£EPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington, DC 20460
EPA/540/R-97/504
May 1997
Proceedings of the
Symposium on
Natural Attenuation of
Chlorinated Organics in
Ground Water
Printed on paper that contains at
least 20 percent postconsumer fiber.
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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
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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.
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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
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Contents (continued)
Page
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 Pittsburgh 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
IV
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Contents (continued)
Page
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(lll)-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. Lige 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
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Contents (continued)
Page
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 Marches!, 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
VI
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Contents (continued)
Page
Assessment of Intrinsic Bioremediation of Chlorinated Aliphatic Hydrocarbons at Industrial Facilities
Marleen A. Troy and C. Michael Swindell 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
VII
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A cknowledgments
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 sympo-
sium was a joint effort of the U.S. Environmental Protection Agency's (EPAs) Biosystems Technol-
ogy 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 EPAs 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
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Executive Summary
The U.S. Environmental Protection Agency (EPA), the
U.S. Air Force Armstrong Laboratory's Environics Direc-
torate, 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 EPAs Office of Research and Devel-
opment (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 natu-
ral 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 Undersec-
retary 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 dis-
posal of remediators.
Herb Ward of Rice University followed with a talk entitled
"Where Are We Now? Moving to a Risk-Based Ap-
proach." 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 useful-
ness of which depends on site-specific characteristics.
Kenneth Lovelace of the Superfund Division of EPAs
Office of Emergency and Remedial Response con-
cluded the introductory talks with a presentation entitled
"Where Are We Now With Public and Regulatory Accep-
tance?" He emphasized EPAs support for natural at-
tenuation 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 reliabil-
ity 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 sympo-
sium highlighted recent achievements in the use of natu-
ral 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
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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 at-
tenuation is the biodegradation and/or chemical destruc-
tion or stabilization of contaminants and can be an
important tool for stabilizing or remediating a contami-
nated 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 us-
ing 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 sci-
entists. Representatives from these groups made pres-
entations at the symposium that covered the following
topics:
• Laboratory studies and field demonstrations con-
ducted in support of natural attenuation at govern-
ment and industry sites.
• Methods for assessing the potential for natural at-
tenuation 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 sol-
vents. 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 sol-
vents. 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 summa-
ries of a meeting and roundtable discussion on state and
federal regulatory issues affecting natural attenuation of
chlorinated solvents. The symposium agenda is in-
cluded in the appendix.
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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 Conserva-
tion and Recovery Acts were implemented in the mid-
1980s, 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 impor-
tant 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 sen-
sitive 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 na-
ively 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 remedia-
tion 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 commit-
tee reviewed the performance of conventional pump-
and-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 nonaqueous-
phase liquids (NAPLs) that are not efficiently re-
moved 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 re-
sults 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. Concen-
trations of contaminants in pumped wells often reached
an asymptote that was above the cleanup goal. In the
instances in which major reductions in contaminant con-
centrations were achieved, the concentrations of con-
taminants 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 technolo-
gies and found that a substantial amount of performance
data existed for three alternative technologies: soil va-
por 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 network
fortetrachloroethylene and trichloroethylene. Bioventing
and soil vapor extraction work only in the vadose zone,
not in aquifers.
The committee also evaluated developing technolo-
gies that still required more controlled field studies
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and implementation at large-scale sites to generate re-
liable performance data. They considered pulsed or
variable pumping, in situ bioremediation designed for
chlorinated solvents, air sparging, steam-enhanced ex-
traction, 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 contami-
nated 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 be-
ing monitored, and many of the ground-water contami-
nants 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-and-
treat containment should thus be compared to the con-
tainment provided by the processes that naturally
attenuate contaminants in ground water. These proc-
esses include biodegradation, abiotic transformation,
sorption, and dilution.
Contribution of Natural Attenuation to
Containment
If natural attenuation can contain the spread of contami-
nation, it is the philosophical equivalent of pump-and-
treat, 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 con-
cern, the characterization is accurate. All they know is
the distribution of contaminants at their site. If site man-
agers 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 predic-
tions about the prospects for the spread of contaminants.
A good characterization study to predict containment by
natural attenuation is the equivalent of reliable perform-
ance data on a proactive technology for containment.
Because site characterizations often require sophisti-
cated sampling techniques, new analytical approaches,
and state-of-the-art ground-water modeling, natural at-
tenuation 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 com-
prehensive risk management strategy. These compre-
hensive 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 sur-
face water. Discharge from chlorinated solvent plumes
has been evaluated at the U.S. Army's Picatinny Arse-
nal, 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 con-
taminants 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 sol-
vents do not bioaccumulate, and they rapidly volatilize
to the atmosphere. As a consequence, there is little
anecdotal evidence that discharge of chlorinated sol-
vents from ground water has damaged surface-water
ecosystems; nonetheless, these issues deserve sys-
tematic evaluation.
The discipline of toxicological assessment of ecosys-
tems has made extensive progress in the last decade.
No established and widely accepted protocol for mak-
ing these assessments exists, however. As a result,
much of the science is not readily available to regula-
tors. This makes it difficult for the regulators to partici-
pate 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
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a protocol would greatly facilitate the task Of determining 2. National Research Council. 1994. Alternatives for ground water
the loadings that surface-water ecosystems can receive cleanuP washing'0". DC
without being damaged. 3. National Research Council. 1993. In situ bioremediation: When
does it work? Washington, DC.
References
1. National Research Council. 1994. Ranking hazardous waste sites
for remedial action. Washington, DC.
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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) re-
mains 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 cir-
cumstances, remedies using natural attenuation can be
more cost effective than "active" remediation ap-
proaches in achieving cleanup objectives equally pro-
tective 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 contin-
ued 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 Con-
tingency Plan (NCP), which is the regulatory framework
for the Superfund program (1). In the NCP, natural at-
tenuation is described as a process that "will effectively
reduce contaminants in the ground water" to concentra-
tions "protective of human health and sensitive ecologi-
cal environments in a reasonable timeframe." The NCP
goes on to recognize that natural attenuation may in-
clude 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 chem-
istry. Other processes not mentioned in the NCP are not
necessarily excluded from the definition (e.g., volatiliza-
tion). 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) pro-
grams.
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 reme-
dial approach for ground-water cleanup under the Su-
perfund, RCRA Corrective Action, and UST remediation
programs. A directive clarifying EPAs policy regarding
the use of natural attenuation for remediation of sites
regulated under these programs is currently in develop-
ment (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
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remedy selection process. Each EPA program offers
guidance suggesting when specific methods of protec-
tion may be more appropriate than others.
EPA recognizes that natural attenuation may be an ap-
propriate remediation method for contaminated ground
water underthe 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 at-
tenuation is not to be considered a default or presump-
tive 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 pream-
ble. The following NCP language specifies the definition
and cleanup expectation for remedies using natural
attenuation:
[Selection of natural attenuation by EPA does not
mean that the ground water has been written off and
not cleaned up but rather that biodegradation, dis-
persion, dilution, and adsorption will effectively re-
duce contaminants in the ground water to
concentrations protective of human health in a time-
frame 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 EPAs respon-
sibility to protect human health and the environment and
to satisfy the cleanup levels and other remediation ob-
jectives 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 remedia-
tion strategies using natural attenuation, such an ap-
proach 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 appropri-
ate 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 at-
tenuation and at a much lower cost than other alterna-
tives.
Sources Controlled or Controllable
Natural attenuation will not be effectively used to reach
desired cleanup levels if the rate of contamination en-
tering ground water exceeds the rate of the natural
attenuation processes. Therefore, the natural attenu-
ation 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 indi-
cate 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 ac-
tive 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.
-------
In some cases, it may be appropriate for natural attenu-
ation to be used as a followup to active remediation. In
this approach, active measures are used to reduce con-
taminant 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 provi-
sions are necessary to convince regulatory agencies
and the public that natural attenuation is a valid reme-
diation approach rather than a "walkaway" and that it will
be sufficiently protective.
Building confidence in the approach can also be pro-
moted by involving the responsible regulatory agencies
as early in the process as possible. For example, up-
front agreement on the type of characterization data
needed to demonstrate the efficacy of natural attenu-
ation 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 at-
tenuation will perform at a given site is essential to show
that natural attenuation will be effective and that poten-
tial adverse impacts to human health and the environ-
ment 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 reason-
able timeframe compared with other remedial alterna-
tives. Such a demonstration can be supported by the
following types of site data:
• Contaminant concentrations have decreased over
time.
• Geochemical or microbiological parameters are char-
acterized 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 attenu-
ation 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. Reme-
dies using natural attenuation should include a monitor-
ing 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 at-
tenuation. Such contingencies provide assurance that
the remedy's protectiveness will be maintained should
natural attenuation not progress as expected. The trig-
ger(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 con-
taminated ground water. Furthering the technical under-
standing 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 at-
tenuation 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 monitor-
ing 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 Emer-
gency Response Directive 9283.1-2 EPA/540/G-88/003 (Decem-
ber).
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.
-------
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 ground-
water contaminants in the country. These solvents, con-
sisting primarily of chlorinated aliphatic hydrocarbons
(CAHs), have been used widely for degreasing of air-
craft engines, automobile parts, electronic components,
and clothing. Only during the past 15 years has it be-
come recognized that CAHs can be transformed biologi-
cally (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 chloro-
form (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 us-
ing for other purposes. In such cases, the transforming
microorganisms must be actively growing, which re-
quires 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 oxy-
gen. 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 environ-
mental 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 chemi-
cal transformation occurs by two different pathways, lead-
ing to the formation of 1,1-DCE and acetic acid (HAc):
CH2=CC12 + H+ + CP
i i rw~*F
(elimination) (Eq.1)
CH3COOH + 3H+
HAc (hydrolysis) (Eq.2)
The rate of each chemical transformation is given by the
first-order reaction:
C = C0e
-kt
(Eq. 3)
where C is the concentration of TCA at any time t, C0
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 transforma-
tion rate constants are functions of temperature:
= A0
-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.
-------
Table 1. Reported First-Order TCA Abiotic Transformation Rates (kTCA)
Ayr-1
E kJ
10°C
15°C
20°C
References
3.47 (1 0)20
6.31 (10)20
1 .56 (1 0)20
Average half-life (yr)
118.0
119.3
116.1
0.058
0.060
0.058
12
0.137
0.145
0.137
4.9
0.32
0.34
0.31
0.95
3
4
5
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, how-
ever, because 1,1-DCE is considered more toxic than
TCA, with an MCL of 7 micrograms per liter (u,g/L)
compared with TCAs MCL of 200 u,g/L. Whenever TCA
is present as a contaminant, 1,1-DCE can also be ex-
pected. 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 harm-
less 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 transfor-
mations are taking place. Co-metabolic aerobic transfor-
mation of TCE is also possible, although if it did occur
the intermediate products formed are unstable and more
difficult, analytically, to measure. Thus, convincing evi-
dence for the latter is more difficult to obtain. Also,
aerobic co-metabolism of TCE would only occur if suffi-
cient dissolved oxygen and a suitable electron donor,
such as methane, ammonia, or phenol, were present.
Since circumstances under which the proper environ-
mental 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 reduc-
tive dehalogenation occurs may be limited by the
amount of these co-contaminants present. Theoretically,
it would require only a 0.4-gram chemical oxygen de-
mand (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 2. Conditions for Biotic and Abiotic Transformations of Chlorinated Solvents
Biotic — Aerobic
Primary substrate
Co-metabolism
Biotic — Anaerobic
Primary substrate3
Co-metabolism
Hazardous intermediates
Abiotic
Carbon Tetrachloride
(CTC)
No
No
Perhaps
Yes
Yes
Perhaps
Trichloroethene
(PCE)
No
No
Yes
Yes
Yes
No
Tetrachloroethene
(TCE)
No
Yes
Yes
Yes
Yes
No
1,1,1-Trichloroethane
(TCA)
No
Perhaps
Perhaps
Yes
Yes
Yes
Can be used as electron acceptor in energy metabolism.
-------
Figure 1. Anaerobic chemical and biological transformation
pathways for chlorinated solvents.
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 fer-
mentation. 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(ll) as
well, if present in significant amounts. If the potential for
natural biological attenuation of CAHs is to be evalu-
ated, then the concentrations of nitrate, nitrite, sulfate,
iron(ll), and methane, as well as organics as indicated
by COD or total organic carbon (TOC), should be deter-
mined. Unfortunately, such analyses are not considered
essential in remedial investigations—they should be.
Several pure cultures of microorganisms are now avail-
able 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 aerobi-
cally, however, with oxygen as an electron acceptor or
even under anoxic conditions with iron(lll) (17). In addi-
tion, 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(lll) 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 chemi-
cal 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. Micro-
cosm 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
Table 3. Environmental Conditions for Reductive Transformations of Chlorinated Solvents
Redox Environment
Chlorinated Solvent
All
Denitrification
Sulfate Reduction
Methanogenesis
Carbon tetrachloride CT -> CF
1 ,1 ,1-Trichloroethane TCA -> 1 ,1-DCE
+ CH3COOH
Tetrachloroethene
Trichloroethene
CT -> CO2+Cr
TCA-
PCE-
TCE-
41,1-DCA
•»1,2-DCE
41,2-DCE
TCA-
PCE-
TCE-
4 C02+C|-
-» ethene
4 ethene
-------
fatty acids and 4.2 mg/L of methanol, organics that
appeared to serve as electron donors for dehalogena-
tion. 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 pre-
sent. This supports conditions suitable for natural biode-
gradation of the chlorinated solvents. While some
chemical transformation of TCA to 1,1-DCE was indi-
cated (0.4 mg/L) biotransformation was extensive, as
indicated by a 1,1-DCA concentration of 7.2 mg/L, com-
pared 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, suggest-
ing that biotransformation was not eliminating the chlo-
rinated 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 Rein-
hard (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 concen-
tration 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 in-
trinsic 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 cis-
DCE, 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 re-
ductive 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 sul-
fate by reduction was most complete.
Since the above early reports, many others have re-
ported 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: Morris, R.D., ed. Handbook of bioreme-
diation. 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.
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on hydrolysis of haloalkanes and epoxides. In: Larson, R.A., ed.
Biohazards of drinking water treatment. Chelsea, Ml: Lewis Pub-
lishers, Inc. pp. 47-56.
5. Jeffers, P., L. Ward, L. Woytowitch, and L. Wolfe. 1989. Homo-
geneous hydrolysis rate constants for selected chlorinated meth-
anes, ethanes, ethenes, and propanes. Environ. Sci. Technol.
23(8):965-969.
6. Vogel, T.M., C.S. Griddle, and PL. McCarty. 1987. Transforma-
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7. McCarty, PL., and J.T Wilson. 1992. Natural anerobic treatment
of a TCE plume, St. Joseph, Michigan, NPL site. In: U.S. EPA.
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nati, OH. pp. 47-50.
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dechlorination of tetrachloroethylene and trichloroethylene to eth-
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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 trans-
formation of tetrachloroethene to ethane. Appl. Environ. Micro-
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10. Holliger, C., G. Schraa, A.J.M. Stams, and A.J.B. Zehnder. 1993.
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dechlorination of tetrachloroethene to growth. Appl. Environ. Mi-
crobiol. 59:2991-2997.
11. Neumann, A., H. Scholz-Muramatsu, and G. Dickert. 1994.
Tetrachloroethene metabolism of Dehalospirillum multivorans.
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10
-------
12. Scholz-Muramatsu, H., A. Neumann, M. MeBmer, E. Moore, and
G. Diekert. 1995. Isolation and characterization of Dehalospiril-
lum multivorans gen. sp. nov., a tetrachloroethene-utilizing,
strictly anaerobic bacterium. Arch. Microbiol. 163:48-56.
13. Holliger, C., and W. Schumacher. 1994. Reductive dehalogena-
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14. Sharma, P., and P.L. McCarty. 1996. Isolation and charac-
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logenates tetrachlorethene to c/s-1,2-dichloroethene. Appl.
Environ. Microbiol. 62:761-765.
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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
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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, PL., and M. Reinhard. 1993. Biological and chemical
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and terrestrial environments in the biochemistry of global change.
In: Oremland, R.S., ed. The biogeochemistry of global change:
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fects. Prepared by Meta Systems, Inc., Cambridge, MA.
24. McCarty, PL., L. Semprini, M.E. Dolan, T.C. Harmon, C. Tiede-
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tion processes for xenobiotic and hydrocarbon treatment. Boston,
MA: Butterworth-Heinemann. pp. 16-40.
25. Semprini, L., PK. 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 Bioreme-
diation of Ground Water. EPA/540/R-94/515. Washington, DC.
27. Hasten, Z.C., PK. 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/R-
94/075. 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., PF. Mazierski, R.J. Buchanan, D.E. Ellis, and L.S.
Sehayek. 1995. Intrinsic in situ anaerobic biodegradation of chlo-
rinated solvents at an industrial landfill. In: Hinchee, R.E., J.T.
Wilson, and D.C. Downey, eds. Intrinsic bioremediation. Colum-
bus, 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 se-
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31. Buchanan, J.R.J., D.E. Ellis, J.M. Odom, PF. Mazierski, and M.D.
Lee. 1995. Intrinsic and accelerated anaerobic biodegradation of
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and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH:
Battelle Press, pp. 245-252.
11
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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 de-
greasing 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 pol-
lutants 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 ac-
ceptable (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 envi-
ronments are indeed anaerobic, reductive dechlorina-
tion to ethene offers promise that natural attenuation
may be exploited in many instances of contamination by
chlorinated ethenes. The completeness of the conver-
sion to ethene is highly variable from site to site, how-
ever, with the responsible factors for this variation not
well understood.
2[H) HC1 2[H] HC1 2[H] HCI 2[H] HC]
PCE —^ * ' TCE —^ * • 1,2-DCEs —^ * » VC —^--^ • ETH
Figure 1. Reductive dechlorination of chlorinated ethenes
(under anaerobic conditions).
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 dechlorina-
tion of PCE and TCE were studies in which the mediat-
ing 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 co-
factors that can mediate the slow, incomplete reductive
dechlorination of PCE and TCE to (usually) DCE iso-
mers (6). This process is co-metabolic in that it happens
more or less accidentally or incidentally as the organ-
isms carry out their normal metabolic functions; the
organisms apparently derive no growth-linked or en-
ergy-conserving benefit from the reductive dechlorina-
tion. Such co-metabolic dechlorinations undoubtedly are
responsible for the incomplete, relatively slow transfor-
mations of chloroethenes observed at many field sites.
The organisms that can mediate such processes are
ubiquitous, but the process is sufficiently slow and in-
complete that a successful natural attenuation strategy
cannot completely rely upon it.
On the other hand, more recent studies have demon-
strated the existence of direct dechlorinators—microorgan-
isms derived from contaminated subsurface environments
and treatment systems—that utilize chlorinated ethenes
as electron acceptors in an energy-conserving, growth-
coupled metabolism termed dehalorespiration(7). Several
species that carry out direct dechlorination of chlorinated
ethenes are described below.
12
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To a large extent, then, success or failure of natural
attenuation can be linked to the specific type of dechlori-
nator 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 dechlorina-
tors must compete for available H2 with hydrogenotro-
phic 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 adminis-
tered electron donor, supplying H2 as it was rapidly
fermented to acetate. During the period of high H2 lev-
els, 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 contin-
ued slowly. In fact, kinetic analysis of mixed cultures
of Dehalococcus ethenogenes and hydrogenotrophic
methanogens showed that this dechlorinator has an
affinity for H210 times greater than that of the methano-
gens in the culture (8). We do not know whetherthis high
affinity for H2 is typical of dechlorinators, but thermody-
namic arguments would suggest it. We also do not yet
Table 1. Properties of Some Direct PCE Dechlorinators
know the differences in relative affinity for H2 between
dechlorinators and sulfate-reducers, important competi-
tors 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 ace-
tate, producing high, short-lived peaks of H2, do not
favor dechlorination as well as would more persistent,
slowly fermented substrates such as benzoate or propion-
ate (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 respira-
tory 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 cis-
DCE are relatively abundant and easier to culture. This
ability seems to have evolved in several different phylo-
genetic groups in the eubacteria, as determined by 16S
rRNA sequence analysis. Many direct dechlorinators
seem to be related to either the gram-negative sulfate-
reducing bacteria (epsilon proteobacteria) or the gram-
positive group, including Desulfotomaculum. Sulfate reducers
Organism
Dehalobacter
restrictus
Dehalospirillum
multivorans
Strain TT4B
Enterobacter
agglomerans
Desulfitobacterium
sp. strain PCE1
Dehalococcus
ethenogenes
strain 195
Dechlorination
Reactions
PCE, TCE -> c/s-DCE
PCE, TCE -> c/s-DCE
PCE, TCE -> c/s-DCE
PCE, TCE -> c/s-DCE
PCE, TCE -> (c/s-DCE)
o-chlorophenols
PCE, others -^ethene
Electron
Donors
H2
H2, formate,
pyruvate, etc.
Acetate
Nonfermentable
substrates
Lactate,
pyruvate,
butyrate,
ethanol, etc.
H2
Other Electron
Acceptors
None
Thiosulfate, nitrate,
f urn a rate, etc.
None
O2, nitrate, etc.
Sulfite, thiosulfate,
fumarate
None
Morphology
Rod
Spirillum
Rod
Rod
Curved rod
Irregular
coccus
Phylogenetic
Position
Gram +
Desulfotomaculum
group
Epsilon
proteobacteria
?
Gamma
proteobacteria
Gram +
Desulfotomaculum
group
Novel eubacterium
References
9-11
12
13
14
15
16a
a Maymo-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
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tend to be versatile at using electron acceptors for an-
aerobic respiration. We know much less about organisms
capable of reducing chloroethenes past DCE. These or-
ganisms 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. etheno-
genes," and strain TT4B apparently can only use a
single electron donor and only chlorinated aliphatic hy-
drocarbons as electron acceptors. These findings raise
questions about what these organisms used as electron
acceptors before widespread chlorinated ethene con-
tamination. The organisms possibly use electron ac-
ceptors 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 dechlori-
nators 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 ace-
tate, vitamin B12, unidentified factors in sewage sludge
(16), and perhaps other factors. Indeed, this organ-
ism's requirement for vitamin B12 allowed a plausible
explanation for methanol's being the best H2-source
for PCE dechlorination by the original mixed dechlori-
nating culture (17), since methanol-utilizing methano-
gens 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 micro-
bial complexities of assessing natural attenuation poten-
tial. It is important to keep in mind the competitive
aspects of electron donor flow. In essence, dechlorina-
tion is in a "foot race" with competing donor uses. If too
little donor is initially present, the pattern of its conver-
sion to H2 is too unfavorable, or there is too much
competition for it, dechlorination may not proceed ade-
quately 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. There-
fore, 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 dechlorina-
tion is limited by physical/chemical factors, nutrients, or
lack of the appropriate dechlorinating organisms. Par-
ticularly 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 proper-
ties of organisms dechlorinating chlorinated ethenes
past DCEs is rudimentary.
In summary, the goal of assessment should be to evalu-
ate 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 po-
tentially 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 trans-
formation of tetrachloroethene to ethane. Appl. Environ. Micro-
biol. 58:1996-2000.
2. Freedman, D.L., and J.M. Gossett. 1989. Biological reductive
dechlorination of tetrachloroethylene and trichloroethylene to eth-
ylene under methanogenic conditions. Appl. Environ. Microbiol.
55:2144-2151.
3. Fathepure, B.Z. and S.A. Boyd. 1988. Dependence of tetrachlo-
roethylene dechlorination on methanogenic substrate consump-
tion by Methanosarcinasp. 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 Micro-
biol. Lett. 49:149-156.
5. Fathepure, B.Z., J.P. Nengu, and S.A. Boyd. 1987. Anaerobic
bacteria that dechlorinate perchloroethene. Appl. Environ. Micro-
biol. 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 dehalogena-
tion 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 bac-
teria. Ph.D. dissertation. Agricultural University, Wageningen, the
Netherlands.
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10. Holliger, C., G. Schraa, A.J.M. Stams, and A.J.B. Zehnder. 1992.
Enrichment and properties of an anaerobic mixed culture reduc-
tively dechlorinating 1,2,3-trichlorobenzene to 1,3-dichloroben-
zene. 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. Mi-
crobiol. 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 tetrachlo-
roethylene 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 charac-
terization of a facultatively aerobic bacterium that reductively de-
halogenates tetrachloroethene to c/s-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 orffto-chlorinated phenols.
Arch. Microbiol. 165:132-140.
16. Maymo-Gatell.X., V. Tandoi, J.M. Gossett, and S.H. Zinder. 1995.
Characterization of an hh-utilizing enrichment culture that reduc-
tively dechlorinates tetrachloroethene to vinyl chloride and ethene
in the absence of methanogenesis and acetogenesis. Appl. En-
viron. 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 an-
aerobic reductive dechlorination. In Hinchee, R.E., A. Leeson,
and L. Semprini, eds. Bioremediation of chlorinated solvents.
Columbus, OH: Battelle Press, pp. 9-16.
15
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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 micro-
organisms. 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 exam-
ples 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 poten-
tial 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 trans-
formations 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 appro-
priate predictions and evaluations, particularly for in situ
transformation processes under intrinsic (native) condi-
tions, where engineered approaches are not available
to influence or dominate in situ hydrogeochemical con-
ditions. 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-dimen-
sional distribution of geochemical and hydraulic condi-
tions 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 physi-
ology 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 bacte-
ria) 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 elec-
tron acceptor may be an organic or inorganic compound.
During this electron transfer process, usable energy is re-
covered through a complex series of oxidation-reduction
(redox) reactions by the formation of energy storage com-
pounds or electrochemical gradients. The oxidation of or-
ganic compounds coupled with the reduction of molecular
oxygen is termed aerobic heterotrophic respiration and has
been the basis of most applications of bioremediation.
16
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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 produces) and an-
other 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 destruc-
tion of a contaminant is often not a single type of micro-
organism. 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, re-
ferred to as reductive dechlorination or biodehalogena-
tion, is a sequential removal of chlorine moieties from
the ethene core during a biologically mediated two-
electron 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 mo-
lecular 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 transfor-
mations. By definition, co-metabolic process do not di-
rectly benefit the organisms buy supplying energy or
material for cellular synthesis. The mono-oxygenase
systems that transform chloroethenes may be inacti-
vated 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 biotreat-
ment process, but it is of limited significant under native
or intrinsic conditions because of the anticompetitive
effects on the microorganisms involved and the environ-
mental conditions needed for significant transformation
to occur.
