidance handbook on intrinsic remeidation of chlorinated solvents. http://www.rtdf.org.

2. Buscheck, T.E., K.T. OReilly, and S.N. Nelson. 1993. Evaluation

of intrinsic bioremediation at field sites. In: Proceedings of the

Conference on Petroleum Hydrocarbons and Organic Chemicals

in Ground Water: Prevention, Detection, and Restoration, Houston, TX. pp. 367-381.

3. De Bruin, W.P., M.J.J. Kotterman, M.A. Posthumus, G. Schraa,

and A.J.B. Zehnder. 1992. Complete biological reductive transformation of tetrachloroethene to ethane. Appl. Environ. Microbiol.

58(6):1966-2000.

Additional Reading

Klecka, G.M., J.T. Wilson, E.J. Lutz, N. Klier, R. West, J. Davis, J.

Weaver, D. Kampbell, and B. Wilson. 1996. Intrinsic remediation of

chlorinated solvents in groundwater. Paper presented at the IBC Conference on Intrinsic Remediation, London, UK. 

132
Environmental Chemistry and the Kinetics of Biotransformation of 

Chlorinated Organic Compounds in Ground Water

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

U.S. Environmental Protection Agency, National Risk Management Research Laboratory,

R.S. Kerr Research Center, Ada, Oklahoma

Introduction

Responsible management of the risk associated with

chlorinated solvents in ground water involves a realistic

assessment of the natural attenuation of these compounds in the subsurface before they are captured by

ground-water production wells or before they discharge

to sensitive ecological receptors. The reduction in risk is

largely controlled by the rate of the biotransformation of

the chlorinated solvents and their metabolic daughter

products. These rates of biotransformation are sensitive

parameters in mathematical models describing the transport of these compounds to environmental receptors.

Environmental Chemistry of

Biodegradation of Chlorinated Solvents

[This section is designed specifically for engineers and

mathematical modelers who have little or no chemistry

background; other readers may wish to proceed directly

to the next section.]

The initial metabolism of chlorinated solvents such as

tetrachloroethylene, trichloroethylene, and carbon tetrachloride in ground water usually involves a biochemical

process described as sequential reductive dechlorination. This process only occurs in the absence of oxygen,

and the chlorinated solvent actually substitutes for oxygen in the physiology of the microorganisms carrying out

the process.

The chemical term “reduction” was originally derived

from the chemistry of smelting metal ores. Ores are chemical compounds of metal atoms coupled with other materials. As the ores are smelted to the pure element, the

weight of the pure metal are reduced compared with the

weight of the ore. Chemically, the positively charged metal

ions receive electrons to become the electrically neutral

pure metal. Chemists generalized the term “reduction”

to any chemical reaction that added electrons to an

element. In a similar manner, the chemical reaction of

pure metals with oxygen results in the removal of electrons from the neutral metal to produce an oxide. Chemists have generalized the term “oxidation” to refer to any

chemical reaction that removes electrons from a material. For a material to be reduced, some other material

must be oxidized.

The electrons required for microbial reduction of chlorinated solvents in ground water are extracted from native

organic matter, from other contaminants such as the

benzene, toluene, ethylene, and xylene compounds released from fuel spills, from volatile fatty acids in landfill

leachate, or from hydrogen produced by the fermentation of these materials. The electrons pass through a

complex series of biochemical reactions that support the

growth and function of the microorganisms that carry out

the process.

To function, the microorganisms must pass the electrons

used in their metabolism to some electron acceptor. This

ultimate electron acceptor can be dissolved oxygen,

dissolved nitrate, oxidized minerals in the aquifer, dissolved sulfate, a dissolved chlorinated solvent, or carbon dioxide. Important oxidized minerals used as electron

acceptors include iron and manganese. Oxygen is reduced to water, nitrate to nitrogen gas or ammonia,

iron(III) or ferric iron to iron(II) or ferrous iron, manganese(IV) to manganese(II), sulfate to sulfide ion, chlorinated solvents to a compound with one less chlorine

atom, and carbon dioxide to methane. These processes

are referred to as aerobic respiration, nitrate reduction,

iron and manganese reduction, sulfate reduction, reductive dechlorination, and methanogenesis, respectively.

The energy gained by the microorganisms follows the

sequence listed above: oxygen and nitrate reduction

provide a good deal of energy, iron and manganese

133
reduction somewhat less energy, sulfate reduction and

dechlorination a good deal less energy, and methanogenesis a marginal amount of energy. The organisms

carrying out the more energetic reactions have a competitive advantage; as a result, they proliferate and exhaust the ultimate electron acceptors in a sequence.

Oxygen and then nitrate are removed first. When their

supply is exhausted, then other organisms are able to

proliferate, and manganese and iron reduction begins.

If electron donor supply is adequate, then sulfate reduction begins, usually with concomitant iron reduction,

followed ultimately by methanogenesis. Ground water

where oxygen and nitrate are being consumed is usually

referred to as an oxidized environment. Water where

sulfate is being consumed and methane is being produced is generally referred to as a reduced environment.

Reductive dechlorination usually occurs under sulfate-reducing and methanogenic conditions. Two electrons are

transferred to the chlorinated compound being reduced.

A chlorine atom bonded with a carbon receives one of

the electrons to become a negatively charged chloride

ion. The second electron combines with a proton (hydrogen ion) to become a hydrogen atom that replaces the

chlorine atom in the daughter compound. One chlorine

at a time is replaced with hydrogen; as a result, each

transfer occurs in sequence. As an example, tetrachloroethylene is reduced to trichlorethylene, then any of the

three dichloroethylenes, then to monochloroethylene

(commonly called vinyl chloride), then to the chlorinefree carbon skeleton ethylene, then finally to ethane.

Kinetics of Transformation in Ground Water

Table 1 lists rate constants for biotransformation of

tetrachloroethylene (P.E.), trichloroethylene (TCE),

cis-dichloroethylene (cis-DCE), and vinyl chloride

extrapolated from field-scale investigations. In some

cases, a mathematical model was used to extract a rate

constant from field data; however, many of the rate

constants were calculated by John Wilson from published raw data. In several cases, the primary authors did

not choose to calculate a rate constant or felt that their

data could not distinguish degradation from dilution or

dispersion.

The data were collected or estimated to build a statistical

picture of the distribution of rate constants, in support of

a sensitivity analysis of a preliminary assessment using

published rate constants. They serve as a point of reference for “reasonable” rates of attenuation; applying

them to other sites without proper site-specific validation

is inappropriate.

Table 1. Apparent Attenuation Rate Constants (Field Scale Estimates)

Location Reference

Distance

From Source

Time From

Source

Residence

Time TCE cis-DCE

Vinyl

Chloride

(meters) (years) (years) Apparent Loss Coefficient (1/year)

St. Joseph, MI 1-3 130 to 390

390 to 550

550 to 855

240 to 460

3.2 to 9.7

 9.7 to 12.5

12.5 to 17.9

2.2 to 4.2

6.5

2.8

5.4

2.0

0.38

1.3 

0.93

1.4 

0.50

0.83

3.1 

Produced

0.18

0.88

2.2 

Produced

Picatinny

Arsenal, NJ

 4, 5 320 to 460

240 to 320

 0 to 250

2.9 to 4.2

2.2 to 2.9

0.0 to 2.3

1.3

0.7

1.2 Produced

1.6 

0.5 

Produced

Sacramento,CA 6 70 to 300 0.5 to 2.3 1.8 1.1 0.86 3.1 

Necco Park, NY 7 0 to 570

 0 to 660

0.0 to 1.6

0.0 to 1.8

1.6

1.8

0.7 

0.7 

Plattsburgh

AFB, NY

Weidemeider,

this volume

 0 to 300

300 to 380

380 to 780

0.0 to 6.7

6.7 to 8.6

 8.6 to 17.7

6.7

1.9

9.1

1.3 

0.23

Absent

Produced

0.6 

0.07

Produced

1.16

0.47

Tibbitt’s Road, NH B. Wilson,

this volume

 0 to 24

 0 to 40

 0 to 55

0.0 to 2.4

0.0 to 6.4

0.0 to 10 

2.4

6.4

10 

0.21

0.42

0.73

Produced

0.68

> 0.73 

San Francisco

Bay Area, CA

 8 4.4 5.11

Perth, Australia 9 0 to 600 0.0 to 14 0.32

Eielson AFB, AK 10 0.73

2.3 

Not identified 11 0.8 0.8 0.8 

Cecil Field

NAS, FL

Chapelle,

this volume

 0 to 140 0.0 to 1.2 1.2 3.3 to 7.3 3.3 to 7.3

134
The estimates of rates of attenuation tend to cluster

within an order of magnitude. Figure 1 compares the

rates of removal of TCE in those plumes that demonstrated evidence of biodegradation. Most of the first-order rates are very close to 1.0 per year, equivalent to a

half life of 8 months. Table 1 also reveals that the rate

of removal of P.E., TCE, and cis-DCE, and vinyl chloride

are similar; they vary by little more than one order of

magnitude.

Table 2 lists first-order and zero-order rate constants

determined in laboratory microcosm studies. The rates

of removal in the laboratory microcosm studies are similar to estimates of removal at field scale for TCE, cisDCE, and vinyl chloride. Rates of removal of

1,1,1-trichloroethane (1,1,1-TCA) are similar to the rates

of removal of the chlorinated alkenes.

Summary

The rates of attenuation of chlorinated solvents and their

less chlorinated daughter products in ground water are

slow as humans experience time. If concentrations of

chlorinated organic compounds near the source are in

the range of 10,000 to 100,000 micrograms per liter,

then a residence time in the plume on the order of a

decade or more will be required to bring initial concentrations to current maximum contaminant levels for

drinking water. Biodegradation as a component of natural attenuation can be protective of ground-water quality

in those circumstances where the travel time of a plume

to a receptor is long. In many cases, it will be necessary

to supplement the benefit of natural attenuation with

some sort of source control or plume management.

Figure 1. The first-order rate constant for biotransformation of

TCE in a variety of plumes of contamination in ground

water.

Table 2. Apparent Attenuation Rate Constants From Laboratory Microcosm Studies

Location of

Material Reference

Distance

From

Source

Time

From

Source

Incubation

Time TCE cis-DCE

Vinyl

Chloride 1,1,1-TCA

(meters) (years) (years)

Apparent First-Order Loss (1/year)

Apparent Zero OrderLoss (μg ⁄ L∗day)

Laboratory Microcosm Studies Done on Material From Field-Scale Plumes

Picatinny

Arsenal, NJ

12

13

240

320

460

2.2

2.9

4.2

0.5

0.5

0.5

0.64

0.42

0.21

0.52

9.4

3.1

St. Joseph, MI 14 0.12, 0.077 1.8, 1.2

Traverse City, MI 15 300 1.8 1.8

Tibbitts Road, NH 16 At Source 4.8

Laboratory Microcosm Studies Done on Material Not Previously Exposed to the Chlorinated Organic Compound

Norman

Landfill, OK

17 Aerobic

material

4.2

10

18 Sulfate

reducing

1.28

1.62 1.75

Methanogenic1.20

1.65 1.42

FL 16 Reducing 3.6

19 Reducing 0.012

135
References

 1. Semprini, L., P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson. Anaerobic Transformation of chlorinated aliphatic hydrocarbons in

a sand aquifer based on spatial chemical distributions. Water

Resour. Res. 31(4):1051-1062.

 2. Weaver, J.W., J.T. Wilson, D.H. Kampbell, and M.E. Randolph.

1995. Field derived transformation rates for modeling natural

bioattenuation of trichloroethene and its degradation products. In:

Proceedings: Next Generation Environmental Models and Computational Methods, August 7-9, Bay City, MI.

 3. Wilson, J.T., J.W. Weaver, D.H. Kampbell. 1994. Intrinsic bioremediation of TCE in ground water at an NPL site in St. Joseph,

Michigan. In: U.S. EPA. Symposium on Natural Attenuation of

Ground Water, Denver, CO, August 30-September 1. EPA/600/R94/162. pp. 116-119.

 4. Ehlke, T.A., B.H. Wilson, J.T. Wilson, and T.E. Imbrigiotta. 1994.

In-situ biotransformation of trichloroethylene and cis-1,2-dichloroethylene at Picatinny Arsenal, New Jersey. In: Morganwalp,

D.W., and D.A. Aronson, eds. Proceedings of the U.S. Geological

Survey Toxic Substances Hydrology Program, Colorado Springs,

Colorado, September 20-24, 1993. Water Resources Investigations Report 94-4014. In press.

 5. Martin, M., and T.E. Imbrigiotta. 1994. Contamination of ground

water with trichloroethylene at the Building 24 site at Picatinny

Arsenal, New Jersey. In: U.S. EPA. Symposium on Natural Attenuation of Ground Water. Denver, CO, August 30-September

1. EPA/600/R-94/162. pp. 109-115.

 6. Cox, E., E. Edwards, L. Lehmicke, and D. Major. 1995. Intrinsic

biodegradation of trichloroethylene and trichloroethane in a sequential anaerobic-aerobic aquifer. In: Hinchee, R.E., J.T. Wilson,

and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH:

Battelle Press. pp. 223-231.

 7. Lee, M.D., P.F. Mazierski, R.J. Buchanan, Jr., D.E. Ellis, and L.S.

Sehayek. 1995. Intrinsic and in situ anaerobic biodegradation of

chlorinated solvents at an industrial landfill. In: Hinchee, R.E., J.T.

Wilson, and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH: Battelle Press. pp. 205-222.

 8. Buscheck, T., and K. O’Reilly. 1996. Intrinsic anaerobic biodegradation of chlorinated solvents at a manufacturing plant. Abstract

presented at the Conference on Intrinsic Remediation of Chlorinated Solvents, Salt Lake City, UT, April 2. Columbus, OH: Battelle

Memorial Institute.

 9. Benker, E., G.B. Davis, S. Appleyard, D.A. Berry, and T.R. Power.

1994. Groundwater contamination by trichloroethene (TCE) in a

residential area of Perth: Distribution, mobility, and implications for

management. In: Proceedings of the Water Down Under 94, 25th

Congress of IAH, Adelaide, South Australia, November 21-25.

10. Gorder, K.A., R.R. Dupont, D.L. Sorensen, and M.W. Kemblowski. 1996. Intrinsic remediation of TCE in cold regions. Abstract presented at the Conference on Intrinsic Remediation of

Chlorinated Solvents, Salt Lake City, UT, April 2. Columbus, OH:

Battelle Memorial Institute.

11. De, A., and D. Graves. 1996. Intrinsic bioremediation of chlorinated aliphatics and aromatics at a complex industrial site. Abstract presented at the Conference on Intrinsic Remediation of

Chlorinated Solvents, Salt Lake City, UT, April 2. Columbus, OH:

Battelle Memorial Institute.

12. Ehlke, T.A., T.E. Imbrigiotta, B.H. Wilson, and J.T. Wilson. 1991.

Biotransformation of cis-1,2-dichloroethylene in aquifer material

from Picatinny Arsenal, Morris County, New Jersey. In: U.S. Geological Survey Toxic Substances Hydrology Program—Proceedings of the Technical Meeting, Monterey, CA, March 11-15. Water

Resources Investigations Report 91-4034. pp. 689-697.

13. Wilson, B.H., T.A. Ehlke, T.E. Imbigiotta, and J.T. Wilson. 1991.

Reductive dechlorination of trichloroethylene in anoxic aquifer

material from Picatinny Arsenal, New Jersey. In: U.S. Geological

Survey Toxic Substances Hydrology Program—Proceedings of

the Technical Meeting, Monterey, CA, March 11-15. Water Resources Investigations Report 91-4034. pp. 704-707.

14. Haston, Z.C., P.K. Sharma, J.N.P. Black, and P.L. McCarty. 1994.

Enhanced reductive dechlorination of chlorinated ethenes. In:

U.S. EPA. Proceedings of the EPA Symposium on Bioremediation

of Hazardous Wastes: Research, Development, and Field Evaluations. EPA/600/R-94/075. pp. 11-14.

15. Wilson, B.H., J.T. Wilson, D.H. Kampbell, B.E. Bledsoe, and J.M.

Armstrong. 1990. Biotransformation of monoaromatic and chlorinated hydrocarbons at an aviation gasoline spill site. Geomicrobiol. J. 8:225-240.

16. Parsons, F., G. Barrio Lage, and R. Rice. 1985. Biotransformation

of chlorinated organic solvents in static microcosms. Environ.

Toxicol. Chem. 4:739-742.

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

of vinyl chloride in groundwater samples. Appl. Environ. Microbiol.

56(12):3878-3880.

18. Klecka, G.M., S.J. Gonsior, and D.A. Markham. 1990. Biological

transformations of 1,1,1-trichloroethane in subsurface soils and

ground water. Environ. Toxicol. Chem. 9:1437-1451.

19. Barrio-Lage, G.A., F.Z. Parsons, R.M. Narbaitz, and P.A. Lorenzo.

1990. Enhanced anaerobic biodegradation of vinyl chloride in

ground water. Environ. Toxicol. Chem. 9:403-415.

136
Future Vision: Compounds With Potential for Natural Attenuation

Jim Spain

U.S. Air Force Armstrong Laboratory, Tyndall Air Force Base, Florida

Introduction

Attenuation of natural organic compounds, such as

those present in hydrocarbon fuels, is predictable because the responsible microorganisms are ubiquitous in

soil and the subsurface. Bacteria able to use hydrocarbons as their source of carbon and energy under either

aerobic or anaerobic conditions have a tremendous selective advantage over other members of the microbial

community. Therefore, the process can be self-sustaining and is limited only by the presence of electron acceptors or inorganic nutrients.

Bacteria able to grow at the expense of chlorinated aliphatic

compounds are not widely distributed; natural attenuation of such compounds is consequently less predictable. The use of a chlorinated compound as a terminal

electron acceptor (chlororespiration or dehalorespiration) can yield energy and thus provides a selective

advantage to a limited range of anaerobic bacteria (1).

