£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.

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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.

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    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.

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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.

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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

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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.
 4. Cline, P.V., and J.J. Delfino. 1989. Effect of subsurface sediment
   on hydrolysis of haloalkanes and epoxides. In: Larson, R.A., ed.
   Biohazards of drinking water treatment. Chelsea, 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-
   tions of halogenated aliphatic compounds. Environ. Sci. Technol.
   21:722-736.
 7. McCarty, PL., and J.T Wilson. 1992.  Natural anerobic treatment
   of a TCE plume, St. Joseph, Michigan, NPL site. In: U.S. EPA.
   Bioremediation of hazardous wastes.  EPA/600/R-92/126. Cincin-
   nati, OH. pp. 47-50.
 8. 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.
 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-
   biol. 58:1996-2000.
10. Holliger, C., G. Schraa, A.J.M. Stams, and A.J.B. Zehnder.  1993.
   A highly purified enrichment   culture couples the  reductive
   dechlorination of tetrachloroethene to growth. Appl.  Environ. Mi-
   crobiol. 59:2991-2997.
11. Neumann, A., H. Scholz-Muramatsu, and G.  Dickert.   1994.
   Tetrachloroethene metabolism  of Dehalospirillum  multivorans.
   Arch. Microbiol. 162:295-301.
                                                     10

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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-
    tion as respiratory process. Antonie Van  Leeuwenhoek 66:239-
    246.

14. Sharma,  P., and  P.L.  McCarty.  1996.  Isolation and  charac-
    terization of facultative aerobic bacterium that reductively deha-
    logenates tetrachlorethene  to   c/s-1,2-dichloroethene.  Appl.
    Environ. Microbiol. 62:761-765.

15. Maymo-Gatell, X., Y.T.  Chien, T. Anguish, J.  Gossett,  and  S.
    Zinder. 1996.  Isolation and characterization of  an  anaerobic
    eubacterium which reductively dechlorinates tetrachloroethene
    (PCE) to ethene. In: Abstracts of the 96th General Meeting of the
    American Society of Microbiology, New Orleans, pp. Q-126.

16. Fennel, D.E.,  M.A. Stover, S.H. Zinder, and J.M.  Gossett. 1995.
    Comparison of alternative electron  donors to sustain PCE  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.

17. Bradley, P.M., and F.H. Chapelle. 1996. Anaerobic mineralization
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18. Dolan, M.E., and P.L. McCarty. 1995.  Small-column microcosm
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19. Hopkins,  G.D., and P.L.  McCarty. 1995. Field  evaluation of in situ
    aerobic  cometabolism  of trichloroethylene  and  three  dichlo-
    roethylene isomers using phenol and toluene  as the primary sub-
    strates. Environ. Sci. Technol. 29:1628-1637.

20. Major, D.W., W.W.  Hodgins, and  B.J. Butler.  1991.  Field and
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    roethene  to ethene and ethane at a chemical transfer facility in
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    site bioreclamation. Stoneham, MA: Butterworth- Heinemann. pp.
    147-171.

21. Fiorenza, S.,  E.L. Hockman, Jr.,  S. Szojka,  R.M. Woeller, and
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22. McCarty, PL., and  M. Reinhard. 1993. Biological  and chemical
    transformations  of halogenated aliphatic compounds in aquatic
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23. Charnley, G., E.A.C.  Crouch, L.C. Green,  and T.L. Lash.  1988.
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24. McCarty, PL., L. Semprini, M.E. Dolan, T.C. Harmon,  C. Tiede-
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    diation for contaminated groundwater at St. Joseph, Michigan. In:
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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.
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    Research,  Development,  and Field  Evaluation. EPA/600/R-
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28. Major, D., E. Cox, E. Edwards,  and  P.  Hare. 1995. Intrinsic
    dechlorination of trichloroethene to ethene in a bedrock aquifer.
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    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.
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30. Cox, E., E.  Edwards,  L. Lehmicke, and D. Major. 1995. Intrinsic
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31. Buchanan, J.R.J., D.E. Ellis, J.M. Odom, PF. Mazierski, and M.D.
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                                                               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.
                                                     14

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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-
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    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
    reactions in  aquatic  sediments.  Geochim. Cosmochim.  Acta
    52:2993-3003.