The lesser-chlorinated DCEs can be reductively trans-
formed, 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 mo-
bile than the parent compounds and is extremely vola-
tile. Due to its toxicity, when vinyl chloride is detected in
the subsurface environment with the other chlo-
roethenes, it is usually the focus of risk-based evalu-
ations 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 underap-
preciated ecosystem. The appropriate application of in-
trinsic remediation at a site requires an understanding
of the ecological processes under site-specific condi-
tions. Subsurface microorganisms will respond to take
advantage of all available resources (i.e., energy, nutri-
ents, 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 advan-
tage of resources and conditions. This concept is re-
ferred to as filling all available ecological niches. The
converse statement is also true, however: subsurface
microorganism do not respond in ways that are counter-
productive 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 subsur-
face (microbial) ecology in terms of energy transforma-
tion and transfer. Under intrinsic conditions, bioavailable
energy is the ecological resource that induces the ob-
served 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 con-
ditions, energy transduction in ground-water environments is
limited by the availability of carbon/energy sources
(electron donors). Energy transduction in contaminated
17
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subsurface systems is usually limited by the availability
of the electron acceptor. When readily degradable or-
ganic matter enters the subsurface in sufficient quanti-
ties, it produces a series of zones defined by the
terminal electron accepting process (TEAR). These
zones are not necessarily mutually exclusive and de-
pend on the availability of electron acceptors (O2, NO3=,
SO4=, Fe3+, CO2). There is no reason to assume a
similarity between the biodegradation potential in differ-
ent metabolic zones. This potential will be based on the
energetics associated with the dominant redox proc-
esses, the metabolic diversity of the microbial commu-
nities, the immediate geochemical conditions, and the
chemical nature of the contaminant of concern.
The reductive dehalogenation process may be thought
of as another TEAR, and the microorganisms involved
compete for the available flow of energy (reducing
equivalents). As noted above, however, PCE orTCE, 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 micro-
organisms. This is why PCE and TCE plumes with de-
tectable levels of dissolved oxygen do not show
evidence of active biodegradation. The presence of dis-
solved oxygen indicates no significant quantities of oxi-
dizable electron donor are present. Vinyl chloride (and
perhaps DCEs) underthe 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 deg-
radation activity. Possible mechanisms of adaptation
include expression of catabolic potential (induction), se-
lection of novel capabilities (mutation), growth of degra-
dative populations, formation of degradative consortia,
and formation of metabolic intermediate pools. Labora-
tory research results indicate that the formation of cat-
abolically 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 trans-
formation rates. Environmental parameters that support
anaerobic microbial transformation processes (both oxi-
dative 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 un-
derstanding of the mechanisms involved may lead to
greater predictive capabilities.
Conclusion
Biotransformations that serve as major mass removal
mechanisms for the intrinsic remediation of chlo-
roethenes and other ground-water contaminants have
been demonstrated in the laboratory and the field. The
use of these processes as remedial technologies, how-
ever, 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 ap-
propriate implementation of intrinsic remediation for
chloroethene sites. A greater understanding of the
mechanisms of adaptation in subsurface microbial com-
munities 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 biotrans-
formation of contaminants in the subsurface. EPA/600/M-90/024.
February.
Vogel, T.M., C.S. Griddle, and P.L. McCarty. 1987. Transformations of
halogenated aliphatic compounds. Environ. Sci. Technol. 22:722-736.
18
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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—per-
chloroethene (PCE), trichloroethene (TCE), dichlo-
roethenes (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. Pe-
troleum hydrocarbons universally serve as electron do-
nors (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 sub-
ject to reductive dechlorination according to the se-
quence PCE -> TCE + Cl -> DCE + 2CI -> VC + 3CI ->
ethylene + 4CI (1). The efficiency of dechlorination,
however, appears to differ under methanogenic, sul-
fate-reducing, iron(lll)-reducing, and nitrate-reducing
con- ditions. Dechlorination of PCE and TCE to DCE is
favored under mildly reducing conditions such as nitrate
or iron(lll) 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 chlorin-
ated ethenes such as VC can be oxidized under oxic
(8) or iron(lll)-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 attenu-
ation of chlorinated ethenes in ground-water systems.
This paper summarizes a methodology for identifying
the zonation of redox conditions in the field. This meth-
odology can serve as an a priori screening tool for
identifying ground-water systems in which redox condi-
tions will favor natural attenuation of chlorinated
ethenes. Conversely, this methodology can identify sys-
tems for which natural attenuation of chlorinated
ethenes is not favored and other remediation technolo-
gies should be considered.
Methodology for Determining Redox
Processes in Ground-Water Systems
Platinum electrode redox potential measurement histori-
cally has been the most widely used method for deter-
mining redox conditions in ground-water systems. While
redox potential measurements can accurately distin-
guish oxic from anoxic ground water, they cannot distin-
guish between different anoxic processes such as
nitrate reduction, iron(lll) reduction, sulfate reduction, or
methanogenesis. One reason is that many redox spe-
cies, 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 chlorin-
ated ethenes, redox potential measurements alone can-
not 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 meas-
urement of concentrations of transient intermediate
19
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products. Molecular hydrogen (H2), the most ubiquitous
intermediate product of anaerobic microbial metabolism,
has proven to be especially useful in this context. Differ-
ent electron-accepting processes have characteristic
H2-utilizing efficiencies. Nitrate reduction, the most en-
ergetically favorable anoxic process, maintains H2 con-
centrations below 0.1 nanomoles (nM) per liter. Iron(lll)
reduction maintains H2 concentrations between 0.2 and
0.8 nM, whereas forsulfate 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 con-
cluded 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 re-
dox 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 confi-
dence 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 exam-
ined in ground-water chemistry investigations. Hydro-
gen concentrations in ground water can be made using
a gas-stripping procedure (13). Astandard 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 head-
space and asymptotically approach equilibrium with the
dissolved phase. After 20 to 25 minutes, equilibrium is
achieved, and the gas bubble is sampled using a syr-
inge. 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 per-
sonnel 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.
Explanation
•
16
monitoring
well location
and number
0 100 feet
scale
Figure 1. Map showing location of fire-training pits and moni-
toring wells, Site 8, NAS Cecil Field, Florida.
Changes in the concentrations of redox-sensitive con-
stituents along the flowpath of the shallow aquifer sys-
tem 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 be-
low 1 mg/L until about 400 feet along the flowpath, then
increase to about 2.5 mg/L, indicating active iron(lll)
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 meth-
ane concentrations are present.
Between 200 and 170 feet downgradient, where sulfate
concentrations decline and sulfide concentrations in-
20
-------
* 6 -
1 5-
* A
8,3-
| 0 -
A
: A
S...A /
"--tf- 1
— •— Dissolved Oxygen
-•-- Fe(ll)
A Methane
V"_ """•-•--* '•
(A)
60
£ 40
I 30
120H
8 10
a °
(B)
100 200 300 400 500 600 700 800
6
4 -
2 -
0
Methanogenesis
. Sulfate Reduction
Fe(lll) Reduction
(C)
200 400 600
Distance along the Flowpath (feet)
800
Figure 2.
Concentration changes of redox-sensitive parameters
along ground-water flowpaths in the shallow aquifer,
Site 8, NAS Cecil Field, Florida.
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(lll) reduction.
Across 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-re-
ducing and iron(lll)-reducing zones further downgradi-
ent. 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 down-
gradient iron(lll)-reducing zone, anoxic oxidation of VC
to carbon dioxide (CO2) can occur (Figure 3).
These biodegradation processes, which can be postu-
lated 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
Fire Pit Area
Ground-Water
Discharge
Area
SCALE
200 feet
Methanogenesis
Sulfate Reduction
Fe(lll) Reduction
Figure 3. Concentrations of dissolved hydrogen (nM) and the
zonation of predominant redox processes, Site 8,
NAS Cecil Field, Florida.
35
3 30
a 25
.1 20
IS 15
I 10
II
m *
(A)
100 200 300 400 500 600 700 800
XAxis
£ 700
| 600
O 500
1 4°°
f 300
"g 20°
-S 100
(B)
—i 1 1 1 1 1 1 1
0 100 200 300 400 500 600 700 800
(C)
~i 1 1 1 1 r
0 100 200 300 400 500 600 700 800
Distance along the Flowpath (feet)
Figure 4. Concentration changes of chloride, dissolved inor-
ganic carbon, and chlorinated ethenes along ground-
water flowpaths.
been so efficient that the best water-chemistry record of
the original contamination is probably the elevated con-
centrations of dissolved inorganic carbon (Figure 4B)
and dissolved chloride (Figure 4C) observed in down-
gradient ground water that currently lacks measurable
chlorinated ethene contamination. These patterns suggest
that most of the chlorinated ethenes have been com-
pletely transformed to CO2 and chloride by the cumula-
21
-------
tive effects of reductive dehalogenation in the methano-
genic and sulfate-reducing zones and oxidative proc-
esses in the downgradient iron(lll)-reducing and oxic
zones.
Conclusion
An understanding of ambient redox conditions is a power-
ful 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 deline-
ation. This methodology was demonstrated at a site at
MAS Cecil Field, Florida. The progression from methano-
genic -» sulfate reduction -»Iron(lll) reduction -» oxygen
reduction has efficiently decreased concentrations of chlo-
rinated 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 imple-
menting 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 Excel-
lence, Brooks Air Force Base, San Antonio, TX. p. 129.
3. Chapelle, F.H., J.M. Landmeyer, and P.M. Bradley. 1996. Assess-
ment 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. Griddle, and PL. McCarty. 1987. Transforma-
tions 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 eth-
ylene 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 trans-
formation of tetrachloroethene to ethene. Appl. Environ. Micro-
biol. 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(lll) reducing aquifer sediments. Environ.
Sci. Technol. 40:2084-2086.
10. McCarty, PL., and L. Semprini. 1994. Groundwater treatment for
chlorinated solvents. In: Handbook of bioremediation. Boca Ra-
ton, 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., PB. 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 micro-
bially catalyzed redox reactions in anoxic groundwater. Environ.
Sci. Technol. 28:1255-1210.
22
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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 contami-
nation 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 con-
junction with aquifer hydrogeological parameters (such
as seepage velocity and dilution) to show that a reduc-
tion 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 contami-
nant concentrations can be directly correlated to in-
creases in metabolic byproduct concentrations. This
evidence can be used to show that concentrations of
electron donors or acceptors in ground water are suffi-
cient to facilitate degradation of the dissolved contami-
nants (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 contami-
nants 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 biode-
gradation of the contaminants of interest. Further, the ap-
pearance 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 applica-
tions. 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 impossi-
ble. In some situations, there are legal or physical im-
pediments to the comparison of concentrations in
monitoring wells along a flow path. In many landscapes,
the direction of ground-water flow (and water-table ele-
vations in monitoring wells) can vary over short periods
due to tidal influences or changes in barometric pres-
sure. 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 uncertain-
ties 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 con-
tinuous monitoring of water-table elevations in the aqui-
fer being evaluated. The product of the microcosm
study, and the continuous monitoring of water-table ele-
vations, will be a yearly or seasonal estimate of the
extent of attenuation along average flow paths. Remov-
als seen at field scale can be attributed to biological
23
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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 geo-
chemical regions for natural attenuation of chlorinated
aliphatic hydrocarbons are regions that are actively
methanogenic, exhibit sulfate reduction and iron reduc-
tion 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 chlo-
ride is consumed but dechlorination of PCE, TCE, or
dichloroethylene (DCE) may not occur (4). Core material
must be acquired from each geochemical region in ma-
jor flow paths represented by the plume, and the hydrau-
lic conductivity of each depth at which core material is
acquired must be measured. If possible, the micro-
cosms should be constructed with the mosttransmissive
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 fordaughter
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 re-
quires an electron donor for the process to proceed. The
electron donor could be soil organic matter, low molecu-
larweight organic compounds (e.g., lactate, acetate, metha-
nol, 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 contami-
nated 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 gener-
ally used to extrapolate rates to the field, can be calcu-
lated by comparing wet and dry weights of the aquifer
material.
To ensure sample integrity and stability during acquisi-
tion, it is important to quickly transferthe 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 incu-
bated 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 geo-
chemical 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 miner-
als, and microbial populations. Major geochemical pa-
rameters that characterize the subsurface include
alkalinity, pH, redox potential, dissolved constituents (in-
cluding electron acceptors), temperature, the physical
and chemical characterization of the solids, and micro-
bial 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 result-
ing from biodegradation of the pollutants. Increases in
both alkalinity and pH have been measured in portions
of an aquifer contaminated by gasoline undergoing ac-
tive 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 microorgan-
isms between 6.0 and 8.0; many microorganisms, how-
ever, 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-buff-
ered systems (9). pH values as low as 3.0 have been
measured for ground waters contaminated with municipal
24
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waste leachates, however, which often contain elevated
concentrations of organic acids (10). In ground waters con-
taminated 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 elec-
trons. Most redox reactions in the subsurface are micro-
bially 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(lll).
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 condi-
tions increasing through the sequence oxygen, nitrate,
iron, sulfate, and carbonate. The redox potential de-
creases in each sequence, with methanogenic (carbon-
ate 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 poten-
tial of many reactions and cannot be used for quantita-
tive 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 predomi-
nant in situ redox reactions (13-15). The evidence sug-
gests 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 geo-
chemical processes occurring in situ at the time of sam-
ple 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 de-
pleted, carbonate becomes the electron acceptor, with
methane gas produced as an end product.
Temperature
The temperature at all monitoring wells should be meas-
ured to determine when the pumped water has stabi-
lized and is ready for collection. Below approximately 30
feet, the temperature in the subsurface is fairly consis-
tent 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 chlo-
ride should increase as the chlorinated solvents de-
grade. 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, con-
taminants, and daughter products that should be meas-
ured during site characterization for natural attenuation.
The tables include the analyses that should be per-
formed, 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/AgCI
Sulfate
Nitrate
Oxygen
Oxygen
Iron(ll)
Sulfide
Hydrogen
Hydrogen
PH
< 20 mg/L
< 1 mg/L
< 0.5 mg/L
> 1 mg/L
> 1 mg/L
> 1 mg/L
> 1 nM
< 1 nM
5 < pH < 9
Reductive pathway possible
Competes at higher
concentrations with reductive
pathway
Competes at higher
concentrations with reductive
pathway
Tolerated; toxic to reductive
pathway at higher
concentrations
Vinyl chloride oxidized
Reductive pathway possible
Reductive pathway possible
Reductive pathway possible;
vinyl chloride may
accumulate
Vinyl chloride oxidized
Tolerated range
25
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Table 2. Contaminants and Daughter Products
Analysis Interpretation
PCE
TCE
1,1,1-Trichloroethane
c/s-DCE
trans-DCE
Vinyl chloride
Ethene
Ethane
Methane
Chloride
Carbon dioxide
Alkalinity
Material spilled
Material spilled or daughter product of
perchloroethylene
Material spilled
Daughter product of trichloroethylene
Daughter product of trichloroethylene
Daughter product of dichloroethylene
Daughter product of vinyl chloride
Daughter product of ethene
Ultimate reductive daughter product
Daughter product of organic chlorine
Ultimate oxidative daughter product
Results from interaction of carbon
dioxide with aquifer minerals
Microcosm Construction
During construction of the microcosms, manipulations
should take place in an anaerobic glovebox. These
gloveboxes exclude oxygen and provide an environ-
ment in which the integrity of the core material may be
maintained, since many strict anaerobic bacteria are sen-
sitive to oxygen. Stringent aseptic precautions are not
necessary for microcosm construction; maintaining the
anaerobic conditions of the aquifer material and solu-
tions 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 micro-
cosm has too much water and the contaminant is pri-
marily 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 sam-
pling interval are preferable to microcosms that are re-
petitively 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 pre-
ferred 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
overtime. Sodium azide is effective in repressing meta-
bolism 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 mi-
crocosms should be inverted in an anaerobic glovebox
as they incubate; anaerobic jars are also available that
maintain an oxygen-free environment. Dry redox indica-
tor 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. Tef-
lon-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 micro-
cosm 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 biode-
gradation 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 biodegrada-
tion 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 mate-
rial. 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 estima-
tion of the field-scale rate of attenuation did not account
for regions of preferential flow in the aquifer. The regions
26
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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 remov-
als of contaminants of concern in the living treatments
are significantly different from zero or significantly differ-
ent 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 re-
spective concentrations of 7,800 u,g/L and 1,100 u,g/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 con-
centrations for benzene, toluene, and TCE were 380
u,g/L, 450 u,g/L, and 330 u,g/L, respectively. The micro-
cosms 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
n TCE Microcosm
• TCE Control
0 5 10 15 20 25 30 35 40 45
Time (Weeks)
Figure 1. TCE concentrations in the Tibbetts Road microcosm
study.
D Benzene Microcosm
• Beiuene Control
D 5 10 15 20 25 30 35 40 45
Time (Weeks)
Figure 2. Benzene concentrations in the Tibbetts Road micro-
cosm study.
a Toluene Micracosi
• Toluene Control
0 5 10 15 20 25 30 35 40 45
Time (Weeks)
Figure 3. Toluene concentrations in the Tibbetts Road micro-
cosm study.
Figure 4. Location of waste piles and flow path segments at
the Tibbetts Road Superfund site.
27
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Table 3. Concentrations of TCE, Benzene, and Toluene in the Tibbetts Road Microcosms
Compound
TCE
Mean ± standard deviation
Benzene
Mean ± standard deviation
Toluene
Mean ± standard deviation
Time Zero
Microcosms
328
261
309
299 ± 34.5
366
280
340
329 ± 44.1
443
342
411
399 ± 51 .6
Time Zero
Controls
337
394
367
366 ± 28.5
396
462
433
430 ± 33.1
460
557
502
506 ± 48.6
Week 23
Microcosms
1
12.5
8.46
7.32 ± 5.83
201
276
22.8
1 67 ± 1 30
228
304
19.9
1 84 ± 1 47
Week 23
Controls
180
116
99.9
132 ±42.4
236
180
152
189 ±42.8
254
185
157
199 ±49.9
Week 42
Microcosms
2
2
2
2.0 ± 0.0
11.1
20.5
11.6
14.4 ±5.29
2
2.5
16.6
7.03 ± 8.29
Week 42
Controls
36.3
54.5
42.3
44.4 ± 9.27
146
105
139
130 ±21 .9
136
92
115
114 ±22.0
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 prob-
ably due to sorption; however, this study exhibited more
sorption than typically seen.
The rate constants determined from the microcosm
study forthe 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
otherwords, the removal that is in excess of the removal
in the controls.
The first-order removal in the living and control micro-
cosms 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 com-
pounds 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, orethene, although removals of TCE and cis-
DCE were extensive. This can be explained by the
observation that iron-reducing bacteria can rapidly oxi-
dize 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 4. First-Order Rate Constants for Removal of TCE, Benzene, and Toluene in the Tibbetts Road Microcosms
Parameter
TCE
95% confidence interval
Minimum rate significant at 95% confidence
Benzene
95% confidence interval
Minimum rate significant at 95% confidence
Toluene
95% confidence interval
Minimum rate significant at 95% confidence
Living
Microcosms
6.31
±2.50
3.87
±1.96
5.49
±2.87
Autoclaved
Controls
First-Order Rate of Removal
2.62
±0.50
1.51
±0.44
1.86
±0.45
Removal Above
Controls
(per year)
3.69
±2.31
1.38
2.36
±1.83
0.53
3.63
±2.64
0.99
28
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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
TCE
c/s-DCE
trans-DCE
1, 1-DCE
Vinyl Chloride
Ethene
Benzene
Toluene
o-Xylene
m-Xylene
p-Xylene
Ethylbenzene
Methane
Iron
SOS 79S
Upgradient g/L Downgradient g/L
200
740
0.41
0.99
< 1
<4
510
10,000
1,400
2,500
1,400
1,300
353
13.7
10.9
<1
< 1
< 1
<4
2.5
< 1
8.4
<1
22
0.7
77
70S
Upgradient g/L
710
220
0.8
< 1
< 1
7
493
3,850
240
360
1,100
760
8
52S
Downgradient g/L
67
270
0.3
1.6
< 1
<4
420
900
71
59
320
310
3
70S
Upgradient g/L
710
220
0.8
< 1
< 1
7
493
3,850
240
360
1,100
760
8
53S
Downgradient g/L
3.1
2.9
< 1
< 1
< 1
<4
<1
< 1
< 1
< 1
<1
<1
<2
27,000
Table 6. First-Order Rate Constants in Segments of the
Tibbetts Road Plume
Flow Path Segments in Length and Time of
Ground-Water Travel
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 esti-
mated at field scale. The agreement between the inde-
pendent 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 7. Comparison of First-Order Rate Constants in a Microcosm Study and in the Field at the Tibbetts Road NPL Site
Microcosms Corrected for Controls Field Scale
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
c/s-DCE
Benzene
Toluene
o-Xylene
m-Xylene
p-Xylene
Ethylbenzene
0.41
0.65
0.82
> 1.42
0.79
> 1.20
0.64
1.16
0.59
Produced
0.04
0.36
0.30
0.45
0.31
0.22
0.54
0.43
>0.62
>0.83
>0.55
>0.59
>0.70
>0.66
rarameier
Trichloroethylene
Benzene
Toluene
Average Rate
3.69
2.36
3.63
Minimum Rate Significant
at 95% Confidence
First-Order Rate
1.38
0.53
0.99
Segment
A
(per year)
0.41
0.82
> 1.42
Segment
B
0.59
0.04
0.36
Segment
C
0.54
>0.62
>0.83
29
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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 trichlo-
roethylene in ground water at Picatinny Arsenal, New Jersey, and
St. Joseph, Michigan. In: U.S. EPA. Symposium on Bioremedia-
tion 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(lll)-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 micro-
cosms 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 micro-
bially 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
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52:2993-3003.
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McGraw-Hill.
30
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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 immiscible-
phase 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 frac-
tured 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 sub-
surface source zone or simply the source zone, whether it
has DNAPL residual or free product or has phased-trans-
ferred DNAPL in low-permeability zones. Significant sol-
vent 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, leak-
ages, or spills have ceased; therefore, the solvent mass
in the source zone is now slowly diminishing and even-
tually 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 main-
taining hazardous concentration levels. These ques-
tions 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 downgradi-
ent receptors the plume can be viewed as having
achieved intrinsic remediation. It is unlikely that deple-
tion of the source zones contributes to intrinsic remedia-
tion of chlorinated solvent plumes; therefore, intrinsic
remediation must depend on attenuation processes op-
erating within the plume.
Intrinsic Remediation
Intrinsic remediation occurrences are well known at pe-
troleum 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 proc-
esses somehow combine to cause the plume front to
achieve steady state or near steady state or perhaps
cause plume shrinkage.
31
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The three processes that can drive the plume front
towards the steady-state condition are mechanical dis-
persion, 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 perma-
nent mass removal from or dilution of the plume. In the
context of chlorinated solvent plumes, biotic or abiotic
processes commonly cause transformations of the par-
ent 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 drinking-
water 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 conclu-
sions on the propensity for chlorinated solvent plumes
to achieve intrinsic remediation. Conclusions about pro-
pensity 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-con-
centration 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 thou-
sand micrograms per liter, is commonly large relative to
the volume of the core, which commonly has concentra-
tions 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 Recov-
ery Act. During the past 15 years, millions of conven-
tional monitoring wells have been used at many
thousands of solvent sites in the United States and
several other countries. Solvent plumes present an ex-
ceptionally 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. Conven-
tional 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 sparse-
ness of data points, conventional networks only rarely
establish the existence of cores, except perhaps close
to the source zones. Thus, such plumes with no ob-
served cores are perceived to have relatively low con-
centrations and therefore small total contaminant mass.
Detailed monitoring of experimental solvent plumes pro-
duced at the Borden field site (an unconfined sand
aquifer located 60 kilometers northwest of Toronto, Can-
ada [1]) using unconventional techniques, as well as
similar monitoring of several plumes at actual industrial
sites, provides exceptional spatial resolution of the dis-
tribution 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 conven-
tional 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 extend-
ing 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 possibili-
ties 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: Re-
search strategies for in-situ remediation. Berlin/New York: Sprin-
ger-Verlag. pp.102-127.
32
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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 accu-
rate logs of horizontal Kversus 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
Klog. These conclusions have been supported by more
recent studies (2-7).
Inadequate performance of many pump-and-treat sys-
tems has been attributed to improper design (8, 9). In
far too many cases, underestimation of aquifer hetero-
geneity plays a significant role in these design fail-
ures. 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 esti-
mates 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 inter-
val carrying a plume of contamination by a factor of
ten or more. Proper characterization of aquifer hy-
draulic 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) ver-
tical flow that exists in most wells, and, through a
small pumpingtest,theflowdistributionenteringthewell
fromthe surrounding formation. If certain conditions are
met, the distributions provide sufficient information to
determine the hydraulicconductivityoftheaquiferzones
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 num-
ber of papers (11-13). By the late 1980s, the electromag-
netic (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 ef-
fect on contaminant migration (3, 6).
Conducting a Flowmeter Test
A flowmeter test may be viewed as a natural generali-
zation 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 verti-
cal 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" re-
fers 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
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PUMP
-0
CAP ROCK-
(Q-DISCHARGE RATE) !
I
I
BOREHOLE FLOW'if
METER 4i
TO LOGGER (Q)
SURFACE
..CASING
I SCREEN
ELEVATION*
DATA
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.
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 conduc-
tivity 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 pro-
duced per test versus thousands or tens of thousands
for conventional pumping tests. Thus, potential treat-
ment and disposal costs are minimal. Atypical 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.
.Ambient Row
0.3 M «.(
FtawCL/mW
Amuwnt now
Pumping Induced Flow
a 4
Row lUmbil
Nut Pumping Flow
z a
FtowlUmbi)
Figure 2. The data actually recorded by a flowmeter are ambi-
ent flow and pumping-induced flow. In both cases,
positive values indicate upward flow. Net pumping
flow is pumping-induced flow minus ambient flow.
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
Kdistribution 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 sam-
ples, 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).
34
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Differential Ambient Flow
-0.4
-0.2
Flow IL/ntin)
Benzene (pg/L.)
2000 4000
6000
836
835
834
833
832
831
Elevation
(meters)
Figure 3. Plots of the differences of neighboring values of am-
bient flow (differential ambient flow) and net flow (dif-
ferential net flow). These values represent the flow
entering (positive) or leaving (negative) the well
from/to the various layers of the aquifer.
100
110
£
Q
130
140
830 i
829 .
828
827
826 ..
825
0.12 0.10 0.08 0.06 0.04 0.02
Hydraulic Conductivity (cm/s)
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 trans-
missivity layer defined by a flowmeter analysis (6).
Figure 4.
0 5 10 15 20
Hydraulic Conductivity (m/day)
Hydraulic conductivity as a function of depth in Well
P26-M1 at the Savannah River site. The measurement
interval (layer thickness) was 1 foot.
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 infor-
mation concerning hydraulic heterogeneity is unprece-
dented 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 reme-
diation, the potential impact of flowmeters on site
35
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characterization and modeling is dramatic. Simultane-
ously, 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 in-
flation 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 pro-
vide 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 heteroge-
neous aquifer, 1. Overview and site description. Water Resour.
Res. 28(l2):3281-3292.
3. Boman, G.K., F.J. Molz, and K.D. Boone. 1996. Borehole flow-
meter 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 forma-
tions. 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 bore-
hole flowmeter: Description and application. Ground Water Moni-
toring 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 Xl:119-124.
9. U.S. EPA. 1990. Basics of pump-and-treat remediation technol-
ogy. EPA/600/8-90/003. Report prepared by Geo Trans Inc.,
Herndon, VA.
10. U.S. EPA. 1990. A new approach and methodologies for charac-
terizing 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. Guven.
1989. The impeller meter for measuring aquifer permeability vari-
ations: evaluations and comparison with other tests. Water Re-
sour. 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 aqui-
fer 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 thermal-
pulse flowmeter in the hydraulic characterization of fractured
rock. ASTM STP 1101. American Society for Testing and Materi-
als, Philadelphia, PA. pp. 99-112.