Many of the transformations of chloroaliphatic compounds, such as trichloroethylene, are co-metabolic and

yield no advantage to the bacteria that catalyze the

reactions. In fact, co-metabolism can select against the

organism because of the wasting of energy and production

of toxic metabolites.

Between the extremes of readily degradable hydrocarbons and chlorinated aliphatic compounds that serve

only as electron acceptors are many other synthetic

organic compounds that can provide sources of carbon

and energy for bacteria. This paper describes compounds that are known to be biodegradable and have

the potential for natural attenuation in the field. Some

synthetic chemicals are expected to be readily susceptible to natural attenuation, others are degraded at a

limited number of sites, and some show only a limited

potential. Where possible, recent review articles rather

than primary literature will be cited. More detailed information on many of the compounds is also available in a

recent book that provides an excellent analysis of the

potential for biodegradation (2).

The first question to be asked when considering the

potential for natural attenuation is whether biodegradation of the chemical contaminant has been reported. The

question could be phrased, “Does the biology exist?”

Biodegradation of some of the compounds has been

studied extensively under field conditions. Transformation of others has only recently been discovered in

laboratory systems or waste streams. Such laboratory

studies should not be ignored because the processes

discovered in such systems are catalyzed by bacteria

obtained from the field. Laboratory studies are essential

for revealing the mechanisms of the reactions and the

conditions required for the process. They can also determine whether the process provides energy or nutrients—and thus a selective advantage—to the bacteria

that catalyze the reactions.

The second question is whether the activity of the necessary specific organisms is present at the site under

consideration. A considerable amount of effort has been

spent on enumerating and identifying bacteria at hydrocarbon-contaminated sites under consideration for

bioremediation. Because such bacteria are ubiquitous,

it is much more useful to assess their activity as revealed by degradation of the hydrocarbons or transformation of electron acceptors. The biology can be

assumed to be present but limited by other factors. In

contrast, bacteria able to degrade specific synthetic

chemicals cannot be assumed to be widely distributed

in the field. Detection of bacteria able to grow on specific

compounds in contaminated sites and failure to detect

them in nearby uncontaminated areas can be taken as

strong evidence for natural attenuation. Absence of bacteria able to catalyze the degradation of compounds

known to be biodegradable could provide an opportunity

for bioaugmentation, a strategy that has earned a poor

reputation because of misapplication in the past.

The third question is whether conditions appropriate for

natural attenuation exist or can be created at the site.

Issues of electron donors and acceptors, bioavailability,

mass transfer, contaminant mixtures, and concentration

137
must be resolved. A good understanding of the biodegradation process, including reaction stoichiometry and

kinetics, is essential for evaluation of the potential for

natural attenuation. Fortunately, such understanding exists for a wide range of synthetic chemical contaminants.

Chloroaromatic Compounds

Bacteria able to degrade all but the most complex chloroaromatic compounds have been discovered during

the past 20 years. Polychlorobenzenes, including hexachlorobenzene, can be sequentially dehalogenated to

monochlorobenzene under methanogenic conditions in

soil slurries (3). Reductive dehalogenation of chlorobenzene has not been reported, but chlorotoluenes are

dehalogenated to toluene in the above methanogenic

systems, and it seems likely that chlorobenzene could

serve as a substrate for reductive dehalogenation.

Chlorobenzenes up to and including tetrachlorobenzene

are readily biodegraded under aerobic conditions. Bacteria

able to grow on chlorobenzene (4), 1,4-dichlorobenzene

(4-6), 1,3-dichlorobenzene (7), 1,2-dichlorobenzene (8),

1,2,4-trichlorobenzene (9), and 1,2,4,5-tetrachlorobenzene

(10) have been isolated and their metabolic pathways

determined. The pathways for aerobic degradation are

remarkably similar and lead to the release of the halogens

as hydrochloride (HCl).

Chlorobenzenes are very good candidates for natural

attenuation under either aerobic or anaerobic conditions. Aerobic bacteria able to grow on chlorobenzene

have been detected at a variety of chlorobenzene-contaminated sites but not at uncontaminated sites (11).

This provides strong evidence that the bacteria are selected for their ability to derive carbon and energy from

chlorobenzene degradation in situ. Removal of multiple

halogens as HCl consumes a large amount of alkalinity

and produces a considerable drop in the pH of unbuffered systems, which could lead to a loss of microbial

activity at some sites.

Chlorophenols and chlorobenzoates are dehalogenated

under anaerobic conditions in sediments and subsurface material (12-13). In some instances, the dehalogenation clearly yields energy for the growth of specific

bacteria. In other examples, the dehalogenation is specific and enriched in the community but has not been

rigorously linked to energy production. The addition of

small fatty acids or alcohols as either electron donors or

sources of carbon can enhance the process of reductive

dehalogenation. Aerobic pathways for the degradation

of chlorophenols and chlorobenzoates are initiated

by an oxygenase catalyzed attack on the aromatic

ring and the subsequent removal of the halogen

after ring fission or hydrolytic replacement of the

halogen with a hydroxyl group. Bacteria able to

grow on chlorophenols and chlorobenzoates are

widely distributed and are readily enriched from a variety

of sources, which indicates a high potential for natural

attenuation. The chlorophenols are unusual among the

synthetic compounds discussed here, however, as they

can be very toxic to microorganisms. They are often

used as biocides, and, therefore, high concentrations

can dramatically inhibit biodegradation. Inoculation with

specific bacteria has been helpful in overcoming toxicity and stimulating degradation of chlorophenols (12).

Pentachlorophenol deserves special consideration because it has been widely used as a wood preservative

and has been released into the environment throughout the world. Reductive dehalogenation of pentachlorophenol under methanogenic conditions can lead to

mineralization (12). Aerobic bacteria catalyze the replacement of the chlorine in the 4 position by a hydroxyl group to form tetrachlorohydroquinone, and

subsequent reductive dehalogenations lead to the formation of ring fission substrates. Bacteria able to degrade pentachlorophenol are widely distributed, and

both experimental and full-scale bioremediation projects

have been successful in field applications (12). In some

instances, the addition of selected strains has been

helpful, whereas in others indigenous strains have been

used. Wood treatment facilities are typically contaminated with complex mixtures of organic compounds, so

investigations of toxicity must be conducted for each site

under consideration. Natural attenuation of pentachlorophenol seems to be possible because specific bacteria able to use it as a growth substrate are enriched

at contaminated sites. Rates seem to be low at the sites

investigated to date, however, due to the toxicity and

bioavailability of the pentachlorophenol.

Polychlorinated biphenyls (PCB) have been studied extensively because of their stability, toxicity, and bioaccumulation potential (14). Anaerobic transformation of

PCB is catalyzed by bacteria in aquatic sediment from

a wide range of both contaminated and uncontaminated

sites. Higher activity in contaminated sites suggests that

the dehalogenation reactions provide a selective advantage to the microbial population, which indicates the

potential for significant natural attenuation. Studies have

clearly demonstrated that natural attenuation of PCB is

taking place in anaerobic sediments at significant rates,

with methanogenic conditions in freshwater sediments

apparently providing the highest rates of reductive dehalogenation. Dehalogenation converts the more highly

chlorinated congeners to less chlorinated products containing one to four chlorine. Complete dehalogenation

does not occur, but the depletion of the more highly

chlorinated congeners dramatically reduces not only the

toxicity and carcinogenicity, but also the bioaccumulation of the mixture.

A variety of different dechlorination patterns have been

identified as a function of the microbial community involved. The patterns are constant within a given microbial

138
community or enrichment, which supports the premise

that dehalogenation provides a selective advantage to

the organisms involved. The results also suggest that a

wide range of different bacteria have the ability to dehalogenate PCB. The electron donors for the dehalogenation in sediment are unknown. The addition of

exogenous carbon sources does not stimulate the reaction. In contrast, “priming” the mixtures with low levels

of bromobiphenyl or specific isomers of tetrachlorobiphenyl (15) seems to selectively enrich a population

of PCB-dechlorinating bacteria and dramatically stimulate the dechlorination of the other congeners.

The lower chlorinated PCB congeners, whether part of

the original Arochlor mixture or derived from reductive

dehalogenation, are biodegraded by aerobic bacteria

(16). The initial attack is catalyzed by a 2,3- or 3,4-dioxygenase, followed by a sequence of reactions that

lead to ring cleavage and the accumulation of chlorobenzoates which are readily degraded by a variety of

bacteria. The enzymes that oxidize PCB are produced

by bacteria grown on biphenyl, and adding biphenyl to

slurry-phase reactors stimulates the growth and activity

of PCB degraders. Such stimulation has been shown to

be effective in the field (17). There is also good evidence

that aerobic PCB degradation is taking place in contaminated river sediments (18).

Clearly, reductive dechlorination is ongoing at a wide

range of PCB contaminated sites. The strategy of anaerobic dehalogenation followed by aerobic degradation

seems to be particularly effective with PCB whether in

an engineered system or in natural systems (e.g., during

resuspension of anaerobic sediments). To date the complete biodegradation of PCB is slow and difficult to

predict or control in the field. Several new strategies,

including construction of novel strains, may increase the

potential for effective PCB biodegradation.

Chloroaliphatic Compounds

Several good reviews have recently appeared on the

biodegradation of small (one- and two-carbon) chloroaliphatic compounds (19-21); therefore, this paper

briefly mentions only some aspects that might otherwise

be overlooked. Among the one- and two-carbon chlorinated compounds, the more highly chlorinated molecules

are subject to reductive dehalogenation under a variety

of conditions. Thus, carbon tetrachloride can be sequentially reduced to chloroform and dichloromethane.

Similarly, perchloroethylene can be reduced to ethylene

via trichloroethylene, dichloroethylene, and vinyl chloride. The degradation of chloroethylenes is discussed in

considerable detail by Gossett and Zinder (this volume).

Most of the work to date has focused on mixed microbial

cultures that use chlorinated solvents fortuitously as

electron acceptors. Such activity is very widely distributed in anaerobic ecosystems and catalyzes the slow and

often partial reduction of chlorinated contaminants. In

contrast, some microbial communities and a few isolated strains can derive energy from the use of chlorinated compounds as terminal electron acceptors

(Gossett and Zinder, this volume). Such processes are

much faster than the co-metabolic processes because

they provide a selective advantage for the bacteria and

are self-sustaining.

Several chloroaliphatic compounds can serve as growth

substrates for aerobic bacteria. The more chlorinated

compounds such as trichloroethylene and chloroform do

not provide energy and carbon for aerobic growth, although they can be degraded co-metabolically. In contrast, methylene chloride can support the growth of both

anaerobes (22) and aerobes (20). 1,2-Dichloroethane

(23) and vinyl chloride (20) similarly can be readily

degraded by aerobic bacteria. Any of these compounds

that serve as growth substrates would be excellent candidates for natural attenuation where oxygen is present.

Aerobic mineralization of the related molecule, ethylene

dibromide, has been reported in soil (24), but the distribution of the responsible bacteria and the corresponding

ability to predict degradation are not well understood.

Nitroaromatic Compounds

The literature on biodegradation of nitroaromatic compounds has been reviewed recently (25). Nitroaromatic

compounds are subject to reduction of the nitro groups

in the environment under either aerobic or anaerobic

conditions. Co-metabolic reduction does not lead to

complete degradation in most instances and could be

considered nonproductive for purposes of natural attenuation. In contrast, aerobic bacteria able to grow on

nitrobenzene, nitrotoluenes, dinitrotoluenes, dinitrobenzene, nitrobenzoates, and nitrophenols have been isolated from a variety of contaminated sites, which suggests

that natural attenuation is taking place at such sites. The

simple nitroaromatic compounds (not including trinitrotoluene) can be considered excellent candidates for

natural attenuation. Some of the compounds, including

3-nitrophenol, nitrobenzene, 4-nitrotoluene, and 4-nitrobenzoate, are degraded via catabolic pathways that

involve a partial reduction of the molecule prior to oxygenative ring fission. The pathways minimize the use of

molecular oxygen and are particularly well suited for

operation in the subsurface, where oxygen is limiting.

Mixtures of the isomeric nitro compounds can be problematic for microbial degradation. For example, the industrial synthesis of polyurethane produces large

amounts of 2,4- and 2,6-dinitrotoluene in a ratio of four

to one. Bacteria able to grow on 2,4-dinitrotoluene have

been studied extensively. Unfortunately, 2,6-dinitrotoluene inhibits the degradation of 2,4-dinitrotoluene and

may prevent natural attenuation. Bacteria able to grow

on 2,6-dinitrotoluene have been isolated recently (26),

139
and insight about the metabolic pathway might allow

better prediction of the mixture’s degradation.

Ketones

Acetone and other ketones are not xenobiotic compounds, but most of the current production is via synthetic routes. They are readily biodegraded by both

aerobic and anaerobic (27) bacteria in soil and have a

very high potential for natural attenuation.

Methyl-tert-butyl Ether

Gasoline oxygenates such as ethanol, methyl-tert-butyl

ether (MTBE), and tert-butyl alcohol are used extensively

as octane enhancers in unleaded gasoline. The ether

bond of MTBE makes it particularly resistant to biodegradation. Its water solubility, low volatility, and high concentrations in gasoline (up to 15 percent) create concerns

about its behavior in the subsurface.

Preliminary studies indicate that it behaves almost as a

conservative tracer in gasoline-contaminated sites.1

Mixed cultures able to grow on MTBE have been enriched from refinery and chemical plant waste treatment

systems (24),2

 so bacteria clearly can successfully attack the ether bond. The degradation rates are slow,

however, and there is no evidence that the bacteria are

widely distributed in soil. MTBE and other oxygenates

containing ether bonds biodegrade very slowly, if at all,

under anaerobic conditions (28). At present, even

though the biological capability for MTBE degradation is

known to exist, the potential for natural attenuation of

MTBE seems low. The problem is sufficiently important

to merit additional study, perhaps involving extensive

acclimation of soil communities or bioaugmentation. The

available evidence indicates that tert-butyl alcohol is

much more readily degradable than MTBE under aerobic

or anaerobic conditions.

Nitrate Esters

A variety of nitrate esters, including glycerol trinitrate,

pentaerythritol tetranitrate, and nitrocellulose, have

been used extensively as explosives. Recent studies

indicate that the nitrate esters can be degraded by

bacteria from a variety of sources (29, 30). Bacterial

metabolism releases nitrite, which can serve as a nitrogen source and yield a selective advantage for the

organisms. The biodegradation of nitrate esters has

only recently been studied extensively, and little is

known about degradation in the environment. Recent

laboratory results strongly suggest that natural attenuation is possible, but more information is needed on the

bioavailability, toxicity, and kinetics of the process.

Pesticides

Most pesticides used in the past 20 years in the United

States have been formulated to degrade in the environment, and a considerable amount of information is available on degradation kinetics in soil and water. The U.S.

Environmental Protection Agency Risk Reduction Engineering Laboratory in Cincinnati, Ohio, has developed

an extensive Pesticide Treatability Database that contains information on a variety of compounds. Many pesticides hydrolyze and yield compounds that serve as

growth substrates for bacteria. For example, carbamates, chlorophenoxyacetates, dinitrocresol, coumaphos, atrazines, and some organophosphates serve

as growth substrates for bacteria and would be good

candidates for natural attenuation. A variety of other

pesticides are hydrolyzed by extracellular enzymes derived from soil bacteria but provide no advantage to the

organisms that produce the enzymes. Similarly, some of

the organohalogen insecticides can be reductively dehalogenated but provide no advantage to specific organisms. Their biodegradation rates are proportional to the

biomass and activity in the soil.

Conclusion

To date, the focus of natural attenuation has been on

hydrocarbon fuels and chlorinated aliphatic solvents. A

wide range of synthetic chemicals released in the environment are known to be biodegradable by bacteria, and

much is known about the processes and their requirements. The potential for natural attenuation of biodegradable contaminants should be considered before

more costly and disruptive treatment options.

References

 1. Mohn, W.W., and J.M. Tiedje. 1992. Microbial reductive dehalogenation. Microbiol. Rev. 56:482-507.

 2. Young, L.Y., and C.E. Cerniglia, eds. 1995. Microbial transformation and degradation of toxic organic chemicals. New York, NY:

Wiley-Liss.

 3. Ramanand, K., M.T. Balba, and J. Duffy. 1993. Reductive dehalogenation of chlorinated benzenes and toluenes under methanogenic conditions. Appl. Environ. Microbiol. 59:3266-3272.

 4. Reineke, W., and H.-J. Knackmuss. 1984. Microbial metabolism

of haloaromatics: Isolation and properties of a chlorobenzene-degrading bacterium. Eur. J. Appl. Microbiol. Biotechnol. 47:395-

402.

 5. Schraa, G., M.L. Boone, M.S.M. Jetten, A.R.W. van Neerven, P.J.

Colberg, and A.J.B. Zehnder. 1986. Degradation of 1,4-dichlorobenzene by Alcaligenes sp. strain A175. Appl. Environ. Microbiol. 52:1374-1381.

 6. Spain, J.C., and S.F. Nishino. 1987. Degradation of 1,4-dichlorobenzene by a Pseudomonas sp. Appl. Environ. Microbiol.

53:1010-1019.

 7. de Bont, J.A.M., M.J.A.W. Vorage, S. Hartmans, and W.J.J. van

den Tweel. 1986. Microbial degradation of 1,3-dichlorobenzene.

Appl. Environ. Microbiol. 52:677-680.

1

 Weaver, J. 1996. Personal communication with the author.

2

 Cowan, R. 1996. Personal communication with the author.

140
 8. Haigler, B.E., S.F. Nishino, and J.C. Spain. 1988. Degradation of

1,2-dichlorobenzene by a Pseudomonas sp. Appl. Environ. Microbiol. 54:294-301.