16.  Glantz,  S.A.  1992.  Primer of  biostatistics.  New York,  NY:
    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

-------
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

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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

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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

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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

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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

-------
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

-------
       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

-------
                       ... 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

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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

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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

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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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                               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

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  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
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(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

-------
                     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

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   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

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     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

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             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

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       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

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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

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      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

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                        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

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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

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                                                                                        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

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                                 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

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    : 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.
                                                      99

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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,
                                                 112

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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
 
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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-
                                                  122

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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-
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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
                                                 126

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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.
                                                      128

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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
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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.
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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.
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•  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

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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
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 3.  Wilson, J.T, J.W. Weaver,  D.H. Kampbell. 1994.  Intrinsic biore-
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 4.  Ehlke, T.A.,  B.H. Wilson,  J.T. Wilson, and I.E. Imbrigiotta. 1994.
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 5.  Martin, M., and T.E. Imbrigiotta.  1994. Contamination of ground
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 6.  Cox, E., E.  Edwards, L. Lehmicke, and D. Major.  1995. Intrinsic
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    Battelle Press, pp. 223-231.

 7.  Lee, M.D., PR Mazierski, R.J. Buchanan, Jr.,  D.E. Ellis, and L.S.
    Sehayek. 1995. Intrinsic and in situ anaerobic biodegradation of
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 8.  Buscheck, T, and K. O'Reilly.  1996. Intrinsic anaerobic biodegra-
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 9.  Benker, E., G.B. Davis, S. Appleyard, D.A. Berry, and T.R. Power.
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10.  Gorder,  K.A., R.R.  Dupont,  D.L.  Sorensen, and  M.W  Kem-
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11.  De, A., and D. Graves. 1996. Intrinsic bioremediation of chlorin-
    ated aliphatics and aromatics at a complex industrial site. Ab-
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    Battelle Memorial Institute.

12.  Ehlke, T.A., T.E. Imbrigiotta, B.H. Wilson, and J.T. Wilson.  1991.
    Biotransformation of cis-1,2-dichloroethylene in aquifer material
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13.  Wilson, B.H., T.A.  Ehlke, T.E. Imbigiotta, and J.T. Wilson.  1991.
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    sources Investigations Report 91-4034. pp. 704-707.

14.  Hasten, Z.C., P.K. Sharma, J.N.P Black, and PL. McCarty.  1994.
    Enhanced reductive dechlorination  of chlorinated  ethenes. In:
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15.  Wilson, B.H., J.T. Wilson, D.H. Kampbell, B.E. Bledsoe, and J.M.
    Armstrong.  1990. Biotransformation of monoaromatic and chlo-
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16.  Parsons,  R, G. Barrio Lage, and R. Rice.  1985. Biotransformation
    of  chlorinated organic solvents in static microcosms. Environ.
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17.  Davis, J.W, and C.L. Carpenter.  1990.  Aerobic biodegradation
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    56(12):3878-3880.

18.  Klecka, G.M.,  S.J.  Gonsior, and D.A.  Markham. 1990. Biological
    transformations of 1,1,1-trichloroethane in  subsurface  soils and
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19.  Barrio-Lage, G.A.,  RZ. Parsons, R.M.  Narbaitz, and PA. Lorenzo.
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                                                              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

-------
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

-------
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.

References

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 3. Ramanand, K., M.T. Balba,  and J. Duffy. 1993. Reductive deha-
   logenation of chlorinated benzenes and toluenes under methano-
   genic conditions. Appl. Environ. Microbiol. 59:3266-3272.
 4. Reineke, W, and H.-J. Knackmuss. 1984. Microbial metabolism
   of haloaromatics: Isolation and properties of a chlorobenzene-de-
   grading bacterium.  Eur. J. Appl. Microbiol. Biotechnol. 47:395-
   402.
 5. Schraa, G., M.L. Boone, M.S.M. Jetten, A.R.W. van Neerven, P.J.
   Colberg, and A.J.B. Zehnder. 1986. Degradation of 1,4-dichlo-
   robenzene by Alcaligenessp.  strain A175. Appl. Environ. Micro-
   biol. 52:1374-1381.
 6. Spain, J.C., and S.F. Nishino. 1987. Degradation of 1,4-dichlo-
   robenzene by  a Pseudomonas sp. Appl. Environ. Microbiol.
   53:1010-1019.
 7. de Bont, J.A.M., M.J.A.W. Vorage, S. Hartmans, and W.J.J. van
   den Tweel. 1986. Microbial  degradation of 1,3-dichlorobenzene.
   Appl. Environ. Microbiol. 52:677-680.
                                                    140

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 8.  Haigler, B.E., S.F. Nishino, and J.C. Spain. 1988. Degradation of
    1,2-dichlorobenzene by a Pseudomonas sp. Appl. Environ.  Mi-
    crobiol. 54:294-301.