36
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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 sta-
bilization of contaminants to effectively reduce con-
taminant 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 charac-
terization 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 collec-
tion 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 ali-
phatic hydrocarbons. In some cases, the information
collected using this protocol will show that natural at-
tenuation processes, with or without source removal, will
reduce the concentrations of these contaminants to be-
low risk-based corrective action criteria or regulatory
standards before potential receptor exposure pathways
are completed. The evaluation should include consid-
eration of existing exposure pathways as well as expo-
sure pathways arising from potential future use of the
ground water.
This protocol is intended to be used within the estab-
lished regulatory framework. It is not the intent of this
document to replace existing EPA or state-specific guid-
ance 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 at-
tenuation mechanisms include sorption, dispersion, di-
lution from recharge, and volatilization. The natural
attenuation of fuel hydrocarbons is described in the
Technical Protocol for Implementing Intrinsic Remedia-
tion 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
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Excellence (AFCEE) (1). This document differs from the
technical protocol for intrinsic remediation of fuel hydro-
carbons because the individual processes of chlorinated
aliphatic hydrocarbon biodegradation are fundamentally
different from the processes involved in the biodegrada-
tion of fuel hydrocarbons.
For example, biodegradation of fuel hydrocarbons, es-
pecially benzene, toluene, ethylbenzene, and xylenes
(BTEX), is mainly limited by electron acceptor availabil-
ity, 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 proc-
ess that requires both electron acceptors (the chlorin-
ated aliphatic hydrocarbons) and an adequate supply of
electron donors. Electron donors include fuel hydrocar-
bons or other types of anthropogenic carbon (e.g., land-
fill leachate, BTEX, or natural organic carbon). If the
subsurface environment is depleted of electron donors
before the chlorinated aliphatic hydrocarbons are re-
moved, reductive dechlorination will cease, and natural
attenuation may no longer be protective of human health
and the environment. This is the most significant differ-
ence between the processes of fuel hydrocarbon and
chlorinated aliphatic hydrocarbon biodegradation.
Forthis 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, disper-
sion, dilution from recharge, and sorption, it is necessary
to better quantify biodegradation. This requires a thor-
ough understanding of the interactions between chlorin-
ated 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 sol-
vent mixtures that are more dense than water, or 2) as
mixtures of fuel hydrocarbons and chlorinated aliphatic
hydrocarbons which, depending on the relative propor-
tion 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 com-
plete remediation of the site. At many sites, however,
mobile NAPL removal is not feasible with available tech-
nology. In fact, the quantity of NAPL recovered by com-
monly 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 di-
oxide, ethene, and water), not just transferred to an-
other phase or location in the environment.
• Natural attenuation is nonintrusive and allows con-
tinuing 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 avail-
able remedial technologies, such as pump-and-treat.
• Natural attenuation is not subject to the limitations of
mechanized remediation equipment (e.g., no equip-
ment 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 anthro-
pogenic changes in local hydrogeologic conditions,
including changes in ground-water gradients and ve-
locity, pH, electron acceptor concentrations, electron
donor concentrations, and/or potential future con-
taminant releases.
• Aquifer heterogeneity may complicate site charac-
terization and quantification of natural attenuation.
• Time frames for complete remediation may be rela-
tively long.
38
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• 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 ac-
tivities that may be performed to support a feasibility
study to include an evaluation of natural attenuation,
natural attenuation modeling using analytical or numeri-
cal solute fate-and-transport models, and the post-
modeling activities that should be completed to ensure
successful support and verification of natural attenu-
ation. The objective of the work described herein is to
quantify and provide defensible data in support of natu-
ral attenuation at sites where naturally occurring subsur-
face 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 commu-
nity (2) summarizes what is required to successfully
implement natural attenuation:
A regulator looks for the data necessary to deter-
mine that a proposed treatment technology, if prop-
erly 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 pro-
ponent 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 attenu-
ation 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, in-
creasing daughter compound concentrations, de-
pletion 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 bio-
logical decay rates at the field scale.
• Microbiological laboratory data that support the oc-
currence of biodegradation and give rates of biode-
gradation.
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 down-
gradient from the NAPL source area along the ground-
water flow path. This line of evidence does not prove
that contaminants are being destroyed because the re-
duction in contaminant concentration could be the result
of advection, dispersion, dilution from recharge, sorp-
tion, 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 concen-
trations of some contaminants, most notably degrada-
tion 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 chemi-
cal 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 cal-
culations 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 do-
nor concentrations in ground water are sufficient to
facilitate degradation of dissolved contaminants. Sol-
ute 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 reduc-
tion 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 contami-
nants 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 scien-
tifically document the occurrence and quantify the rates
of natural attenuation. Ideally, the first two lines of evidence
39
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should be used in the natural attenuation demonstration.
To further document natural attenuation, or at sites with
complex hydrogeology, obtaining a field-scale biodegra-
dation rate may not be possible; in this case, microbi-
ological laboratory data can be used. Such a
"weight-of-evidence" approach will greatly increase the
likelihood of successfully implementing natural attenu-
ation 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 charac-
terization should provide data on the location, nature,
and extent of contaminant sources. Contaminant sour-
ces generally consist of hydrocarbons present as mobile
NAPL (i.e., NAPL occurring at sufficiently high satura-
tions 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 infor-
mation 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 attenu-
ation. 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, disper-
sion, sorption, and biodegradation; 2) to assess the po-
tential 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 ef-
fort, 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 negotiat-
ing 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 Divi-
sion; the Bioremediation Research Team at EPAs Na-
tional Risk Management Research Laboratory in Ada,
Oklahoma (NRMRL), Subsurface Protection and Reme-
diation Division; and Parsons Engineering Science, Inc.
(Parsons ES). This compilation is designed to facilitate
implementation of natural attenuation at chlorinated ali-
phatic 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 biodegra-
dation is important to obtain when determining whether
ground-water contamination presents a substantial
threat to human health and the environment. This infor-
mation 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 micro-
organisms 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 fortui-
tous and there is no benefit to the microorganism. At a
given site, one or all of these processes may be operat-
ing, although at many sites the use of chlorinated ali-
phatic 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 hydrocar-
bons 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 elec-
tron 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 elec-
tron 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 oxidation-
reduction 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 biodegrada-
tion of the more highly chlorinated solvents is reductive
dechlorination. During this process, the chlorinated hy-
drocarbon 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, c/s-1,2-di-
chloroethene is a more common intermediate than
frans-1,2-dichloroethene, and that 1,1-dichloroethene is
the least prevalent intermediate of the three dichlo-
roethene isomers. Reductive dechlorination of chlorin-
ated 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, perchlo-
roethene is the most susceptible to reductive dechlori-
nation because it is the most oxidized. Conversely, vinyl
chloride is the least susceptible to reductive dechlorina-
tion because it is the least oxidized of these compounds.
The rate of reductive dechlorination also has been ob-
served to decrease as the degree of chlorination de-
creases (24, 25). Murray and Richardson (26) have
postulated that this rate decrease may explain the ac-
cumulation 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 methano-
genic conditions (24). Because chlorinated aliphatic hy-
drocarbon compounds are used as electron acceptors
during reductive dechlorination, there must be an appro-
priate source of carbon in order for microbial growth to
occur (24). Potential carbon sources include natural
organic matter, fuel hydrocarbons, or other organic com-
pounds 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 sub-
strate (i.e., electron donor). Under aerobic and some
anaerobic conditions, the less-oxidized chlorinated ali-
phatic hydrocarbons (e.g., vinyl chloride) can be used as
the primary substrate in biologically mediated redox re-
actions (22). In this type of reaction, the facilitating micro-
organism 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 investi-
gations in which vinyl chloride and 1,2-dichloroethane
were shown to serve as primary substrates under aero-
bic conditions. These authors also document that dichlo-
romethane has the potential to function as a primary
substrate under either aerobic or anaerobic environ-
ments. In addition, Bradley and Chapelle (27) show
evidence of mineralization of vinyl chloride under iron-
reducing conditions so long as there is sufficient
bioavailable iron(lll). Aerobic metabolism of vinyl chlo-
ride 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 biode-
graded 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 environ-
ments, although it could occur under anaerobic condi-
tions. It has been reported that under aerobic conditions
41
-------
chlorinated ethenes, with the exception of perchlo-
roethene, are susceptible to co-metabolic degradation
(22, 23, 26). Vogel (23) further elaborates that the co-
metabolism rate increases as the degree of dechlorina-
tion 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 degra-
dation of BTEX or other carbon sources, nor will its co-me-
tabolism 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 differ-
ent 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 dechlori-
nation. 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(lll), and sulfate)?
3. Is vinyl chloride oxidized, or is it reduced?
Type 1 behavior results in the rapid and extensive deg-
radation of the highly chlorinated solvents such as per-
chloroethene, trichloroethene, and dichloroethene.
Type 2 Behavior
Type 2 behavior dominates in areas that are charac-
terized by relatively high concentrations of biologically
available native organic carbon. This natural carbon
source drives reductive dechlorination (i.e., is the pri-
mary substrate for microorganism growth). When evalu-
ating 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 chlorin-
ated 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 charac-
terized by low concentrations of native and/or anthropo-
genic 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 dichlo-
roethene. The most significant natural attenuation
mechanisms for these compounds is advection, disper-
sion, and sorption. However, vinyl chloride can be rap-
idly 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 chlori-
nated 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 be-
havior) either aerobically or via iron reduction. Vinyl
chloride is oxidized to carbon dioxide in this type of
plume and does not accumulate. The following se-
quence 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-re-
ducing conditions.
A less desirable scenario—but one in which all contami-
nants may be entirely biodegraded— involves a plume
in which all chlorinated aliphatic hydrocarbons are re-
ductively 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 de-
grades 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 investi-
gation is to show that natural processes of contaminant
degradation will reduce contaminant concentrations in
ground waterto below risk-based corrective action or regu-
latory 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 inher-
ent uncertainty associated with such predictions, the
investigator must provide sufficient evidence to demon-
strate 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 in-
put 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 reduc-
tion and removal of the dissolved contaminant plume.
Figure 1 outlines the steps involved in the natural at-
tenuation 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 sol-
ute transport and transformation processes. Quantifica-
tion of contaminant migration and attenuation rates and
successful implementation of the natural attenuation re-
medial 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-path way 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 forthe 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, hydrologi-
cal, climatological, and analytical data for the site. This
type of conceptual model differs from the conceptual site
models that risk assessors commonly use that qualita-
tively consider the location of contaminant sources, re-
lease mechanisms, transport pathways, exposure
points, and receptors. The ground-water system con-
ceptual 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 addi-
tional data collection points (as necessary) to aid in the
natural attenuation investigation and to develop the sol-
ute 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 de-
signed 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 natu-
ral 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
-------
Review Available Site Data and
Develop Preliminary Conceptual Model
Collect More Screening Data
Screen the Site using the Procedure
Presented in Figure 3
Are
Sufficient Data
Available to Properly
Screen the Site?
Are
Screening Criteria
Met?
Engineered Remediation Required
Implement Other Protocols
Perform Site Characterization
to Support Remedy Decision Making
Does it
Appear That
Natural Attenuation Alone
Will Meet Regulate
Criteria?
Evaluate Use of
Selected Additional
Remedial Options
Along with
Natural Attenuation
Assess Potential For
Natural Attenuation
With Remediation
System Installed
Perform Site Characterization
to Support Natural Attenuation
Refine Conceptual Model and
Complete Pre-Modeling
Calculations
Refine Conceptual Model and
Complete Pre-Modeling
Calculations
Simulate Natural Attenuation
Combined with Remedial
Option Selected Above
Using Solute Transport Models
Simulate Natural Attenuation
Using Solute Fate and
Transport Models
Initiate Verification of
Natural Attenuation
using Long-Term Monitoring
Initiate Verification of
Natural Attenuation
using Long-Term Monitoring
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
Use Results of Modeling and
Site-Specific Information in an
Exposure Pathways Analysis
Does
Revised Remediation
Strategy Meet Remediation
Objectives Without Posing
Unacceptable Risks
To Potential
Receptors ?
Figure 1. Natural attenuation of chlorinated solvents flow chart.
44
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Table 1A. Soil and Ground-Water Analytical Protocol3
Recommended Sample Volume, Field or
Frequency of Sample Container, Fixed-Base
Matrix
Soil
Soil
Soil
gas
Soil
gas
Water
Water
Analysis
Volatile
organic
compounds
Total
organic
carbon
(TOO)
02, C02
Fuel and
chlorinated
volatile
organic
compounds
Volatile
organic
compounds
Poly cyclic
aromatic
hydro-
carbons
(PAHs)
(optional;
intended
for diesel
and other
heavy oils)
Method/Reference b'e
SW8260A
SW9060, modified
for soil samples
Field soil gas
analyzer
EPA Method
TO-14
SW8260A
Gas chromatography/
mass spectroscopy
Method SW8270B;
high-performance
liquid chromatography
Method SW831 0
Comments''3
Handbook
method
modified for
field extraction
of soil using
methanol
Procedure
must be
accurate over
the range of
0.5 to 1 5%
TOO
Handbook
method;
analysis may
be extended to
higher
molecular-
weight alkyl
benzenes
Analysis
needed only
when required
for regulatory
compliance
Data Use
Useful for determining
the extent of soil
contamination, the
contaminant mass
present, and the need
for source removal
The amount of TOC
in the aquifer matrix
influences
contaminant migration
and biodegradation
Useful for determining
bioactivity in the
vadose zone
Useful for determining
the distribution of
chlorinated and BTEX
compounds in soil
Method of analysis for
BTEX and chlorinated
so Ive nts/by prod ucts
PAHs are components
of fuel and are
typically analyzed for
regulatory compliance
Analysis
Each soil
sampling round
At initial
sampling
At initial
sampling and
respiration
testing
At initial
sampling
Each sampling
round
As required by
regulations
Sample Preservation
Collect 100 g of soil
in a glass container
with Teflon-lined cap;
cool to 4°C
Collect 100 g of soil
in a glass container
with Teflon-lined cap;
cool to 4°C
Reuseable 3-L
Tedlar bags
1-L Summa canister
Collect water
samples in a 40-mL
volatile organic
analysis vial; cool to
4°C; add hydrochloric
acid to pH 2
Collect 1 L of water
in a glass container;
cool to 4°C
Laboratory
Fixed-base
Fixed-base
Field
Fixed-base
Fixed-base
Fixed-base
Water Oxygen
DO meter
Water Nitrate
Water Iron(ll)
(Fe+2)
Iron chromatography
Method E300; anion
method
Colorimetric HACH
Method 8146
Refer to
Method A4500
for a
comparable
laboratory
procedure
Method E300
is a handbook
method; also
provides
chloride data
Filter if turbid
Concentrations less
than 1 mg/L generally
indicate an anaerobic
pathway
Substrate for microbial
respiration if oxygen
is depleted
May indicate an
anaerobic degradation
process due to
depletion of oxygen,
nitrate, and
manganese
Each sampling
round
Each sampling
round
Each sampling
round
Measure DO on site
using a flow-through
cell
Collect up to 40 ml
of water in a glass or
plastic container; add
H2SO4 to pH less
than 2; cool to 4°C
Collect 100 ml of
water in a glass
container
Field
Fixed-base
Field
45
-------
Table 1A. Soil and Ground-Water Analytical Protocol3 (Continued)
Matrix Analysis Method/Reference13"6 Comments''3 Data Use
Recommended Sample Volume, Field or
Frequency of Sample Container, Fixed-Base
Analysis Sample Preservation Laboratory
Water
Water
Water
Water
Water
Water
Water
Water
Sulfate
(S04-2)
Methane,
ethane,
and ethene
Alkalinity
Oxidation-
reduction
potential
PH
Temperature
Conductivity
Chloride
Iron chromatography
Method E300 or
HACH Method 8051
Kampbell et al. (35)
or SW3810, modified
HACH alkalinity test
kit Model AL AP MG-L
A2580B
Field probe with
direct reading meter
Field probe with
direct reading meter
E120.1/SW9050,
direct reading meter
Mercuric nitrate
titration A4500-CI' C
Method E300
is a handbook
method, HACH
Method 8051
is a
colorimetric
method; use
one or the
other
Method
published by
EPA
researchers
Phenolphtalein
method
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
Field
Field only
Protocols/
Handbook
methods
Ion
chromatography
Method E300;
Method
SW9050 may
also be used
Substrate for
anaerobic microbial
respiration
The presence of CH4
suggests
biodegradation of
organic carbon via
methanogensis;
ethane and ethane
are produced during
reductive
dechlorination
Water quality
parameter used to
measure the buffering
capacity of ground
water; can be used to
estimate the amount
of CO2 produced
during biodegradation
The oxidation-
reduction 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
Aerobic and
anaerobic processes
are pH-sensitive
Well development
Water quality
parameter used as a
marker to verify that
site samples are
obtained from the
same ground-water
system
Final product of
chlorinated solvent
reduction; can be
used to estimate
dilution in calculation
of rate constant
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Each sampling
round
Collect up to 40 ml
of water in a glass or
plastic container; cool
to 4°C
Collect water
samples in 50 ml
glass serum bottles
with butyl
gray/Teflon-lined
caps; add H2SO4 to
pH less than 2; cool
to4°C
Collect 100 ml of
water in glass
container
Collect 100 to
250 ml of water
in a glass container
Collect 100 to
250 ml of water
in a glass or plastic
container; analyze
immediately
Not applicable
Collect 100 to 250
ml of water in a
glass or plastic
container
Collect 250 ml of
water in a glass
container
E300 =
Fixed-base
HACH
Method
8051 = Field
Fixed-base
Field
Field
Field
Field
Field
Fixed-base
46
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Table 1A. Soil and Ground-Water Analytical Protocol3 (Continued)
Matrix
Water
Water
Analysis
Chloride
(optional;
see data
use)
Total
organic
carbon
Method/Reference b'e
HACH chloride test
kit Model 8-P
SW9060
Comments''3
Silver nitrate
titration
Laboratory
Data Use
As above, and to
guide selection of
additional data points
in real time while in
the field
Used to classify
plumes and to
determine whether
Recommended
Frequency of
Analysis
Each sampling
round
Each sampling
round
Sample Volume,
Sample Container,
Sample Preservation
Collect 100 ml of
water in a glass
container
Collect 100 ml of
water in a glass
container; cool
Field or
Fixed -Base
Laboratory
Field
Laboratory
anaerobic metabolism
of chlorinated solvents
is possible in the
absence of
anthropogenic carbon
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).
'"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 1B. Soil and Ground-Water Analytical Protocol: Special Analyses Under Development and/or Consideration3 b
Recommended Sample Volume, Field or
Frequency Container, Fixed-Base
Matrix
Soil
Water
Water
Water
Analysis
Biologically
available iron(lll)
Nutritional
quality of native
organic matter
Hydrogen (H2)
Oxygenates
(including
methyl-ferf-butyl
ether, ethers,
acetic acid,
methanol, and
acetone)
Method/Reference
Under development
Under development
Equilibration with
gas in the field;
determined with a
reducing gas
detector
SW8260/801 5C
Comments
HCI
extraction
followed by
quantification
of released
iron(lll)
Spectro-
photometric
method
Specialized
analysis
Laboratory
Data Use
To predict the
possible extent of
iron reduction in
an aquifer
To determine the
extent of reductive
dechlorination
allowed by the
supply of electron
donor
To determine the
terminal electron
accepting process;
predicts the
possibility for
reductive
dechlorination
Contaminant or
electron donors
for dechlorination
of solvents
of Analysis
One round of
sampling in
five borings,
five cores
from each
boring
One round of
sampling in
two to five
wells
One round of
sampling
At least one
sampling
round or as
determined
by regulators
Preservation
Collect minimum
1-inch diameter
core samples into
a plastic liner; cap
and prevent
aeration
Collect 1,000 mL
in an amber glass
container
Sampling at well
head requires the
production of 100
mL per minute of
water for 30
minutes
Collect 1 L of
water in a glass
container;
preserve with HCI
Laboratory
Laboratory
Laboratory
Field
Laboratory
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 A ttenuation
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 expo-
sure pathways, or from reaching a predetermined point
of compliance, in concentrations above applicable regu-
latory or risk-based corrective action standards. Deter-
mining the likelihood of exposure pathway completion is
an important component of the natural attenuation in-
vestigation. 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 or-
ganic carbon (TOC) content; and the location of the
plume and contaminant source relative to potential re-
ceptors (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 effec-
tiveness 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 charac-
terization activities in support of this remedial option
should be performed. If exposure pathways have al-
ready been completed and contaminant concentrations
exceed regulatory levels, or if such completion is likely,
other remedial measures should be considered, possi-
bly 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 exam-
ple, 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 attenu-
ation is likely to be a viable remedial alternative before
additional time and money are expended. The data re-
quired to make the preliminary assessment of natural
attenuation can also be used to aid the design of an
engineered remedial solution, should the screening proc-
ess 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 collec-
tion 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 sum-
marize 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. Bio-
degradation 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
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Analyze Available Site Data
to Determine if Biodegradation
is Occurring
Biodegradation
Occurring?
H
Collect More Screening Data
Yes
Determine Groundwater Flow and
Solute Transport Parameters using
Site-Specific Data; Porosity and
Dispersivity May be Estimated
Are
Sufficient Data
Available ?
Engineered
Remediation Required,
Implement Other
Protocols
Locate Source(s)
and Receptor(s)
Estimate Biodegradation
Rate Constant
Compare the Rate of Transport
to the Rate of Attenuation using
Analytical Solute Transport Model
Are
Screening Criteria
Met?
Does it
Appear that Natural
Attenuation Alone will Meet
Regulatory Criteria?
Evaluate use of Selected
Additional Remedial Options
along with Natural Attenuation
Proceed to
Figure 1
Perform Site Characterization
to Support Natural Attenuation
Proceed to
Figure 1
Figure 2. Initial screening process flow chart.
49
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Table 2. Analytical Parameters and Weighting for Preliminary Screening
Concentration in Most
Analyte Contaminated Zone Interpretation
Oxygen3
Oxygen3
Nitrate3
Iron (II)3
Sulfate3
Sulfide3
Methane3
Oxidation reduction
potential3
pH3
DOC
Temperature3
Carbon dioxide
Alkalinity
Chloride3
Hydrogen
Hydrogen
Volatile fatty acids
BTEX3
Perchloroethene3
Trichloroethene3
Dichloroethene3
Vinyl chloride3
Ethene/Ethane
Chloroethane3
< 0.5 mg/L
> 1 mg/L
< 1 mg/L
> 1 mg/L
< 20 mg/L
> 1 mg/L
> 0.1 mg/L
> 1
< 1
< 50 mV against Ag/AgCI
5 20 mg/L
>20°C
> 2x background
> 2x background
> 2x background
> 1 nM
< 1 nM
> 0.1 mg/L
> 0.1 mg/L
< 0.1 mg/L
Tolerated; suppresses reductive dechlorination at higher
concentrations
Vinyl chloride may be oxidized aerobically, but reductive
dechlorination will not occur
May compete with reductive pathway at higher
concentrations
Reductive pathway possible
May compete with reductive pathway at higher
concentrations
Reductive pathway possible
Ultimate reductive daughter product
Vinyl chloride accumulates
Vinyl chloride oxidizes
Reductive pathway possible
Tolerated range for reductive pathway
Carbon and energy source; drives dechlorination; can be
natural or anthropogenic
At T > 20EC, biochemical process is accelerated
Ultimate oxidative daughter product
Results from interaction of carbon dioxide with aquifer
minerals
Daughter product of organic chlorine; compare chloride
in plume to background conditions
Reductive pathway possible; vinyl chloride may
accumulate
Vinyl chloride oxidized
Intermediates resulting from biodegradation of aromatic
compounds; carbon and energy source
Carbon and energy source; drives dechlorination
Material released
Material released or daughter product of perchloroethene
Material released or daughter product of trichloroethene;
if amount of c/s-1 ,2-dichloroethene is greater than 80%
of total dichloroethene, it is likely a daughter product of
trichloroethene
Material released or daughter product of dichloroethenes
Daughter product of vinyl chloride/ethene
Daughter product of vinyl chloride under reducing
Points
Awarded
3
-3
2
3
2
3
2
3
< 50 mV = 1
<-100 mV = 2
2
1
1
1
2
3
2
2
2b
2b
2b
> 0.01 mg/L= 2
>0.1 =3
2
conditions
1,1,1-Trichloroethane3 Material released
1,1-dichloroethene3 Daughter product of trichloroethene or chemical reaction
of 1,1,1 -trichloroethane
3 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
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... Helps Define
®^ Lateral Extent
F of Contamination
Helps Define
Downgradierit Extent
of Contamination
Direction of
Plume Migration
^Dissolved
Contaminant
Plume
® Required Data Collection Point
Not To Scale
Figure 3. Data collection points required for screening.
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 parame-
ters 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 deter-
mination of the dominant terminal electron-accepting
processes at the site. In conjunction with samples col-
lected 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 do-
nors and metabolic byproducts might be along the flow
path. The sample collected downgradient from the dis-
solved 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 con-
centrations of the electron acceptors and donors.
After these samples have been analyzed for the pa-
rameters listed in Table 2, the investigator should ana-
lyze 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 exam-
ple, 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.
Table 3. Interpretation of Points Awarded During Screening Step 1
Score Interpretation
0 to 5
6 to 14
15 to 20
>20
Inadequate evidence for biodegradation
of chlorinated organics
Limited evidence for biodegradation of
chlorinated organics
Adequate evidence for biodegradation of
chlorinated organics
Strong evidence for biodegradation of
chlorinated organics
Consider the following two examples. Example 1 con-
tains 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.
Example 1. Strong Evidence for Biodegradation of
Chlorinated Organics
Analyte
DO
Nitrate
Iron(ll)
Sulfate
Methane
Oxidation-reduction
potential
Chloride
Perchloroethene
(released)
Trichloroethene
(none released)
c/s-1 ,2-Dichloroethene
(none released)
Vinyl chloride
(none released)
Concentration in Most
Contaminated Zone
0.1 mg/L
0.3 mg/L
10 mg/L
2 mg/L
5 mg/L
-190mV
3x background
1 ,000 ng/L
1 ,200 ng/L
500 ng/L
50 ng/L
Total points awarded
Points
Awarded
3
2
3
2
3
2
2
0
2
2
2
23
51
-------
In this example, the investigator can infer that biodegra-
dation is occurring and may proceed to Step 2.
Example 2. Biodegradation of Chlorinated Organics Unlikely
Analyte
DO
Nitrate
Iron(ll)
Sulfate
Methane
Oxidation-reduction
potential
Chloride
Trichloroethene
(released)
c/s-1 ,2-Dichloroethene
Vinyl chloride
Concentration in Most
Contaminated Zone
3 mg/L
0.3 mg/L
Not detected
1 0 mg/L
ND
100 mV
Background
1 ,200 ng/L
Not detected
ND
Total points awarded
Points
Awarded
-3
2
0
2
0
0
0
0
0
0
1
In this example, the investigator can infer that biodegra-
dation is probably not occurring or is occurring too slowly
to be a viable remedial option. In this case, the investi-
gator 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 investi-
gator must have site-specific hydraulic gradient and hy-
draulic 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 lit-
erature 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 model-
ing 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 po-
tential downgradient or cross-gradient receptors.