 9. van der Meer, J.R., W. Roelofsen, G. Schraa, and A.J.B. Zehnder.

1987. Degradation of low concentrations of dichlorobenzenes

and 1,2,4-trichlorobenzene by Pseudomonas sp. strain P51 in

nonsterile soil columns. FEMS Microbiol. Lett. 45:333-341.

10. Sander, P., R.-M. Wittaich, P. Fortnagel, H. Wilkes, and W.

Francke. 1991. Degradation of 1,2,4-trichloro- and 1,2,4,5-

tetrachlorobenzene by Pseudomonas strains. Appl. Environ. Microbiol. 57:1430-1440.

11. Nishino, S.F., J.C. Spain, and C.A. Pettigrew. 1994. Biodegradation of chlorobenzene by indigenous bacteria. Environ. Toxicol.

Chem. 13:871-877.

12. Haggblom, M.M., and R.J. Valo. 1995. Bioremediation of chlorophenol wastes. In: Young, L.Y., and C.E. Cerniglia, eds. Microbial transformation and degradation of toxic organic chemicals.

New York, NY: Wiley-Liss. pp. 389-434.

13. Suflita, J.M., and G.T. Townsend. 1995. The microbial ecology

and physiology of aryl dehalogenation reactions and implications

for bioremediation. In: Young, L.Y., and C.E. Cerniglia, eds. Microbial transformation and degradation of toxic organic chemicals. New York, NY: Wiley-Liss. pp. 243-268.

14. Bedard, D.L., and J.F. Quensen. 1995. Microbial reductive

dechlorination of polychlorinated biphenyls. In: Young, L.Y., and

C.E. Cerniglia, eds. Microbial transformation and degradation of

toxic organic chemicals. New York, NY: Wiley-Liss. pp. 127-216.

15. Bedard, D.L., S.C. Bunnell, and L.A. Smullen. 1996. Stimulation

of microbial para-dechlorination of polychlorinated biphenyls that

have persisted in Housatonic River sediment for decades. Environ. Sci. Technol. 30:687-694.

16. Bedard, D.L., R. Unterman, L. Bopp, M.J. Brennan, M.L. Haberl,

and C. Johnson. 1986. Rapid assay for screening and characterizing microorganisms for the ability to degrade polychlorinated

biphenyls. Appl. Environ. Microbiol. 51:761-768.

17. Harkness, M.R., J.B. McDermott, D.A. Abramowicz, J.J. Salvo,

W.P. Flanagan, M.L. Stephens, F.J. Mondello, R.J. May, J.H. Lobos, K.M. Carrol, M.J. Brennan, A.A. Bracco, K.M. Fish, G.L.

Warner, P.R. Wilson, D.K. Dietrich, D.T. Lin, C.B. Morgan, and

W.L. Gately. 1993. In situ stimulation of aerobic PCB biodegradation in Hudson River sediments. Science 259:503-507.

18. Flanagan, W.P., and R.J. May. 1993. Metabolite detection as

evidence for naturally occurring aerobic PCB biodegradation in

Hudson River sediments. Environ. Sci. Technol. 27:2207-2212.

19. Adriaens, P., and T.M. Vogel. 1995. Biological treatment of chlorinated organics. In: Young, L.Y., and C.E. Cerniglia, eds. Microbial transformation and degradation of toxic organic chemicals.

New York, NY: Wiley-Liss. pp. 435-486.

20. Fetzner, S., and F. Lingens. 1994. Bacterial dehalogenases: Biochemistry, genetics, and biotechnological applications. Microbiol.

Rev. 58:641-685.

21. Wackett, L.P. 1995. Bacterial co-metabolism of halogenated organic compounds. In: Young, L.Y., and C.E. Cerniglia, eds. Microbial transformation and degradation of toxic organic

chemicals. New York, NY: Wiley-Liss. pp. 217-242.

22. Magli, A., F.A. Rainey, and T. Leisinger. 1995. Acetogenesis from

dichloromethane by a two-component mixed culture comprising

a novel bacterium. Appl. Environ. Microbiol. 61:2943-2949.

23. Stucki, G., U. Krebser, and T. Leisinger. 1983. Bacterial growth

on 1,2-dichloroethane. Experientia 39:1271-1273.

24. Salanitro, J.P., L.A. Diaz, M.P. Williams, and H.L. Wisniewski. 1994.

Isolation of a bacterial culture that degrades methyl t-butyl ether.

Appl. Environ. Microbiol. 60:2593-2596.

25. Spain, J.C. 1995. Biodegradation of nitroaromatic compounds.

Ann. Rev. Microbiol. 49:523-55.

26. Nishino, S.F., and J.C. Spain. 1996. Degradation of 2,6-dinitrotoluene by bacteria. In: Proceedings of the Annual Meeting, Ame

rican Society for Microbiology. pp. Q-381.

27. Janssen, P.H., and B. Schink. 1995. Catabolic and anabolic

enzyme activities and energetics of acetone metabolism of the

sulfate-reducing bacterium Desulfococcus biacutus. J. Bacteriol. 177:277-282.

28. Mormile, M.R., S. Liu, and J.M. Suflita. 1994. Anaerobic biodegradation of gasoline oxygenates: Extrapolation of information to

multiple sites and redox conditions. Environ. Sci. Technol.

28:1727-1732.

29. White, G.F., and J.R. Snape. 1993. Microbial cleavage of nitrate

esters: Defusing the environment. J. Gen. Microbiol. 139:1947-

1957.

30. White, G.F., J.R. Snape, and S. Nicklin. 1996. Biodegradation of

glycerol trinitrate and pentaerythritol tetranitrate by Agrobacterium radiobacter. Appl. Environ. Microbiol. 62:637-642.

141
Natural Attenuation of Chlorinated Compounds in Matrices 

Other Than Ground Water: The Future of Natural Attenuation

Robert E. Hinchee

Parsons Engineering Science, Salt Lake City, Utah

Introduction

To date, natural attenuation study and application have

focused on the dissolved phase in ground water in

unconsolidated sediments. There are good reasons for

this. Ground-water transport is the primary pathway of

concern at many sites, our understanding of aqueousphase processes with relatively short half-lives (1 year

or less) in ground water is relatively well developed, and

we have a better understanding of ground-water processes in unconsolidated media than in rock. This paper

addresses the potential importance of both natural attenuation in other media and of slower processes. Specifically, natural attenuation in fractured rock, of

nonaqueous-phase liquids (NAPLs), and in the vadose

zone, as well as low-rate processes, will be discussed.

Fractured Rock

Ground water in fractured rock presents a special problem. In most rock formations, surface area is limited and

flow paths tend to be complex when compared with

unconsolidated sediments. Many of the same processes

occur, but they tend to be more difficult to monitor. The

Test Area North (TAN) site, located at the Idaho National

Engineering Laboratory (INEL), contains a trichloroethene

(TCE) ground-water plume approximately 9,000 feet

long. The geology is characterized by basalt flows with

sedimentary interbeds that consist primarily of low permeability, fine-grained sediments. The basalt flows are

highly variable, and ground-water flow appears to occur

primarily in the fractured basalt. The basalt varies from

massive to highly fractured. The source of contamination appears to be an abandoned waste disposal well.

In addition to chlorinated solvents, the well received a

variety of wastes including nonchlorinated sludges and

some radioactive materials. TCE appears to be the only

significant chlorinated solvent in the source material, yet

in ground water near the source, dichloroethene (DCE)

concentrations are in the same range as TCE. The DCE:TCE

ratio then declines downgradient to a distance of about

6,000 feet beyond which only TCE is found. All the TAN

site data can be found in INEL (1).

What appears to be happening is anaerobic dechlorination near the source, very likely driven by the carbon

source in the nonchlorinated sludge. Downgradient conditions appear to be aerobic, and no evidence of

dechlorination is seen more then a few hundred feet

from the source well. One possible explanation for the

smaller DCE plume is aerobic degradation. This site

also has a tritium plume originating from the same

source. If we assume that all of the plumes are of the

same age, we can estimate the kinetics of the aerobic

degradation of DCE and make some inferences concerning the TCE.

The DCE plume is approximately 6,000 feet, the tritium

plume 7,500 feet, and the TCE plume 9,000 feet long.

Ignoring retardation and assuming a 12 year half-life for

tritium, the DCE half-life would be approximately 10

years. If the TCE is degrading aerobically, its half-life is

probably greater than 14 years. Based on these field

observations, it appears that the same processes that

have been observed at many sites in consolidated sediments are occurring in the fractured basalt at the TAN

site. Therefore, anaerobic dechlorination and aerobic

oxidation of the less chlorinated solvents should occur

in fractured rock. The significant challenge presented by

fractured rock will be the accurate determination of flow

paths, the same as for any ground-water investigation.

Nonaqueous-Phase Liquid

When NAPL is present on a site, the mass of contaminant in the NAPL is typically orders of magnitude greater

than that dissolved in ground water. With the exception

of dissolution (and evaporation in the vadose zone), little

is known about natural attenuation processes that occur

in or near the NAPL phase. While evaporation can be a

significant attenuation mechanism and should certainly

be considered whenever vadose-zone contamination is

of concern, the NAPL below the water table is normally

the greatest concern. Dissolution is the mechanism by

which the ground water is initially contaminated, and

although rates are high enough to create a ground-water

142
problem at many sites, dissolution is often quite slow

when compared with the mass of contaminant present.

At the Hill Air Force Base (AFB) OU-2 site, thousands

of gallons of NAPL (primarily TCE) have been recovered

and many thousands of gallons remain below the water

table, yet the rate of dissolution (based on mass of

dissolved contaminant migrating off site) is tens of gallons per year.1 This is not unusual. It is difficult based

on our current understanding of the fate and behavior of

chlorinated solvents to postulate mechanisms for the

biotic or biotic degradation of NAPLs, but a few years

ago the same would have been concluded about dissolved TCE or even benzene. The rates of any such

degradation would not have to be high to be significant.

In the fractured rock discussion above, aerobic DCE

degradation with a half life on the order of 9 years is

noted. This phenomenon is rarely observed in laboratories or in short-term field studies, yet such a process

could be quite important. It is conceivable that an as yet

unidentified process exists that degrades NAPL in situ

with a half life of 10 years, which could result in a much

more significant mass removal than the dissolution process followed by degradation in ground water. This is an

area which has been largely overlooked, and the research needed to evaluate these mechanisms will require a longer-term and significantly different approach

than is current practice; however, to achieve a reasonable understanding of the long-term effects of natural

attenuation, it should not be overlooked.

Vadose and Discharge Processes

One of the primary practical values of natural attenuation

is plume stability. In many plumes, at some point the rate

of degradation of dissolved contaminants is more or less

equal to the rate of dissolution, and the plume achieves a

quasi-steady state. To date, most of the work on natural

attenuation has focused on degradation in the aqueousphase in the aquifer. Little attention has been given to the

vadose zone or discharge points. Any attenuation mechanism that contributes to plume stability is important, and it

appears that other mechanisms such as volatilization to

the vadose zone and ground-water discharge can be important mechanisms in creating plume stability, although

volatilization from ground water to the vadose zone is

probably only important where net evaporation exceeds

infiltration to ground water. The process of contaminant

diffusion to the water table and through the capillary fringe

into ground water is likely too slow to be of much significance at most sites.

There are sites in the western United States, however, at

which net ground-water evaporation occurs. The obvious

manifestation of this is the caliche found in many western

soils. This appears to be happening at Hill AFB. For

example, there is a TCE plume approximately 5,000 feet

long at the OU-6 site. In the first several thousand feet

of plume, the depth to ground water is about 100 feet,

and net infiltration appears to be occurring. Near the

downgradient extreme of the plume, ground water is

much shallower (10 feet or less), and net evaporation

appears to be occurring. Significant TCE concentrations

have been observed in soil gas above the downgradient

end of the plume, possible evidence of volatilization to

the vadose zone.

Discharge is an obvious attenuation mechanism, and

the nature of the discharge will determine its usefulness.

For example, if the discharge is to a surface-water

stream where the result is unacceptably high contaminant concentrations, this would not be a helpful mechanism. Frequently, however, discharge may not result in

unacceptable exposure. At Hill AFB there are six plumes

that vary in length, but all are in the thousands of feet.

In five of the plumes, TCE is the predominant contaminant; in one DCE predominates. Although the plumes

are miles apart and their source elevations vary, all of

the plumes end at approximately the same elevation,

and most of these plumes appear stable.

One reason for the stability is discharge. An old, low-permeability deposit from Lake Bonneville occurs just below

this depth that causes the ground water to discharge. This

discharge takes several forms: evaporation, evaportranspiration, discharge into field drains which in turn discharge

to ditches, and discharge into seeps or springs. To the

author’s knowledge, no contamination reaches a water

supply, a permanent surface-water body, or a stream. At

the Hill AFB sites, plume stability appears to have been

achieved by a combination of mechanisms. There is certainly evidence of conventional degradation in ground

water, and at all of the sites some anaerobic dechlorination

is occurring. Plume stability appears to have been

achieved as a result of this degradation, coupled with

volatilization and discharge.

Summary

Natural attenuation of chlorinated compounds is an important process, and a full understanding will require looking

beyond the conventional aqueous-phase processes at

many sites. This will probably include both very slow

mechanisms we do not yet understand and a more careful

consideration of physical, chemical, and evapotransporative processes we have not often quantified in natural

attenuation studies.

Reference

1. INEL. 1995. Record of decision, declaration for the technical support

facility injection well (TSF-05) and surrounding groundwater contamination (TSF-23) and miscellaneous no action sites, final remedial

action. Idaho National Engineering Laboratory, Idaho Falls, ID. 1

 Parsons Engineering Science. 1996. Unpublished compilation of

data from six chlorinated solvent sites at Hill Air Force Base, UT.

143
Poster Session

Anaerobic Mineralization of Vinyl Chloride in Iron(III)-Reducing Aquifer Sediments

Paul M. Bradley and Francis H. Chapelle

U.S. Geological Survey, Columbia, South Carolina

In anaerobic aquifer systems, intrinsic bioremediation of

chlorinated ethenes is considered problematic because

of both the production of vinyl chloride during microbial

reductive dechlorination of higher chlorinated contaminants and the apparent poor biodegradability of vinyl

chloride under anaerobic conditions. Previous investigations have suggested that reductive dechlorination of

vinyl chloride to ethene may represent an environmentally significant pathway for in situ bioremediation of

vinyl chloride contamination. This poster provides laboratory evidence for an alternative mechanism of vinyl

chloride degradation: anaerobic oxidation of vinyl chloride under iron(III)-reducing conditions.

Microcosm experiments conducted with material collected from two geographically isolated, chlorinatedethene-contaminated aquifers demonstrated oxidation of [1,2-14C]vinyl chloride to 14CO2 by indigenous

microorganisms under iron(III)-reducing conditions.

Addition of chelated iron(III) (as Fe-EDTA) to aquifer

microcosms resulted in mineralization of up to 34

percent of [1,2-14C]vinyl chloride within 84 hours. The

results indicate that vinyl chloride can be mineralized

under iron(III)-reducing conditions, and that the

bioavailability of iron(III) is an important factor affecting the rates of mineralization. The microcosm results

are consistent with the attenuation of vinyl chloride

concentrations observed in the field and suggest that

contaminant oxidation coupled to microbial iron(III) reduction may be an environmentally significant mechanism contributing to intrinsic bioremediation of vinyl

chloride in anaerobic ground-water systems.

146
Intrinsic Biodegradation of Chlorinated Aliphatics Under Sequential

Anaerobic/Co-metabolic Conditions

Evan E. Cox and David W. Major

Beak Consultants, Guelph, Ontario

Leo L. Lehmicke

Beak Consultants, Kirkland, Washington 

Elizabeth A. Edwards

McMaster University, Hamilton, Ontario

Richard A. Mechaber

GEI Consultants, Inc., Concord, New Hampshire

Benjamin Y. Su

GEI Consultants, Inc., Winchester, Massachusetts

Tetrachloroethene (PCE) and trichloroethene (TCE) are

being biodegraded under naturally occurring sequential

anaerobic/co-metabolic conditions in ground water at an

inactive landfill in New Hampshire. Ground water in the

vicinity of the landfill is predominantly aerobic, with the

exception of an anaerobic zone that has developed at

the landfill source area where significant historical

biodegradation of dichloromethane, ketones, and aromatic hydrocarbons has occurred. Acetogenesis,

methanogenesis, sulphate reduction, and iron reduction

are the dominant microbial processes occurring in the

anaerobic zone. PCE and TCE have been sequentially

dechlorinated to cis-1,2-dichloroethene (cis-1,2-DCE) in

the anaerobic zone, to the extent that PCE and TCE are

no longer present at significant concentrations in the site

ground water. Cis-1,2-DCE concentrations attenuate

more rapidly (e.g., from 20 to less than 1 milligrams per

liter) than can be predicted based on physical processes

(i.e., advection, dispersion, retardation) alone. Vinyl

chloride (VC) and ethene concentrations do not account

for the extent of cis-1,2-DCE attenuation occurring.

Degradation of VC and ethene to carbon dioxide under

aerobic conditions (1) or anaerobic iron-reducing conditions (2) may result in an underestimation of cis-1,2-

DCE reductive dechlorination. Toluene and methane are

present in the downgradient aerobic ground water, however, and are likely promoting co-metabolic biodegradation of cis-1,2-DCE. Preliminary laboratory microcosm

studies have confirmed that the indigenous microorganisms can co-metabolize cis-1,2-DCE (and VC) in the

presence of toluene and methane at the concentrations

found in the site ground-water.

References

1. Cox, E.E., E.A. Edwards, L.L. Lehmicke, and D.W. Major. 1995.

Intrinsic biodegradation of trichloroethene and trichloroethane in a

sequential anaerobic-aerobic aquifer. In: Hinchee, R.E., J.T. Wilson, and D.C. Downey, eds. Intrinsic bioremediation. Columbus,

OH: Battelle Press. pp. 223-231.

