EPA/540/R-96/509
 xvEPA
                 3d States
               Environmental Protection
               Agency
              Office of Research and
              Development
              Washington, DC 20460
EPA/540/R-96/509
September 1996
Symposium on
Natural Attenuation of
Chlorinated Organics in
Ground Water
               Hyatt Regency Dallas
               Dallas, TX
               September 11-13,1996
                                           Printed on paper that contains at
                                           least 20 percent postconsumer fiber.

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                                                                 EPA/540/R-96/509
                                                                  September 1996
         Symposium on Natural Attenuation of Chlorinated Organics
                                 in  Ground Water
                                  Hyatt Regency Dallas
                                      Dallas, TX
                                 September 11-13, 1996
14

CJ
•-^ j
                            EPA REGION 6 LIBRARY 6MD-II

                            1445ROSSAVE, STE. 1200

                            DALLAS, TX 75202-2733
                             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

Introductory Talk: Where Are We Now? Moving to a Risk-Based Approach
       C.H. Ward	  1

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	  4

Biotic and Abiotic Transformations of Chlorinated Solvents in Ground Water
       Perry L. McCarty	  5

Microbiological Aspects Relevant to Natural Attenuation of Chlorinated Ethenes
       James  M. Gossett and Stephen H. Zinder	  10

Microbial Ecology of Adaptation and Response in the Subsurface
       Guy W. Sewell and Susan A.  Gibson	  14

Identifying Redox Conditions That Favor the Natural Attenuation of Chlorinated Ethenes in
Contaminated Ground-Water Systems
       Francis H. Chapelle	  17

Design and Interpretation of Microcosm Studies for Chlorinated Compounds
       Barbara H. Wilson, John T. Wilson, and Darryl Luce	  21

Conceptual Models for Chlorinated Solvent Plumes and Their Relevance to Intrinsic Remediation
       John A. Cherry	  29

Site Characterization Tools: Using a Borehole Flowmeter To Locate and Characterize the
Transmissive Zones of an Aquifer
       Fred  Molz and Gerald Boman	  31

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	  35

The BIOSCREEN Computer  Tool
       Charles J. Newell, R. Kevin McLeod, and James R.  Gonzales	  60

Case Study: Naval Air Station Cecil Field, Florida
       Francis H. Chapelle and Paul M. Bradley	  64
                                                 II!

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                                       Contents (continued)

                                                                                               Page

Case Study of Natural Attenuation of Trichloroethene at St. Joseph,  Michigan
       James W. Weaver, John T. Wilson, and Donald H. Kampbell	  65

Extraction of Degradation Rate Constants From the St. Joseph, Michigan, Trichloroethene Site
       James W. Weaver, John T. Wilson, and Donald H. Kampbell	  69

Natural Attenuation of Chlorinated Aliphatic Hydrocarbons at Pittsburgh Air Force Base, New York
       Todd H. Wiedemeier, John T. Wilson, and Donald H. Kampbell	  74

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	  83

Case Study: Plant 44, Tucson, Arizona
       Hanadi S. Rifai, Philip B. Bedient, and Kristine S. Burgess	  90

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	  93

Case Study: Wurtsmith Air Force Base, Michigan
       Michael J. Barcelona	  98

Case Study: Eielson Air Force Base, Alaska
       R. Ryan  Dupont, K. Gorder, D.L. Sorensen, M.W. Kemblowski,  and Patrick Haas	  104

Considerations and Options for Regulatory Acceptance of Natural Attenuation in Ground Water
       Mary Jane Nearman	  110

Lessons Learned: Risk-Based Corrective Action
       Matthew C. Small	  114

Informal Dialog on Issues of Ground-Water and Core Sampling
       Donald H. Kampbell	  116

Introductory Remarks: Appropriate Opportunities for Application—Civilian Sector (RCRA and CERCLA)
       Fran Kremer	  118

Introductory Remarks: Appropriate Opportunities for Application—U.S. Air Force and
Department of Defense
       Patrick Haas	  119

Intrinsic Remediation in the Industrial Marketplace
       David E. Ellis	  120
                                                 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	  124

Future Vision: Compounds With Potential for Natural Attenuation
       Jim Spain	  128

Natural Attenuation of Chlorinated Compounds in Matrices Other Than Ground Water:
The Future of Natural Attenuation
       Robert E. Hinchee and Donald H. Kampbell	  133

Poster Session

Degradation of Chloroform Under Anaerobic Soil Conditions
       Frances Y. Saunders and Van Maltby	  137

Anaerobic Mineralization of Vinyl Chloride in Iron(lll)-Reducing Aquifer Sediments
       Paul M.  Bradley and Francis  H. Chapelle	  138

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	  139

Analysis of Methane and Ethylene Dissolved in Ground Water
       Steve Vandegrift, Bryan Newell, Jeff Hickerson, and Donald H. Kampbell	  140

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	  141

Measurement of Dissolved Hydrogen in Ground Water
       Mark Blankenship, Francis H. Chapelle, and Donald H. Kampbell	  143