Step 4: Estimate the Biodegradation Rate
Constant
Biodegradation is the most important process that de-
grades contaminants in the subsurface; therefore, the
biodegradation rate is one of the most important model
input parameters. Biodegradation of chlorinated ali-
phatic hydrocarbons can commonly be represented as
a first-order rate constant. Site-specific biodegradation
rates generally are best to use. Calculation of site-spe-
cific biodegradation rates is discussed by Wiedemeier
et al. (1, 36, 37). If determining site-specific biodegrada-
tion 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 biodegra-
dation 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 demonstra-
tion, 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 advection-
dispersion equation (38), and allows use of either a
first-order biodegradation rate or an instantaneous reac-
tion between contaminants and electron acceptors to
simulate the effects of biodegradation. To model trans-
port 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
52
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residence time along the flow path is adequate to be
protective of human health and the environment (i.e., to
qualitatively estimate whetherthe contaminant is attenu-
ating at a rate fast enough to allow degradation of the
contaminant to acceptable concentrations before recep-
tors are reached). It is important to perform a sensitivity
analysis to help evaluate the confidence in the prelimi-
nary screening modeling effort. If modeling shows that
receptors may not be exposed to contaminants at con-
centrations 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 hydrau-
lic conductivity and hydraulic gradient, as well as esti-
mates 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 be-
fore 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 envi-
ronment. The second is to provide sufficient site-specific
data to allow prediction of the future extent and concen-
tration of a contaminant plume through solute fate-and-
transport 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 charac-
terization 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, agricul-
tural, and domestic water wells; patterns of aquifer
use (current and future); lithology; site stratigraphy,
including identification of transmissive and nontrans-
missive units; grain-size distribution (sand versus silt
versus clay); aquifer hydraulic conductivity; ground-
water 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 charac-
terization in support of natural attenuation. Additional infor-
mation can be obtained from Wiedemeier et al. (1, 37).
Soil Characterization
To adequately define the subsurface hydrogeologic sys-
tem 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, ex-
tensive soil characterization must be completed. De-
pending 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 subsur-
face distribution of hydrostratigraphic units and the dis-
tribution 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 col-
lected, described, analyzed, and disposed of in accord-
ance 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
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The analytical protocol to be used for soil sample analy-
sis 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 regu-
latory 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 sam-
ples should be collected from a background location in
the stratigraphic horizon(s) where most contaminant
transport is expected to occur. Oxygen and carbon di-
oxide 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-dimen-
sional 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 quantita-
tive evaluation of natural attenuation's importance at a
site are possible.
Ground-water sampling is conducted to determine the
concentrations and distribution of contaminants, daugh-
ter products, and ground-water geochemical parame-
ters. Ground-water samples may be obtained from
monitoring wells or with point-source sampling devices
such as a Geoprobe, Hydropunch, or cone penetrome-
ter. All ground-water samples should be collected in
accordance with local, state, and federal guidelines.
Wiedemeier et al. (1) suggest procedures for ground-
water 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 analy-
sis is presented in Table 1. This analytical protocol in-
cludes 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 pa-
rameter and the use of each analyte in the natural
attenuation demonstration.
Volatile organic compound analysis (by Method
SW8260a) is used to determine the types, concentra-
tions, and distributions of contaminants and daughter
products in the aquifer. DO is the electron acceptor most
thermodynamically favored by microbes for the biode-
gradation of organic carbon, whether natural or anthro-
pogenic. 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, an-
aerobic microbes will use nitrate as an electron ac-
ceptor, followed by iron(lll), then sulfate, and finally
carbon dioxide (methanogenesis). Each sequential re-
action drives the oxidation-reduction potential of the
ground water further into the realm where reductive
dechlorination can occur. The oxidation-reduction po-
tential range of sulfate reduction and methanogenesis is
optimal, but reductive dechlorination may occur under
nitrate- and iron(lll)-reducing conditions as well. Be-
cause reductive dechlorination works best in the sulfate-
reduction and methanogenesis oxidation-reduction
potential range, competitive exclusion between micro-
bial sulfate reducers, methanogens, and reductive
dechlorinators can occur.
After DO has been depleted in the microbiological treat-
ment zone, nitrate may be used as an electron acceptor
for anaerobic biodegradation via denitrification. In some
cases iron(lll) is used as an electron acceptor during
anaerobic biodegradation of electron donors. During this
process, iron(lll) is reduced to iron(ll), which may be
soluble in water. Iron(ll) concentrations can thus be used
as an indicator of anaerobic degradation of fuel com-
pounds. After DO, nitrate, and bioavailable iron(lll) have
been depleted in the microbiological treatment zone,
sulfate may be used as an electron acceptor for anaero-
bic biodegradation. This process is termed sulfate re-
duction and results in the production of sulfide. During
methanogenesis (an anaerobic biodegradation proc-
ess), carbon dioxide (or acetate) is used as an electron
acceptor, and methane is produced. Methanogenesis
generally occurs after oxygen, nitrate, bioavailable
iron(lll), 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, po-
54
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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 dur-
ing 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 micro-
bially 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 dif-
ferent 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 biodegrada-
tion. 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-re-
duction potential measurements should be taken during
well purging and immediately before and after sample
Redox Potential (Eh°)
in Millivolts @ pH = 7
and J=25rC
I
I
S3
£
•o
1
UL
I
HI
•5
1000 -r
Aerobic
Possible Range
for Reductive—
Dechlorination
Anaerobic
500-K
0 --
Optimal Range
for Reductive
Dechlorination
I
-500 -*-
I + HCO; + 3H* + 2e
(Eh° = + 740)
MnCOs(s) + 2H3O
FeOOH(s) + HCOS + 2H* + e
(Eh° = -50)
(Eh° = -220)
(Eh' = -240)
Modified From Bouwer (1994)
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-watersamples (which can affect oxidation-reduction
potential measurements), it is important to minimize
potential aeration.
Dissolved hydrogen concentrations can be used to de-
termine the dominant terminal electron-accepting proc-
ess 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(lll)-re-
ducing, or denitrifying conditions, it is helpful to have
hydrogen concentrations to help delineate redox condi-
tions when evaluating the potential for natural attenu-
ation of chlorinated ethenes in ground-water systems.
Collection and analysis of ground-water samples for
Table 4. Range of Hydrogen Concentrations for a Given
Terminal Electron-Accepting Process
Terminal
Electron-Accepting Process
Hydrogen Concentration
(nanomoles per liter)
Denitrifi cation
Iron(lll) reduction
Sulfate reduction
Methanogenesis
0.2 to 0.8
1 to 4
>5
dissolved hydrogen content is not yet commonplace or
standardized, however, and requires a relatively expen-
sive 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 meas-
ured 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 capa-
ble 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 tem-
perature 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"i0 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 repre-
sentative of the water in the saturated zone containing
the dissolved contamination. If the conductivities of
samples taken from different sampling points are radi-
cally different, the waters may be from different hydro-
geologic zones.
Elemental chlorine is the most abundant of the halo-
gens. 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). Chlo-
ride 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 chem-
istry of fresh water (45). The chemical behavior of chlo-
ride is neutral. Chloride ions generally do not enter into
oxidation-reduction reactions, form no important solute
complexes with other ions unless the chloride concen-
tration is extremely high, do not form salts of low solu-
bility, 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 hydrocar-
bons 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. Be-
cause of the neutral chemical behavior of chloride, it can
be used as a conservative tracer to estimate biodegra-
dation 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 im-
pact on contaminant fate and transport include hydraulic
conductivity, hydraulic gradient, porosity, and dispersiv-
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 hy-
draulic conductivity and hydraulic gradient and on litera-
ture values for porosity and dispersivity for the types of
sediments present at the site. Methods for field meas-
urement of aquifer hydraulic parameters are described
by Wiedemeier et al. (1, 37).
Microbiological Laboratory Data
Microcosm studies are used to show that the microor-
ganisms necessary for biodegradation are present and
to help quantify rates of biodegradation. If properly de-
signed, implemented, and interpreted, microcosm stud-
ies can provide very convincing documentation of the
occurrence of biodegradation. Such studies are the only
"line of evidence" that allows an unequivocal mass bal-
ance determination based on the biodegradation of en-
vironmental 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 un-
dertaken only at sites where there is considerable skep-
ticism concerning the biodegradation of contaminants.
Biodegradation rate constants determined by micro-
cosm 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, sorp-
tion, dilution, and biodegradation. The results of these
calculations are used to scientifically document the occur-
rence 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 prelimi-
nary 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. Con-
ceptual model refinement consists of several steps, in-
cluding preparation of geologic logs, hydrogeologic
sections, potentiometric surface/water table maps, con-
taminant contour (isopleth) maps, and electron acceptor
and metabolic byproduct contour (isopleth) maps. Re-
finement of the conceptual model is described by
Wiedemeier et al. (1).
Premodeling Calculations
Several calculations must be made prior to implementa-
tion of the solute fate-and-transport model. These cal-
culations 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 simu-
late accurately the fate and transport of contaminants
dissolved in ground water. In many cases, biodegrada-
tion of contaminants can be approximated using first-or-
der kinetics. To calculate first-order biodegradation rate
constants, the apparent degradation rate must be nor-
malized for the effects of dilution and volatilization. Two
methods for determining first-order rate constants are
described by Wiedemeier et al. (36). One method in-
volves the use of a biologically recalcitrant compound
found in the dissolved contaminant plume that can be
used as a conservative tracer. The other method, pro-
posed by Buscheck and Alcantar (47) involves interpre-
tation 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
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Simulate Natural Attenuation Using Solute
Fate-and-Transport Models
Simulating natural attenuation using a solute fate-and-
transport model allows prediction of the migration and
attenuation of the contaminant plume through time. Natu-
ral 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 biologi-
cal processes. Hence, the results of the modeling effort
are not in themselves sufficient proof that natural attenu-
ation 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 thor-
ough site characterization will improve the validity of the
modeling results. In some cases, straightforward ana-
lytical 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 influ-
ence of advection, dispersion, sorption, and biodegra-
dation. 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 docu-
mented and predictions of the future extent and concen-
trations 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 imple-
mentation of this remedial option. Supporting the natural
attenuation option generally will involve performing a
receptor exposure-pathway analysis. This analysis in-
cludes identifying potential human and ecological recep-
tors 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 in-
put parameters are used, the solute fate-and-transport
model should give conservative estimates of contami-
nant plume migration. From this information, the poten-
tial 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 re-
duce plume expansion if the exposure-pathway analysis
suggests that one or more exposure pathways may be
completed before natural attenuation can reduce chemi-
cal concentrations below risk-based levels of concern.
Further, some regulators may require source removal in
conjunction with natural attenuation. Several technolo-
gies 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 require-
ments. 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 im-
pact of source removal can readily be evaluated by
modifying the contaminant source term if a solute fate-
and-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 contami-
nant concentrations in ground water to acceptable lev-
els. Experience at 40 Air Force sites contaminated with
fuel hydrocarbons using the protocol presented by
Wiedemeier et al. (1) suggests that many fuel hydrocar-
bon plumes are relatively stable or are moving very
slowly with respect to ground-water flow. This informa-
tion 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' experi-
ence chlorinated solvent plumes are likely to migrate
further downgradient than fuel hydrocarbon plumes be-
fore 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 suffi-
cient to protect potential downgradient receptors. The
long-term monitoring plan should be developed based
on site characterization data, the results of solute fate-
and-transport modeling, and the results of the exposure-
pathway analysis.
The long-term monitoring plan includes two types of
monitoring wells: long-term monitoring wells are in-
tended 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
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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
LNAPL
Source Area.,
Plume Migration
Anaerobic Treatment Zone
Extent of Dissolved
BTEX Plume
Aerobic Treatment
Zone
LEGEND
• Point-of-Compliance Monitoring Well
O Long-Term Monitoring Well Not To Scale
Note: Complex sites may require more wells The final
number and placement should be determined in conjunction
with the appropriate regulators.
Figure 5. Hypothetical long-term monitoring strategy.
treatment, along the periphery of the plume, 5) a well
located downgradient from the plume where contami-
nant concentrations are below regulatory acceptance
levels and soluble electron acceptors are depleted with
respect to unaffected ground water, and 6) three point-
of-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 prop-
erty line. The final number and location of point-of-
compliance 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-of-
compliance wells. To provide a valid monitoring system,
all monitoring wells must be screened in the same hy-
drogeologic unit as the contaminant plume. This gener-
ally requires detailed stratigraphic correlation. To
facilitate accurate stratigraphic correlation, detailed vis-
ual 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(ll), sulfate,
and methane. For point-of-compliance wells, ground-
water analyses should be limited to determining volatile
organic compound and DO concentrations. Any state-
specific analytical requirements also should be ad-
dressed in the sampling and analysis plan to ensure that
all data required for regulatory decision-making are col-
lected. 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 rec-
ommended 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 sam-
pling in the quarter showing the greatest extent of the
plume. Sampling frequency depends on the final place-
ment 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 attenu-
ation as the most appropriate remedial option for a given
site. All available site-specific data and information de-
veloped during the site characterization, conceptual
model development, premodeling calculations, biode-
gradation rate calculation, ground-water modeling,
model documentation, and long-term monitoring plan
preparation phases of the natural attenuation investiga-
tion should be presented in a consistent and comple-
mentary 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 envi-
ronment. The regulators must be presented with a
"weight-of-evidence" argument in support of this reme-
dial 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 com-
mitment to proving the effectiveness of natural attenu-
ation as a remedial option. Because long-term
monitoring and contingency plans are very site specific,
they should be addressed in the individual reports gen-
erated using this protocol.
59
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49. Rice, D.W, R.D. Grose, J.C. Michaelsen, B.P. Dooher, D.H. Mac-
Queen, 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 Con-
trol Board.
61
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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 simu-
lating the natural attenuation of dissolved hydrocarbons
at petroleum fuel release sites. The software, pro-
grammed in the Microsoft Excel spreadsheet environ-
ment 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 attenu-
ation 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 com-
pleted. Some data, such as electron acceptor concen-
trations, 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 oc-
curring at a site. In addition, BIOSCREEN is an excel-
lent 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 reme-
diation 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 Super-
fund 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 remedia-
tion study.
BIOSCREEN Input and Output
To run BIOSCREEN, the user enters site data in the
following categories: hydrogeologic, dispersion, adsorp-
tion, biodegradation, general information, source char-
acteristics, and observed data. For several parameters
(e.g., seepage velocity), the user can either enter the
value directly or use supporting data (hydraulic conduc-
tivity, hydraulic gradient, and effective porosity) to calcu-
late 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 reac-
tion 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 simu-
late intrinsic remediation at petroleum release sites,
the system can be modified to simulate intrinsic reme-
diation of chlorinated hydrocarbons. Current plans call
for converting the BIOSCREEN model to BIOCHLOR.
Key changes are:
• Biodegradation using first-order decay only: Micro-
bial constraints on kinetics are much more important
for chlorinated solvents than for petroleum com-
pounds. 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: Chlorin-
ated 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 gaso-
line and JP-4. The source terms will be discussed in
more detail to ensure that model input data and pre-
liminary calculations are representative of DNAPL
sites.
• Evaluation of biodegradation products: The genera-
tion of products of chlorinated solvent biodegradation
will be discussed. Simple analytical tools may be
developed and incorporated into BIOCHLOR.
BIOSCREEN is available by contacting EPAs 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
BIOSCREEN
Remediation
ftKrni.«R rtt JiJtTiin ftrtk-M*^
* Data used dtmcft/
VarOceaMana Sautca
_ - - - Section and Input Cane
itrws I. 2 and J
•1. DISPERSION
tl F>n:-Hikjbit,' ttfrt f
apni y
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tf No data ^aa/D Btnk or (Erfw TT
4. BKJOECVRADATION
a
W
Hlttlf
Fnncui Iron'
Figure 1. BIOSCREEN input screen.
63
-------
iii'-.ui vi.u •,n,-::••! \i. .;UM: UNC trmWTKJN UQNGntMiCSKnaUNI •».: i •<"
.,
in
MO
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ii rvi
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D»(incr Frora Source I fu
,i->
-
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tapU
Beowoulnte Thrt
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Figure 2. BIOSCREEN centerline output screen.
DISSOLVED HYDUOCAAiOtt CON'CtNTRATICIKSIS FLU*ietmtfL tH
Tine [
f PtolJttfrrti )
ixn • .••
Figure 3. BIOSCREEN concentration array output screen.
1*1
Pun« Man H No Btod»omi*aQfi| ITT o
- Actual njma I
« Hum Mat Hamoptfod tv I
Change In ftetticn A£.««ptaj9ypKXtx:t Mosssa
Cfrypgfi rvunjtu *m.T SuTgtp Mfttiarm
•W" ' "r ' "^
Mem H 9ouw« Now
-------
(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 un-
der 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 multi-
ple 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, Ver-
sion 1.3, U.S. Air Force Center for Environmental Excellence,
Brooks AFB, San Antonio, TX.
65
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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 contami-
nation, methanogenesis predominates. As ground water
flows downgradient, distinctsulfate-reducing, iron(lll)-re-
ducing, and oxygen-reducing zones are encountered. This
naturally occurring sequence favors the reductive dehalo-
genation 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 po-
tential 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 manufac-
turing 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 eleva-
tion 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 ground-
water 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 oc-
curring 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 an-
aerobically under differing redox conditions and to pro-
duce 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 west-
ern 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 inter-
vals 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 chro-
matography ensured that the entire width of the con-
taminant 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 geo-
probe (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 supple-
mented by Figures 6 through 9, which show contaminant
distribution with depth in single boreholes from repre-
sentative 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 accu-
rately represented, as waters of differing chemistry may
mix upon sampling. This may explain why a few data
points simultaneously have high methane and high oxy-
gen concentrations. Most importantly, the figure indi-
cates that a large number of sample locations at the site
had the necessary strong reducing conditions for reduc-
tive dechlorination to occur.
67
-------
TRANSECT 5
fs-sf
LftKE
MICHIGAN
.?*/ d
TRANSECT 4
-. (
of I / I r\ I I I TRAN
7 //////, 34
587 588 >S89 590 591 592 Sp3 / 594 _ 5,95
' ' J> l' "~~
Figure 1. St. Joseph Superfund site plan.
o
03
c:
o
-------
_g
O
2.5
2
1.5
1
0.5
0
0 100 200 300 400
Micromolar Oxygen Concentration
Figure 4. Composited chlorine number plotted against oxygen
concentration.
3
2.5
2
1.5
1
0.5
» 1.2.3
• 4.5
• lake
i 1.2.3
n 4.5
o lake
0 200 400 600 800 1000
Micromolar Methane Concentration
Figure 5. Composited chlorine number plotted against methane
concentration.
were devoid of contaminants and were oxygenated.
Sulfate concentrations in the range of 300 to 500 u,M 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, Na, that is defined by
Nc, = -
w, C,
where W| is the number of chlorine atoms in molecule i
and C| is the molar concentration of each ethene spe-
cies. The chlorine number composites the ethene con-
centrations 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, how-
ever, 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),
10"
10'
40 50 60 70 80
Depth (feet)
90
Figure 6. Distribution of chloride, sulfate, dissolved oxygen,
and methane with depth in Borehole T23.
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 corre-
sponded to methanogenic locations. The latter condi-
tion, 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 asso-
ciated 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 con-
tamination, 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
o
O
500
500
400
300
200
100
-D TCE
-o c-DCE
-4 VC
-a ETHENE
40 50 50 70
Depth (feet)
BO
90
Figure 7. Distribution of ethenes with depth at Borehole T23.
69
-------
103
101
10"
A SOI
• OXVGEN
* METHANE
"40 SO 60 70 80 90 100
Depth (feet)
Figure 8. Distribution of chloride, sulfate, dissolved oxygen,
and methane with depth in Borehole T42.
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 u,M 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 con-
centrations recorded for Transect 4, and it also repre-
sents 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 con-
centrations 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 in-
creased 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.
600
500
— 400
o
2 300
c
CD
" 200
o
100
0
40 50 60 70 80 90 100
Depth (feet)
Figure 9. Distribution of ethenes with depth in Borehole T42.
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 excep-
tion 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 asso-
ciated with declines in oxygen and sulfate concentra-
tions 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, sug-
gesting 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, Michi-
gan, Superfund site. In: U.S. EPA. Symposium on Bioremediation
of Hazardous Wastes: Research, Development, and Field Evalu-
ations. EPA/600/R-93/054. pp. 57-60.
2. McCarty, PL., 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., PK. Kitanidis, D. Kampbell, and J.T. Wilson. 1995.
Anaerobic transformation of chlorinated aliphatic hydrocarbons in
a sand aquifer based on spatial chemical distributions. Water Re-
source. 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 hy-
drogen in sediments. Appl. Environ. Microbiol. 43(6):1373-1379.
70
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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) oc-
curs 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 theo-
retical 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 con-
centrations, 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 aque-
ous phase, one would expect that contaminants re-
leased 18 years or longer ago would by now have
discharged into the lake. The observed contaminant distri-
bution 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
Table 1. Attenuation of the Chlorinated Ethenes Along the
Length of the Plume
Average Concentration (fig/L)
Highest Concentration (fig/L)
Distance
From Transport Transect
Source
(m)
130
390
550
855
Time
(y)
3.2
9.7
12.5
17.9
Width
(m)
108
150
192
395
TCE
6,500
68,000
520
8,700
15
56
<1
0.4
cis-DCE
8,100
128,000
830
9,800
18
870
<1
0.8
Vinyl
Chloride
930
4,400
450
1,660
106
205
<1
0.5
that crossed the plume (Weaver et al. a, this volume).
These range from 130 to 855 meters from the sus-
pected source of contamination. From the borings, a
three-dimensional view of the contamination was devel-
oped. Afield 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 ap-
proach 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 devel-
oped by using the SITE-3D graphics package (5). The
data were represented as sets of blocks that are cen-
tered 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
71
-------
by summing over the blocks and dividing by the area of
the transects.
In Table 2, concentration estimates are presented forthe
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 downgradi-
ent; thus, all of the rate coefficients developed below reflect
a net loss of the species. The chloride concentrations
increase downgradient as expected from the dechlori-
nation 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
Table 2. Transect-Averaged Concentrations (ug/L) From the
Anaerobic Zone
Lake
Chemical Transect 2 Transect 4 Transect 5 Transect
TCE
cis-DCE
t-DCE
1,1-DCE
VC
Ethene
Sum of the
ethenes
Chloride
7,411
9,117
716
339
998
480
19,100
65,073
864
1,453
34.4
24.3
473
297
3,150
78,505
30.1
281
5.39
2.99
97.7
24.2
442
92,023
(1.4)
(0.80)
(1.1)
blq
(0.16)
No data
(3.5)
44,41 8
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 ve-
locity 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
,3c_n #c _
'3f~ ^A2* yy: ""
dc
-re
(Eq. 1)
Note: Values in parentheses were based on one or more estimated
values; blq indicates no detection above the limit of quantitation.
where R is the retardation coefficient; c is the concen-
tration; t is time; Dxx and Dyy are the longitudinal and
transverse dispersion coefficients, respectively; x is lon-
gitudinal distance; y is the distance transverse to the
plume centerline in the horizontal plane; v is the seep-
age velocity; and X* 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 chlorin-
ated hydrocarbons (6). This form of the transport equa-
tion assumes that ground-water flow is uniform and
aligned with the axis of the plume, as observed forthe
plume. This assumption also allows application of ana-
lytic 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 appar-
ent physical dilution in aquifers. Dispersion is currently
Table 3. Flux Estimates for Transects 1, 2, 4, and 5
Mass Flux (kg/y)
Transect
1 (August-September 1991)
2 (August-September 1991)
4 (March 1992)
5 (April 1992)
Reduction ratio
TCE
50.0
117
30.9
0.95
123
cis-DCE
45.2
133
41.7
10.0
13
VC
16.8
16.8
3.87
1.68
10
Ethene
7.95
7.60
10.8
0.164
46
Total
Ethenes
125
283
88.4
13.1
22
Methane
49.2
65.7
101
46.7
Chloride
545
1,456
4,610
5,290
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 physio-
chemical and biological processes. Observed concen-
trations 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 multi-
ple 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, retarda-
tion, longitudinal and transverse dispersion, and first-
order 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 trans-
verse dispersivity on the estimated rate constants, how-
ever, 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).
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
Table 5. Net Apparent Degradation Rate Constants
(One Per Year) From the Two-Dimensional Model
(Equation 3)
Chemical
TCE
cis-DCE
Vinyl chloride
Transect
2 to 4
0.30
0.54
2.6
Transect
4 to 5
1.7
1.1
3.1
Transect
5 to Lake
1.7
4.0
20
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 proce-
dure 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, be-
cause production of a compound must be balanced by
high gross rates to attain the observed net rate.
Chemical
TCE
cis-DCE
Vinyl chloride
Transect
2 to 4
0.30
0.26
0.15
Transect
4 to 5
1.7
0.58
0.78
Transect
5 to Lake
1.7
3.3
2.6
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 contami-
nant declined to values below the respective maximum
contaminant levels when sampled from the lake sedi-
ments. 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 equa-
tions, the apparent reduction in concentration is only
accounted for by loss of mass. These apparent degra-
dation rate constants were calculated from the St.
Joseph data set through application of a two-dimen-
sional analytical solution of the transport equation. Since
transverse spreading of the plume reduces the contami-
nant 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) =
-y2
c (°o,y,f) = c (x,-oo,f) = c (x,°o,f) = 0
(Eq. 2)
73
-------
has the approximate steady state solution (9)
Net and Gross Decay Rates
c«exp [T?f~(' ~
2, D
2 +•
•U'RD,,
1 +-
i + 4rgo-
(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 up-
gradient peak concentration, c0, and the standard devia-
tion, o, 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 = 60. A mass balance can then be solved for the
peak concentration of the Gaussian distribution, c0, from
\ncdy = J
= nc0
--) dy= nc0 o
(Eq. 4)
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:
;+LTo) (Eq. 7)
where ^(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), Vi(n) 's tne 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 ^j(g) is the gross apparent decay rate of daughter
j. For the DCE isomers, fj is approximated by the aver-
age 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 concen-
tration 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 deter-
mined from the reaction rate model (8).
where n is the porosity, c is the vertically averaged concen-
tration, 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
U
yy
V-
1 cfo2
(Eq. 5)
where ayy is the transverse dispersivity. By applying Equa-
tion 5 in a discrete form and substituting A t = A xR/v, an
expression for ayy is obtained in terms of the seepage
velocity, retardation coefficient, distance between tran-
sects (Ax), and change in variance of the Gaussian
distributions for the transect concentrations (A o2):
1 Ao2
ayy~2RAx
(Eq. 6)
The only remaining unknown in Equation 3 is the decay
constant ^*, which is determined through a bisection
search. Table 5 gives the rate constants from the two-
dimensional model.
References
1. McCarty, P.L., and L. Semprini. 1994. Ground-water treatment for
chlorinated solvents. In: Morris 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, Ml: Lewis Pub-
lishers, 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, Michi-
gan, Superfund site. In: U.S. EPA. Symposium on Bioremediation
of Hazardous Wastes: Research, Development, and Field Evalu-
ations. EPA/600/R-93/054. pp. 57-60.
3. McCarty, PL., 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. Intrin-
sic 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. 8. Smith, V.J., and R.J. Charbeneau. 1990. Probabilistic soil contami-
1995. Field-derived transformation rates for modeling natural nation exposure assessment procedures. J. Environ. Engineer.
bioattenuation of trichloroethene and its degradation products. 116(6):1143-1163.