2. Bradley, P.M., and F.H. Chapelle. Anaerobic mineralization of vinyl

chloride in Fe(III)-reducing, aquifer sediments. Environ. Sci. Technol. In press.

147
Analysis of Methane and Ethylene Dissolved in Ground Water

Steve Vandegrift, Bryan Newell, and Jeff Hickerson

ManTech Environmental Research Services Corporation, Ada, Oklahoma

Donald H. Kampbell

U.S. Environmental Protection Agency, 

National Risk Management Research Laboratory, Ada, Oklahoma

A headspace equilibrium technique and gas chromatography can be used to measure dissolved methane and

ethylene in water. A water sample is collected in a

50-milliliter (mL) glass serum bottle. Several drops of 1:1

diluted sulfuric acid are added. The bottles are then

capped using Teflon-lined butyl rubber septa. Later at

the analytical laboratory, a headspace is prepared by

replacing 10 percent of the water sample by helium. The

bottle is then shaken for 5 minutes. Aliquots of headspace, usually 300 microliters, are removed using a

gas-tight syringe. The subsample is injected into a gas

chromatograph with a Porapak Q stainless-steel column

and a flame ionization detector. The gaseous components are separated, and chromatogram peak retention

times and areas are compared with calibration standards. The concentration of the aqueous gas components

can be calculated using sample temperature, bottle volume, headspace concentrations, and Henry’s Law.

Limits of detection for methane and ethylene are

0.001 and 0.003 milligrams per liter (mg/L), respectively.

Determination of precision and accuracy for a 19.8

mg/L methane prepared sample using six replicates

was a standard deviation of 0.6 mg/L, risk-specific

dose (RSD) = 3.2 percent, and average recovery of

87 percent. Similar statistics for 118 mg/L ethylene using three replicates was a standard deviation of

8.8 mg/L, RSD=7.5 percent, and an average recovery of 90 percent. Typical dissolved methane

and ethene concentrations at natural attenuation

field sites have been less than 1 and less than 0.1

mg/L, respectively. Methane levels have always been

higher.

The method can also be adapted to determine ethane,

nitrous oxide, vinyl chloride, carbon dioxide, and possibly other dissolved gases in ground water.

148
Estimation of Laboratory and In Situ Degradation Rates for 

Trichloroethene and cis-1,2-Dichloroethene in a Contaminated Aquifer at

Picatinny Arsenal, New Jersey

Theodore A. Ehlke and Thomas E. Imbrigiotta

U.S. Geological Survey, West Trenton, New Jersey

Natural attenuation of chlorinated organic compounds in

aquifers includes apparent loss mechanisms, such as

biodegradation, advective transport, volatilization, sorption, and diffusion. Determination of quantitative degradation rates for the different processes is an important

step in planning cost-effective site remediation. Soil and

ground water at Picatinny Arsenal, New Jersey, have

been studied by the U.S. Geological Survey since 1986

to determine fate and transport of chlorinated ethenes

in a shallow unconfined aquifer. This poster describes

the methods used to quantify the major processes affecting fate and transport of trichloroethene (TCE) and

cis-1,2-dichloroethene (cis-DCE) in the aquifer.

Analyses of water and soil core samples, collected at a

series of locations within and outside a contaminant

plume, were used to identify the lateral and vertical

distribution of organic contaminants in the aquifer, major

electron acceptors, background geochemistry, and dissolved chemicals that affect biodegradation of chlorinated ethenes within the plume. Results indicated that

ground water within the plume contained TCE concentrations ranging up to 20 mg/L-1, methane concentrations generally less than 85 mg/L-1, and dissolved oxygen

and nitrate concentrations of less than 0.5 mg/L-1,

the major terminal electron accepting processes were

sulfate and iron(III) reduction; and anaerobic in situ

biodegradation of TCE and cis-DCE was occurring.

Following initial site characterization, soil cores were

collected from a series of locations along the major

ground-water flow path within the plume for determination of TCE and cis-DCE biodegradation rates in a

laboratory study.

Static batch microcosms were constructed under anaerobic conditions to determine the rates of TCE and

cis-DCE biodegradation. Sterilized 50-milliliter serum vials were filled to the base of the neck with composited

core materials and amended with a 2-milliliter sterile

aqueous solution of TCE or cis-DCE to bring the porewater chlorinated ethene concentration to 1,100 mg/L-1.

Pore-water samples from duplicate serum vials were

periodically assayed by gas chromatography to quantify

the chlorinated ethene concentrations. The results were

used to determine the first-order biodegradation rate

constants for TCE and cis-DCE, after compensation for

abiotic losses. First-order biodegradation rate constants

for TCE ranged from -0.004 wk-1 to -0.035 wk-1 and were

greater near the plume origin and the discharge point

(Green Pond Brook) than in the plume center. Geochemical results indicated that natural organic acids

leached from shallow peat deposits in the vadose zone

probably were a major electron donor for biodegradation

of chlorinated ethenes in situ. In general, cis-DCE was

degraded more slowly than TCE. First-order biodegradation rate constants for cis-DCE ranged from less than

-0.01 wk-1 to -0.05 wk-1. Biodegradation of cis-DCE was

most rapid in soils underlying a peat layer near the

plume discharge point.

Degradation of TCE in situ also was estimated using the

concentrations of chlorinated ethenes determined for a

series of monitoring wells along the major ground-water

flow path within the plume. Chlorinated ethene concentrations in ground water at up- and downgradient wells

measured at time intervals corresponding to the estimated TCE solute transport time between sites were

used to estimate first-order TCE removal rate constants

in the aquifer. In situ first-order rate constants for TCE

removal generally ranged from -0.012 wk-1 to -0.02 wk-1.

The close approximation of these in situ removal estimates to laboratory biodegradation rates indicated that

biodegradation in situ probably was a major removal

process for TCE at Picatinny Arsenal.

Results of in situ geochemistry and ground-water modeling were used to quantify the removal of TCE by major

processes in the unconfined aquifer at Picatinny Arsenal.

149
Diffusive, sorptive, and volatilization losses were estimated separately and used to correct for apparent in situ

TCE removal. Biodegradation is probably the major removal process for chlorinated ethenes in the aquifer,

removing about 400 kg/y-1 of TCE from the contaminant

plume. Advective transport removes about 47 kg/y-1 of

TCE in ground water, discharging to Green Pond Brook.

Volatilization and lateral diffusive losses of TCE from the

plume are estimated to total 10 to 50 kg/y-1. Sorptive

losses of TCE to aquifer soils are minor because of the

low organic carbon concentration of sediments in the

saturated zone.

150
Measurement of Dissolved Hydrogen in Ground Water

Mark Blankenship

ManTech Environmental Research Services Corporation, Ada, Oklahoma

Francis H. Chapelle

U.S. Geological Survey, Columbia, South Carolina

Donald H. Kampbell

U.S. Environmental Protection Agency, 

National Risk Management Research Laboratory, Ada, Oklahoma

A gas stripping procedure and reduction gas detector

can be used to measure aqueous concentrations of

hydrogen in ground water. Polyethylene tubing is placed

near the center of the screen in a well casing with the

other end connected to a peristaltic pump. After purging

several well volumes, a 250-milliliter (mL) glass sampling bulb is placed in the water sampling line. The bulb

is completely filled with water. Then the pump is

stopped, and nitrogen gas is injected into the bulb to

create a 20-mL headspace. The bulb outlet is placed at

a lower level then the inlet, and the pump is turned on.

A water flow of 200 mL per minute is maintained for 20

minutes to equilibrate the dissolved hydrogen with the

nitrogen gas phase. Duplicate 2-mL gas samples are

then removed with a gas-tight syringe for analysis by

the hydrogen detector. The hydrogen analyzer operates on the reaction principle of X + HgO (solid) → XO

+ Hg (vapor), where X represents any reducing gas. An

ultraviolet photometer quantitatively measures the resultant mercury vapor. Reduction gas species are identified

as chromatograph peaks at different retention times.

Retention time for hydrogen is less than 1 minute. The

limit of detection is 0.01 parts per million (ppm) hydrogen. A standard calibration curve over the range of 0.01

to 1.26 ppm hydrogen has a linear correlation coefficient

of R2

=1.00. A 1.0-ppm hydrogen in the gas phase corresponds to 0.8 nanomoles per liter of dissolved hydrogen for fresh water in equilibrium with a gas phase at 1

atmosphere. Typical dissolved hydrogen concentrations

detected at four different natural attenuation sites were

less than 10 nanomoles and most frequently less than

1 nanomole.

Successful sample assays depend on careful following

of procedure detail and overnight stabilization of the

detector.

151
Evidence of Natural Attenuation of Chlorinated Organics at Ft. McCoy, Wisconsin

Jason Martin

Rust Environment and Infrastructure, Sheboygan, Wisconsin

Ft. McCoy is a Resource Conservation and Recovery

Act regulated U.S. Army facility located in western

Wisconsin. Fire Training Burn Pit 1 (FTBP1) on the

site was operated from approximately 1973 to 1987.

Operations at the pit consisted of filling the pit with a

layer of water and fuel, then repeatedly igniting and

extinguishing the contents until the fuel was consumed. The soil beneath the 3-foot deep and 30-foot

diameter pit is a well sorted sand (low organic content)

with an average hydraulic conductivity of 0.0048 centimeters per second. The water table is generally 12 feet

below the ground surface.

Sampling activities conducted in 1993 and 1994 indicated significant concentrations of chlorinated organics

(1,2-dichloroethene [1,2-DCE], trichloroethene, and perchloroethene) in the soil and ground water. Chlorinated

organic contamination in the soil was limited to the area

under the former fire pit. Based on the local hydraulic

gradient and hydraulic conductivity, ground water present under FTBP1 when operations were initiated in

1973 has traveled an estimated 7,000 feet. Evidence of

natural attenuation of ground-water contamination is

provided by the short travel distance (approximately 600

feet) of the leading edge (1 microgram per liter) of the

chlorinated organics relative to the ground water over

the 20-year period and the decrease in size and concentration of the chlorinated organic contaminant plume

during the period of sampling (e.g., peak 1,2-DCE concentrations decreased from 2,100 to 700 micrograms

per liter during the sampling period).

Natural attenuation mechanisms potentially active on

ground-water contamination at the site include dispersion, sorption, volatilization, and biological degradation.

The bulk of the ground-water contamination was recently remediated using air sparging/soil vapor extraction. Based on the evidence of natural attenuation

present at the site and information in the U.S. Air Force

technical protocol on intrinsic remediation (1), natural

attenuation will be included as a component of the recommended remedial alternative for remaining groundwater contamination at this site.

Reference

1. Wiedemeier, T.H., J.T. Wilson, D.H. Kampbell, R.N. Miller, and J.E.

Hansen. 1995. Technical protocol for implementing intrinsic remediation with long-term monitoring for natural attenuation of fuel

contamination dissolved in groundwater. U.S. Air Force Center for

Environmental Excellence, San Antonio, TX.

152
Challenges in Using Conventional Site Characterization Data To Observe

Co-metabolism of Chlorinated Organic Compounds in the Presence of an

Intermingling Primary Substrate

Ian D. MacFarlane, Timothy J. Peck, and Joy E. Ligé

EA Engineering, Science, and Technology, Inc., Sparks, Maryland

Site characterization data from a leaking underground

storage tank (LUST) site and adjacent dry cleaners were

retrospectively analyzed for evidence of chlorinated solvent biodegradation. The sites are in the path of a wide

chlorinated solvent ground-water plume emanating from

the Dover Air Force Base (DAFB) in Dover, Delaware.

Discrete hydrocarbon and tetrachlorethene plumes

originate from the aforementioned LUST and dry cleaner

sources and mingle with the DAFB plume (1, 2). From

the chlorinated organics natural attenuation program at

DAFB (3) and our own laboratory studies using DAFB

sediment (4), evidence abounds regarding the potential

for natural biodegradation of chlorinated compounds in

the shallow Columbia aquifer.

We hypothesized that the subsurface, containing gasoline product, gasoline vapors, and high levels of dissolved hydrocarbons, was a likely area for co-metabolism

of chlorinated compounds derived from either DAFB or

the dry cleaners. In this case, the hydrocarbons would

serve as the primary substrate for co-metabolism of

chlorinated compounds mixing within the hydrocarboncontaminated zone. Soil vapor, multilevel hydropunch,

and monitoring well data from the LUST and dry cleaner

investigations were reviewed, looking specifically for

relationships between concentrations of hydrocarbons

(the presumed primary substrate), chlorinated solvents

(e.g., tetrachloroethene [PCE], trichloroethene [TCE],

and carbon tetrachloride), and chlorinated solvent

breakdown products (e.g., vinyl chloride, dichloroethene

[DCE], TCE, and chloroform).

Although some patterns of intrinsic biodegradation were

evident, the data did not make a compelling case for

co-metabolism in or near the hydrocarbon plume. The

most promising data were the soil gas concentrations,

which generally showed a decrease in the PCE:TCE

ratio with increase in hydrocarbon concentration, implying degradation of PCE to TCE in the presence of

hydrocarbon vapors. Even though numerous groundwater samples were obtained for the site characterization studies, no relationships could be established

for the ground-water regime.

We conclude that the data from this conventional site

characterization effort were either too limited in quality

(e.g., not enough analytes) or quantity to adequately

discern patterns, or that co-metabolism was not occurring in the saturated zone. Perhaps vapor diffusion in the

unsaturated zone promotes better substrate mixing than

in the saturated zone, where slow dispersion may limit

the effects of co-metabolism. This retrospective analysis

points out the need for careful development of a natural

attenuation conceptual model while planning site characterization efforts. Sampling and analysis not conventionally used in contaminant site assessments, particularly

for chlorinated natural attenuation assessments, may be

required to test the hypothesized conceptual model.

References

1. Peck, T.J., and I.D. MacFarlane. 1991. Multiphased environmental

assessment of intermingling subsurface contamination: A case

study. Proceedings of the 1991 Environmental Site Assessments:

Case Studies and Strategies Conference. Association of Ground

Water Scientists and Engineers. July.

2. Peck, T.J., and I.D. MacFarlane. 1993. Characterization of intermingling organic plumes from multiple sources. Presented at the

NGWA Annual Convention and Exposition. October.

3. Klecka, G.M. 1995. Chemical and biological characterization of

intrinsic bioremediation of chlorinated solvents: The RTDF Program at Dover Air Force Base. Presented at the IBC Intrinsic

Bioremediation/Biological Dehalogenation Conference, Annapolis,

MD. October 16-17.

4. Ligé, J.E., I.D. MacFarlane, and T.R. Hundt. 1995. Treatability

testing to evaluate in situ chlorinated solvent and pesticide bioremediation. In: Hinchee, R.E., A. Leeson, and L. Semprini, eds.

Bioremediation of chlorinated solvents. Columbus, OH: Battelle

Press.

153
Development of an Intrinsic Bioremediation Program for Chlorinated Solvents at

an Electronics Facility

Michael J. K. Nelson

Nelson Environmental, Kirkland, Washington

Anne G. Udaloy

Udaloy Environmental Services, Lake Forest Park, Washington

Frank Deaver

Deaver Environmental Group, Portland, Oregon

From 1963 to 1978, an area on a manufacturing facility

for electronic components was used to dispose of residual sludge from cleaning baths. The sludge contained chlorinated solvents, including trichloroethylene

(TCE), tetrachloroethylene (PCE) and 1,1,1-trichloroethane (TCA). An initial investigation of the site during

the mid to late 1980s revealed substantial levels of TCE

(up to 5,700 micrograms per liter) and TCA (up to 6,900

micrograms per liter) in the ground water. A corrective

measures study was performed, and corrective action

was implemented in the form of standard pump-andtreat activities.

After 5 years of pumping, it was evident that this method

was removing very little chlorinated solvent mass, and

alternative remediation methods were assessed. During a review of the historical data, it was determined

that the concentration of chlorinated solvents had

greatly decreased before implementation of the pumpand-treat program and that site soils were likely to be

anaerobic, potentially allowing natural biodegradation

of TCE and related solvents. Discussions held with the

regulatory agencies, the U.S. Environmental Protection

Agency and the Oregon Department of Environmental

Quality, resulted in a program designed to investigate

intrinsic bioremediation as a viable remedial option for

the site.

Information was obtained using Geoprobe sampling techniques; evidence of anaerobic conditions and of the anaerobic breakdown products of the contaminants was

sought. The results indicated anaerobic conditions; this

was based on low to nondetectable dissolved oxygen,

dissolved nitrogen predominantly as ammonia, high levels

of ferrous iron (up to 85 milligrams per liter), and significant

levels of methane (up to 1.2 milligrams per liter). The

results also indicated that TCE was being biodegraded by

sequential, reductive dechlorination to nonchlorinated

products prior to reaching the site boundary. Both cis-1,2-

dichloroethylene and vinyl chloride were detected near the

source area at maximum concentrations of 130 and 12

micrograms per liter, respectively, then decreased to near

or below the detection level at the site boundary. Low

levels of the nonchlorinated product, ethylene, were detected downgradient of the source area.

Subsequent discussions with the agencies led to an

agreement that intrinsic bioremediation was a viable

remedial alternative for contaminant containment and

eventual cleanup. The ground-water pump-and-treat

system is being decommissioned, and a monitoring program is being implemented to track and ensure that

adequate remediation of the site continues by intrinsic

bioremediation. Implementation of this program is allowing redevelopment of this site.