Evidence of Natural Attenuation of Chlorinated Organics at Ft. McCoy, Wisconsin
       Jason Martin	  144

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	  145

Development of an Intrinsic Bioremediation Program for Chlorinated Solvents at an Electronics Facility
       Michael  J. K.  Nelson, Anne G. Udaloy, and Frank Deaver	  146

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	  147

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                                        Contents (continued)

                                                                                                 Page

Incorporation of Biodegradability Concerns Into a Site Evaluation Protocol for Intrinsic Remediation
       Robert M. Cowan, Keun-Chan Oh, Byungtae Kim, and Gauri Ranganathan	  148

Intrinsic Remediation of Chlorinated Solvents as an Effective Containment Strategy
       Ronald Hughen,  Randall Hicks, and  Leon Grain	  149

A Field Evaluation of Natural Attenuation of Chlorinated Ethenes in a Fractured Bedrock Environment
       Peter Kunkel, Chris  Vaughan, and Chris Wallen	  150

Intrinsic Bioattenuation of Chlorinated Solvents in a Fractured Bedrock System
       William R. Mahaffey and K. Lyle Dokken	  151

Modeling Natural Attenuation of Selected Explosive Chemicals at a Department of Defense Site
       Mansour Zakikhani and Chris J. McGrath	  152

Long-Term Application of Natural Attenuation at Sierra Army Depot
       Jerry T. Wickham and Harry R. Kleiser	  153

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	  154

Natural Attenuation as a Cleanup Alternative for Tetrachloroethylene-Affected Ground Water
       Steve Nelson	  155

Natural Attenuation of Trichloroethene in a Sandy Unconfined Aquifer
       Neale Misquitta,  Dale Foster, Jeff Hale, Prime Marchesi, and Jeff Blankenship	  156

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	  157

Involvement of Dichloromethane in the Intrinsic Biodegradation of Chlorinated Ethenes and Ethanes
       Leo  L. Lehmicke, Evan E. Cox, and  David W. Major	  158

Intrinsic Bioremediation of 1,2-Dichloroethane
       Michael D. Lee, Lily S. Sehayek, and Terry D. Vandell	  159

A Practical Evaluation of Intrinsic Biodegradation of Chlorinated Volatile  Organic Compounds
       Frederick W. Blickle, Patrick N. McGuire, Gerald Leone, and Douglas D. Macauley	  160

Using Evidence of Natural Attenuation To Locate the Source of a Chlorinated Volatile Organic
Compound Plume
       John M. Armstrong, John J.  D'Addona, Charles W. Dittmar II, Greg  M. Tatara, and
       Joel W. Parker	  161
                                                  VI

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                                       Contents (continued)

                                                                                                Page

New Jersey's Natural Remediation Compliance Program: Practical Experience at a Site Containing
Chlorinated Solvents and Aromatic Hydrocarbons
       James Peterson and Martha Mackie	  162

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	  163

Assessment of Intrinsic Bioremediation of Chlorinated Aliphatic Hydrocarbons at Industrial Facilities
       Marleen A. Troy and C. Michael Swindell	  164

Natural Attenuation as Remedial Action: A Case Study
       Andrea Putscher and Betty Martinovich	  165

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	  166

Applying Natural Attenuation of Chlorinated Organics in Conjunction With Ground-Water Extraction
for Aquifer Restoration
       W. Lance Turley and Andrew Rawnsley	  167

Natural Attenuation of Chlorinated Organics in Ground Water Based on Studies Conducted at
Naval Amphibious Base Little Creek Sites  12 and 13
       Scott Park and Nitin Apte	  168

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	  170
                                                 VII

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                                 Acknowledgments
The papers abstracted in this book were presented at the Symposium on Natural Attenuation of
Chlorinated Organics in Ground Water, held September 11-13, 1996, in Dallas, Texas. The sympo-
sium was a joint effort of the U.S. Environmental Protection Agency's (EPA's) 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 EPA's Office
of Research and Development, Cathy Vogel of USAF AL/EQ, and Marty Faile and Patrick Haas of
USAF AFCEE served as co-organizers of the symposium.
                                         VIII

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     Introductory Talk: 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 not work
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 Pittsburgh 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- Washington, 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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 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
                U.S. Environmental Protection Agency, Washington, DC
(Paper unavailable at press time.)

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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):
"DCE
                                  + + CP

                                   (elimination) (Eq. 1)
CH3CC13
  TCA
                        CHjCOOH + 3H+ + 3d"
                          HAc      (hydrolysis)  (Eq. 2)


The rate of each chemical transformation is given by the
first-order reaction:
     C = C0e'
                                           (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:
                                           (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
EkJ
10°C
15°C
20°C
References
3.47 (10)20
6.31 (10)20
1.56(10)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 TCA's MCL of 200 (ig/L. Whenever TCA
is present as a contaminant, 1,1-DCE  can also be ex-
pected. In general, TCA is probably the main source of
1,1-DCE contamination found in aquifers.