Presented at the Next Generation of Computational Models Com-
putational Methods, August 17-19, Bay City, Ml. Society of Indus-
trial and Applied Mathematics.
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), c/s-1,2-dichloroethene (c/s-1,2-DCE), vinyl chlo-
ride, and benzene, toluene, ethylbenzene, and xylenes
(BTEX). Table 1 contains contaminant data for selected
wells at the site.
Contaminant plumes formed by chlorinated aliphatic hy-
drocarbons (CAHs) dissolved in ground water can ex-
hibit 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 typi-
cally 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 con-
centrations and anaerobic conditions. Under these con-
ditions, 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 behav-
ior occurs in areas characterized by low natural organic
carbon concentrations, low anthropogenic carbon con-
centrations, 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, how-
ever, 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 (Fig-
ure 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, ap-
proximately 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 dis-
charges to a swamp (Figure 2). Ground-water flow at the
site is to the southeast, with the average gradient ap-
proximately 0.010 foot per foot (ft/ft). Hydraulic conduc-
tivity of the upper sand aquifer was measured using
constant drawdown tests and rising head tests. Hydrau-
lic conductivity values for the unconfined sand aquifer
76
-------
Table 1. Analytical Data, Plattsburgh Air Force Base
Point Date
A
B
C
D
E
F
Aug.
95
May
96
Aug.
95
May
96
Aug.
95
May
96
Aug.
95
May
96
Aug.
95
May
96
Aug.
95
May
96
Distance
From
Source 1MB
(feet) Oig/L)
0 1 ,757
828
970 491
463
1 ,240 488
509
2,050 NA
9
2,560 0
0
3,103 0
0
BTEX TCE
Uig/L) (jig/L)
16,790
6,598
3,060
4,198
3,543
3,898
NA
89
40
40
2
2
25,280
580
2
1
3
1
NA
0
24
17
1
0
Total Vinyl
DCEa Chloride
(H9/L) (jig/L)
51,412
12,626
1 4,968
9,376
10,035
10,326
NA
1,423
2,218
1,051
226
177
0
0
897
1,520
1,430
1,050
NA
524
8
12
5
4
Methane
(ng/L)
1,420
1,600
305
339
1,010
714
NA
617
3,530
1,800
115
44
Ethene Chloride
(Hg/L) (mg/L)
< 0.001
< 0.001
35.00
13.00
182.00
170.00
NA
4.00
< 0.001
< 0.001
< 0.001
< 0.001
63
82
48
43
46
57
NA
14
20
18
3
3
Dissolved
Oxygen Nitrate
(mg/L) (mg/L)
0.1
0.5
0.5
0.1
0.4
0.2
NA
0.2
0.9
0.1
0.4
0.2
0.2
0.0
0.2
0.0
0.2
0.0
NA
0.1
0.3
0.0
10.4
9.5
Iron(ll)
(mg/L)
4.0
45.6
15.3
16.0
13.8
19.3
NA
2.5
0.7
0.0
0.0
0.1
Total
Hydro- Organic
Sulfate gen Carbon
(mg/L) (nM) (mg/L)
5.5
1.0
0.0
0.0
0.0
0.0
NA
1.5
0.5
1.0
14.7
14.4
6.70
2.00
1.66
1.40
NA
11.13
NA
NA
NA
0.81
0.22
0.25
80
94
30
31
21
24
NA
14
8
8
NA
NA
Greater than 99% of DCE is c/s-1,2-DCE.
NA = Not analyzed.
Point A = MW-02-108, B = MW-02-310, C = 84DD, D = 84DF, E = 34PLTW12, F = 35PLTW13.
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 gradi-
ent 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 approxi-
mately 142 feet per year. Because of low background
total organic carbon (TOC) concentrations at the site,
retardation is not considered to be an important trans-
port 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 chlo-
ride are not present in measurable concentrations. For
the most part, ground water beneath and downgradient
from the LNAPL is contaminated with dissolved fuel-re-
lated compounds and solvents consistent with those
identified in the LNAPL. The most notable exceptions
are the presence of c/s-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 approxi-
mately 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 contamina-
tion 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 steady-
state equilibrium and is no longer expanding.
Detectable concentrations of dissolved TCE, DCE, and
vinyl chloride currently extend approximately 4,000 feet
77
-------
T
Extent of
Dissolved
Contaminant
Plume (1996)
1,800
47PUW22
4SPI1W72 »
49P11W2?
0 450 900
—
FEET
Figure 1. Site map.
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 c/s-1,2-DCE iso-
mer. Figure 3 shows the extents of CAM 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 histori-
cal data for the site show that the dissolved CAM 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 Fig-
ure 4. Electron acceptors displayed in this figure include
dissolved oxygen, nitrate, and sulfate. There is a strong
correlation between areas with elevated BTEX concen-
trations and areas with depleted dissolved oxygen, ni-
trate, and sulfate. The absence of these compounds in
contaminated ground water suggests that aerobic respi-
ration, 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(ll) and
methane (Figure 5). There is a strong correlation be-
tween areas with elevated BTEX concentrations and
areas with elevated iron(ll) and methane. The presence
of these compounds in concentrations above back-
ground in contaminated ground water suggests that
iron(lll) reduction and methanogenesis are working to
biodegrade fuel hydrocarbons at the site. Background
iron(ll) 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, indi-
cating that biologically mediated oxidation-reduction re-
actions are occurring in the area with ground-water
contamination.
78
-------
A
Northwest
F
Southeast
260-
•3
CO
ro
0)
£
ra
180 -
Well-Sorted, Fine- to
Medium-Grained Sand
Discharge
to
Wetlands
HORIZ 0 300 600
1200
VERT 0 15 30 60
Vertical Exaggeration = 20x
Figure 2. Hydrogeologic section.
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 concentra-
tions 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 de-
termine the dominant terminal electron-accepting proc-
ess 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)
Table 2. Range of Hydrogen Concentrations for a Given
Terminal Electron-Accepting Process
Terminal Electron
Accepting Process
Hydrogen Concentration
(nM/L)
Denitrification
Iron(lll) reduction
Sulfate reduction
Methanogenesis
0.2 to 0.8
1 to 4
>5
for trimethylbenzene (TMB). A modified version of this
method that takes into account the production of chlo-
ride 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.
79
-------
TOTAL BTEX
\
TRICHLOROETHENE
DICHLOROETHENE
• > 4,000 |jg/L
\ m 2,000 - 4,000 pg/L
\ Si ND - 2,000 pg/L
>10,000 Mg/L
1,000-10,000 pg/L
ND -1,000 |jg/L
CHLORIDE
>1,OOOng/L
500-1,000 |ig/L
ND-SOO|ig/L
>500 |ig/L
100 - 500 |jg/L
ND -100 Mg/L
>5,000 (jg/L
1,000 - 5,000 |ig/L
ND-1,000|jg/L
CHLORIDE
/> \ \ / \
! >100 mg/L
i 50 -100 mg/L
: ND - 60 mg/L
Figure 3. Chlorinated solvents and byproducts (1995).
TOTAL BTEX
D/SSOLVED OXYGEN
\
> 4,000
2,000 - 4,M» (i^L
ND - 2.0M
I -sng.i
-20 men.
in ngu
Figure 4. BTEX and electron acceptors (1995).
80
-------
Table 3. Approximate First-Order Biodegradation Rate Constants
Compound
TCE
DCE
Vinyl chloride
BTEX
Correction
Method
Chloride
1MB
Average
Chloride
1MB
Average
Chloride
1MB
Average
Chloride
1MB
Average
A-B
0 to
970 feet
(1/year)
1.27
1.20
1.24
0.06
0.00
0.03
0.00
0.00
0.00
0.13
0.06
0.10
B-C
970 to
1,240 feet
(1/year)
0.23
0.52
0.38
0.60
0.90
0.75
0.14
0.43
0.29
0.30
0.60
0.45
C -E
1,240 to
2,560 feet
(1/year)
-0.30
NA
-0.30
0.07
NA
0.07
0.47
NA
0.47
0.39
NA
0.39
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 (approxi-
mated as CH) to carbon dioxide during reduction of
carbon to chlorine bonds (represented as C-CI) to carb-
on to hydrogen bonds (represented as C-H).
CH + 2H20 + 2.5C-CI -» CO2 + 2.5H+ +
2.5CI- + 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 lib-
eration of 6.8 mg of biogenic chloride. PCE loses two
C-CI 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 com-
pared 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 ele-
mental composition of this material was used to calcu-
late an empirical formula for the dissolved organic
matter, and to estimate the number of moles of C-CI
bonds required to reduce one mole of dissolved organic
carbon in this material:
Ci.oHi.osiOo.6i9 + 1.38H2O + 1.91 C-CI -» CO2 +
1.91CI-+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 re-
quires 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 c/s-1,2-DCE slows or ceases
between points C and E. This correlates with the ex-
haustion 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.
Table 4. Comparison of the Estimated Electron Donor
Demand To Support Reductive Dechlorination to the
Supply of BTEX and Native Organic Carbon
Organic
Organic BTEX BTEX TOC Carbon
Chloride Chloride Available Demand Supply Demand
Point (mg/L) (mg/L) (mg/L) (mg/L) (mg/L) (mg/L)
A
B
C
D
E
63
43
57
13.6
18.4
58.1
7.72
8.26
1.34
0.78
16.8
4.2
3.9
0.09
0.04
8.5
1.13
1.21
0.20
0.114
80.4
31.1
24.3
13.8
8.2
10.3
1.37
1.46
0.24
0.14
Discussion and Conclusions
Available geochemical data indicate that the geochem-
istry 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 downgradi-
ent from the source. Near the source the plume exhibits
Type 1 behavior. At about 1,500 feet downgradient from
81
-------
TOTAL BTEX
:\\
J.WO-4,000
ND • 2,000 Mfli L
METHANE
f
4M
FEET
Figure 5. BTEX and metabolic byproducts (1995).
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 down-
gradient from the former fire- training pit (source area),
the dissolved contaminant plume consists of commin-
gled BTEX and TCE and is characterized by anaerobic
conditions that are strongly reducing (i.e., Type 1 behav-
ior). Dissolved oxygen concentrations are on the order
of 0.1 mg/L (background = 10 mg/L), nitrate concentra-
tions are on the order of 0.1 mg/L (background = 10
mg/L), iron(l I) concentrations are on the order of 15 mg/L
(background = less than 0.05 mg/L), sulfate concentra-
tions 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 concentra-
tions in the source area range from 1.4 to 11 nanomoles
(nM). As shown by Table 2, these hydrogen concentra-
tions are indicative of sulfate reduction and methano-
genesis, 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 c/s-1,2-
>
J - Id
• 5 man.
. •
1.KB
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 pro-
duced by the biogenic reductive dechlorination of TCE.
Furthermore, the dominant isomer of DCE found at the
site is c/s-1,2-DCE, the isomer preferentially produced
during reductive dechlorination. Average calculated first-
order biodegradation rate constants in this zone are as
high as 1.24, 0.75, and 0.29 per year for TCE, c/s-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 accumu-
lating 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 biode-
graded and the system begins to exhibit Type 3 behav-
ior. 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 CAM plume. Iron(ll) concen-
trations have significantly decreased and are on the
82
-------
Zone of
Type3
Behavior
Figure 6. Zonation of CAM plume.
60000
0 SOO 1000 1500 2000 2500 3000 3500
Distance From Source (feet)
Figure 7. Plot of TCE, DCE, and ethene versus distance down-
gradient.
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 CAM plume. Methane con-
centrations are the highest in this area but could have
migrated from upgradient locations. The hydrogen con-
centrations at Points E and F are 0.8 nM and 0.25 nM,
respectively, suggesting that the dominant terminal elec-
tron-accepting process in this area is iron(lll) reduction.
These conditions are not optimal for reductive dechlori-
nation, 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 in-
creases, probably because vinyl chloride is being oxi-
dized. Because biodegradation of vinyl chloride is faster
under Type 3 geochemical conditions than the biodegra-
dation of other CAM 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).
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 reac-
tion 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 reme-
diation 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 ground-
water. 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 micro-
cosms 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. En-
viron. Quality 23:792-798.
84
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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 chlo-
rinated solvents typically have relied on engineered re-
mediation systems that were costly to build and operate.
Recently, environmental regulatory agencies have be-
gun 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 un-
derstanding 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 con-
ducting an interdisciplinary research study of ground-
water contamination by chlorinated solvents at Picatinny
Arsenal, New Jersey. The objectives of the study are to
identify and quantify the physical, chemical, and biologi-
cal processes that affect the transport and fate of chlo-
rinated solvents, particularly trichloroethene (TCE), in
the subsurface; determine the relative importance of
these processes at the site; and develop predictive mod-
els of chlorinated-solvent transport that may have trans-
fer 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 re-
move TCE to and from the plume at Picatinny Arsenal,
and to determine which natural TCE-attenuation mecha-
nisms are the most important on a plume-wide basis.
Geohydrology
Picatinny Arsenal is a weapons research and develop-
ment 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 down-
ward flow near the valley walls and upward flow near
Green Pond Brook. Estimated ground-water flow veloci-
ties 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 asso-
ciated 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 (u,g/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 concen-
trations greater than 1,000 u,g/L are found in the source
area, the TCE concentrations are highest (greater than
10,000 u,g/L) near the base of the aquifer midway be-
tween 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
-------
EXPLANATION
Area in which trichloroethene concentration
exceeds 10 micrograms per liter
10 LINE OF EQUAL TRICHLOROETHENE
CONCENTRATION-Shows trichloroethene
concentration, in micrograms per liter.
Dashed where approximate
A' Line of section
. 41-9 Ground-water sampling site location
and local identifier
Sear -Swamp 8rooX Dry wei
METERS
215 i
lagcona
13-2 \
13-1 3-c
'I'" CAF-7 92-1
9"a CAF-2 92-5
/|I§ CAF-8 92Z4 112.8
A'
METERS
215
100 0 100 200 300 400 SCO
DISTANCE FROM BUILDING 24 CB-24). IN METERS
B.
EXPLANATION
-J2io 1Qfl LINE OF EQUAL TRICHLOROETHENE
CONCENTRATION-Shows trichloroethene
concentration, in micrograms per liter.
Dashed where approximate
[ Well screen and trichloroethene concentration,
110 in micrograms per liter
MS Not sampled
< Less than
CAF-7 Location of well and local identifier
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.)
Results of water-quality analyses indicate that the pH of mg/L) are very low. Concentrations of iron(ll) are greater
ground water in the plume is near neutral (6.5 to 7.5), than 1 mg/L in some areas of the plume, whereas sulfate
and concentrations of both dissolved oxygen (less than and carbon dioxide are consistently plentiful (greater than
0.5 milligrams per liter [mg/l]) and nitrate (less than 1 40 mg/L and 100 mg/L as bicarbonate, respectively) as
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 u,g/L.
These findings indicate the plume is primarily anaerobic
and contains a variety of reducing redox environments
controlled in different areas by iron(lll) reduction, sulfate
reduction, and methanogenesis. Under these condi-
tions, reductive dechlorination of TCE can take place if
sufficient electron donors are available. Dissolved or-
ganic carbon (DOC), consisting primarily of humic and
fulvic acids, may fulfill the electron donor requirement in
this system. Concentrations of DOC are highest imme-
diately 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 com-
pounds 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 concentra-
tions 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 repre-
sents 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 concen-
trations are greater than 10,000 u,g/L.
The mass of sorbed TCE within the plume was esti-
mated 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 calcu-
lated 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 Ar-
senal (9, 10). The TCE removal processes that were
considered include advective transport, lateral disper-
sion, anaerobic biotransformation, diffusion-driven vola-
tilization, 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 concen-
tration of 1,200 u,g/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 u,g/L (Figure
1). Using Fick's Law, the lateral TCE-concentration gra-
dient, 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 un-
dergoes the sequential replacement of the chlorine at-
oms 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 vola-
87
-------
tilization was estimated using Pick's Law, field-meas-
ured unsaturated-zone soil-gas TCE gradients, bulk dif-
fusion 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 vola-
tilization 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 cal-
culated to be approximately 50 kg/yr.
Sorption is the partitioning of TCE from the ground water
into the organic-carbon fraction of the aquifer sedi-
ments. 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 concen-
trations. 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 char-
acterized 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 experi-
ments. 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 labora-
tory soil columns, field infiltration experiments, and mul-
tiphase 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 nonaqueous-
phase 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 repre-
sented 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 annu-
ally, is removed as a result of anaerobic biotransforma-
tion. This is by far the most important TCE removal
process operating in the Picatinny Arsenal plume. Re-
moval 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 mi-
nor 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 natu-
ral 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
-------
ADVECTION-DRIVEN
VOLATILITZATION
(50 kg/yr)
DIFFUSION-DRIVEN
VOLATILIZATION
kg/yr)
ADVECTIVE
TRANSPORTTO
GREEN POND BROOK
(50 kg/yr)
.•-••••.-.--•• •.--•-'-. INFILTRATION •.-.•-•-••.-
•'••'-'•"•-'•"•"••"•'•'• <1 k/r "-'•"•••'
TRIG H LO R 0 ETH EN E PLUME
LATERAL
DISPERSION
kg/yr)
ANAEROBIC
BIOTRANSFORMATION
(360 kg/yr)
DISSOLUTION
OF DNAPL
(not estimated)
Estimated top of confining unit
NOT TO SCALE
GAINS
TRICHLOROETHENE MASS-BALANCE COMPONENTS
[kg/yr, kilograms per year; <, less than]
LOSSES
DESORPTION 550 kg/yr
INFILTRATION <1 kg/yr
DISSOLUTION OF DENSE not estimated
NONAQUEOUS PHASE LIQUID
TOTAL
550 kg/yr
ANAEROBIC BIOTRANSFORMATON 360 kg/yr
ADVECTIVE TRANSPORT TO BROOK 50 kg/yr
ADVECTION-DRIVEN VOLATILIZATION 50 kg/yr
LATERAL DISPERSION <1 kg/yr
DIFFUSION-DRIVEN VOLATILIZATION <1 kg/yr
SQRPTION <1 ka/vr
TOTAL
460 kg/yr
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.
The fact that long-term desorption is a significant con-
tinuing source of TCE to the aquifer may explain why
the TCE concentrations are still relatively high in the
source area (greater than 1,000 u,g/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. Ad-
ditional study of the individual processes would be nec-
essary to refine the mass balance further. Because
confidence in the output-process mass-flux estimates is
89
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high and the TCE desorption rate constants used prob-
ably 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 labo-
ratory microcosm studies. Assuming first-order kinetics
and considering the decrease in TCE concentrations
from the source area to the discharge area (1,900 u,g/L
to 760 u,g/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 experi-
ments. 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 meas-
urements 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 strip-
ping 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 trans-
port and advection-driven volatilization.
Conclusion
The relative importance of all naturally occurring proc-
esses that introduce or remove TCE to or from a con-
tamination 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 biotrans-
formation accounts for an estimated 78 percent of the
total mass of TCE removed from the plume annually.
Other removal processes—lateral dispersion, diffusion-
driven volatilization, and sorption—are minor in com-
parison. Desorption is the most significant TCE input
process evaluated. A mass-balance analysis shows that
the removal of TCE from the plume by natural attenu-
ation 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 an-
aerobic 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 Pi-
catinny Arsenal site. The pump-and-treat system re-
moves 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 trichlo-
roethylene 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, Mon-
terey, CA, March 11-15. U.S. Geological Survey Water-Re-
sources 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 Hydrol-
ogy 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., TV. 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, PA., 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: Descrip-
tions of and sampling plans for remedial investigation sites. Ar-
gonne 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 Arse-
nal, New Jersey. Hydrological Sci. Technol. 11(1-4):26-50.
90
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8. Koller, D., I.E. Imbrigiotta, A.L. Baehr, and J.A. Smith. 1996.
Desorption of trichloroethylene from aquifer sediments at Picat-
inny Arsenal, New Jersey. In: Morganwalp, D.W., and D.A. Aron-
son, 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, I.E., and M. Martin. 1991. Overview of research ac-
tivities 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 Sub-
stances Hydrology Program—proceedings of the technical meet-
ing, Monterey, CA, March 11-15. U.S. Geological Survey
Water-Resources Investigations Report 91-4034. pp. 673-680.
10. Imbrigiotta, I.E., and M. Martin. 1996. Overview of research ac-
tivities on the transport and fate of chlorinated solvents in ground
water at Picatinny Arsenal, New Jersey, 1991-93. In: Morgan-
walp, 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 I.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 Biore-
mediation of Ground Water, Denver, CO, August 30-September
1, 1994. EPA/540/R-94/515. pp. 143-153.
12. Parsons, F.Z., PR. Wood, and J. DeMarco. 1984. Transforma-
tions of tetrachloroethene and trichloroethene in microcosms and
ground water. J. Am. Waterworks Assoc. 76(2):56-59.
13. Vogel, T.M., C.S. Griddle, and PL. McCarty. 1987. Transforma-
tions of halogenated aliphatic compounds. Environ. Sci. Technol.
21 (8):722-736.
14. Wilson, B.H., T.A. Ehlke, I.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 Sub-
stances Hydrology Program—proceedings of the technical meet-
ing, Monterey, CA, March 11-15. U.S. Geological Survey
Water-Resources Investigations Report 91-4034. pp. 704-707.
15. Ehlke, T.A., I.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: Morgan-
walp, 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., PR. Jaffe, and J.A. Smith. 1993. Simulating the vola-
tilization 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 Bioreme-
diation of Ground Water, Denver, CO, August 30-September 1.
EPA/540/R-94/515. pp. 154-160.
91
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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 pump-
and-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 concentra-
tions after pump turnoff and the emergence of the "tail-
ing" 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 be-
tween 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 conduc-
tivity 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 ground-
water velocity ranges from 250 to 800 feet per year (2).
Activities at Plant 44 include development, manufactur-
ing, 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-dichlo-
roethene (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 mil-
lion (ppm) was measured in 1986, although concentra-
tions 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 re-
ported 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 concentra-
tion 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. Aqueous-
phase flow in the upper zone was simulated along with
the pump-and-treat remediation system. Source loca-
tions used in the modeling were based on the location
92
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EXPLANATION „?„
E-4 • Extraction wells
R-8A Recharge wells
R-9
Figure 1. TCE plume prior to remediation, December 1986 (ppb).
Site VII
SiteV
Site III
EXPLANATION
^ On-site disposal
*.' Current surface impoundment area
Site IX
Site VIII
MB Site II
• Site XV
I Site I
Site IV
Note; Sites II, III, VII, and VIII were reportedly used for disposal ofDNAPL related wastes.
Figure 2. Historical onsite disposal locations.
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 ana-
lyzed assuming the following four potential configura-
tions 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 depend-
ing 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 as-
sociated dissolution times ranged from 100 to 1,000
years. The fate-and-transport modeling results, assum-
ing no continuing sources of TCE into the aquifer, indi-
cate 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 re-
mediation 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 unsatu-
rated zone and the potential dissolution from TCE plumes.
93
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Complete dissolution of the DNAPL pools may take as References
long as 100 years under pumped conditions, while dis-
solution of unsaturated residual by infiltrating ground r Hargisand Montgomery, inc. 1982. Phase n investigation of sub-
water may continue for thousands of years. The ground- surface conditions in the vicinity of abandoned waste disposal
water extraction System at the Site has contained the sites, Hughes Aircraft Company manufacturing facility, Tucson,
dissolved plume and removed significant amounts of Arizona' VoL' Tucson' AZ
Chlorinated compounds. If DNAPL is present at the Site, 2 Groundwater Resources Consultants, Inc. 1992. Quarterly ground
however, complete removal Of TCE using pump-and-treat water monitoring report, well field reclamation system, July through
will require a very lengthy and costly operation period. September 1991, u.s. Air Force Plant 44.
94
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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 sol-
vents, such as tetrachloroethylene (PCE) and trichlo-
roethene (TCE). The intrinsic bioremediation program is
part of an integrated study that also includes co-
metabolic bioventing and accelerated anaerobic treat-
ment. 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 vali-
date protocols for implementing intrinsic remediation at
other sites.
Awide variety of geological, geochemical, and biological
research is being integrated into this study. This presen-
tation 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 ex-
pertise lies in geochemistry. The participants who fo-
cused on the biological aspects of this study will undoubtedly
be presenting their conclusions at future meetings.
95
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Background
The RTDF Bioremediation Consortium initiated this
study in February 1995. Dover AFB was chosen overthe
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 geo-
chemical markers, and the presence of the soluble chlo-
ride 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
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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 concentra-
tions 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 biodegrada-
tion 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 reduc-
tive dehalogenation of TCE does occur in the deep
zone. The amount of ethylene is small, however: 50
micrograms per liter (u,g/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 chlo-
ride 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 concen-
tration increases to over 40 mg/L. Background chloride
levels are approximately 10 mg/L. The dissolved chlo-
ride (Figure 3) increases in the deep zone of the aquifer
but not the shallow zone. This eliminates other, extrane-
ous chloride sources such as road salt. This evidence
clearly supports the hypothesis that solvents are being
destroyed by an intrinsic process.
7/7
' Monitoring well
SCALE
500
1000 FEET
Figure 2. Total chlorinated compounds (mg/L) in the deep zone at Dover AFB.
97
-------
: Monitoring well
Figure 3. Dissolved chloride in the deep zone at Dover AFB.
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 labo-
ratory studies.
In the vicinity of the plume, the dissolved oxygen con-
centration 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, dis-
solved 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 u,g/L are found
within the contaminated zone, while no methane is ob-
served 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 High-
way 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
500
SCALE
5
0 500 1000 FEET
that bacterial reductive dehalogenation of PCE and TCE
to DCE can occur in relatively oxidizing conditions, re-
quiring 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), proc-
esses that occur at redox levels below -200 millivolts.
These low oxidation states are probably found in mi-
croenvironments but do not dominate the aquifer. Fi-
nally, ongoing RTDF microcosm studies of Dover AFB
samples are showing clear production of 14CO2 from
14C-labeled DCE under oxygenated conditions. There-
fore, 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
Table 1. Half-Life Calculations for Dover AFB
Method PCE -> TCE TCE -> DCE DCE^VC VC -> ETH
Buscheck
r2
Graphical
extraction
2.4
0.99
2.80
2.8
0.94
4.19
1.4
0.94
2.81
2.2
0.93
1.84
98
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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 chlorin-
ated 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, isocon-
centration maps can be used to calculate that about
250 pounds of chlorinated solvents are being biode-
graded 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 de-
stroyed 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 dehalo-
genation and oxidation, and possible evidence for co-oxi-
dation.
References
1. Sharma, P.K., and P.L. McCarty, 1996. Isolation and charac-
terization of a facultative bacterium that reductively dechlorinates
tetrachloroethene to cis-1,2 dichloroethene. Appl. Environ. Micro-
biol. 62(3):761-765.
2. Kastner, M. 1991. Reductive dechlorination of tri- and tetrachlo-
roethylenes depends on transition from aerobic to anaerobic con-
ditions. 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, I.E., K.T. OReilly, S.N. Nelson. 1993. Evaluation of
intrinsic bioremediation at field sites. Proceedings of the confer-
ence Petroleum Hydrocarbons and Organic Chemicals in Ground
Water: Prevention, Detection and Restoration, Houston, TX, pp. 367-
381.