154
Overview of the U.S. Air Force Protocol for Remediation of Chlorinated Solvents

by Natural Attenuation

Todd H. Wiedemeier

Parsons Engineering Science, Inc., Denver, Colorado

John T. Wilson and Donald H. Kampbell

U.S. Environmental Protection Agency, National Risk Management Research Laboratory,

Subsurface Protection and Remediation Division, Ada, Oklahoma

Jerry E. Hansen and Patrick Haas

U.S. Air Force Center for Environmental Excellence, Technology Transfer Division,

Brooks Air Force Base, Texas

The U.S. Air Force Center for Environmental Excellence, Technology Transfer Division (AFCEE/ERT), in

conjunction with personnel from the U.S. Environmental

Protection Agency’s National Risk Management Research Laboratory (NRMRL) and Parsons Engineering

Science, Inc. (Parsons ES), has developed a technical

protocol to document the effects of natural attenuation

of fuel hydrocarbons dissolved in ground water. This

same group is currently developing a similar protocol for

confirming and quantifying natural attenuation of chlorinated solvents. The intended audience for the new

protocol is U.S. Air Force personnel and their contractors, scientists, and consultants, as well as regulatory

personnel and others charged with remediating ground

water contaminated with chlorinated solvents.

Mechanisms of natural attenuation of chlorinated solvents include biodegradation, hydrolysis, volatilization,

advection, dispersion, dilution from recharge, and sorption. Patterns and rates of natural attenuation can vary

markedly from site to site depending on governing physical and chemical processes. The proposed protocol

presents a straightforward approach based on state-ofthe-art scientific principles that will allow quantification

of the mechanisms of natural attenuation. In this way,

the effectiveness of each mechanism can be evaluated

in a cost-effective manner, allowing a decision to be

made regarding the effectiveness of natural attenuation

as a remedial approach.

155
Incorporation of Biodegradability Concerns Into a Site Evaluation 

Protocol for Intrinsic Remediation

Robert M. Cowan, Keun-Chan Oh, Byungtae Kim, and Gauri Ranganathan

Rutgers University, Cook College, Department of Environmental Sciences, 

New Brunswick, New Jersey

A project is being conducted to develop a site evaluation

protocol for determining the potential applicability of intrinsic remediation at industrial sites with soil and

ground-water contamination. The project is sponsored

by an industry-supported research center because the

sponsor industries are interested in extending the applicability of currently available intrinsic remediation protocols (e.g., the U.S. Air Force guidance document by

Wiedemeier et al. [1]) to include any biodegradable contaminant, not just benzene, toluene, ethylbenzene, and

xylenes and related (fuel-derived) compounds. To extend the protocol in this manner, the biodegradability of

any contaminants that may exist at these sites must be

addressed because the knowledge of contaminant biodegradability can be an absolute requirement for application of intrinsic remediation. How to go about this is

the focus of the work.

Progress on the project to date has been the development of a preliminary site screening document and a

draft of the protocol to determine biodegradability. In

addition, information has been collected concerning

contamination at several industrial sites, and one site

has been selected for more detailed study. The site

selected contains a contaminated fractured bedrock

aquifer so we are experiencing difficulty concerning the

predictability of contaminant transport in addition to the

contaminant biodegradability issues that were initially

the focus of the project.

This poster will:

• Present an overview of intrinsic remediation technology and definitions for related terminology.

• Discuss preliminary site screening using existing

data to make an initial determination as to whether

intrinsic bioremediation is likely to be suitable for a

given site; the goal is to decide whether a more

detailed look at the site should be taken.

• Describe the biodegradability assessment protocol,

which contains two sections: assessment of biodegradability through a search of existing databases

and the literature, and experimental methods for the

determination of in situ biodegradability.

• Depict a flow chart, based on biodegradability concerns, that can be used to select and implement the

appropriate approach for making a detailed assessment of the potential for intrinsic remediation.

• Give the current status of the industrial site study.

Reference

1. Wiedemeier, T.H., J.T. Wilson, D.H. Kampbell, R.N. Miller, and J.E.

Hansen. 1995. Technical protocol for implementing intrinsic remediation with long-term monitoring for natural attenuation of fuel

contamination dissolved in groundwater. U.S. Air Force Center for

Environmental Excellence, San Antonio, TX.

156
A Field Evaluation of Natural Attenuation of Chlorinated Ethenes 

in a Fractured Bedrock Environment

Peter Kunkel and Chris Vaughan

ABB Environmental Services, Inc., Portland, Maine

Chris Wallen

Hazardous Waste Remedial Actions Program, Oliver Springs, Tennessee

Before a long-term ground-water monitoring program

was conducted in support of natural attenuation as a

remedial remedy for halogenated organic contamination, a focused evaluation of ground-water chemistry

provided valuable insight into attenuative mechanisms

in areas where remedial options were being evaluated.

This poster describes a field investigation and data analysis at Loring Air Force Base, Limestone, Maine, and presents the results of the evaluation of natural attenuation.

Chemical data collected prior to this evaluation indicated

the presence of chlorinated ethenes (tetrachloroethene

[PCE] and trichloroethene [TCE], cis-1,2-dichloroethene

[cis-1,2-DCE] and vinyl chloride [VC]) in the fractured

bedrock ground-water environment at several locations

on site. Samples representative of the interior and exterior

of the chlorinated hydrocarbon plumes were collected at

pre-existing basewide remedial investigation locations.

The analytical protocol included hydrocarbon target

compounds, ground-water quality parameters, indicator

parameters, electron acceptors, and microbial community evaluations. Several of the contaminant plumes

demonstrated characteristics of reductive dehalogenation, indicating a potential for natural degradation of

PCE and TCE to cis-1,2-DCE and VC in a fractured

bedrock environment. Dissolved oxygen and nitrate

concentrations were depleted, oxidation/reduction potential values and sulfate concentrations decreased,

and methane concentrations were observed at locations

where chlorinated ethenes were detected.

157
Intrinsic Bioattenuation of Chlorinated Solvents in a Fractured Bedrock System

William R. Mahaffey and K. Lyle Dokken

Walsh Environmental Scientists & Engineers, Inc., Boulder, Colorado

Spent chlorinated solvents were released from two underground storage tanks at the Colorado Department of

Transportations materials testing laboratory in Denver,

Colorado. An estimated 5,000 to 15,000 liters of trichloroethene, 1,1,1-trichloroacetic acid (TCA), 1,1,2-TCA,

dichloromethane, benzene, toluene, ethylbenzene,

xylene, and asphaltic compounds were released into a

highly fractured bedrock consisting of interbedded

claystone, siltstone, and fine-grained sandstone. The

resulting dense, nonaqueous-phase liquid resides between 20 and 30 feet below ground surface (bgs).

Downward migration has been impeded by a relatively

massive claystone at 30 to 40 feet bgs, although some

solvents are present at a depth of more than 50 feet in

a siltstone. The ground-water plume, consisting of

source compounds and products of reductive dechlorination (e.g., 1,1-DCE, 1,2-DCE, 1,1-DCA and 1,2-DCA),

has migrated in excess of 4,500 feet off site.

This site is being characterized for intrinsic bioattenuation to establish baseline conditions prior to the potential implementation of a source removal action,

recognizing that substantial residuals would likely remain. An anaerobic core in the source area has been

characterized on the basis of water chemistry differences between the plume and inflowing upgradient

ground water. Downwell probe sondes were used to

measure dissolved oxygen, pH, redox potential, and

temperature. Zero headspace ground-water samples

were collected into 160 milliliter serum bottles using a

Grundfos submersible pump and were immediately

capped with Teflon lined caps. Analysis for methane,

ethane, ethene, and hydrogen was performed by headspace analysis after displacing a fixed volume of water

by nitrogen gas displacement. Aqueous samples were

analyzed for NO3-, PO4-3, SO4-2, Cl-

, S-2, and Fe+2

using Hach methodologies; total organic carbon, chemical oxygen demand, and bicarbonate were analyzed

using standard methods. An evaluation of microbial

populations in ground water was performed using the

phospholipid fatty acid procedure.

Low levels of dissolved oxygen, in conjunction with

the identification of elevated methane, ferrous iron,

and chloride levels in the source area of the plume,

indicate the presence of anaerobic activity. Significant

reductions in the levels of inflowing nitrate within the

source area of the plume have been observed and

appear to be coincident with the reductions in the levels

of aromatic hydrocarbon constituents within and downgradient of the source area. Intrinsic bioattenuation of

dichloromethane (DCM) appears to be occurring based

on contaminant transport model predictions (MT3D)

and actual field measurements of the DCM plume

dimensions. Further indication of intrinsic bioattenuation has been the identification of low levels (15 parts

per billion) of vinyl chloride immediately downgradient

of the source area. 

158
Modeling Natural Attenuation of Selected Explosive Chemicals

at a Department of Defense Site

Mansour Zakikhani and Chris J. McGrath

U.S. Army Engineers Waterways Experiment Station, Vicksburg, Mississippi

Natural attenuation of explosives in the subsurface has

received considerable attention during recent years.

The idea behind natural attenuation is that within a

reasonable time natural processes can degrade effectively some explosive chemicals. One site selected to

evaluate natural attenuation is located at the Louisiana

Army Ammunition Plant (LAAP) in northwest Louisiana

approximately 22 miles east of Shreveport. The study

site is the area including the former Area P lagoons, 16

unlined lagoons covering approximately 25 acres. The

Area P lagoons were used sporadically between 1940

and 1981. Untreated, explosive-laden wastewater from

munition packing operations within LAAP was collected

in concrete sumps at each of several facilities and

hauled by tanker to Area P. The site also was used as a

burning ground for many years.

LAAP was placed on the National Priority List (NPL) in

March 1989 due to detection of measurable explosive

chemicals in the soil and ground water and its proximity

to water supply wells. As part of an interim remediation,

the wastewaters at Area P were removed and the soil

was excavated to a depth of 5 feet. The total explosives

concentration in untreated soil was in excess of 100

milligrams per kilogram. Excavated soil was incinerated,

and treated soil was used to backfill the area. The

concentration of treated soil was below a detection limit

(BDL). A natural cap of low permeability was placed over

the site to inhibit infiltration and further migration of

residual explosives below the excavation depth.

The monitoring wells at LAAP have been sampled and

analyzed for explosives since 1982. The results of these

analyses are maintained in the U.S. Army Environmental

Center database (IRDMIS). A comparison between 1990

and 1994 data for trinitrotoluene (TNT) and RDX concentrations within and adjacent to Area P showed a general

decrease during this period. The concentration of TNT in

1990 ranged from 16,000 to 55.6 micrograms per liter

(μg/L); by 1994, the concentration ranged from 11,000 μg/L

to BDL. The RDX concentration ranged from 7,600 to 33.8

μg/L in 1990 and from 8,400 to 14.4 μg/L in 1994. Although

these two data sets indicated a general downward trend

in contamination at Area P due to remedial measures

and/or natural attenuation, a few monitoring wells showed

the opposite trend. To clarify the conflicting results and

provide a better understanding of explosives attenuation,

eight additional monitoring wells have been installed at the

site since 1995.

This poster discusses the feasibility of applying three-dimensional ground-water flow and transport to this heterogeneous aquifer. The capability of a comprehensive

computer graphical system—Groundwater Modeling

System (GMS)—which is used in the modeling of the

site, also will be discussed and illustrated.

159
Long-Term Application of Natural Attenuation at Sierra Army Depot

Jerry T. Wickham

Montgomery Watson, Walnut Creek, California

Harry R. Kleiser

U.S. Army Environmental Center, Aberdeen Proving Ground, Maryland

A record of decision (ROD) for Sierra Army Depot selecting natural attenuation and degradation for treatment

of ground water was signed by the state of California

and the Army on September 8, 1995—the first approved

ROD in the United States selecting natural attenuation as

a primary remedial alternative for trichloroethene (TCE)

and explosives in ground water. The natural attenuation

alternative consists of institutional controls to eliminate

future use of ground water in the area surrounding the

site, long-term ground-water monitoring, and evaluation

of contaminant migration and degradation rates.

Explosives and volatile organic compounds (VOCs) are

present in shallow ground water over a 26-acre area of

Sierra Army Depot, which is located approximately 50

miles northwest of Reno, Nevada. No surface water

features or water supply wells exist within the area of

the site. A ground-water plume of explosive compounds

originates from the TNT Leaching Beds, a facility used

during the 1940s for percolation of waste water from a

shell washout facility. Dissolved explosive compounds

in the ground water include RDX, 1,3,5-trinitrobenzene,

HMX, and minor concentrations of numerous other explosive compounds. The highest concentration of total

explosive compounds detected is 1,200 micrograms per

liter within the vicinity of the former leaching beds. A

VOC plume originates from a former paint shop used

during the 1940s and 1950s for the renovation of ammunition. Dissolved VOCs present in the highest concentrations are TCE, chloroform, and carbon tetrachloride.

The highest concentration of trichloroethene detected is

1,000 micrograms per liter in a monitoring well 175 feet

downgradient from the former paint shop.

Under current conditions, the plumes appear to migrate

at slow rates. Estimated ground-water flow velocities

across the site range from 1 to 140 feet per year, with

average estimated ground-water velocities of 2 to 6 feet

per year. The shallow aquifer is highly stratified, with

numerous fine-grained layers in the upper 25 feet. Because contaminants have diffused into the fine-grained

layers over approximately a 50-year period, restoration

of ground water to background or drinking-water quality

by pump-and-treat or other active remediation does not

appear feasible. Long-term ground-water monitoring of

the plumes is expected to provide data on degradation

reactions that may occur at slow rates over extended

periods. In addition to providing these data, this action

could save the Army up to $10 million in ground-water

remediation costs at Sierra Army Depot.

160
When Is Intrinsic Bioremediation Cost-Effective? Financial-Risk Cost-Benefit

Analysis at Two Chlorinated Solvent Sites

Bruce R. James, Evan E. Cox, and David W. Major

Beak Consultants, Guelph, Ontario

Katherine Fisher

Beak Consultants, Brampton, Ontario

Leo G. Lehmicke

Beak Consultants, Kirkland, Washington

Interest in intrinsic bioremediation and natural attenuation as remediation alternatives for chlorinated solvent

sites is rapidly growing because the methods are significantly more cost-effective than conventional remediation alternatives (e.g., pump-and-treat). When evaluating

the long-term cost-effectiveness of intrinsic bioremediation and natural attenuation alternatives, however, many

analysts and decision-makers consider direct engineering costs, such as capital, operation and maintenance,

and monitoring costs, but fail to adequately assess the

potential legal and corporate costs that may arise from

choosing an intrinsic-based remediation alternative. If

the alternative fails, for example, additional costs would

be incurred to address remediation with a new method

or to deal with the land’s decrease in value or marketability or possible legal action. Financial-risk cost-benefit

analysis, which incorporates a more comprehensive set

of costs in the cost-effectiveness analysis, is a tool that

analysts and decision-makers can use to evaluate objectively whether intrinsic bioremediation and natural

attenuation are in fact the most cost-effective remediation alternatives in the long run.

This poster presents the results of using financial-risk

cost-benefit analysis to examine the impact of cost factors other than engineering costs on the long-term costeffectiveness of intrinsic bioremediation versus other

remediation alternatives under various scenarios at two

chlorinated solvent sites. At both sites, chlorinated volatile organic compounds are currently being intrinsically

bioremediated to environmentally acceptable end products (e.g., ethene and ethane).

161
Natural Attenuation of Chlorinated Organics in Ground Water:

The Dutch Situation

Lex W.A. Oosterbaan and Hans Rovers

Tebodin, Environment and Safety Department, The Hague, The Netherlands

National Policy in General

The official goal of the policy on soil remediation in the

Netherlands is future “multifunctional use” of the sites in

question. This goal is triggered by the shortage of suitable sites for development (including the development

of nature reserves and ecological values) and by consciousness and care for the environment.

A recent chapter of the Soil Protection Act, however,

presents guidelines for decision-makers to deviate from

multifunctional use. These guidelines are based on

“specific local conditions.” Thus, the goal in the field has

shifted to future “functional use” of sites.

Through a remedial option analysis, which is comparable to the U.S. Environmental Protection Agency’s remedial investigation/feasibility study, remedial technologies

are identified and remedial action alternatives evaluated. The evaluation is based on technical and financial

feasibility on one hand and environmental impact on the

other. Environmental impact is substantiated by analyses of human risks and by considerations of ecological

deterioration and migration of contaminants into the

environment.

National Policy on Natural Attenuation

Chlorinated volatile organic compounds (VOCs) have

been extensively used in metal processing and dry

cleaning industries in the Netherlands and are a major

component in many soil and ground-water contamination problems. This is due to the chemical and physical

properties of the contaminants in relation to permeable

soils, high water tables, and regional aquifers, which

contain valuable ground-water resources for the drinking water supply.

Target and intervention values for these compounds in

ground water are set low, as illustrated below.

Target Value

From the Soil

Protection Act

Intervention

Value From the

Soil Protection

Act

Trichloroethylene (TCE)

Perchloroethylene (PCE)

Vinylchloride (VC)

0.01 μg/L

0.01 μg/L

Not available

500 μg/L

40 μg/L

0.7 μg/L

Provisional

target value

Provisional

intervention value

cis-1,2-Dichloroethene (DCE) 0.01 μg/L 10 μg/L

To date, no protocols have been developed for natural

attenuation. Information on the behavior and fate of

VOCs and on the survey of these items is available,

although not readily so. Practical experience is emerging. Target and intervention values are absolute values,

that is, definitive and ultimate. The benefit of a time span

favoring ongoing attenuation is not considered fully.

Presently much energy and funding are put into the

NOBIS research program, a government-subsidized

program to develop in situ bioremediation techniques

for a great variety of contamination problems. The

program focuses on investigation and remediation and

consists of laboratory and field experiments and of

the enhancement of potentially promising technologies,

feasibility studies, and methodologies for decision

analysis. Unique in this program is the collaboration

between government, industrial polluters, research institutes, and environmental consultancy firms/contractors.

The results of the various projects are accessible for

NOBIS participants immediately and will become public

domain later.