 9.  van der Meer, J.R., W. Roelofsen, G. Schraa, and A.J.B. Zehnder.
    1987. Degradation of low concentrations of dichlorobenzenes
    and 1,2,4-trichlorobenzene by Pseudomonas sp. strain P51 in
    nonsterile soil columns. FEMS Microbiol. Lett. 45:333-341.

10.  Sander,  P., R.-M.  Wittaich,  P.  Fortnagel,  H. Wilkes,  and  W.
    Francke.  1991.  Degradation of  1,2,4-trichloro- and  1,2,4,5-
    tetrachlorobenzene by Pseudomonas strains. Appl. Environ.  Mi-
    crobiol. 57:1430-1440.

11.  Nishino,  S.F., J.C. Spain, and C.A. Pettigrew. 1994. Biodegrada-
    tion of chlorobenzene  by indigenous  bacteria. Environ. Toxicol.
    Chem. 13:871-877.

12.  Haggblom,  M.M., and  R.J. Valo. 1995.  Bioremediation of chlo-
    rophenol wastes. In: Young,  L.Y., and C.E. Cerniglia, eds.  Micro-
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13.  Suflita,  J.M., and G.T. Townsend.  1995. The microbial ecology
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14.  Bedard,  D.L.,  and  J.F. Quensen.  1995.  Microbial  reductive
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15.  Bedard,  D.L., S.C. Bunnell, and  L.A. Smullen. 1996. Stimulation
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    ron. Sci. Technol. 30:687-694.

16.  Bedard,  D.L., R.  Unterman, L. Bopp, M.J. Brennan, M.L. Haberl,
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17.  Harkness, M.R.,  J.B. McDermott, D.A. Abramowicz,  J.J.  Salvo,
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    WL. Gately. 1993. In situ stimulation  of aerobic PCB biodegra-
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19. Adriaens, P., and T.M. Vogel. 1995. Biological treatment of chlo-
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20. Fetzner, S., and F. Lingens. 1994. Bacterial dehalogenases: Bio-
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21. Wackett, L.P 1995.  Bacterial co-metabolism of halogenated or-
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    crobial  transformation   and   degradation  of   toxic organic
    chemicals. New York, NY: Wiley-Liss.  pp. 217-242.

22. Magli, A., F.A. Rainey, and T Leisinger. 1995. Acetogenesis from
    dichloromethane by a two-component mixed culture  comprising
    a novel bacterium. Appl. Environ. Microbiol. 61:2943-2949.

23. Stucki, G., U. Krebser,  and T. Leisinger. 1983.  Bacterial growth
    on 1,2-dichloroethane. Experientia 39:1271-1273.

24. Salanitro, J.P, L.A. Diaz, M.P Williams, and  H.L. Wisniewski. 1994.
    Isolation of a bacterial culture that degrades methyl t-butyl ether.
    Appl. Environ. Microbiol. 60:2593-2596.

25. Spain, J.C.  1995. Biodegradation  of nitroaromatic compounds.
    Ann. Rev. Microbiol. 49:523-55.

26. Nishino, S.F., and J.C.  Spain. 1996. Degradation of 2,6-dinitro-
    toluene by bacteria.  In: Proceedings of the Annual Meeting, Ame
    rican Society for Microbiology, pp. Q-381.

27. Janssen, PH.,  and  B.  Schink. 1995. Catabolic and anabolic
    enzyme activities and energetics of acetone  metabolism of the
    sulfate-reducing bacterium Desulfococcus biacutus. J. Bac-
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28. Mormile, M.R., S. Liu, and J.M. Suflita. 1994. Anaerobic biodegra-
    dation  of gasoline oxygenates:  Extrapolation of  information to
    multiple sites  and  redox  conditions.  Environ.  Sci.  Technol.
    28:1727-1732.

29. White, G.F.,  and J.R. Snape. 1993.  Microbial cleavage of nitrate
    esters: Defusing the environment. J. Gen. Microbiol.  139:1947-
    1957.

30. White, G.F.,  J.R. Snape, and S. Nicklin. 1996. Biodegradation of
    glycerol trinitrate and  pentaerythritol  tetranitrate  by Agrobac-
    terium radiobacter.  Appl. Environ. Microbiol. 62:637-642.
                                                              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.
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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.
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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.
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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
                                                  180

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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.
                                                  183

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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.
                                                   184

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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
                                 187

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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
                                             188

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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
                                           189

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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
                                                 190

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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
                                           191

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