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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 thereaf-
ter, the U.S. Environmental Protection Agency (EPA),
the Strategic Environmental Research and Develop-
ment Program (SERDP) of the Department of Defense
(DoD), the University of Michigan, and the Michigan
Department of Environmental Quality contributed re-
sources 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 well-
defined and controlled research and development plat-
form for the evaluation of in situ site characterization and
remediation technologies. The emphasis is on bioreme-
diation techniques applied to subsurface and sediment
contamination problems. In situ biological technologies
with the potential to remediate unsaturated- and saturated-
zone fuel and organochlorine solvent contamination in
subsurface and sediment systems are of particular inter-
est. NCIBRD focused its early activities on the de-
velopment 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 losco 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 Develop-
ment Commission. The U.S. Air Force Base Conversion
Authority (BCA) is charged with remediating contami-
nated 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-mile-
wide 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 ap-
proximately 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 quater-
nary 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 miner-
als, 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
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I Mission
Fire Training J?rive
ne I
EXPLANATION
IRPSite
—• - - — Base Boundary
Groundwater Plume
Direction of Groundwater Row
S"
EOD Range Safety Zone
run ft
0 750 1500 3000 Feet
Figure 1. Map of Installation Restoration Program sites at WAFB.
average of approximately 100 feet per day (31 meters AuSable River discharge areas at average rates of 1.0
per day) based on selected slug or pump tests and to 0.3 feet per day (0.3 to 0.1 meter per day). In general,
estimations from particle size distributions. Flow in the vertical flow gradients are negligible except in zones of
sand and gravel upper aquifer is generally eastward ground-water discharge to surface-water bodies or near
towards Lake Van Etten and south-southeast to the pumping centers.
101
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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); organo-
chlorine solvents (e.g., trichloroethylene [TCE], dichlo-
roethylene [DCE]); fire-fighting compounds; combustion
products (e.g., naphthalene and phenanthrene); and
chlorinated aromatic compounds (e.g., dichloroben-
zenes). Soil, aquifer solids, sediments, and ground
water are the major environmental media involved. Of
the 58 high-priority sites at the base, 13 include chlorin-
ated solvents or partial microbial degradation products
as primary contaminants. Twelve of these 13 sites iden-
tify perchloroethylene (PCE) and TCE as primary con-
taminants 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 mono-
mer). These sites have abundant levels of nonchlori-
nated 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 oxicto 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 Re-
sources, and the BCA actively cooperate in the ongoing
IRP activities as well as the efforts of NCIBRD. Cur-
rently, 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 in-
field 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 measure-
ments at more than 500 wells. Site-wide water balance
and refined ground-water transport models exist for
sites of current or future technology demonstration ac-
tivity as well as for a controlled in situ injection experi-
mental 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 poly-
cylic 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). Focus-
ing on the dissolved volatile organic compounds (i.e.,
aromatics and chlorinated alkenes), the plume was de-
lineated 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-dichlo-
roethylene, trichloroethylene, and perchloroethylene)
plume, which was restricted to the upper 6 feet (approxi-
mately 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 moni-
toring in the ground water was undertaken as the initial
part of a demonstration of intrinsic bioremediation.
Quarterly monitoring results since that time have dis-
closed variable dissolved concentrations of the chlorin-
ated 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. Fig-
ure 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. Far-
field wells have generally shown diminished concentra-
tions 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 investi-
gation. 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 hh measurements for delineat-
ing redox processes in a contaminated aquifer. In preparation.
102
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Figure 2. Plan view of Fire Training Area 2 showing dimensions of major chlorinated VOC plume in ground water in 1993.
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 com-
pound mass from the plume, 13 borings were made in
1994 along the axis of the plume coincident with domi-
nant ground-water flow direction. A total of 300 core
subsamples were taken by Geoprobe techniques col-
lecting 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 determi-
nations at the near- and mid-field locations of the moni-
toring wells. It should be noted that, in contrast to the
water samples, which were contaminated by reductive
dechlorination metabolites, the solid-associated chlorin-
ated hydrocarbon distributions were dominated by par-
ent compounds, principally PCE and TCE. It is clear
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)3
Location in Major
Plume
Near field
(Approximately
200 feet
downgradient
from Pad Boring 6;
Well 4S)
Mid-field
(Approximately
450 feet
downgradient from
Pad Boring 12;
Wells 8S and 8M)
Ground Aquifer
Water Solids
(mg/L) (mg/L)
0.08 10.5
0.003 6.3
% of Total
Associated
With
Solids
99%
99%
Volume
% of
Oily
Phaseb
2.9
0.006
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.
' 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
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1
TCE
Deep Wells
(At Midfield and Farfield)
TCE
Shallow Wells
(At Midfield, and Farfield)
15 -
8/87 4/88 8/90 12/93 4/94 6/94 9/94 1/95 4/95 7/95 10/95 1/96
Time
8/37 4/88 8/90 12/93 4;94 6/94 9/94 1/95 4/95 7/95 10/95 t/96
Time
cis-1,2-DCE
Shallow Wells
(At Midfield)
cis-1,2-DCE
Deep Wells
(At Nearfield and Midfield)
16
00
|
1400 -••
1200 —--
' 1000 1
e I
_o
« 800 L
E -
I 60° r
O ;
400 - —
1 1
350 -
300 -••••
250 -
a 200 -•
g '5° _-
o
\
V
12/93 4/94 6/94 9/94 1/95 4/95 7/95 10/95 1/96
Time
12/93 4/94 6/94 9/94 1/95 4/95 7/95
Time
Figure 3. Concentration trends over time for TCE and DCE at Fire Training Area 2 wells.
from these data that aqueous concentration variability in
determinations of metabolite concentrations are a
negligible portion of the total mass of chlorinated hydro-
carbon 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 re-
ductive 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 pre-
sumed 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, AmirSalezedeh, Debbie
Patt, and Patty Laird. NCIBRD staff and collaborators
welcome the contributions and participation of interested
104
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groups in future investigations Of promising Site Charac- 2- Barcelona, M.J. 1995. Verification of active and passive ground-
terization and bioremediation Cleanup technologies. water 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
Serjes NQ 364 Wjen/New York: Springer-Verlag. pp. 161-175.
1. U.S. Geological Survey. 1993. Data submission via memo to U.S. 3. Hess, K.M., W.N. Herkelrath, and H.I. Essaid. 1992. Determination
Air Force Base Conversion Agency, Wurtsmith Air Force Base, of subsurface fluid contents at a crude-oil spill site. J. Contam.
U.S. Geological Survey Lansing Regional Office, Michigan. Hydrol. 10:75-96.
105
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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 proc-
ess of site assessment and data reduction and interpre-
tation that focuses on the quantification of the capacity
of a given aquifer system to assimilate ground-water
contaminants through physical, chemical, and/or bio-
logical 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 re-
garding source removal activities or estimates of source
lifetime has not generally been presented in the litera-
ture. 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 proc-
ess outlined in Figure 1. This process provides a logical
approach to evaluating the feasibility and appropriate-
ness of implementing intrinsic remediation at a given
site and includes: 1) determining whether steady-state
Intrinsic Remediation
Assessment Approach
1. Steady-State Plume
Conditions?
2. Estimate
Contaminant
Degradation Rate
3. Estimate Source
Mass
4. Estimate Source
Lifetime
- 5. Long -Term Behavior
6. ntrtnsic
Remediation for Site?
7. Long-Term
Monitoring for She
Figure 1. Components of the intrinsic remediation assessment
approach.
plume conditions exist; 2) estimating contaminant deg-
radation rates; 3) estimating the source mass; 4) esti-
mating 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 manage-
ment. 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 over-
lying a low-permeability bed rock formation (4). The aquifer
106
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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. An-
aerobic dechlorination reactions, evident as dechlorina-
tion 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
forthe 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 overtime in
the plume. With estimated dissolved TCE concentra-
tions 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 = -In (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
EIELSONAFB SITE 45/57
Overlay plot sh owing TCE
•and reduction by-products
Jut/ ] 995
Figure 2. Overlay plot of TCE and its degradation products measured in July 1995 at Site 45/57, Eielson AFB, Alaska.
107
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-PNL-92
Moy-94
-Sept-'94
-July-'95
40 60 80 100 120
Dlstanc* Downgradlcnt from Sourc* Ar*o at 45MW08 (m)
20
40 60 80 100 120
Distance Downgradfentfrom Saurca Ar*a at 45MW08 (m)
Figure 3. Plume centerline TCE concentrations measured from fall 1992 through July 1995 at Site 45/57, Eielson AFB, Alaska.
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 esti-
mate 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 degrada-
tion rates. Through this process, a mean TCE degrada-
tion 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 Hard-
ing 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 overtime
strengthens the argument that a residual phase does not
exist at the site. If it is assumed that a distinct free-
product phase does not exist in the source area, an
estimate of source mass can be made assuming con-
taminated soil in equilibrium with the measured source
area dissolved TCE concentration, Co. Using this ap-
proach, the source area mass was estimated using the
following equation:
Msource = Co (Y) (L) (b) (R) (9) (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 9 = aquifer total porosity =
0.38. Source dimensions were estimated based on inter-
polation of ground-water data collected within and outside
the source area, while R and 9 were based on aquifer-spe-
cific characteristics determined from cores collected from
the site. Using these values, a source mass of 37.5 kilo-
grams 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/Land a ground-water
108
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impact concentration (maximum contaminant level [MCL])
of 5 micrograms per liter (u,g/L) established for TCE, an
estimated source lifetime for TCE at this site is:
Source lifetime = In(5/90,000)/k1 =
ln(5.6 x 1(r5)/(-0.0026) = 10.3 years
(Eq. 3)
A worst-case scenario can be formulated for contamina-
tion 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, con-
tained within the sandy aquifer at Site 45/57 is approxi-
mated to be 25 percent of the pore volume (9), or 9.5
percent of the source area volume. Based on a meas-
ured 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 con-
centration of TCE in the residual phase, CTCE:
(MaSSTcE)/(MaSSResidual Phase) =
0.065 (MWTCE/MWResidua| 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) (9) (Sr) (pResidual 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) (9) (Co) =
(15 m) (3 m) (0.1 m/d) (0.38) (0.09 kg/m3) =
0.15kg/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 kilo-
grams exists at the site:
Source lifetime =
Mass-pee 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, in-
creasing the overall cost of plume management at the
site. More information regarding residual-phase distribu-
tion 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 con-
taminants 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 deci-
sion 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 assimi-
lated 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 life-
time 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 recom-
mended at Site 45/57, a long-term monitoring network
is required. To have this network serve multiple purposes,
a combination of upgradient, downgradient, and within-
plume 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
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TP3
Modeling Parameters:
v = 0.07 m/d, vr = 0.03 m/d,
Longitudinal dispersivity = 2.12 m
Transverse dispersivity = 0.53 m
Vertical dispersivity = 0.001 m
TCE Decay rate = 0.00026/d
SP7
Eielson AFB Site 45/57-Plume Area
Predicted TCE Contours
0, 3 and 8 years Following
Source Removal/Depletion
Max = 39,000 ppb. Interval = variable
SP43 .
SP44 «
SP45 .
SP10-
SP41
SP38
SP12
SP35
SP32
SP29
SP42 *
Approximate Scale
100 meters
SP40
45MW02
GP02
SP8
MW08
45MW07 *
Existing Long-Term Monitoring Wells
Proposed Long-Term Monitoring Wells
Proposed Point-of-Compliance Wells
t = 0 years after source removal/depletion
t = 3 years after source removal/depletion
t = 8 years after source removal/depletion
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.
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 al-
low updating of the conceptual model for plume and
source area configuration overtime. The second set of
monitoring wells consists of a transect of three wells
perpendicular to the direction of plume migration, ap-
proximately 250 feet (75 meters) downgradient from Moni-
toring 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 u,g/L)
migrates beyond the area under institutional control.
A sampling frequency of 1 to 2 year intervals was rec-
ommended for this site. This interval provides sufficient
data overtime 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 re-
mediation protocol to a hydrocarbon/solvent contami-
nated 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) as-
sessment of aquifer assimilative capacity and the desir-
ability of source removal at the site; and 7) development
of a long-term monitoring strategy for verification of
intrinsic remediation process performance and regula-
tory compliance purposes.
Intrinsic remediation of solvent contaminated ground
water was demonstrated at Site 45/57 through the iden-
tification of TCE dechlorination products in the plume
(Figure 1), the recognition of decreasing TCE dissolved
plume mass overtime, and calibration of field data to a
110
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fate-and-transport model. No residual-phase product
was identified within the source area based on historical
and recent site investigation activities; however, a worst-
case 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 proc-
esses without source removal, long-term monitoring with-
out 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 Protec-
tion Agency, Analytical Sciences Branch, Characterization Re-
search Division, Las Vegas, NV.
2. Dupont, R.R., D.L. Sorensen, M. Kemblowski, K. Gorder, and G.
Ashby. 1996. Assessment and quantification of intrinsic remedia-
tion at a chlorinated solvent/hydrocarbon contaminated site,
Eielson AFB, Alaska. Paper presented at the Conference on In-
trinsic 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 Rl Report, Vol. 1. Eielson AFB, AK
5. CH2M-HNI. 1982. Installation restoration program records search,
Eielson Air Force Base, AK.
6. Utah Water Research Laboratory. 1995. Intrinsic remediation en-
gineering 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, PA. 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 po-
rous media. Water Resour. Res. 23:618-624.
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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 contamina-
tion, both technical and regulatory options must be iden-
tified and evaluated to ensure compliance with state and
federal regulatory requirements. A strong technical de-
fense 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 encoun-
tered under the Superfund program and poses a difficult
dilemma from both technical and regulatory perspec-
tives for compliance with the U.S. Environmental Pro-
tection Agency's (EPAs) 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 contami-
nated 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 ap-
proach to accommodate the proposed technical solu-
tion. 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 antici-
pated fate and transport of the contamination. Once this
is accomplished, alternatives for addressing the con-
tamination can be evaluated.
In the feasibility study, a range of alternatives are devel-
oped 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 sub-
surface soils and smear zone combined with institu-
tional controls and ground-water monitoring.
• Ground-water extraction and physical/chemical treat-
ment 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,
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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 res-
toration required. Considerations for remedy selection
were the amount of contamination remaining in the
unsaturated and saturated zones and the availabil-
ity of the contamination for removal and treatment.
The technical evaluation is largely an issue of a
balance between the need for and feasibility of con-
taminant removal in the unsaturated and/or saturated
zones and the efficiency of natural attenuation. The
NCR (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 pre-
sent, 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 cate-
gory, the source was continuing to overwhelm the rate
of natural attenuation, resulting in an expanding con-
taminant 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 contami-
nant mass would return the aquifer to its beneficial
uses throughout the plume in a significantly greater
timeframe than natural attenuation alone. This evalu-
ation was not a trivial task given the difficulties in esti-
mating the source term, accurately evaluating the
contaminant fate and transport in the subsurface, and
assessing the effectiveness of source removal. In evalu-
ations conducted at Eielson AFB, modeling was gener-
ally the mechanism chosen to evaluate the benefits of
source removal. Generally, these modeling efforts used
conservative assumptions for fate-and-transport analy-
sis 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 re-
moval, therefore, was not expected to significantly re-
duce 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 deter-
mine the most relevant approach for the specific situ-
ation. 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 contami-
nated 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 investi-
gation and to ensure that the remedy remains pro-
tective.
• Statutory 5-year reviews are required whenever the
selected remedy will leave contamination on site
above levels that allow for unlimited use and unre-
stricted exposure (NCR §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 statisti-
cally significant increases in contaminant concentration
in the surface water. Natural attenuation is the mecha-
nism for cleanup in ground water between the contami-
nation 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 FR8754).
For Eielson AFB, ACLs were not applicable because
contaminated ground water did not discharge into sur-
face water on base.
Alternate Points of Compliance
As stated previously, remediation levels should gener-
ally 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, alter-
nate points of compliance may be considered in combi-
nation 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. Consis-
tent 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 cir-
cumstances if the remedy can be demonstrated to be
protective. One of the six ARAR waivers provided by the
Comprehensive Environmental Response, Compensa-
tion, and Liability Act (CERCLA §121(d)(4)) is technical
impracticability (Tl). The use of the Tl waiver requires a
demonstration that compliance with ARARs, including
maximum contaminant levels (MCLs) or non-zero maxi-
mum contaminant level goals (MCLGs), is technically
impracticable from an engineering perspective. A dem-
onstration 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 Tl 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 con-
trol 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), estab-
lishing a less-stringent cleanup level, and/or using natu-
ral attenuation.
At Eielson AFB, Tl 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 inor-
ganic lead. Ground-water contamination is confined to
areas approximately 600 feet in length. Ground-water
remediation is technically impracticable because the in-
organic 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 waterwill 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-ef-
fective, 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). In-
stitutional controls may be necessary to ensure that
such ground waters are not used before levels protec-
tive 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 con-
taminated with both petroleum compounds and chlorin-
ated 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 ap-
prehensions associated with the selection of natural
attenuation.
The uncertainty associated with environmental deci-
sions, specifically the selection of natural attenuation,
was addressed at Eielson AFB through the use of the
observational method. Key components of the obser-
vational method are 1) a decision based on the most
probable site conditions; 2) identification of reasonable
deviations from those conditions; 3) identification of pa-
rameters to monitor to detect deviations; and 4) prepa-
ration of contingency plans for each potential deviation
(1). The conceptual site model developed through the
investigation will be tested and confirmed through con-
tinued ground-water monitoring. A phased approach
with contingencies for additional remediation was estab-
lished in the event that the conceptual site model is not
confirmed.
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In addition, statutory 5-year reviews require an evalu-
ation 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 institu-
tional 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 ground-
water 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 pro-
tectiveness 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.
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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 Pro-
tection Agency (EPA) regulations (2) created guidelines
and requirements for safe and responsible operation of
USTs, with provisions for early leak detection, leak re-
porting, 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 re-
sponsibility 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 (petro-
leum 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 com-
pleted by the end of 1995, and some state cleanup funds
were almost exhausted, bordering on insolvency.
The American Society for Testing and Materials' docu-
ment "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 path-
ways from contaminant source to potential receptors are
evaluated. Risk levels are used to back-calculate ac-
ceptable 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 proc-
ess. 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 inter-
ested parties, including state regulators, environmental
consultants, UST owners and operators, and the gen-
eral public. All of these interested parties orstakeholders
must be involved in the process up front to avoid mis-
conceptions and misunderstandings. It is especially im-
portant 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, path-
ways 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.
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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 back-
ground 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 bal-
anced 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 com-
pliance, 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 particu-
lar, 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 con-
cern left behind at LUST sites following property transfer
may still cause uncertainty and potential problems. Asite
closed using cleanup levels determined through RBCA
or by previous standards will leave some level of com-
pounds 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 consult-
ants and regulators. Sites will have to be closed even-
tually, 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 ques-
tion early in the cleanup process to better utilize re-
sources 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 pro-
gram 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 re-
lease sites. E-1739-95. September 10.
4. Partnership in RBCA Implementation (PIRI). 1996. RBCA imple-
mentation summary graph from EPA/ASTM data. February 6.
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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 CRT push technology be used for well
installation?
• What are the advantages and disadvantages of con-
ventional wells, mini wells, or water samples col-
lected with a Hydropunch and a CRT 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, tempera-
ture, 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 oxygen-
sensing electrodes and indicator dye kits?
- What level of training is required to use the equip-
ment 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(ll) 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 eth-
ane 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 sam-
ples 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?
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What are the cost tradeoffs of these analyses com- • How many wells or cores samples are needed?
pared with the benefit of improved understanding of _ j0 examine plume flow velocity?
plume behavior? _,. . . .. . ... . _
- To examine proximity to sensitive receptors?
What should be the relative investment in sample
acquisition, sample analysis, data reduction, mathe-
matical modeling, and report preparation?
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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 attenu-
ation policies and guidelines for natural attenuation in
both ground water and soils. Finally, it briefly addresses
the factors that are important for promoting greater ac-
ceptance of natural attenuation as a remediation ap-
proach 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
20
Q
O
o 10
-------
Figure 2. Number of Natural Attenuation RODs per state by EPA region (1982-1994).
Coal Gasification
Fuel Storage
Junkyard
Wood Treatment
Oil Reclamation
Metal Plating/Mining
Chemical/Industrial Mfg.
Industrial Landfill
Municipal Landfill
I
5
I
10
I
15
I
20
Number of Sites*
*Some sites have more than one use
Figure 3. Types of sites at which natural attenuation was selected.
PCBs, Pesticides
PAHs, Phenols
BTEX, MTBE
Solvents
Inorganics
Number of Sites*
*Some sites have more than one contaminant
Figure 4. Contaminants present at sites for which natural attenuation was specified.
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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 treat-
ment at such sites. As of 1995, natural attenuation of
soils at such sites constituted 28 percent of the treat-
ment scenarios utilized, second to landfilling (34 per-
cent), 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 un-
derground storage tank sites is shown in Figure 5.
Air Sparging
(13%)
In Situ Bioremediaton
(4%)
Biosparging
(4%)
Dual Phase Extraction
(2%)
Pump-and-Treat
(29%)
Adapted from Dana lulls,
EPA UST/LUST National Conference Talk,
March 11, 1996
Figure 5. Use of ground-water cleanup technology at UST sites.
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 attenu-
ation, 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 petro-
leum 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 decid-
ing on natural attenuation; generally the states have
requirements for peer-reviewed models and will con-
sider 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 leach-
ing 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 recep-
tors, 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 guid-
ance 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 accep-
tance of natural attenuation involves several key com-
ponents. First, good communication is critical in getting
natural attenuation implemented at a site. It should be
conveyed that natural attenuation is a responsible, man-
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10
Written policies and guidelines
Developing policies and guidelines
Informal policies and guidelines
Informal guidelines/Do not consider non-UST waste
Do not have any formal policy
Figure 6. State policies regarding natural attenuation in ground water.
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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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. Chlorin-
ated solvents, such as trichloroethylene (TCE), are the
second most common group of Air Force site contami-
nants. Sites contaminated with heavy metals are a dis-
tant 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 attenu-
ation 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 trans-
port, specific quantification of potentially major proc-
esses, such as biodegradation, often is overlooked in
traditional site characterizations—an egregious over-
sight at sites contaminated with petroleum hydrocar-
bons.
The Air Force Center for Environmental Excellence (AF-
CEE), Technology Transfer Division, has completed 50
site characterizations, inclusive of natural attenuation
parameters. Each of these investigations has demon-
strated 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 pa-
rameters gives the owner and regulator the ability to
discern which sites will pose a potential future risk and
which sites do not orwill 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 ap-
proaches. 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) pro-
vides 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 remedi-
ated, statewide water basin capacity would increase
by 5/10,OOOths 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 ground-
124
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water plumes. A growing body of evidence supports the
contention that natural degradation processes are im-
portant in understanding the fate and transport of chlo-
rinated 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 attenu-
ation 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 charac-
terization process improves the quality and accuracy of
the site investigation and facilitates defensible, risk-
based 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 evalu-
ation and design of other remedial systems.
Given that the contribution of natural attenuation proc-
esses to chlorinated organics cleanup is less estab-
lished 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 attenu-
ation 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 or-
ganics, the AFCEE Technology Transfer Division is con-
vinced that the current focus should be on "How can we
measure natural attenuation processes, like biodegra-
dation, during the site characterization process?" and
"How can we reduce the cost of sampling and analysis
procedures?" Current efforts center around the develop-
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.
ment 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 pro-
tocol for chlorinated organics is under development and
is currently in draft form. Development and field valida-
tion of this protocol are the first and most necessary
steps toward the development of an integrated risk man-
agement strategy for this class of sites. Natural attenu-
ation is not selected presumptively. In fact, the selection
of natural attenuation alone or in concert with supple-
mental 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., degra-
dation byproduct formation, electron acceptor usage,
or redox potential).
In some cases, it may be beneficial to supplement the
above evidence with laboratory microcosm data. Labo-
ratory microcosms can be used to demonstrate that
under specific conditions the constituent of concern is
biodegraded, as well as to provide estimates of degra-
dation rates. Microcosm data are particularly relevant if
the laboratory conditions closely approximate field con-
ditions. Confirmed contaminant removal and contain-
ment 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 elec-
tron 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, natu-
ral 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 dehalogena-
tion of chlorinated organics requires the presence of a
substrate or electron donor. As a result of this depend-
ence, the draft Technical Protocol for Natural Attenu-
ation of Chlorinated Solvents in Ground Water (4)
outlines different types of plume behavior.
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Types of Plume Behavior
Type 1 behavior is associated with areas within the
plume where anthropogenic carbon (e.g., petroleum hy-
drocarbons 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 behav-
iors can be exhibited within a single plume simultane-
ously or in various combinations. The behavior profile
could also change overtime.
These plume behaviors or profiles are important from
the perspective of 1) determining whether site constitu-
ents 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 con-
tainment can be achieved with natural attenuation alone
or in concert with supplemental remedial systems. An-
other primary remedial objective is to achieve reductions
in the concentrations of constituents of concern to risk-
based 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 pres-
ence 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 com-
pounds and to be biodegradable under aerobic and/or
iron-reducing conditions. Thus, with respect to risk man-
agement, a point of critical importance would be to
determine whether the vinyl chloride will enter into a
zone where the geochemical conditions support biode-
gradation. In certain instances, the zone where the con-
stituent^) of concern is located or the area directly
downgradient can be viewed and managed as an effec-
tive treatment zone.
Natural Attenuation and Alternative Plume
Management
Understanding natural attenuation effects allows the
consideration and implementation of alternative reme-
dial approaches. Pump and treat systems are tradition-
ally 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 biodegrada-
tion of vinyl chloride. Alternative methods of ground-
water aeration are also possible. Pumping upgradient of
source zones could slow ground-water velocities suffi-
ciently so that the kinetics of biodegradation are com-
patible 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 ap-
proaches (e.g., soil vapor extraction [SVE]). This com-
bination 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 re-
quirements, and operation and maintenance require-
ments 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 environ-
ment often contained a combination of degreasing agent
and solvated petroleum hydrocarbons. Chlorinated sol-
vents 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 petroleum-
contaminated 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
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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 as-
similative capacity of the ground water. This concept of
supplemental source removal should be directly appli-
cable to sites where chlorinated organics are the single
predominant species. Source removal processes that
preferentially remove petroleum hydrocarbons or in ef-
fect 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 de-
commissioned fire training area where waste chlorin-
ated solvents, primarily TCE, are codistributed with
petroleum hydrocarbons. Free-phase petroleum hydro-
carbons are present and detectable in significant appar-
ent 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 recov-
ery 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. In-
itial 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 dehalo-
genation 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 do-
nor supply. This concept is being extended for discus-
sion 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 commu-
nity; 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 re-
Wiedemeier, T.H. 1996. Personal communication between the
author and Todd H. Wiedemeier, Parsons Engineering Science, Inc.,
Denver, CO.
moved 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 or-
ganics depends on substrate or electron donor availabil-
ity, any alteration could dramatically affect transport. The
main point of this discussion is to underline the impor-
tance 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 Proto-
col 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 parame-
ters 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 empha-
sis will be on the inclusion of natural attenuation as a
remedial alternative alongside and in concert with con-
ventional remedial alternatives—to include the "no ac-
tion alternative." Thus, this approach is an inclusive
analysis, not an exclusive analysis. The AFCEE hopes
to promote the proper consideration of the natural at-
tenuation alternative since this alternative, to date, has
probably been excluded more often than it has been
included.