162
Natural Attenuation as a Cleanup Alternative for Tetrachloroethylene-Affected

Ground Water

Steve Nelson

EMCON, Bothell, Washington

A chlorinated solvent storage and transfer facility operated in an industrial area of Seattle, Washington,

from the mid-1940s to the mid-1970s. Historical releases of tetrachloroethylene (PCE) at fill pipes and

underground storage tanks have migrated into a shallow

sand aquifer underlying the site. A recently completed

field screening and ground-water sampling investigation characterized the nature and extent of a local

PCE ground-water plume and a more extensive plume

of cis-1,2-dichloroethylene and vinyl chloride. Additional ground-water chemistry data, including nitrogen, phosphorus, iron, sulfur, dissolved oxygen, and

permanent gas (methane, ethane, ethene) concentrations, were collected. Elevated concentrations of ferrous iron (11 parts per million [ppm]), sulfide (0.39

ppm), and ammonia (14 ppm), and low concentrations

of dissolved oxygen (0.25 ppm) indicate anaerobic

conditions in the source area that are conducive to

natural attenuation of PCE. Methane, ethane, ethene,

cis-1,2-dichloroethylene, and vinyl chloride concentrations increase by one to two orders of magnitude

150 feet downgradient of the source area. Nearsaturation concentrations of PCE decrease by several

orders of magnitude over the same distance. Preliminary estimates indicate a half-life of 150 to 200 days

for PCE degradation.

There are no beneficial uses of ground water in the

industrial area, and ground-water discharges to a surface-water body 2,500 feet from the site. Concentrations

of the contaminants of concern at the property boundary

are lower than Washington State surface-water quality

criteria. Because natural attenuation appears to effectively remediate the chlorinated hydrocarbons, the proposed remedial action for the site will be limited to

ground-water monitoring.

163
Natural Attenuation of Trichloroethene in a Sandy Unconfined Aquifer

Neale Misquitta, Dale Foster, and Jeff Hale

Key Environmental, Inc., Carnegie, Pennsylvania

Primo Marchesi and Jeff Blankenship

American Color and Chemical Corporation, Lockhaven, Pennsylvania

The natural attenuation of dissolved-phase trichloroethene (TCE) in ground water was evaluated at a

state-regulated, operating chemical plant in South Carolina. Natural attenuation was documented via the observed attenuation and loss of TCE within a sandy

unconfined aquifer (approximately 25 feet thick with

Kh=10-3 centimeters per second), 8 years of groundwater monitoring data, and modeling with site-specific

retardation coefficients.

The evaluation and demonstration of natural attenuation

of TCE was part of a successful technical argument that

considered the natural microbial and/or geochemical

attenuation processes in the establishment of downgradient ground-water quality compliance points, obviating the need for containment or other remedial actions.

The recently promulgated South Carolina Groundwater

Mixing Zone Regulations require that, under very specific and stringent attenuation conditions, alternate

ground-water protection standards are addressed in

zones where attenuation of dissolved-phase chemicals

is demonstrated.

No relationship between TCE, electron acceptors, and

biodegradation byproduct isopleth maps was observed,

suggesting that TCE was not degrading via aerobic or

anaerobic pathways. Elevated microtoxicity levels were

observed, and minimal quantities of both aerobic and

anaerobic TCE degraders were identified. The empirically

calculated Kd (using soil total organic carbon [TOC]) was

estimated to be 0.4 liters per kilogram. The resulting

empirically calculated retardation factor did not correlate

with the observed attenuation of TCE at the site, indicating that non-TOC related mechanisms were contributing to TCE attenuation. Consequently, a site-specific

Kd was estimated via batch adsorption tests employing toxicity characteristic leaching procedure extraction techniques, using ground water and soils from the

site area of interest. A site-specific Kd of 10 liters per

kilogram was estimated through these tests. The sitespecific retardation factor correlated with the observed

natural attenuation of TCE. Differences in the sitespecific retardation factor and the empirically calculated estimate may be attributed to soil/ground-water

geochemical interactions, such as low pH-induced

bonding of the TCE to the soil matrix, which are unrelated to TOC.

Subsequent ground-water modeling using the site-specific retardation factor (and Kd) indicates that dissolvedphase TCE would not migrate to the downgradient

receptor for a minimum period of 100 years. The final

natural attenuation remedy for the site, recommended

in the mixing zone application submitted to South Carolina, included a time-weighted “monitoring only” component with no active remediation. This application is

currently under review.

164
Analysis of Intrinsic Bioremediation of Trichloroethene-Contaminated

Ground Water at Eielson Air Force Base, Alaska

Kyle A. Gorder, R. Ryan Dupont, Darwin L. Sorensen,

Maria W. Kemblowski, and Jane E. McLean

Utah Water Research Laboratory, Utah State University, Logan, Utah

A simple ground-water model was used to determine the

apparent rate of trichloroethene (TCE) transformation,

to estimate the mass of TCE and its transformation

products, and to predict the effects of active treatment

options, such as source removal, at Eielson Air Force

Base, Alaska.

A modification of the three-dimensional solution to the

advection-dispersion-reaction equation (ADRE) proposed by Domenico (1) was used to estimate the rate

of TCE transformation. The model was calibrated using

the spatial distribution of TCE observed during a field

sampling event conducted in July 1995. TCE concentrations as high as 90,000 micrograms per liter were observed at the site and utilized in the model calibration

effort. The calibrated model showed that intrinsic remediation of TCE is occurring at the site. The estimated

first-order degradation rate for TCE ranged from 0.0020

to 0.0064 day-1.

TCE mass and apparent mass degraded were also estimated using the calibrated model. TCE mass predictions

using the model closely matched TCE mass calculated

from observed ground-water data. The TCE mass degraded was used to estimate the mass of TCE products

that would be present in the system, assuming these

products are accumulating. A comparison of observed

product mass to the estimated mass of these compounds showed that the mass of these compounds

present was significantly less than estimated, suggesting rapid transformation of the compounds to nonchlorinated compounds.

The calibrated model was also used to predict the effects of source removal on the lifetime of the dissolved

TCE plume. These predictions, along with source lifetime estimations, suggest that source removal activities

may not significantly reduce the time required to meet

cleanup goals for the site.

Reference

1. Domenico, P.A. 1987. An analytical model for multidimensional

transport of a decaying contaminant species. J. Hydrol. 91:49-58.

165
Involvement of Dichloromethane in the Intrinsic Biodegradation of 

Chlorinated Ethenes and Ethanes

Leo L. Lehmicke

Beak Consultants, Kirkland, Washington

Evan E. Cox and David W. Major

Beak Consultants, Guelph, Ontario

The metabolism of dichloromethane (DCM) by acetogenic

microorganisms has resulted in the production of an electron donor (acetic acid) that is stimulating reductive

dechlorination of tetrachloroethene (PCE), trichloroethene

(TCE), and 1,1,1-trichloroethane (TCA) to ethene and ethane in a shallow aquifer beneath a bulk chemical transfer

facility in Oregon. DCM, TCE, and toluene releases as well

as de minimis losses of PCE, TCA, ethylbenzene, and

xylene have occurred at the site. DCM concentrations in

the source area decreased by an order of magnitude (from

2,300 milligrams per liter [mg/L] to 190 mg/L) between

1990 and 1995, with corresponding production of acetic

acid. The distribution of DCM attenuates two orders of

magnitude to less than 1 mg/L within 100 meters from the

source area, far more rapidly than predicted by its mobility in the site ground water. PCE, TCE, and TCA concentrations also attenuate more rapidly downgradient from

the source area than would be predicted by their mobilities

relative to the ground-water velocity at the site. The distributions of 1,2-dichloroethene, vinyl chloride (VC), 1,1-dichloroethane, and chloroethane (CA) increase downgradient

from the source area. Ethene and ethane are present in

the ground water downgradient from the source area, in

association with VC and CA, indicating that the chlorinated

volatile organic compounds are being dechlorinated to

environmentally acceptable end products. Intrinsic bioremediation is being considered as a remediation alternative

for this site.

166
Intrinsic Bioremediation of 1,2-Dichloroethane

Michael D. Lee

DuPont Central Research and Development, Newark, Delaware

Lily S. Sehayek

DuPont Environmental Remediation Services, Wilmington, Delaware

Terry D. Vandell

Conoco, Ponca City, Oklahoma

Spills of 1,2-dichloroethane, also known as ethylene

dichloride (EDC), resulted in free- phase contamination

of a Gulf Coast site. There are two aquifers beneath the

site, as well as peat, clay, and silt layers. An ongoing

recovery and hydraulic containment program in the shallow aquifer is recovering nonaqueous-phase liquid

(NAPL) and dissolved-phase EDC. Degradation products of EDC, including 2-chloroethanol, ethanol, ethene,

and ethane, were detected in both the highly contaminated upper aquifer as well as in the deeper, less contaminated aquifer. Possibly as a result of cross

contamination during drilling operations, low concentrations (less than 1.0 parts per million) of dissolved EDC

were detected in the deeper aquifer. 

EDC concentrations in wells in the deeper aquifer have

decreased greatly over the last year, to between less

than 0.005 parts per million (the detection limit) and 0.05

parts per million. First-order decay half-lives for loss of

EDC from wells in this aquifer range from 64 to 165

days. Laboratory microcosm studies demonstrated that

microbes from the deeper aquifer can transform EDC

under anaerobic conditions. A geochemical evaluation

demonstrated that microbes at the site are capable of

using oxygen, nitrate, sulfate, iron, manganese, and

carbon dioxide as electron acceptors; elevated methane

concentrations indicate carbon dioxide is the major electron acceptor. 

Modeling efforts with DuPonts comprehensive multiphase NAPL model revealed that free-phase EDC will

not reach the underlying aquifer because of retention of

the free-phase EDC in the overlying silt and clay zones

and ongoing intrinsic biodegradation of the dissolvedphase EDC. The three-dimensional, three-phase finite

difference model includes simultaneous flow of water,

gas, and organic phases; energy transport; temperature-, pressure-, and composition-dependent interphase

partitioning; and dispersive transport within phases. The

model was originally developed by Sleep and Sykes (1,

2) and modified by Sehayek. The modified model is not

commercially available.

References

1. Sleep, B.E., and J.F. Sykes. 1993. Compositional simulation of

groundwater contamination by organic compounds, 1. Model development and verification. Water Resour. Res. 29(6):1697-1708.

2. Sleep, B.E., and J.F. Sykes. 1993. Compositional simulation of

groundwater contamination by organic compounds, 2. Model applications. Water Resour. Res. 29(6):1709-1718.

167
A Practical Evaluation of Intrinsic Biodegradation of 

Chlorinated Volatile Organic Compounds

Frederick W. Blickle and Patrick N. McGuire

Blasland, Bouck & Lee, Inc., Boca Raton, Florida

Gerald Leone

Waste Management, Inc., Atlanta, Georgia

Douglas D. Macauley

Reynolds Metals Corporation, Richmond, Virginia

At a former industrial site, intrinsic bioremediation was

evaluated to address low levels (less than 100 micrograms per liter [μg/L]) of chlorinated volatile organic

compounds (VOCs) in ground water. The VOCs detected in ground water include chlorinated ethanes, 1,1-

dichloroethene, vinyl chloride, chlorobenzene, benzene,

toluene, and ethylbenzene. Total VOC concentrations

ranged from not detected to 530 μg/L. Historically, the

site was mined for rock and subsequently used for the

disposal of tailing sands and clay waste from ore processing. As a result, a complicated ground-water system

consisting of at least five water-bearing units exists at

the site. Although current remedial activities at the site,

including a ground-water pump-and-treat system, have

been effective at reducing VOC levels to their present

concentrations, continued pumping does not appear to

be effective at further concentration reduction.

To reevaluate remedial options, an assessment of naturally

occurring transformation processes was performed. Initially, the assessment included VOC data over time,

collected to monitor the ground-water pump-and-treat system. Long-term ground-water monitoring results indicate

that concentrations of parent VOC compounds have

been reduced in all water-bearing units; after an associated temporary increase, a reduction in concentrations

of reduction dehalogenation breakdown products was

observed.

To further determine whether the natural attenuation

observed at the site is a result of intrinsic bioremediation, a study was implemented involving field monitoring

and ground-water sampling and analysis for select geochemical indicator compounds and dissolved permanent gases. The geochemical indicator compounds

included NO3/N, total and dissolved iron, and SO4/S.

Dissolved permanent gases include oxygen, CH4, and

CO2. Redox potential and pH were field measured. Concentrations of organic compounds were evaluated over

time, and trends in inorganic indicator compound and

dissolved permanent gas concentrations were evaluated spatially. Results of this study strongly suggest that

intrinsic bioremediation is responsible for transformation

of the VOCs present in site ground water. This poster

discusses the study and provides results for evaluating

bioremediation of chlorinated VOCs in ground water.

168
New Jersey’s Natural Remediation Compliance Program:

Practical Experience at a Site Containing Chlorinated Solvents and

Aromatic Hydrocarbons

James Peterson and Martha Mackie

McLaren/Hart Environmental Engineering Corporation, Warren, New Jersey

In recent years, regulatory agencies have begun to

place increasing emphasis on understanding the natural

mechanisms of contaminant degradation/attenuation in

ground-water at sites undergoing remediation. Guidelines and criteria for natural remediation assessments

have been established at both the state and federal

level, providing the regulated community with an improved ability to determine site conditions under which

a natural remediation approach is feasible and will be

acceptable to regulators. These initiatives reflect transition from conventional remedy selection to consideration and appropriate implementation of alternate

remedies that incorporate considerations of risk and

cost-effectiveness.

One good example of this evolving regulatory process

is the Natural Remediation Compliance Program

(NRCP), developed by the New Jersey Department of

Environmental Protection (NJDEP). General guidelines

(termed “minimum requirements”) for natural remediation proposals were defined concurrent with NJDEP’s

establishment of the NRCP in 1994 and have been

augmented recently by detailed technical suggestions

for screening of sites (1).

In late 1994, McLaren/Hart Environmental Engineering

Corporation conducted investigative and remedial activities that led to a proposal to implement the NRCP at a

New Jersey industrial site with soil and ground-water

affected by chlorinated solvents and aromatic hydrocarbons. The NRCP proposal, submitted as part of a remedial action workplan for site ground water and

concurrent with a remedial action report for source area

soils, addressed the following NJDEP prerequisite conditions (“minimum requirements”) for natural remediation proposals: delineation and remediation of sources;

contaminant migration assessment to confirm receptors

not at risk; documentation of degradibility and/or attenuation capacity; identification of site-specific characteristics favorable to natural degradation and/or

attenuation; establishment of a sentinel well system;

development of a ground-water monitoring program;

documentation regarding current and potential future

ground-water uses; and written notification to potentially

affected downgradient property owners.

Specific activities conducted to address the requirements included delineation of source area soils using a

Geoprobe, source area soil excavation/disposal,

postexcavation sampling, ground-water sampling, in situ

measurement of ground-water field parameters, a well

search, and an evaluation of potential receptor impacts

through modeling. The results of these investigations

suggested that “steady state” conditions of contaminant

influx and attenuation were in effect, and that, even in

the absence of the source remediation conducted, receptor impacts were not expected. Accordingly, the

NRCP proposal was submitted, requesting approval to

implement the monitoring program outlined therein.

This program could save the property owner significant

remediation costs. Given NJDEP’s rigorous minimum

requirements for NRCP implementation, the costs to

demonstrate applicability of the NRCP should be compared with costs for active ground-water plume management. This cost evaluation will allow a property owner to

make informed decisions regarding remedial options

and cash flow management.

Reference

1. New Jersey Department of Environmental Protection. 1996. Site

remediation news. March.

169
Field and Laboratory Evaluations of Natural Attenuation

of Chlorinated Organics at a Complex Industrial Site

M. Alexandra De, Julia Klens, Gary Gaillot, and Duane Graves

IT Corporation, Knoxville, Tennessee

Natural attenuation of tetrachloroethene (PCE), trichloroethene (TCE), carbon tetrachloride, and hexachlorobenzene (HCB) is under investigation at a large

industrial site. The site has a number of complexities

due to past manufacturing activities, topography, hydrology, and the presence of several surface water bodies

that are connected to shallow ground water. Perched

ground water, shallow and deep aquifers, creeks, rivers,

and a manufactured impoundment all contribute to site

hydrology and affect ground-water flow direction and

velocity. Regions of high ground-water pH (pH 10 to 14)

are found beneath settling ponds that contain high pH

waste liquors from past manufacturing processes.

Ground-water microbes have been shown to be inactive

when the pH exceeds 9.5. The aquifer apparently has

significant buffering capacity to neutralize the ground

water as it migrates away from the impoundments.

Contaminant concentration varies across the site,

with dense nonaqueous-phase liquid contributing a

high concentration of dissolved contaminants in a few

locations. Contaminant concentration changes and the

occurrence of anaerobic biodegradation products of

PCE and TCE support the conclusion that intrinsic

biodegradation is occurring. Attenuation rates for PCE,

TCE, and cis-1,2-dichloroethene indicate a half-life of

approximately 300 days for each. Preliminary evidence

of intrinsic biodegradation has also been derived from

ground-water geochemistry data. Increased concentrations of iron(II) and dissolved manganese correspond

with neutral ground-water pH, chemicals of concern,

and biodegradation products. Nitrate is found in very

low concentrations in the general area, and this respiratory substrate does not significantly contribute to

biodegradation. Ground-water alkalinity is affected by

site activities and pH, which mask changes in alkalinity

due to intrinsic biodegradation. Dissolved methane

concentrations, oxidation reduction potential, sulfate

concentrations, and dissolved oxygen in the ground

water are currently being examined.

This poster discusses the intrinsic remediation of PCE

and TCE with respect to contaminant concentrations

and ground-water geochemistry. Laboratory studies

conducted to evaluate the rate of biodegradation are

also described.

170
Assessment of Intrinsic Bioremediation of Chlorinated Aliphatic Hydrocarbons 

at Industrial Facilities

Marleen A. Troy and C. Michael Swindoll

DuPont Environmental Remediation Services, Wilmington, Delaware

Intrinsic bioremediation of chlorinated aliphatics at several industrial sites was evaluated to determine its significance and whether it could be used as a corrective

action alternative for reducing potential environmental

impact. The premise behind the implementation of an

intrinsic bioremediation approach was that naturally occurring microorganisms present in subsurface environments of each site were capable of degrading the

contaminants of interest and that the contaminant concentrations would be degraded to acceptable levels.

A variety of chlorinated aliphatic hydrocarbons were

detected in ground water at the sites, including tetrachloroethene (PCE), trichloroethene (TCE), dichloroethene

(DCE), vinyl chloride (VC), trichloroethane (TCA), and

methylene chloride. Concentrations of individual chlorinated aliphatics typically ranged from nondetect to less

than 500 micrograms per liter (mg/L), with the highest

concentrations in the 1,000 to 3,000 mg/L range.

Ground-water data from each site were examined for

indicators of intrinsic bioremediation and the existence

of conditions favorable for bioremediation. Indicators of

intrinsic bioremediation included changes in contaminant concentrations, detection of biodegradation metabolites, and changes in geochemical measurements.

Indicators of conditions favorable for bioremediation that

were evaluated included pH, oxidation-reduction potential, concentrations of electron acceptors, nutrients, primary substrates sufficient to support microbial activity,

and the lack of inhibitory concentrations of toxicants.

The data from each site were collectively evaluated

through a “weight-of-evidence” approach to determine

whether intrinsic bioremediation was a viable remedial

alternative for each site. Based on these evaluations, it

was concluded that intrinsic bioremediation was occurring under anaerobic conditions at each site, nonchlorinated co-contaminants served as primary substrates,

and microbial activity was limited by nutrient availability.

Although the data indicated that intrinsic bioremediation

was occurring, the existing data were insufficient to

support intrinsic bioremediation as the sole remedial

alternative. Ground-water monitoring for indicator parameters continued to allow further evaluation of the

potential application of intrinsic bioremediation as a remedial alternative for the sites.

171
Natural Attenuation as Remedial Action: A Case Study

Andrea Putscher

Camp Dresser & McKee, Woodbury, New York

Betty Martinovich

Polytechnic University, Farmingdale, New York, and Camp Dresser & McKee,

Woodbury, New York

The subject site is an informative case study of factors

leading to a decision by state regulators to acknowledge

natural attenuation as the principal action to remediate

trichloroethene (TCE), cis-1,2-dichloroethene (cis-1,2-

DCE), and vinyl chloride.

A team of hydrogeologists and engineers, under contract

with the New York State Department of Environmental Conservation (NYSDEC), completed a remedial investigation

and feasibility study for a site in Rockland County, New

York. The site is in the glaciated northeast, with a 100-

foot thick glacial till underlying the site. The till overlies

Brunswick (Passaic) formation fractured silty sandstone

bedrock, which comprises the principal aquifer system

in the site vicinity. Heterogeneity in the till and fractured

rock ground-water hydraulics have resulted in a complex array of potential contaminant migration pathways.

A lighting fixture manufacturing operation discharged an

unknown volume of liquid waste containing TCE-dominated mixed volatile organic compounds (VOCs), in

concentrations ranging from 1 to 1,000 parts per million

total VOC, into a shallow, ephemeral stream/drainage

ditch on site for an unknown period, ending in 1980. The

remedial investigation was initiated in 1994, 14 years

after the discharge was eliminated, and implemented in

two phases over a 1.5-year period. The timing and duration of the investigation facilitated identification and

characterization of natural degradation and attenuation

of the chlorinated constituents (TCE, cis-1,2-DCE, and

vinyl chloride) in the site subsurface.

Project personnel used conventional techniques, including soil gas survey and stream sediment, soil, surface

water, and ground-water sampling and analysis, during

the initial Phase I remedial investigation. The Phase I

Remedial Investigation and Phase I and II Feasibility

Study lasted 1 year. The investigation and study results

suggested that concentrations of chlorinated constituents were naturally attenuating to levels below NYSDEC

established cleanup standards (in the parts per billion

range). Furthermore, the rates of natural attenuation

appeared to be sufficient to preclude offsite migration via

most of the potential pathways.

Due to the indications that natural attenuation was functioning on site, project personnel designed and implemented a focused Phase II remedial investigation that,

in part, addressed natural attenuation related issues.

The latter phase of the remedial investigation was implemented over a period of 6 months and included

modified techniques and strategies for stream sediment,

soil, surface water, and ground-water sampling and

analysis, in addition to a treatability study at the field and

laboratory scale. The remedial investigation/feasibility

study RI/FS was completed in February 1996. The record of decision was signed in March 1996, and the

selected remedy allows for limited (near-surface hotspot removal) soils remedial action and continued

ground-water monitoring to demonstrate the efficacy of

natural attenuation in the subsurface as the principal

ground-water remedial action for the site. 

172
Patterns of Natural Attenuation of Chlorinated Aliphatic Hydrocarbons

at Cape Canaveral Air Station, Florida

Matt Swanson, Todd H. Wiedemeier, and David E. Moutoux

Parsons Engineering Science, Inc., Denver, Colorado

Donald H. Kampbell

U.S. Environmental Protection Agency, National Risk Management Research Laboratory,

Subsurface Protection and Remediation Division, Ada, Oklahoma

Jerry E. Hansen

U.S. Air Force Center for Environmental Excellence, Technology Transfer Division,

Brooks Air Force Base, Texas

Activities at a former fire training area (Site CCFTA-2

[FT-17]) at Cape Canaveral Air Station in Florida resulted in contamination of shallow soils and ground

water with a mixture of chlorinated aliphatic hydrocarbons (CAHs) and fuel hydrocarbons. The dissolved

contaminant plume, beneath and at least 1,200 feet

downgradient from a body of mobile, light nonaqueous

phase liquid (LNAPL) containing commingled petroleum and chlorinated solvents, consists of commingled benzene, toluene, ethylbenzene, and xylenes

(BTEX) and CAHs. Before construction of a horizontal

air sparging system, contaminated ground water discharged to surface water in a canal downgradient of

the source area. The desire for a long-term approach

to address the dissolved contaminant mass prompted

an assessment of the potential for natural attenuation

mechanisms to reduce the mass, toxicity, and mobility

of trichloroethene (TCE), dichloroethene (DCE), vinyl

chloride (VC), and BTEX dissolved in ground water at

CCFTA-2 (FT-17).

Several lines of chemical and geochemical evidence

indicate that dissolved CAHs at the site are undergoing

reductive dehalogenation, facilitated by microbial oxidation

of BTEX compounds and native organic matter. Data on

the distributions of TCE, cis-1,2-DCE, VC, and ethene

indicate that TCE dissolved from the LNAPL body is

being sequentially dehalogenated, with VC accumulating near the terminus of the CAH plume. While the

ground-water system outside of the plume is nearly

anaerobic due to microbial degradation of native organic

matter, petroleum hydrocarbons released at the site

have fostered additional microbial activity and created

conditions that favor reductive dehalogenation of CAHs.

Distribution of electron acceptors and metabolic byproducts, along with dissolved hydrogen concentrations,

indicate that biodegradation mechanisms operating at

the site include aerobic respiration, iron reduction, sulfate reduction, and methanogenesis.

Approximation of field-scale biodegradation rates at the

site suggests that TCE and cis-1,2-DCE have a half-life

of approximately 2.4 to 3.2 years. Because reducing

conditions persist from the source area to the canal, VC

has accumulated and therefore affected surface water.

Air sparging near the canal, however, will serve to both

physically remove dissolved contaminants and foster

more rapid (aerobic) biodegradation of VC. 

173
Applying Natural Attenuation of Chlorinated Organics

in Conjunction With Ground-Water Extraction for Aquifer Restoration

W. Lance Turley and Andrew Rawnsley

Hull & Associates Engineering, Inc., Austin, Texas

Natural attenuation of dissolved chlorinated organics

(primarily tetrachloroethene, trichloroethene, and 1,1,1-

trichloroethane) via dilution is being successfully employed near the end of a 2,200-foot long plume at the

South Municipal Water Supply Well Superfund site in

Peterborough, New Hampshire. The U.S. Environmental Protection Agency’s (EPA’s) record of decision

required that the entire plume be remediated through

pumping a network of extraction wells. Installation and

operation of an extraction well near the end of the plume

was not practical, however, because of property access

difficulties, the presence of a flood plain, and anticipated

problems in conveying extracted water via a forcemain

due to expected low flows and a large head differential.

Field measurements were supported by finite-difference,

three-dimensional flow modeling and indicated that the

aquifer at the end of the plume discharges into the Contoocook River. Furthermore, modeling indicated that discharge would occur, although at a lower rate, when the

aquifer was pumped in upgradient portions of the plume.

Modeled flux through the end of the plume was

compared with projected removal rates by an extraction

well and was found to be similar. Concentrations

of water discharging into the river were conservatively

estimated based on the highest concentration detected

in a monitoring well within the proposed attenuation

zone. Dilution factors were calculated based on the

flux of contaminated water from the aquifer versus

the river’s 7-day low flow over a period of 10 years

(7Q10). Dilution factors were applied to discharge concentrations, and results were compared with healthbased water quality criteria (for water and fish ingestion)

and found to be acceptable. Finally, a flushing model

was used to determine that the attenuation zone would

be reduced to cleanup levels within the time frame

stated in the record of decision. EPA accepted the

technical arguments for integrating natural attenuation

into the ground-water remediation system, and the record of decision was modified accordingly through issuance of an explanation of significant difference.

174
A Modular Computer Model for Simulating Natural Attenuation of

Chlorinated Organics in Saturated Ground-Water Aquifers

Yunwei Sun and James N. Petersen

Chemical Engineering Department, Washington State University, Pullman, Washington

T. Prabhakar Clement and Brian S. Hooker

Pacific Northwest National Laboratory, Richland, Washington

Although several field-scale natural attenuation projects

have already been considered for managing benzene,

toluene, ethylbenzene, and xylene (BTEX) plumes, a

rational basis for implementing natural attenuation technology has yet to be formulated for chlorinated solvent

plumes. Successful validation of natural attenuation involves significant upfront and followup field characterization to ensure that intrinsic processes are indeed

destroying contaminants of concern at reasonable rates.

Given the extensive amount of data collected during this

type of effort, computer-aided design and data analysis

tools are needed. These tools must facilitate the interpretation of these data so that the design engineer can

determine whether intrinsic remediation can achieve the

cleanup objectives and assess the risks associated with

the action. Computer models are also useful for forecasting the influence of natural attenuation processes

over long periods.

To adequately analyze natural attenuation processes,

models should also consider simultaneous multispecies

transport and bio- and geochemical interactions. This

poster describes a newly developed computational tool,

designated RT3D (Reactive Transport in Three Dimensions). This tool can simulate natural attenuation of

various subsurface contaminants and their decay products in saturated ground-water aquifers.

RT3D was developed from the U.S. Environmental Protection Agency’s public domain computer code MT3D.

The MT3D model simulates single-species transport

with or without sorption and first-order reaction. Contaminant transport velocities are calculated from the

head distribution computed by the U.S. Geological Survey’s model MODFLOW. 

We have extended MT3D to describe multispecies

transport and reactions. The present version of RT3D

can simulate three-dimensional transport of multiple

aqueous-phase species and the fate of multiple solidphase species, along with the physical, chemical, and

biological interactions among them. The code is organized in a modular fashion to ensure flexibility. The reactive portion of the code is a separate module using an

operator-split strategy; hence, any type of reaction kinetics can be accommodated through an appropriate reaction module. The present version has four separate

reaction modules: aerobic, instantaneous BTEX reactions (similar to BIOPLUME II); multiple-electron acceptor, kinetic-limited BTEX reactions (similar to

BIOPLUME III); denitrification-based carbon tetrachloride transformation reactions; and chlorinated ethene

reactions.

This poster describes the numerical details of the RT3D

code and the chlorinated ethene reaction module. An

example problem is solved to illustrate the potential use

of this code for planning natural attenuation of chlorinated organics in saturated ground-water aquifers.

175
State and Federal Regulatory Issues
Federal and State Meeting on Issues Impacting the Use of Natural Attenuation for

Chlorinated Solvents in Ground Water—An Overview

Introduction

On September 10, 1996, a meeting took place in Dallas,

Texas, to discuss state and federal regulatory issues

affecting natural attenuation of chlorinated solvents.

This meeting was held in conjunction with EPA’s Symposium on Natural Attenuation of Chlorinated Organics

in Ground Water. The major topics of discussion were:

• Site characterization

• Monitoring

• Degradation rates

• Risk management and closure issues

• Technical impracticability (TI) waivers

Fran Kremer and John Wilson of EPA’s Office of Research and Development chaired the meeting; an overview of the topics discussed at the meeting is presented

below.

Site Characterization

The discussion of site characterization addressed four

major issues:

• The identification of important indicator parameters

for natural attenuation, and the development of

standard operating procedures, quality assurance/quality control plans, and data quality objectives for the measurement of those parameters.

• An overall protocol for assessing the redox chemistry

of the subsurface environment.

• The reliability of rapid field estimation techniques.

• The need for methods to characterize nonaqueous

phase liquids (NAPLs).

The redox chemistry of the subsurface environment

plays an important role in governing the rate and extent

of natural attenuation. A representative of the U.S. Geological Survey (USGS) stated that the 10 parameters

currently measured by the Survey as indicators of natural attenuation are dissolved oxygen, nitrate, nitrite, ferrous sulfate, sulfite, methane, chloride, carbon dioxide,

and hydrogen. The need for an overall protocol for assessing the redox chemistry was recognized. It was

suggested that USGS and EPA collaborate in developing such a protocol.

The discussion emphasized the need for rapid and reliable field techniques. Participants debated the pros and

cons of field and laboratory analytical methods, and

discussed, for example, the need to consider the tradeoffs between sample integrity (collection, preservation,

and laboratory analysis) and the relative accuracy of

simple Hach kits. Also important is the need for trained

personnel to perform onsite analytical measurements.

The importance of identifying the sources of contamination due to NAPLs was stressed. Participants urged

consideration of push-technologies such as cone penetrometer because they can provide quick and dependable estimates of parameters. Further work needs to be

undertaken to develop improved methods to characterize subsurface NAPLs.

Monitoring

Natural attenuation is not a “walk-away” remedy. In fact,

it requires a rigorous and long-term monitoring strategy

to indicate that natural processes are effective in reducing contaminants to acceptable levels before potential

receptors are reached. Since improper monitoring may

lead to misconceptions about the extent and/or the types

of biotic and abiotic processes occurring at a contaminated site, a thorough remediation strategy should include a complete site characterization investigation

followed by performance monitoring. The need to select

proper monitoring locations and sampling frequencies to

indicate spatial and temporal variability of a site was a

recurring theme in the discussion. Developing a conceptual model is essential for the interpretation of data

collected during the monitoring process.

The discussion reemphasized that compared to other

remedial efforts, natural attenuation requires a more

intensive hydrogeologic characterization. During the implementation of other proactive remedial technologies,

such as pump and treat, the natural gradients of groundwater flow are significantly altered. Since natural attenuation is a passive remedy, however, understanding the

existing flow conditions is very important.

179
In addition to establishing points of compliance to

achieve a primary goal of monitoring, it is important to

show that a plume will not migrate beyond accepted

boundaries. Achieving these goals entails performing a

detailed geochemical study prior to the selection of natural attenuation as a remedial alternative and conducting

periodic sampling to assure that remediation is occurring. The duration and frequency of sampling is dictated

by the velocity of flow, location of monitoring wells, and

various site-specific factors. Monitoring should occur

more frequently initially and then be reduced when remediation rates, pattern of migration, and seasonal influences are established.

Participants discussed the need for validation of data

from conventional wells and mass balance calculations,

as well as the role and importance of numerical models

versus field data. The extent to which numerical modeling could replace field data and assist regulators in

predicating degradation of chlorinated compounds

and/or behavior of a plume was the topic of interest.

Participants expressed a need for standard operating

procedures for analyzing different classes of compounds, in particular dissolved gases. They also discussed the importance and role of hydrogen levels in the

hierarchical scheme of terminal electron acceptor processes. In general, the discussion indicated a need for

information about potential parameters for evaluating

biodegradation of chlorinated solvents in ground water.

Degradation Rates

Estimating the rate of natural attenuation requires a

good understanding of both degradation rates and hydrogeology. Participants expressed concerns about the

evaluation of natural attenuation in the presence of a

mixture of contaminants, both organic and inorganic.

Although the degradation of dissolved petroleum hydrocarbons has been observed to follow first-order kinetics,

it is unclear whether the degradation of chlorinated hydrocarbons is generally a first-order process. Toxicity,

competitive inhibition, and nutrient availability are

among the factors that can cause the kinetics of the

microbial degradation process to significantly deviate

from the first-order model.

Participants discussed three methods for estimating the

first-order degradation rate constant (k): conservative

tracer study, microcosm study, and transport modeling.

Tracer studies are not commonly performed because

they are time consuming and the results have a great

deal of uncertainty associated with them. Laboratory

microcosm studies are more commonly used to estimate k, but they too are prone to uncertainty because

the laboratory setup cannot truly mimic field conditions.

If field data cannot be obtained to determine degradation

rates, microcosm studies have to be relied upon for the

estimation of the rates. A properly calibrated flow and

transport model can be a valuable tool for the estimation

of first-order rate constants. Participants suggested that

an extensive catalogue of information on rate constants

according to contamination scenarios and ground-water

geochemistry would be very useful. This would facilitate

the estimation of k for a given scenario by simply extrapolating k values from similar, but known, conditions.