The AFCEE Technology Transfer Division is also inves-
tigating the technical and economic feasibility of other
remedial technologies and their applicability. Also, the
compatibility of each technology with the natural attenu-
ation alternative is being considered. For Example, AF-
CEE has installed and is currently evaluating a zero
valence iron passive treatment wall to determine
whether this technology can provide long-term site re-
mediation. Source reduction approaches such as sur-
factant-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 ap-
plicability, cost effectiveness, and reliability.
The above examples are not meant to be an exhaustive
list of all technologies to be considered in feasibility
127
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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 per-
formance 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 manage-
ment strategy shall emphasize plume management
concepts within current regulatory frameworks to in-
clude 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 regula-
tory framework. The use of natural attenuation as a
treatment alternative, however, is considered scientifi-
cally supportable, and when properly applied it is con-
sidered 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, WE. 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 Re-
mediation 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, Ver-
sion 1.30. San Antonio, TX: U.S. Air Force Center for Environ-
mental Excellence.
4. Weidemeier, T.H., M.A. Swanson, D.E. Moutoux, J.T. Wlson, D.H.
Kampbell, J.E. Hansen, P. Haas, and F.H. Chapelle. 1996. Tech-
nical Protocol for Natural Attenuation of Chlorinated Solvents in
Ground Water. San Antonio, TX: U.S. Air Force Center for Envi-
ronmental Excellence. In preparation.
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Intrinsic Remediation in the Industrial Marketplace
David E. Ellis
DuPont Specialty Chemicals-CRG, Wilmington, Delaware
Introduction
Intrinsic remediation of chlorinated solvents is a com-
mon 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 remedia-
tion is a safe and effective remedy at individual sites.
Intrinsic remediation is an important development for
industry because it protects human health and the envi-
ronment yet is more cost-effective than the competing,
intrusive ground water remediation techniques.
When To Consider Intrinsic Remediation
Decision-makers should determine whether the follow-
ing 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 en-
vironmental media or disrupting adjacent ecosystems.
• The rate of degradation balances the rate of migra-
tion 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 re-
mediation demonstration considers the extent of the
data-gathering effort and the cost of the resources re-
quired. DuPont has developed the following list of mini-
mum 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, cal-
cium, 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 greaterthe 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 rela-
tively small. This incremental cost may be more than
offset if intrinsic remediation is chosen over a technology
129
-------
Table 1. Tiered Site Characteristics
Tier 1 (Easy Sites)
Tier 2 (Intermediate Sites)
Tier 3 (Difficult Sites)
Simple hydrogeology
Single parent compound
Size known (areal extent)
Source and mass known
Highest ground-water concentration < 10 mg/L
Static or shrinking plume
Bioindicators obvious
Receptors very far away
Analytical model sufficient
Moderately complex hydrogeology
Few contaminants
Plume size questionable
Source and mass not well defined
Highest ground-water concentration < 100 mg/L
Plume-size trend not known
Some bioindicators
Receptors are "not too far"
Flow-and-transport model needed
Hydrogeology complex
Confusing mixture
Large plume
Very large or inaccessible source
Mobil DNAPLs
Growing plume or trend not known
No bioindicators
Receptors close or affected
Needs a detailed fate-and-transport
model
that would be more expensive to implement. This con-
clusion 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 dis-
count 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 monitor-
ing is approximately $100,000 over 30 years. The sim-
plest pump-and-treat remedy (air stripping and vapor-
phase 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 biode-
gradation 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 litera-
ture. To be included, the sites needed to have either
SW846 Method 8240 analyses for VOCs, good geologi-
cal 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 c/s-1,2-DCE isomer, that is conclu-
sive evidence of the biological degradation of trichlo-
roethene (TCE). The biodegradation results are
presented in Table 3. The data showed that:
• 88 percent of the sites have bacteria that can biode-
grade PCE and TCE to DCE.
Table 2. Present Cost of Intrinsic Remediation Versus Investigation and Long-Term Monitoring
Cost Element
Up front
Annual
Present cost (30 years)
Investigation
and Long-Term
Monitoring Cost
$95,000
$62,000
$800,000
Intrinsic
Remedy
Cost
$35,000
$68,000
$900,000
Incremental
Cost — Intrinsic
Versus Investigation
and Monitoring
$40,000
$6,000
$100,000
Simple
Pump-and-
Treat Cost
$650,000
$35,000
$2,100,000
Incremental Cost —
Pump-and-Treat
Versus Intrinsic
$515,000
$67,000
$1 ,200,000
Note: 12 percent discount rate, 3 percent annual inflation.
130
-------
Table 3. Biodegradation Results at Survey Sites
Reaction
PCE to TCE
TCE to DCE
DCE to VC
VC to ethane
Number of
Sites Present
27
39
28
18
Total
Sites
31
44
37
31
Percentage
87
88
75
58a
Ethane data often are unavailable.
• 75 percent of the sites have bacteria that can biode-
grade DCE to vinyl chloride (VC) or ethylene.
Some sites may not have enough bioavailable substrate
to complete the degradation reactions. Insufficient sub-
strate should always be suspected at sites where biode-
gradation stops at either TCE or VC. Sites without the
full bacterial population needed for complete degrada-
tion 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, disper-
sion, and sorption factors are constant; measures con-
centration declines along the centerline of plumes on
high-quality isoconcentration maps; and calculates the
time for the water package to move each of those dis-
tances. 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 aver-
age solvent half-lives are shown in Table 4.
Table 4. Half-Lives Calculated by Graphical Extraction
Reaction Average Half-Life (years) Number of Sites
PCE to TCE
TCE to DCE
DCE to VC
VC to ethane
1.20
1.19
1.05
1.22
7
15
12
9
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 reac-
tions. Based on this assumption, one can calculate the
amount of chlorinated solvent that an aquifer can biode-
grade, although this estimate can only apply to sites at
which the soils contain bacteria that can degrade chlo-
rinated solvents.
Here is an example calculation. Typical aquifers contain
between 0.3 percent and 1 percent natural organic carb-
on. 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 degra-
dation. 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 hypo-
thetical 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 re-
quired for VC. Why shouldn't chlorinated solvent pump-
and-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 sys-
tems have already reached their useful lifetime for con-
taminant removal.
All of the following criteria should be met before intrinsic
remediation can replace an existing, operating pump-
and-treat system that treats chlorinated solvents:
• It can be demonstrated that intrinsic activity is al-
ready occurring in the aquifer.
• It is possible to predict how far the plume might
extend if the pump-and-treat system was not oper-
ating, and it can be shown that no receptor will be
affected.
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 effec-
tiveness of intrinsic remediation are plume residence
time and the half-lives of the sequential biodegrada-
tion reactions.
• Most aquifers contain much more organic carbon
than necessary to support intrinsic bioremediation.
While anthropogenic carbon may help support intrin-
sic 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 hand-
book on intrinsic remeidation of chlorinated solvents, http://www.rtdf.org.
2. Buscheck, I.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, Hous-
ton, 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 transfor-
mation 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 Con-
ference on Intrinsic Remediation, London, UK.
132
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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
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 com-
pounds 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 trans-
port 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 tetra-
chloride in ground water usually involves a biochemical
process described as sequential reductive dechlorina-
tion. This process only occurs in the absence of oxygen,
and the chlorinated solvent actually substitutes for oxy-
gen 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 chemi-
cal compounds of metal atoms coupled with other materi-
als. 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 elec-
trons from the neutral metal to produce an oxide. Chem-
ists have generalized the term "oxidation" to refer to any
chemical reaction that removes electrons from a mate-
rial. For a material to be reduced, some other material
must be oxidized.
The electrons required formicrobial reduction of chlorin-
ated solvents in ground water are extracted from native
organic matter, from other contaminants such as the
benzene, toluene, ethylene, and xylene compounds re-
leased from fuel spills, from volatile fatty acids in landfill
leachate, or from hydrogen produced by the fermenta-
tion 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, dis-
solved sulfate, a dissolved chlorinated solvent, or carb-
on dioxide. Important oxidized minerals used as electron
acceptors include iron and manganese. Oxygen is re-
duced to water, nitrate to nitrogen gas or ammonia,
iron(lll) or ferric iron to iron(ll) or ferrous iron, manga-
nese(IV) to manganese(ll), sulfate to sulfide ion, chlo-
rinated 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, reduc-
tive 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 methano-
genesis a marginal amount of energy. The organisms
carrying out the more energetic reactions have a com-
petitive advantage; as a result, they proliferate and ex-
haust 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 reduc-
tion 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 pro-
duced is generally referred to as a reduced environment.
Reductive dechlorination usually occurs under sulfate-re-
ducing 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 (hydro-
gen 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, tetrachlo-
roethylene is reduced to trichlorethylene, then any of the
three dichloroethylenes, then to monochloroethylene
(commonly called vinyl chloride), then to the chlorine-
free 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 publish-
ed 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 ref-
erence 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)
St. Joseph, Ml
Picatinny
Arsenal, NJ
Sacramento,CA
Necco Park, NY
Pittsburgh
AFB, NY
Tibbitt's Road, NH
San Francisco
Bay Area, CA
Perth, Australia
Eielson AFB, AK
1-3
4, 5
6
7
Weidemeider,
this volume
B. Wilson,
this volume
9
10
130 to 390
390 to 550
550 to 855
240 to 460
320 to 460
240 to 320
0 to 250
70 to 300
0 to 570
0 to 660
0 to 300
300 to 380
380 to 780
0 to 24
0 to 40
0 to 55
0 to 600
3.2 to 9.7
9.7 to 12.5
12.5 to 17.9
2.2 to 4.2
2.9 to 4.2
2.2 to 2.9
0.0 to 2.3
0.5 to 2.3
0.0 to 1.6
0.0 to 1.8
0.0 to 6.7
6.7 to 8.6
8.6 to 17.7
0.0 to 2.4
0.0 to 6.4
0.0 to 10
0.0 to 14
(years)
6.5
2.8
5.4
2.0
1.3
0.7
1.8
1.6
1.8
6.7
1.9
9.1
2.4
6.4
10
Apparent Loss Coefficient (1/year)
0.38
1.3
0.93
1.4
1.2
0.50
0.83
3.1
Produced
Produced
1.6
0.5
0.18
0.88
2.2
Produced
Produced
1.1
0.7
0.7
1.3
0.23
Absent
4.4
0.86
5.11
0.32
0.73
2.3
3.1
Produced
0.6
0.07
0.21
0.42
0.73
Produced
1.16
0.47
Produced
0.68
>0.73
Not identified
Cecil Field
NAS, FL
11 0.8 0.
Chapelle, 0 to 140 0.0 to 1.2 1.2 3.3 to 7.3
this volume
8 0.8
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 demon-
strated evidence of biodegradation. Most of the first-or-
der 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 simi-
lar to estimates of removal at field scale for TCE, cis-
DCE, and vinyl chloride. Rates of removal of
1,1,1-trichloroethane (1,1,1-TCA) aresimilartothe rates
of removal of the chlorinated alkenes.
TCE Removal in Field
I
I
"ni
o
o
S
a
tf.
I
O
M
10 11 12 13 14 15 16 17
Sites
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 con-
centrations to current maximum contaminant levels for
Figure 1. The first-order rate constant for biotransformation of
TCE in a variety of plumes of contamination in ground
water.
drinking water. Biodegradation as a component of natu-
ral 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.
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)
Laboratory Microcosm Studies Done on Material From Field-Scale Plumes
Picatinny
Arsenal, NJ
St. Joseph, Ml
Traverse City, Ml
Tibbitts Road, NH
12
13
14
15
16
240
320
460
300
At Source
2.2
2.9
4.2
0.5
0.5
0.5
0.12, 0.077
1.8
Apparent First-Order Loss (1/year)
Apparent Zero OrderLoss
0.64
0.42
0.21
1.8, 1.2
1.8
4.8
0.52
9.4
3.1
Laboratory Microcosm Studies Done on Material Not Previously Exposed to the Chlorinated Organic Compound
Norman
Landfill, OK
FL
17
18
16
19
Aerobic
material
Sulfate
reducing
Methan-
ogenic
Reducing
Reducing
4.2
10
1.28
7.62
1.20
7.65
3.6
0.012
7.75
7.42
135
-------
References
1. Semprini, L, P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson. An-
aerobic 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 Com-
putational Methods, August 7-9, Bay City, Ml.
3. Wilson, J.T, J.W. Weaver, D.H. Kampbell. 1994. Intrinsic biore-
mediation 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/R-
94/162. pp. 116-119.
4. Ehlke, T.A., B.H. Wilson, J.T. Wilson, and I.E. Imbrigiotta. 1994.
In-situ biotransformation of trichloroethylene and cis-1,2-dichlo-
roethylene at Picatinny Arsenal, New Jersey. In: Morganwalp,
D.W, and D.A. Aronson, eds. Proceedings of the U.S. Geological
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5. Martin, M., and T.E. Imbrigiotta. 1994. Contamination of ground
water with trichloroethylene at the Building 24 site at Picatinny
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tenuation 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 se-
quential 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., PR Mazierski, R.J. Buchanan, Jr., D.E. Ellis, and L.S.
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transformations of 1,1,1-trichloroethane in subsurface soils and
ground water. Environ. Toxicol. Chem. 9:1437-1451.
19. Barrio-Lage, G.A., RZ. Parsons, R.M. Narbaitz, and PA. Lorenzo.
1990. Enhanced anaerobic biodegradation of vinyl chloride in
ground water. Environ. Toxicol. Chem. 9:403-415.
136
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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 be-
cause the responsible microorganisms are ubiquitous in
soil and the subsurface. Bacteria able to use hydrocar-
bons as their source of carbon and energy under either
aerobic or anaerobic conditions have a tremendous se-
lective advantage over other members of the microbial
community. Therefore, the process can be self-sustain-
ing and is limited only by the presence of electron ac-
ceptors or inorganic nutrients.
Bacteria able to grow at the expense of chlorinated aliphatic
compounds are not widely distributed; natural attenu-
ation of such compounds is consequently less predict-
able. The use of a chlorinated compound as a terminal
electron acceptor (chlororespiration or dehalorespira-
tion) can yield energy and thus provides a selective
advantage to a limited range of anaerobic bacteria (1).
Many of the transformations of chloroaliphatic com-
pounds, such astrichloroethylene, 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 hydrocar-
bons 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 com-
pounds that are known to be biodegradable and have
the potential for natural attenuation in the field. Some
synthetic chemicals are expected to be readily suscep-
tible 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 infor-
mation 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 biodegrada-
tion 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. Transforma-
tion 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 de-
termine whether the process provides energy or nutri-
ents—and thus a selective advantage—to the bacteria
that catalyze the reactions.
The second question is whether the activity of the nec-
essary specific organisms is present at the site under
consideration. A considerable amount of effort has been
spent on enumerating and identifying bacteria at hydro-
carbon-contaminated sites under consideration for
bioremediation. Because such bacteria are ubiquitous,
it is much more useful to assess their activity as re-
vealed by degradation of the hydrocarbons or transfor-
mation 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 bac-
teria 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
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must be resolved. A good understanding of the biode-
gradation process, including reaction stoichiometry and
kinetics, is essential for evaluation of the potential for
natural attenuation. Fortunately, such understanding ex-
ists for a wide range of synthetic chemical contaminants.
Chloroaromatic Compounds
Bacteria able to degrade all but the most complex chlo-
roaromatic compounds have been discovered during
the past 20 years. Polychlorobenzenes, including hex-
achlorobenzene, can be sequentially dehalogenated to
monochlorobenzene under methanogenic conditions in
soil slurries (3). Reductive dehalogenation of chloroben-
zene 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 (HCI).
Chlorobenzenes are very good candidates for natural
attenuation under either aerobic or anaerobic condi-
tions. Aerobic bacteria able to grow on chlorobenzene
have been detected at a variety of chlorobenzene-con-
taminated sites but not at uncontaminated sites (11).
This provides strong evidence that the bacteria are se-
lected for their ability to derive carbon and energy from
chlorobenzene degradation in situ. Removal of multiple
halogens as HCI consumes a large amount of alkalinity
and produces a considerable drop in the pH of unbuf-
fered systems, which could lead to a loss of microbial
activity at some sites.
Chlorophenols and chlorobenzoates are dehalogenated
under anaerobic conditions in sediments and subsur-
face material (12-13). In some instances, the dehalo-
genation clearly yields energy for the growth of specific
bacteria. In other examples, the dehalogenation is spe-
cific 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 toxic-
ity and stimulating degradation of Chlorophenols (12).
Pentachlorophenol deserves special consideration be-
cause it has been widely used as a wood preservative
and has been released into the environment through-
out the world. Reductive dehalogenation of pentachlo-
rophenol under methanogenic conditions can lead to
mineralization (12). Aerobic bacteria catalyze the re-
placement of the chlorine in the 4 position by a hy-
droxyl group to form tetrachlorohydroquinone, and
subsequent reductive dehalogenations lead to the for-
mation of ring fission substrates. Bacteria able to de-
grade 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 contami-
nated with complex mixtures of organic compounds, so
investigations of toxicity must be conducted for each site
under consideration. Natural attenuation of pentachlo-
rophenol seems to be possible because specific bacte-
ria 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 ex-
tensively because of their stability, toxicity, and bioaccu-
mulation 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 advan-
tage 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 de-
halogenation. Dehalogenation converts the more highly
chlorinated congeners to less chlorinated products con-
taining 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 bioaccumula-
tion of the mixture.
A variety of different dechlorination patterns have been
identified as a function of the microbial community in-
volved. The patterns are constant within a given microbial
138
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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 deha-
logenate PCB. The electron donors for the dehalogena-
tion in sediment are unknown. The addition of
exogenous carbon sources does not stimulate the reac-
tion. In contrast, "priming" the mixtures with low levels
of bromobiphenyl or specific isomers of tetrachloro-
biphenyl (15) seems to selectively enrich a population
of PCB-dechlorinating bacteria and dramatically stimu-
late 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-di-
oxygenase, followed by a sequence of reactions that
lead to ring cleavage and the accumulation of chlo-
robenzoates 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 contami-
nated river sediments (18).
Clearly, reductive dechlorination is ongoing at a wide
range of PCB contaminated sites. The strategy of an-
aerobic 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 com-
plete 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) chlo-
roaliphatic compounds (19-21); therefore, this paper
briefly mentions only some aspects that might otherwise
be overlooked. Among the one- and two-carbon chlorin-
ated compounds, the more highly chlorinated molecules
are subject to reductive dehalogenation under a variety
of conditions. Thus, carbon tetrachloride can be se-
quentially reduced to chloroform and dichloromethane.
Similarly, perchloroethylene can be reduced to ethylene
via trichloroethylene, dichloroethylene, and vinyl chlo-
ride. 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 distrib-
uted in anaerobic ecosystems and catalyzes the slow and
often partial reduction of chlorinated contaminants. In
contrast, some microbial communities and a few iso-
lated strains can derive energy from the use of chlorin-
ated 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, al-
though they can be degraded co-metabolically. In con-
trast, 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 can-
didates for natural attenuation where oxygen is present.
Aerobic mineralization of the related molecule, ethylene
dibromide, has been reported in soil (24), but the distri-
bution of the responsible bacteria and the corresponding
ability to predict degradation are not well understood.
Nitroaromatic Compounds
The literature on biodegradation of nitroaromatic com-
pounds 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 at-
tenuation. In contrast, aerobic bacteria able to grow on
nitrobenzene, nitrotoluenes, dinitrotoluenes, dinitroben-
zene, nitrobenzoates, and nitrophenols have been iso-
lated from a variety of contaminated sites, which suggests
that natural attenuation is taking place at such sites. The
simple nitroaromatic compounds (not including trinitro-
toluene) can be considered excellent candidates for
natural attenuation. Some of the compounds, including
3-nitrophenol, nitrobenzene, 4-nitrotoluene, and 4-ni-
trobenzoate, are degraded via catabolic pathways that
involve a partial reduction of the molecule prior to oxy-
genative 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 prob-
lematic for microbial degradation. For example, the in-
dustrial 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-dinitrotolu-
ene 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
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and insight about the metabolic pathway might allow
better prediction of the mixture's degradation.
Ketones
Acetone and other ketones are not xenobiotic com-
pounds, but most of the current production is via syn-
thetic 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-terf-butyl
ether (MTBE), and terf-butyl alcohol are used extensively
as octane enhancers in unleaded gasoline. The ether
bond of MTBE makes it particularly resistant to biodegra-
dation. Its water solubility, low volatility, and high concen-
trations 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 en-
riched from refinery and chemical plant waste treatment
systems (24),2 so bacteria clearly can successfully at-
tack 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 terf-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 nitro-
gen 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 attenu-
ation is possible, but more information is needed on the
bioavailability, toxicity, and kinetics of the process.
1 Weaver, J. 1996. Personal communication with the author.
2 Cowan, R. 1996. Personal communication with the author.
Pesticides
Most pesticides used in the past 20 years in the United
States have been formulated to degrade in the environ-
ment, and a considerable amount of information is avail-
able on degradation kinetics in soil and water. The U.S.
Environmental Protection Agency Risk Reduction Engi-
neering Laboratory in Cincinnati, Ohio, has developed
an extensive Pesticide Treatability Database that con-
tains information on a variety of compounds. Many pes-
ticides hydrolyze and yield compounds that serve as
growth substrates for bacteria. For example, car-
bamates, chlorophenoxyacetates, dinitrocresol, cou-
maphos, 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 de-
rived from soil bacteria but provide no advantage to the
organisms that produce the enzymes. Similarly, some of
the organohalogen insecticides can be reductively de-
halogenated but provide no advantage to specific organ-
isms. 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 envi-
ronment are known to be biodegradable by bacteria, and
much is known about the processes and their require-
ments. The potential for natural attenuation of biode-
gradable contaminants should be considered before
more costly and disruptive treatment options.
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141
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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 aqueous-
phase 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 proc-
esses in unconsolidated media than in rock. This paper
addresses the potential importance of both natural at-
tenuation in other media and of slower processes. Spe-
cifically, 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 prob-
lem. 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 per-
meability, 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 contamina-
tion 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 dechlorina-
tion near the source, very likely driven by the carbon
source in the nonchlorinated sludge. Downgradient con-
ditions 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 con-
cerning 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 sedi-
ments 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 contami-
nant 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
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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 gal-
lons 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 dis-
solved 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 laborato-
ries 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 proc-
ess followed by degradation in ground water. This is an
area which has been largely overlooked, and the re-
search needed to evaluate these mechanisms will re-
quire a longer-term and significantly different approach
than is current practice; however, to achieve a reason-
able 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 aqueous-
phase in the aquifer. Little attention has been given to the
vadose zone or discharge points. Any attenuation mecha-
nism 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 im-
portant 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 signifi-
cance 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
1 Parsons Engineering Science. 1996. Unpublished compilation of
data from six chlorinated solvent sites at Hill Air Force Base, UT.
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 contami-
nant concentrations, this would not be a helpful mecha-
nism. 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 contami-
nant; 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-per-
meability deposit from Lake Bonneville occurs just below
this depth that causes the ground water to discharge. This
discharge takes several forms: evaporation, evaportran-
spiration, 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 cer-
tainly 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 impor-
tant 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 evapotranspora-
tive 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 contami-
nation (TSF-23) and miscellaneous no action sites, final remedial
action. Idaho National Engineering Laboratory, Idaho Falls, ID.
143
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Poster Session
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Anaerobic Mineralization of Vinyl Chloride in Iron(lll)-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 contami-
nants and the apparent poor biodegradability of vinyl
chloride under anaerobic conditions. Previous investiga-
tions have suggested that reductive dechlorination of
vinyl chloride to ethene may represent an environmen-
tally significant pathway for in situ bioremediation of
vinyl chloride contamination. This poster provides labo-
ratory evidence for an alternative mechanism of vinyl
chloride degradation: anaerobic oxidation of vinyl chlo-
ride under iron(lll)-reducing conditions.
Microcosm experiments conducted with material col-
lected from two geographically isolated, chlorinated-
ethene-contaminated aquifers demonstrated oxida-
tion of [1,2-14C]vinyl chloride to 14CO2 by indigenous
microorganisms under iron(lll)-reducing conditions.
Addition of chelated iron(lll) (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(lll)-reducing conditions, and that the
bioavailability of iron(lll) is an important factor affect-
ing 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(lll) re-
duction may be an environmentally significant mecha-
nism contributing to intrinsic bioremediation of vinyl
chloride in anaerobic ground-water systems.
146
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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
Me Master 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 aro-
matic 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 condi-
tions (2) may result in an underestimation of cis-1,2-
DCE reductive dechlorination. Toluene and methane are
present in the downgradient aerobic ground water, how-
ever, and are likely promoting co-metabolic biodegrada-
tion of cis-1,2-DCE. Preliminary laboratory microcosm
studies have confirmed that the indigenous microorgan-
isms 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. Wil-
son, 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(lll)-reducing, aquifer sediments. Environ. Sci. Tech-
nol. In press.
147
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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 chromatog-
raphy 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 head-
space, 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 compo-
nents are separated, and chromatogram peak retention
times and areas are compared with calibration stand-
ards. The concentration of the aqueous gas components
can be calculated using sample temperature, bottle vol-
ume, headspace concentrations, and Henry's Law.
Limits of detection for methane and ethylene are
0.001 and 0.003 milligrams perliter(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 us-
ing three replicates was a standard deviation of
8.8 mg/L, RSD=7.5 percent, and an average recov-
ery 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 possi-
bly other dissolved gases in ground water.
148
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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, sorp-
tion, and diffusion. Determination of quantitative degra-
dation 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 af-
fecting 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 dis-
solved chemicals that affect biodegradation of chlorin-
ated ethenes within the plume. Results indicated that
ground water within the plume contained TCE concen-
trations ranging up to 20 mg/L"1, methane concentra-
tions 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(lll) 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 determi-
nation of TCE and cis-DCE biodegradation rates in a
laboratory study.
Static batch microcosms were constructed under an-
aerobic conditions to determine the rates of TCE and
cis-DCE biodegradation. Sterilized 50-milliliter serum vi-
als 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 pore-
water 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. Geo-
chemical 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 biodegra-
dation rate constants for cis-DCE ranged from less than
-0.01 wk"1 to-0.05wk"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 concen-
trations in ground water at up- and downgradient wells
measured at time intervals corresponding to the esti-
mated 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 esti-
mates 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 mod-
eling were used to quantify the removal of TCE by major
processes in the unconfined aquifer at Picatinny Arsenal.
149
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Diffusive, sorptive, and volatilization losses were esti- TCE in ground water, discharging to Green Pond Brook.
mated separately and used to correct for apparent in situ Volatilization and lateral diffusive losses of TCE from the
TCE removal. Biodegradation is probably the major re- plume are estimated to total 10 to 50 kg/y"1. Sorptive
moval process for chlorinated ethenes in the aquifer, losses of TCE to aquifer soils are minor because of the
removing about 400 kg/y"1 of TCE from the contaminant low organic carbon concentration of sediments in the
plume. Advective transport removes about 47 kg/y"1 of saturated zone.