Risk Management and Closure Criteria

This segment of the meeting focused on guidelines for

the types of information regulators need to perform risk

assessment and management at sites using natural

attenuation of chlorinated solvents. Regulators stressed

the immediate need for better methodologies for assessing environmental risks and determining action levels. Better methodologies are needed to ensure that

institutional controls can be implemented effectively.

The current cleanup guidelines are based only on concentration and not on real site-specific health and ecological risks of available contaminants. Although the

National Contingency Plan is explicit about achieving

maximum contaminant levels (MCLs) within a plume, it

also allows for alternate concentration limits (ACLs).

The ACLs, however, are seldom used because the associated caveats restrict their use.

Participants concluded that there is a need to better

describe and quantify the interactions between surface

water and ground water. It is important not only to assess the risk of ground-water discharge into surface

water, but also to consider the contamination of ground

water by surface water infiltration. Although the presence of microorganisms as health hazards does not

require monitoring under Superfund regulations, the issue is handled very differently in the proposed groundwater disinfection rule. It is clear that the implications of

the different ground-water rules need to be clarified.

Another concern repeatedly expressed by participants

was the current data gap in determining whether biodegradation is actually taking place. There is a need for a

better data base on biodegradation rates to improve

judgments on whether apparent contaminant reduction

results from retardation or degradation. Risk management and assessment are more difficult for sites contaminated with chlorinated contaminants than for those

contaminated with petroleum products. This is because

of a lack of data bases constructed from case studies

documenting both successes and failures of natural attenuation under field conditions.

Other important factors to consider in the selection of

natural attenuation as a remediation option are cost

effectiveness and cleanup time. Decision-makers and

regulators need to be certain that natural attenuation is

achievable within a “reasonable time frame.” In some

cases, natural attenuation may be more cost effective

180
than other alternatives, while in other cases, active remediation may be more cost effective.

Clear, simple information about the appropriate situations for selecting natural attenuation over other remedial alternatives should be made available. Participants

suggested developing protocols to indicate the parameters to be selected and the analytical methodologies to

be used. They also stressed the critical need for decision support systems to assist site coordinators in each

step of the process. Specifically, they requested training

for the federal and state decision-makers in evaluating

environmental risks.

There is a need to convince both regulators and the

public that natural attenuation is not a “walk-away” remedy. Some participants suggested that the selection of

natural attenuation may be more appropriate at sites

where contamination is relatively low, the risk to drinking

water is negligible, and sources of contamination have

been removed.

The final critical issue discussed in this segment dealt

with closure criteria. Participants considered it paramount to have information that can assist in establishing

closure criteria and in determining when a site can be

considered clean. They questioned the adequacy of the

criteria in current use. It would be of enormous value to

generate information about toxicological testing, presented in clear, simple terms, that addresses specifics

on toxicity of the daughter products and by-products of

the degradation process.

Technical Impracticability (TI) Waiver

Participants were concerned that there is considerable

confusion regarding natural attenuation and TI waivers

as options for remediation of ground water. Both are

erroneously regarded as a “do-nothing” approach and

both typically rely on monitoring and institutional controls to ensure continued protection of public health and

the environment. A TI waiver is used when it is not

technically feasible to return contaminated ground water

to relevant and appropriate cleanup levels (e.g., federal

or state drinking water standards). When attainment of

required cleanup levels is feasible, natural attenuation

can be selected as the cleanup method, as a alternative

to pump and treat or other methods, if natural processes

can achieve the cleanup goals in a time frame that is

reasonable compared to other alternatives. Thus, a TI

waiver is not needed in order to select natural attenuation as the appropriate remedy because natural processes are expected to attain the required level of

cleanup in a reasonable time frame. At some sites it will

be appropriate to use a TI waiver over a portion of the

plume where highly contaminated ground water cannot

be cleaned up but will be contained, and to clean up the

remainder of the plume using natural attenuation or

other methods.

Monitoring of remedy performance is a critical component of all ground-water remedies, to ensure that remediation remedies are being met and that there are no

adverse impacts to public health or the environment.

The monitoring data collected to ensure that a natural

attenuation remedy is progressing as expected may be

different than that for other types of remedies. As discussed above, geochemical indicators, degradation byproducts, and other parameters may need to be

measured in addition to the contaminants of concern to

determine if biotic and/or abiotic processes are continuing as expected.

Summary

The meeting identified the following needs as being of

paramount importance:

• Standard operating procedures for the indicator parameters of natural attenuation and protocols for the

demonstration of natural attenuation of chlorinated

organics.

• Rapid and reliable field techniques for characterizing

the sources of contamination.

• Improved methods for the estimation of field degradation rate constants.

• Development of guidance for the use of natural attenuation.

• Training for federal and state decision-makers in

evaluating data supporting the use of natural attenuation.

181
Summary of Roundtable Discussion on Regulatory Issues

Panelists

• Michael Barden, Hydrogeologist, Wisconsin Department of Natural Resources

• Mark Ferrey, Research Scientist, Minnesota Pollution Control Agency

• Ruth Izraeli, EPA Hydrogeologist, Co-Chair of Superfund/RCRA Ground-water Forum

• Tim Larson, Professional Engineer, Florida Department of Environmental Protection

• Kenneth Lovelace, Environmental Engineer, Office

of Emergency and Remedial Response, Superfund

Division, EPA

• Kay Wischkaemper, EPA Hydrogeologist, Co-Chair

of Superfund/RCRA Ground-water Forum

Summary

A roundtable discussion on how natural attenuation fits

into state and federal regulatory frameworks concluded

the symposium. Panelists from state and federal regulatory agencies answered questions from the audience.

The following points were made during the discussion:

• Public acceptance of natural attenuation. Public acceptance of natural attenuation as a tool for risk management and site closure is as critical as regulatory

acceptance. It is important to stress that natural attenuation is one technical approach that often is used

in combination with other technical approaches such

as source control, hydraulic containment, and pump

and treat. In other words, stakeholders need to convey to the public that natural attenuation is an engineered application of a technology, not a “do-nothing”

approach. Another public misconception is that contaminated sites can be cleaned up in several years

if enough money is allocated. It is therefore important

to emphasize that complete cleanup of a site over a

short time period is often not the most cost-effective

strategy and may not always be possible.

One person in the audience asked panelists how to

convince regulators and the public that natural attenuation is a viable alternative for a large, chlorinated

plume in the Phoenix area. Panelists responded that

key issues in deciding between various alternatives

and presenting a decision to the public include comprehensive site characterization, integration of other

remediation tools, ground-water demand, and future

plans for aquifer use.

• Regulatory flexibility with a technical impracticability

(TI) waiver. TI waivers can waive ARAR-based

cleanup levels for some ground-water plumes, such

as in those areas with a fractured rock substrate,

where the ARAR levels are impossible to achieve. TI

waivers can also be specific to the source zone of a

plume, with ARAR-based cleanup levels applying to

outer parts of the plume.

• Alternative Concentration Limits (ACLs). ACLs are a

separate regulatory tool that has not been used frequently. ACLs might be appropriate in a situation

where an aquifer under the land between a site and

a river is discharging contaminants into the surface

water of that river. If there is a direct point of discharge and if people are not and will not be using

the aquifer for drinking water as guaranteed by institutional controls, ACLs that are protective of the surface water could be set for the ground-water plume.

In many places, however, the public will not accept

long-term or permanent restrictions on aquifer use.

• Site closure via natural attenuation. Natural attenuation is an ongoing remedial strategy, the use of

which does not necessarily lead to site closure. The

real issue involved in site closure decisions at natural

attenuation sites is how long the site should be monitored before ultimate closure. Trend analysis for

plumes of chlorinated solvents is more difficult than

for hydrocarbon plumes, and site closure decisions

are accordingly more difficult.

• Ground-water classification at the state and federal

level. EPA has various classifications for ground

water in place, but states are encouraged to develop

their own classification schemes to reflect state priorities for ground-water protection. For example, a

state might declare a certain area as a kind of nonattainment zone if that area is subject to contamination

from a number of sources and potential for cleanup

is low. The federal government is working to develop

a tool for states to apply consistently in such situations.

183
One person in the audience asked how to classify

aquifers in the arid west that have high alkalinity, high

total dissolved solids, and high concentrations of

heavy metals and how to classify surface water bodies that have extremely high pH levels. EPA has

ground-water and surface water classification systems, and these aquifers and surface water bodies

might fall into a lower-priority class, depending on the

concentrations involved and the quantity of water (for

the aquifers) and on ecological sensitivity (for the

surface water bodies). States, however, may also

develop their own classification schemes for ground

water and surface water.

• Natural attenuation at sites that do not fit the classic

“sandbox” model. Natural attenuation is an option for

sites that do not fit the classic “sandbox” model, such

as sites with a high soil clay content and low rates

of ground-water flow. There are benefits and drawbacks to applying natural attenuation at these sites.

The largest drawback is often that natural attenuation

may take a long time to achieve cleanup goals. As

with any site, site characterization is the most important step in determining the benefits of natural attenuation.

• Cost-benefit analyses for restoring water resources.

Cost-benefit analysis can be an important technique

for setting priorities for cleanup of contaminated

sites. Economists and policy-makers have not yet

agreed on a model for conducting cost-benefit analysis, although a number of economists are developing

such models. One major difficulty is that it is difficult

to place a value on water resources.

• Natural attenuation and the “land ban.” If natural attenuation is occurring in ground water, the ban on

land disposal of untreated hazardous waste does not

apply.

• Natural attenuation and natural resources damage

claims. States will probably not pursue natural resource damage claims under CERCLA on groundwater issues as often as on soil or surface water

issues. Sites at which states might consider pursuing

such a claim include aquifers used by municipalities

that become so contaminated that they are no longer

available for use.

• Natural attenuation and deed restrictions. When

deed restrictions are required for controls on future

ground-water use, states are generally concerned

only with the end result and let the involved parties

negotiate these restrictions according to whatever

process they choose.

• Permeable walls as an alternative remediation strategy. Researchers and regulators in a number of

states are conducting pilot tests on permeable walls.

Permeable walls appear to be a promising tool for

remediation of sites contaminated with chlorinated

organic solvents.

184
Appendix A
187

Symposium on Natural Attenuation of

Chlorinated Organics in Ground Water

Hyatt Regency Dallas

Dallas, TX

September 11-13, 1996

Final Agenda

Wednesday, September 11

Session Chair: Fran Kremer, U.S. EPA

BACKGROUND/PROCESS DESCRIPTION

8:00AM Introductory Remarks ........................................ Patricia Rivers

Office of the Deputy Undersecretary of Defense,

(Environmental Cleanup),

Washington, DC

8:20AM Welcome and Overview Remarks ................................. Fran Kremer

Office of Research and Development (ORD),

U.S. Environmental Protection Agency (U.S. EPA), Cincinnati, OH

Catherine Vogel, U.S. Air Force (USAF), Tyndall AFB, FL

8:50AM Introductory Talk: Where Are We Now?

Moving to a Risk-Based Approach ............................... C. Herb Ward

Rice University, Houston, TX

9:15AM Introductory Talk: Where Are We Now With Public and 

Regulatory Acceptance? (Resource Conservation and Recovery 

Act [RCRA] and Comprehensive Environmental Response, 

Compensation, and Liability Act [CERCLA]) .................... Kenneth Lovelace

Office of Emergency and Remedial Response,

Superfund Division, U.S. EPA, Washington, DC

9:45AM Biotic and Abiotic Transformations of 

Chlorinated Solvents in Ground Water ........................... Perry McCarty

Stanford University, Stanford, CA

10:15AM BREAK
188

Wednesday, September 11 (cont'd)

10:45AM Microbiological Aspects Relevant to Natural Attenuation

of Chlorinated Ethenes ....................................... James Gossett

Cornell University, Ithaca, NY

11:00AM Microbiological Aspects Relevant to Natural Attenuation

of Chlorinated Ethenes (cont'd) ................................... Steve Zinder

Cornell University, Ithaca, NY

11:15AM Microbial Ecology of Adaptation and Response 

in the Subsurface .............................................. Guy Sewell 

National Risk Management Research Laboratory (NRMRL),

U.S. EPA, Ada, OK 

11:45AM Identifying Redox Conditions That Favor the Natural 

Attenuation of Chlorinated Ethenes in Contaminated

Ground-Water Systems ..................................... Francis Chapelle

U.S. Geological Survey (USGS), Columbia, SC

12:15PM LUNCH

Session Chair: Catherine Vogel, U.S. Air Force

1:30PM Design and Interpretation of Microcosm Studies for 

Chlorinated Compounds ..................................... Barbara Wilson

NRMRL, U.S. EPA, Ada, OK

2:00PM Conceptual Models for Chlorinated Solvent Plumes and Their 

Relevance to Intrinsic Remediation ................................ John Cherry

University of Waterloo, Waterloo, Ontario, Canada

2:30PM Site Characterization Tools: Using a Borehole

Flowmeter To Locate and Characterize the 

Transmissive Zones of an Aquifer ................................... Fred Molz

Clemson University, Anderson, SC

3:00PM BREAK

3:30PM Evaluation of Natural Attenuation: Current Practices and 

Opportunities for Ground Water ................................ Joseph Salvo

General Electric,

Corporate Research & Development, Schenectady, NY

4:00PM Overview of the Technical Protocol for Natural 

Attenuation of Chlorinated Aliphatic Hydrocarbons in Ground 

Water Under Development for the U.S. Air Force 

Center for Environmental Excellence .............................. John Wilson

NRMRL, U.S. EPA, Ada, OK
189

Wednesday, September 11 (cont'd)

4:30PM The BIOSCREEN Computer Tool .............................. Charles Newell

Groundwater Services, Inc., Houston, TX

5:00PM ADJOURN

5:30PM to 7:30PM POSTER SESSION AND CASH BAR

Posters will be on display of projects involving natural attenuation of chlorinated organics and other

xenobiotic compounds. Authors will be present to discuss their projects and answer questions during

the Poster Session. Posters will remain on display throughout the symposium.

Thursday, September 12

Session Chair: John Wilson, U.S. EPA

CASE STUDIES

8:00AM Introductory Remarks ......................................... Patrick Haas

AFCEE, USAF, Brooks AFB, TX

8:15AM Case Study: Naval Air Station, Cecil Field, Florida ................. Francis Chapelle

USGS, Columbia, SC

9:00AM Case Study of Natural Attenuation of Trichloroethene at 

St. Joseph, Michigan .......................................... James Weaver

NRMRL, U.S. EPA, Ada, OK

9:45AM Natural Attenuation of Chlorinated Aliphatic Hydrocarbons at

Plattsburgh Air Force Base, New York ........................ Todd Wiedemeier

Parsons Engineering Science, Inc., Denver, CO

10:30AM BREAK

11:00AM Case Study: Natural Attenuation of a Trichloroethene Plume at 

Picatinny Arsenal, New Jersey ............................. Thomas Imbrigiotta

USGS, West Trenton, NJ

11:45AM Case Study: Plant 44, Tucson, Arizona ............................. Hanadi Rifai

Rice University, Houston, TX

12:30PM LUNCH
190

Thursday, September 12 (cont'd)

Session Chair: Patrick Haas, U.S. Air Force

1:45PM Remediation Technology Development Forum, Intrinsic Remediation 

Project at Dover Air Force Base, Delaware .......................... David Ellis

DuPont Specialty Chemicals-CRG,Wilmington, DE;

Joe Salvo, General Electric, Schenectady, NY;

and Gary Klecka, Dow Chemical Company, Midland, MI

2:30PM Case Study: Wurtsmith Air Force Base, Michigan ............... Michael Barcelona

University of Michigan, Ann Arbor, MI

3:15PM BREAK

3:45PM Case Study: Eielson Air Force Base, Alaska ...................... R. Ryan Dupont

Utah State University, Logan, UT

4:30PM Considerations and Options for Regulatory Acceptance of 

Natural Attenuation in Ground Water ...................... Mary Jane Nearman

U.S. EPA, Seattle, WA

4:45PM Lessons Learned: Risk-Based Corrective Action .................. Matthew Small

Office of Underground Storage Tanks (UST),

U.S. EPA, San Francisco, CA

5:00PM ADJOURN

7:00PM - 9:00PM EVENING DISCUSSION 

Informal Dialog on Issues of Ground Water 

and Core Sampling ........................................ Donald Kampbell

NRMRL, U.S. EPA, Ada, OK

Friday, September 13

Session Chair: Marty Faile, U.S. Air Force

8:00AM Introductory Remarks: Appropriate Opportunities for 

Application—Civilian Sector (RCRA and CERCLA) ................... Fran Kremer

ORD, U.S. EPA, Cincinnati, OH

8:20AM Introductory Remarks: Appropriate Opportunities for 

Application—U.S. Air Force and Department of Defense ............. Patrick Haas

AFCEE, USAF, Brooks Air Force Base, TX

8:40AM Intrinsic Remediation in the Industrial Marketplace ..................... David Ellis

DuPont Specialty Chemicals-CRG, Wilmington, DE
191

Friday, September 13 (cont'd)

9:00AM Environmental Chemistry and the Kinetics of Biotransformation of 

Chlorinated Organic Compounds in Ground Water .................. John Wilson

NRMRL, U.S. EPA, Ada, OK

9:20AM Future Vision: Compounds With Potential for 

Natural Attenuation .............................................. Jim Spain

USAF, Tyndall AFB, FL

9:40AM Natural Attenuation of Chlorinated Compounds in Matrices 

Other Than Ground Water:

The Future of Natural Attenuation ............................. Robert Hinchee

Parsons Engineering Science, Inc., South Jordan, UT

10:00AM BREAK

10:30AM Report From the Regulatory Work Session .................... Kenneth Lovelace

Office of Emergency and Remedial Response, U.S. EPA, Washington, DC

Michael Barden

Wisconsin Department of Natural Resources, Madison, WI

11:10AM Roundtable Discussion ....................................... Panel of Speakers

12:15PM ADJOURN