150
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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 sam-
pling 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 oper-
ates on the reaction principle of X + HgO (solid) ->XO
+ Hg (vapor), where X represents any reducing gas. An
ultraviolet photometer quantitatively measures the resul-
tant 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) hydro-
gen. 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 cor-
responds to 0.8 nanomoles per liter of dissolved hydro-
gen 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
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Evidence of Natural Attenuation of Chlorinated Organ ics 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 con-
sumed. 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 centi-
meters per second. The water table is generally 12 feet
below the ground surface.
Sampling activities conducted in 1993 and 1994 indi-
cated significant concentrations of chlorinated organics
(1,2-dichloroethene [1,2-DCE], trichloroethene, and per-
chloroethene) 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 pre-
sent 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 concen-
tration of the chlorinated organic contaminant plume
during the period of sampling (e.g., peak 1,2-DCE con-
centrations 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 disper-
sion, sorption, volatilization, and biological degradation.
The bulk of the ground-water contamination was re-
cently remediated using air sparging/soil vapor extrac-
tion. 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 rec-
ommended remedial alternative for remaining ground-
water 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 reme-
diation 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
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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. Lige
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 sol-
vent 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 gaso-
line product, gasoline vapors, and high levels of dis-
solved 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 hydrocarbon-
contaminated 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, imply-
ing degradation of PCE to TCE in the presence of
hydrocarbon vapors. Even though numerous ground-
water samples were obtained for the site charac-
terization 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 occur-
ring 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 char-
acterization efforts. Sampling and analysis not conven-
tionally 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 inter-
mingling 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 Pro-
gram at Dover Air Force Base. Presented at the IBC Intrinsic
Bioremediation/Biological Dehalogenation Conference, Annapolis,
MD. October 16-17.
4. Lige, J.E., I.D. MacFarlane, and T.R. Hundt. 1995. Treatability
testing to evaluate in situ chlorinated solvent and pesticide biore-
mediation. In: Hinchee, R.E., A. Leeson, and L. Semprini, eds.
Bioremediation of chlorinated solvents. Columbus, OH: Battelle
Press.
153
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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 re-
sidual sludge from cleaning baths. The sludge con-
tained chlorinated solvents, including trichloroethylene
(TCE), tetrachloroethylene (PCE) and 1,1,1-trichlo-
roethane (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-and-
treat 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. Dur-
ing a review of the historical data, it was determined
that the concentration of chlorinated solvents had
greatly decreased before implementation of the pump-
and-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 tech-
niques; evidence of anaerobic conditions and of the an-
aerobic 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 c/s-1,2-
dichloroethylene and vinyl chloride were detected nearthe
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 de-
tected 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 pro-
gram is being implemented to track and ensure that
adequate remediation of the site continues by intrinsic
bioremediation. Implementation of this program is allow-
ing redevelopment of this site.
154
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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 Excel-
lence, Technology Transfer Division (AFCEE/ERT), in
conjunction with personnel from the U.S. Environmental
Protection Agency's National Risk Management Re-
search 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 chlo-
rinated solvents. The intended audience for the new
protocol is U.S. Air Force personnel and their contrac-
tors, 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 sol-
vents include biodegradation, hydrolysis, volatilization,
advection, dispersion, dilution from recharge, and sorp-
tion. Patterns and rates of natural attenuation can vary
markedly from site to site depending on governing physi-
cal and chemical processes. The proposed protocol
presents a straightforward approach based on state-of-
the-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
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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 in-
trinsic 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 appli-
cability of currently available intrinsic remediation proto-
cols (e.g., the U.S. Air Force guidance document by
Wiedemeier et al. [1]) to include any biodegradable con-
taminant, not just benzene, toluene, ethylbenzene, and
xylenes and related (fuel-derived) compounds. To ex-
tend the protocol in this manner, the biodegradability of
any contaminants that may exist at these sites must be
addressed because the knowledge of contaminant bio-
degradability can be an absolute requirement for appli-
cation of intrinsic remediation. How to go about this is
the focus of the work.
Progress on the project to date has been the develop-
ment 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 technol-
ogy 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 biode-
gradability 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 con-
cerns, that can be used to select and implement the
appropriate approach for making a detailed assess-
ment 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 reme-
diation 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
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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 contamina-
tion, 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 analy-
sis at Loring Air Force Base, Limestone, Maine, and pre-
sents 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 commu-
nity evaluations. Several of the contaminant plumes
demonstrated characteristics of reductive dehalogena-
tion, 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 po-
tential values and sulfate concentrations decreased,
and methane concentrations were observed at locations
where chlorinated ethenes were detected.
157
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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 un-
derground storage tanks at the Colorado Department of
Transportations materials testing laboratory in Denver,
Colorado. An estimated 5,000 to 15,000 liters of trichlo-
roethene, 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 be-
tween 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 dechlori-
nation (e.g., 1,1-DCE, 1,2-DCE, 1,1-DCAand 1,2-DCA),
has migrated in excess of 4,500 feet off site.
This site is being characterized for intrinsic bioattenu-
ation to establish baseline conditions prior to the poten-
tial implementation of a source removal action,
recognizing that substantial residuals would likely re-
main. An anaerobic core in the source area has been
characterized on the basis of water chemistry differ-
ences 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 head-
space analysis after displacing a fixed volume of water
by nitrogen gas displacement. Aqueous samples were
analyzed for NO3-, PO4-3, SO4-2, CI', S'2, and Fe+2
using Hach methodologies; total organic carbon, chemi-
cal 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 down-
gradient 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 bioattenu-
ation has been the identification of low levels (15 parts
per billion) of vinyl chloride immediately downgradient
of the source area.
158
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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 effec-
tively 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 concen-
trations 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
(u,g/L); by 1994, the concentration ranged from 11,000 u,g/L
to BDL. The RDX concentration ranged from 7,600 to 33.8
u,g/L in 1990 and from 8,400 to 14.4 u,g/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-di-
mensional ground-water flow and transport to this het-
erogeneous aquifer. The capability of a comprehensive
computer graphical system—Groundwater Modeling
System (CMS)—which is used in the modeling of the
site, also will be discussed and illustrated.
159
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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 se-
lecting natural attenuation and degradation fortreatment
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. Aground-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 ex-
plosive 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 am-
munition. Dissolved VOCs present in the highest concen-
trations 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. Be-
cause 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
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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 attenu-
ation as remediation alternatives for chlorinated solvent
sites is rapidly growing because the methods are signifi-
cantly more cost-effective than conventional remedia-
tion alternatives (e.g., pump-and-treat). When evaluating
the long-term cost-effectiveness of intrinsic bioremedia-
tion and natural attenuation alternatives, however, many
analysts and decision-makers consider direct engineer-
ing 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 market-
ability 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 ob-
jectively whether intrinsic bioremediation and natural
attenuation are in fact the most cost-effective remedia-
tion alternatives in the long run.
This poster presents the results of using financial-risk
cost-benefit analysis to examine the impact of cost fac-
tors other than engineering costs on the long-term cost-
effectiveness of intrinsic bioremediation versus other
remediation alternatives under various scenarios at two
chlorinated solvent sites. At both sites, chlorinated vola-
tile organic compounds are currently being intrinsically
bioremediated to environmentally acceptable end prod-
ucts (e.g., ethene and ethane).
161
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Natural Attenuation of Chlorinated Organ ics 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 suit-
able sites for development (including the development
of nature reserves and ecological values) and by con-
sciousness 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 compara-
ble to the U.S. Environmental Protection Agency's reme-
dial investigation/feasibility study, remedial technologies
are identified and remedial action alternatives evalu-
ated. The evaluation is based on technical and financial
feasibility on one hand and environmental impact on the
other. Environmental impact is substantiated by analy-
ses 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 contamina-
tion 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 drink-
ing water supply.
Target and intervention values for these compounds in
ground water are set low, as illustrated below.
Intervention
Target Value Value From the
From the Soil Soil Protection
Protection Act Act
Trichloroethylene (TCE) 0.01 |ig/L
Perchloroethylene (PCE) o.01 |ig/L
Vinylchloride (VC) Not available
Provisional
target value
cis-1 ,2-Dichloroethene (DCE) 0.01 |ig/L
500 ng/L
40 ng/L
0.7 ng/L
Provisional
intervention value
10ng/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 emerg-
ing. 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 insti-
tutes, and environmental consultancy firms/contractors.
The results of the various projects are accessible for
NOBIS participants immediately and will become public
domain later.
162
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Natural Attenuation as a Cleanup Alternative for Tetrachloroethylene-Affected
Ground Water
Steve Nelson
EMCON, Bothell, Washington
A chlorinated solvent storage and transfer facility op-
erated in an industrial area of Seattle, Washington,
from the mid-1940s to the mid-1970s. Historical re-
leases 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 investiga-
tion 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. Addi-
tional ground-water chemistry data, including nitro-
gen, phosphorus, iron, sulfur, dissolved oxygen, and
permanent gas (methane, ethane, ethene) concentra-
tions, were collected. Elevated concentrations of fer-
rous 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 concen-
trations increase by one to two orders of magnitude
150 feet downgradient of the source area. Near-
saturation concentrations of PCE decrease by several
orders of magnitude over the same distance. Prelimi-
nary 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 sur-
face-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 effec-
tively remediate the chlorinated hydrocarbons, the pro-
posed remedial action for the site will be limited to
ground-water monitoring.
163
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Natural Attenuation of Trichloroethene in a Sandy Unconfined Aquifer
Neale Misquitta, Dale Foster, and Jeff Hale
Key Environmental, Inc., Carnegie, Pennsylvania
Primo Marches! and Jeff Blankenship
American Color and Chemical Corporation, Lockhaven, Pennsylvania
The natural attenuation of dissolved-phase trichlo-
roethene (TCE) in ground water was evaluated at a
state-regulated, operating chemical plant in South Caro-
lina. Natural attenuation was documented via the ob-
served attenuation and loss of TCE within a sandy
unconfined aquifer (approximately 25 feet thick with
Kh=10"3 centimeters per second), 8 years of ground-
water 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 down-
gradient ground-water quality compliance points, obvi-
ating the need for containment or other remedial actions.
The recently promulgated South Carolina Groundwater
Mixing Zone Regulations require that, under very spe-
cific 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, indicat-
ing that non-TOC related mechanisms were contribut-
ing to TCE attenuation. Consequently, a site-specific
Kd was estimated via batch adsorption tests employ-
ing toxicity characteristic leaching procedure extrac-
tion 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 site-
specific retardation factor correlated with the observed
natural attenuation of TCE. Differences in the site-
specific retardation factor and the empirically calcu-
lated 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 unre-
lated to TOC.
Subsequent ground-water modeling using the site-spe-
cific retardation factor (and Kd) indicates that dissolved-
phase 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 Caro-
lina, included a time-weighted "monitoring only" compo-
nent with no active remediation. This application is
currently under review.
164
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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) pro-
posed 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 concentra-
tions as high as 90,000 micrograms per liter were ob-
served at the site and utilized in the model calibration
effort. The calibrated model showed that intrinsic reme-
diation of TCE is occurring at the site. The estimated
first-order degradation rate for TCE ranged from 0.0020
to 0.0064 day1.
TCE mass and apparent mass degraded were also esti-
mated using the calibrated model. TCE mass predictions
using the model closely matched TCE mass calculated
from observed ground-water data. The TCE mass de-
graded 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 com-
pounds showed that the mass of these compounds
present was significantly less than estimated, suggest-
ing rapid transformation of the compounds to nonchlori-
nated compounds.
The calibrated model was also used to predict the ef-
fects of source removal on the lifetime of the dissolved
TCE plume. These predictions, along with source life-
time 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
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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 elec-
tron donor (acetic acid) that is stimulating reductive
dechlorination of tetrachloroethene (PCE), trichloroethene
(TCE), and 1,1,1-trichloroethane (TCA) to ethene and eth-
ane 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 mo-
bility in the site ground water. PCE, TCE, and TCA con-
centrations 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 distri-
butions of 1,2-dichloroethene, vinyl chloride (VC), 1,1-di-
chloroethane, 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 biore-
mediation is being considered as a remediation alternative
for this site.
166
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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 shal-
low aquifer is recovering nonaqueous-phase liquid
(NAPL) and dissolved-phase EDC. Degradation prod-
ucts of EDC, including 2-chloroethanol, ethanol, ethene,
and ethane, were detected in both the highly contami-
nated upper aquifer as well as in the deeper, less con-
taminated aquifer. Possibly as a result of cross
contamination during drilling operations, low concentra-
tions (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 elec-
tron acceptor.
Modeling efforts with DuPonts comprehensive multi-
phase 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 dissolved-
phase EDC. The three-dimensional, three-phase finite
difference model includes simultaneous flow of water,
gas, and organic phases; energy transport; tempera-
ture-, 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 de-
velopment 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 ap-
plications. Water Resour. Res. 29(6):1709-1718.
167
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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 micro-
grams per liter [u,g/L]) of chlorinated volatile organic
compounds (VOCs) in ground water. The VOCs de-
tected 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 u,g/L. Historically, the
site was mined for rock and subsequently used for the
disposal of tailing sands and clay waste from ore proc-
essing. 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. In-
itially, the assessment included VOC data over time,
collected to monitor the ground-water pump-and-treat sys-
tem. Long-term ground-water monitoring results indicate
that concentrations of parent VOC compounds have
been reduced in all water-bearing units; after an asso-
ciated 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 bioremedia-
tion, a study was implemented involving field monitoring
and ground-water sampling and analysis for select geo-
chemical indicator compounds and dissolved perma-
nent 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. Con-
centrations of organic compounds were evaluated over
time, and trends in inorganic indicator compound and
dissolved permanent gas concentrations were evalu-
ated 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
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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. Guide-
lines and criteria for natural remediation assessments
have been established at both the state and federal
level, providing the regulated community with an im-
proved ability to determine site conditions under which
a natural remediation approach is feasible and will be
acceptable to regulators. These initiatives reflect transi-
tion from conventional remedy selection to considera-
tion 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 remedia-
tion 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 activi-
ties 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 hydrocar-
bons. The NRCP proposal, submitted as part of a reme-
dial action workplan for site ground water and
concurrent with a remedial action report for source area
soils, addressed the following NJDEP prerequisite con-
ditions ("minimum requirements") for natural remedia-
tion proposals: delineation and remediation of sources;
contaminant migration assessment to confirm receptors
not at risk; documentation of degradibility and/or attenu-
ation capacity; identification of site-specific charac-
teristics 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 require-
ments included delineation of source area soils using a
Geoprobe, source area soil excavation/disposal,
postexcavation sampling, ground-watersampling, 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, re-
ceptor 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 com-
pared with costs for active ground-water plume manage-
ment. 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
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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), trichlo-
roethene (TCE), carbon tetrachloride, and hexachlo-
robenzene (HCB) is under investigation at a large
industrial site. The site has a number of complexities
due to past manufacturing activities, topography, hydrol-
ogy, 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 concentra-
tions of iron(ll) 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 respi-
ratory 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
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Assessment of Intrinsic Bioremediation of Chlorinated Aliphatic Hydrocarbons
at Industrial Facilities
Marleen A. Troy and C. Michael Swindell
DuPont Environmental Remediation Services, Wilmington, Delaware
Intrinsic bioremediation of chlorinated aliphatics at sev-
eral industrial sites was evaluated to determine its sig-
nificance 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 oc-
curring microorganisms present in subsurface environ-
ments of each site were capable of degrading the
contaminants of interest and that the contaminant con-
centrations would be degraded to acceptable levels.
A variety of chlorinated aliphatic hydrocarbons were
detected in ground water at the sites, including tetrachlo-
roethene (PCE), trichloroethene (TCE), dichloroethene
(DCE), vinyl chloride (VC), trichloroethane (TCA), and
methylene chloride. Concentrations of individual chlorin-
ated 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 contami-
nant concentrations, detection of biodegradation meta-
bolites, and changes in geochemical measurements.
Indicators of conditions favorable for bioremediation that
were evaluated included pH, oxidation-reduction poten-
tial, concentrations of electron acceptors, nutrients, pri-
mary 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 occur-
ring under anaerobic conditions at each site, nonchlori-
nated 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 pa-
rameters continued to allow further evaluation of the
potential application of intrinsic bioremediation as a re-
medial alternative for the sites.
171
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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.
Ateam of hydrogeologists and engineers, under contract
with the New York State Department of Environmental Con-
servation (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 com-
plex array of potential contaminant migration pathways.
Alighting fixture manufacturing operation discharged an
unknown volume of liquid waste containing TCE-domi-
nated 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 dura-
tion 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, includ-
ing 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 constitu-
ents 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 func-
tioning on site, project personnel designed and imple-
mented a focused Phase II remedial investigation that,
in part, addressed natural attenuation related issues.
The latter phase of the remedial investigation was im-
plemented 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 re-
cord of decision was signed in March 1996, and the
selected remedy allows for limited (near-surface hot-
spot 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
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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 re-
sulted in contamination of shallow soils and ground
water with a mixture of chlorinated aliphatic hydrocar-
bons (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 petro-
leum and chlorinated solvents, consists of commin-
gled benzene, toluene, ethylbenzene, and xylenes
(BTEX) and CAHs. Before construction of a horizontal
air sparging system, contaminated ground water dis-
charged 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, c/s-1,2-DCE, VC, and ethene
indicate that TCE dissolved from the LNAPL body is
being sequentially dehalogenated, with VC accumulat-
ing 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 bypro-
ducts, along with dissolved hydrogen concentrations,
indicate that biodegradation mechanisms operating at
the site include aerobic respiration, iron reduction, sul-
fate reduction, and methanogenesis.
Approximation of field-scale biodegradation rates at the
site suggests that TCE and c/s-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
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Applying Natural Attenuation of Chlorinated Organ ics
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 em-
ployed 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. Environ-
mental 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 nearthe 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 Con-
toocook River. Furthermore, modeling indicated that dis-
charge 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 con-
centrations, and results were compared with health-
based 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 re-
cord of decision was modified accordingly through issu-
ance of an explanation of significant difference.
174
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A Modular Computer Model for Simulating Natural Attenuation of
Chlorinated Organ ics 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 tech-
nology has yet to be formulated for chlorinated solvent
plumes. Successful validation of natural attenuation in-
volves significant upfront and followup field charac-
terization 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 inter-
pretation 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 fore-
casting 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 Dimen-
sions). This tool can simulate natural attenuation of
various subsurface contaminants and their decay prod-
ucts in saturated ground-water aquifers.
RT3D was developed from the U.S. Environmental Pro-
tection Agency's public domain computer code MT3D.
The MT3D model simulates single-species transport
with or without sorption and first-order reaction. Con-
taminant transport velocities are calculated from the
head distribution computed by the U.S. Geological Sur-
vey'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 solid-
phase species, along with the physical, chemical, and
biological interactions among them. The code is organ-
ized in a modular fashion to ensure flexibility. The reac-
tive portion of the code is a separate module using an
operator-split strategy; hence, any type of reaction kinet-
ics can be accommodated through an appropriate reac-
tion module. The present version has four separate
reaction modules: aerobic, instantaneous BTEX reac-
tions (similar to BIOPLUME II); multiple-electron ac-
ceptor, kinetic-limited BTEX reactions (similar to
BIOPLUME III); denitrification-based carbon tetrachlo-
ride 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 chlorin-
ated organics in saturated ground-water aquifers.
175
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State and Federal Regulatory Issues
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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 Sym-
posium 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 (Tl) waivers
Fran Kremer and John Wilson of EPA's Office of Re-
search and Development chaired the meeting; an over-
view 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 assur-
ance/quality control plans, and data quality objec-
tives 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. Geo-
logical Survey (USGS) stated that the 10 parameters
currently measured by the Survey as indicators of natu-
ral attenuation are dissolved oxygen, nitrate, nitrite, fer-
rous sulfate, sulfite, methane, chloride, carbon dioxide,
and hydrogen. The need for an overall protocol for as-
sessing the redox chemistry was recognized. It was
suggested that USGS and EPA collaborate in develop-
ing such a protocol.
The discussion emphasized the need for rapid and reli-
able field techniques. Participants debated the pros and
cons of field and laboratory analytical methods, and
discussed, for example, the need to consider the trade-
offs 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 contamina-
tion due to NAPLs was stressed. Participants urged
consideration of push-technologies such as cone pene-
trometer because they can provide quick and depend-
able estimates of parameters. Further work needs to be
undertaken to develop improved methods to charac-
terize 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 reduc-
ing contaminants to acceptable levels before potential
receptors are reached. Since improper monitoring may
lead to misconceptions about the extent and/orthe types
of biotic and abiotic processes occurring at a contami-
nated site, a thorough remediation strategy should in-
clude 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 concep-
tual 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 im-
plementation of other proactive remedial technologies,
such as pump and treat, the natural gradients of ground-
water flow are significantly altered. Since natural attenu-
ation is a passive remedy, however, understanding the
existing flow conditions is very important.
179
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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 natu-
ral attenuation as a remedial alternative and conducting
periodic sampling to assure that remediation is occur-
ring. 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 re-
mediation rates, pattern of migration, and seasonal in-
fluences 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 model-
ing 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 com-
pounds, in particular dissolved gases. They also dis-
cussed the importance and role of hydrogen levels in the
hierarchical scheme of terminal electron acceptor proc-
esses. 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 hy-
drogeology. 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 hydro-
carbons has been observed to follow first-order kinetics,
it is unclear whether the degradation of chlorinated hy-
drocarbons 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 esti-
mate 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 ex-
trapolating 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 as-
sessing environmental risks and determining action lev-
els. Better methodologies are needed to ensure that
institutional controls can be implemented effectively.
The current cleanup guidelines are based only on con-
centration and not on real site-specific health and eco-
logical 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 as-
sociated 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 as-
sess the risk of ground-water discharge into surface
water, but also to consider the contamination of ground
water by surface water infiltration. Although the pres-
ence of microorganisms as health hazards does not
require monitoring under Superfund regulations, the is-
sue is handled very differently in the proposed ground-
water 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 biode-
gradation 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 manage-
ment and assessment are more difficult for sites con-
taminated 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 at-
tenuation 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
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than other alternatives, while in other cases, active re-
mediation may be more cost effective.
Clear, simple information about the appropriate situ-
ations for selecting natural attenuation over other reme-
dial alternatives should be made available. Participants
suggested developing protocols to indicate the parame-
ters to be selected and the analytical methodologies to
be used. They also stressed the critical need for deci-
sion 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" rem-
edy. 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 para-
mount 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, pre-
sented in clear, simple terms, that addresses specifics
on toxicity of the daughter products and by-products of
the degradation process.
Technical Impracticability (Tl) Waiver
Participants were concerned that there is considerable
confusion regarding natural attenuation and Tl 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 con-
trols to ensure continued protection of public health and
the environment. A Tl 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 Tl
waiver is not needed in order to select natural attenu-
ation as the appropriate remedy because natural proc-
esses are expected to attain the required level of
cleanup in a reasonable time frame. At some sites it will
be appropriate to use a Tl 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 compo-
nent of all ground-water remedies, to ensure that reme-
diation 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 dis-
cussed above, geochemical indicators, degradation by-
products, and other parameters may need to be
measured in addition to the contaminants of concern to
determine if biotic and/or abiotic processes are continu-
ing as expected.
Summary
The meeting identified the following needs as being of
paramount importance:
• Standard operating procedures for the indicator pa-
rameters 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 degra-
dation rate constants.
• Development of guidance for the use of natural at-
tenuation.
• Training for federal and state decision-makers in
evaluating data supporting the use of natural attenu-
ation.
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Summary of Roundtable Discussion on Regulatory Issues
Panelists
• Michael Barden, Hydrogeologist, Wisconsin Depart-
ment of Natural Resources
• Mark Ferrey, Research Scientist, Minnesota Pollu-
tion Control Agency
• Ruth Izraeli, EPA Hydrogeologist, Co-Chair of Su-
perfund/RCRA Ground-water Forum
• Tim Larson, Professional Engineer, Florida Depart-
ment 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 regu-
latory agencies answered questions from the audience.
The following points were made during the discussion:
• Public acceptance of natural attenuation. Public ac-
ceptance of natural attenuation as a tool for risk man-
agement and site closure is as critical as regulatory
acceptance. It is important to stress that natural at-
tenuation 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 con-
vey to the public that natural attenuation is an engi-
neered application of a technology, not a "do-nothing"
approach. Another public misconception is that con-
taminated 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 attenu-
ation 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 com-
prehensive site characterization, integration of other
remediation tools, ground-water demand, and future
plans for aquifer use.
• Regulatory flexibility with a technical impracticability
(Tl) waiver. Tl 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. Tl
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 fre-
quently. 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 dis-
charge and if people are not and will not be using
the aquifer for drinking water as guaranteed by insti-
tutional controls, ACLs that are protective of the sur-
face 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 attenu-
ation 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 moni-
tored 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 pri-
orities for ground-water protection. For example, a
state might declare a certain area as a kind of nonat-
tainment 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 situ-
ations.
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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 bod-
ies that have extremely high pH levels. EPA has
ground-water and surface water classification sys-
tems, 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 draw-
backs 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 impor-
tant step in determining the benefits of natural at-
tenuation.
• 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 analy-
sis, 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 at-
tenuation 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 re-
source damage claims under CERCLA on ground-
water 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 strat-
egy. 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.
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Appendix A
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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:OOAM 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
9:1 SAM 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,
Super fund 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
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Wednesday, September 11 (cont'd)
10:45AM Microbiological Aspects Relevant to Natural Attenuation
of Chlorinated Ethenes James Gossett
Cornell University, Ithaca, NY
11 :OOAM Microbiological Aspects Relevant to Natural Attenuation
of Chlorinated Ethenes (cont'd) Steve Zinder
Cornell University, Ithaca, NY
11:1 SAM 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:OOPM 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:OOPM BREAK
3:30PM Evaluation of Natural Attenuation: Current Practices and
Opportunities for Ground Water Joseph Salvo
General Electric,
Corporate Research & Development, Schenectady, NY
4:OOPM 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
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Wednesday, September 11 (cont'd)
4:30PM The BIOSCREEN Computer Tool
5:OOPM ADJOURN
5:30PM to 7:30PM
Charles Newell
Groundwater Services, Inc., Houston, TX
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:OOAM Introductory Remarks Patrick Haas
AFCEE, USAF, Brooks AFB, TX
8:1 SAM Case Study: Naval Air Station, Cecil Field, Florida Francis Chapelle
USGS, Columbia, SC
9:OOAM 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
Pittsburgh Air Force Base, New York Todd Wiedemeier
Parsons Engineering Science, Inc., Denver, CO
10:30AM BREAK
11 :OOAM 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
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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
DuPontSpecialty Chemicals-CRG,Wilmington, DE;
Joe Salvo, General Electric, Schenectady, NY;
and Gary Klecka, Dow Chemical Company, Midland, Ml
2:30PM Case Study: Wurtsmith Air Force Base, Michigan Michael Barcelona
University of Michigan, Ann Arbor, Ml
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:OOPM ADJOURN
7:OOPM - 9:OOPM 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:OOAM 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
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Friday, September 13 (cont'd)
9:OOAM 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:OOAM 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, I/I//
11:10AM Roundtable Discussion Panel of Speakers
12:15PM ADJOURN
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