CA, formed through  biological transformation of TCA,
can also be chemically transformed with  a half-life on
the order of months by hydrolysis to ethanol, which can
then be biologically converted to acetic acid and harm-
less products (6).

Biological Transformation

CAHs  can be oxidized  or reduced, generally  through
co-metabolism, as noted in Table 2.  In ground water,
reductive transformations are most often noted, perhaps
because the presence of intermediate products that are
formed provide strong evidence that reductive transfor-
mations are taking place. Co-metabolic aerobic transfor-
mation of TCE is also possible, although  if it did occur
the intermediate products formed are unstable and more
difficult, analytically, to  measure.  Thus, convincing evi-
dence for  the latter is  more difficult  to  obtain.  Also,
                         aerobic co-metabolism of TCE would only occur if suffi-
                         cient dissolved oxygen and a suitable  electron donor,
                         such as methane, ammonia, or phenol, were present.
                         Since circumstances  under which the proper environ-
                         mental conditions for significant aerobic co-metabolism
                         are not likely to occur often, natural attenuation by aerobic
                         co-metabolism of TCE is probably of little significance.

                         Ample  evidence suggests that anaerobic reductive
                         transformation of CAHs occurs frequently, however, and
                         this process is of importance to the transformation of all
                         chlorinated solvents and  their transformation products.
                         The major environmental  requirement is the presence of
                         sufficient concentrations of other organics that can serve
                         as electron donors for energy metabolism, which is often
                         the case in aquifers. Indeed, the extent to which reduc-
                         tive  dehalogenation  occurs may  be  limited  by the
                         amount of these co-contaminants present. Theoretically,
                         it  would require only a 0.4-gram chemical oxygen de-
                         mand (COD) equivalent of primary substrate to convert
                         1  gram  of PCE to ethene  (7), but many times more than
                         this is actually required because of competition by other
                         microorganisms for the electron donors present.

                         Figure 1 illustrates the potential chemical and biological
                         transformation pathways  for the four major chlorinated
                         solvents under anaerobic environmental conditions (6).
                         Freedman and Gossett (8) provided the first evidence
                         for conversion of PCE and TCE to ethene, and de Bruin
                         et al. (9) reported complete reduction to ethane. Table 3
Table 2.  Conditions for Biotic and Abiotic Transformations of Chlorinated Solvents

Biotic — Aerobic
Primary substrate
Co-metabolism
Biotic— Anaerobic
Primary substrate3
Co-metabolism
Hazardous intermediates
Abiotic
Carbon Tetrachloride
(CTC)

No
No

Perhaps
Yes
Yes
Perhaps
Trichloroethene
(PCE)

No
No

Yes
Yes
Yes
No
Tetrachloroethene
(TCE)

No
Yes

Yes
Yes
Yes
No
1,1,1 -Trichloroethane
(TCA)

No
Perhaps

Perhaps
Yes
Yes
Yes
  Can be used as electron acceptor in energy metabolism.

-------
Figure 1.  Anaerobic  chemical  and  biological transformation
         pathways for chlorinated solvents.
indicates that while some transformations, such as that
of CT to CF and carbon dioxide, may take place under
mildly reducing conditions such  as those associated
with denitrification, complete reductive transformation to
inorganic end products and of PCE and TCE to ethene
generally requires conditions suitable for methane fer-
mentation.  Extensive  reduction can also  occur under
sulfate-reducing conditions. For methane fermentation
to occur in an aquifer, the presence of sufficient organic
co-contaminant is required to reduce all of the oxygen,
nitrate, nitrite, and sulfate present. Some organics will
be required to reduce the CAHs, and perhaps iron(ll) as
well, if present in significant amounts.  If the potential for
natural biological attenuation of CAHs  is to be  evalu-
ated, then the concentrations of nitrate, nitrite, sulfate,
iron(ll), and methane, as well as organics as indicated
by COD or total organic carbon (TOC), should be deter-
mined. Unfortunately, such analyses are not considered
essential in remedial investigations—they should be.

Several pure cultures of microorganisms are now avail-
able that can also reduce PCE to cis-DCE (10-14). Only
one has been reported that can  convert PCE completely
to ethene (15). Most of the isolates are strict anaerobes
and use hydrogen  as an electron donor,  with  CAHs
being used  as electron acceptors in energy metabolism.
One isolate, however, is a facultative aerobe  (14) that
can use many organics, such as acetate, as the electron
donor and  oxygen,  nitrate, PCE,  or  TCE as  electron
acceptors, which it does in that  order of preference. It is
                          now believed that the CAH  reducers compete for the
                          hydrogen they use, which is formed as an intermediate
                          in anaerobic organic oxidation, with  sulfate reducers,
                          methanogens, and holoacetogens (16). This may explain
                          the excessive donor requirements for CAH reduction.

                          Concerns are frequently expressed over the  VC formed
                          as an intermediate in reductive dehalogenation of PCE,
                          TCE, and DCE in ground water, because VC is a known
                          human carcinogen. It is possible to oxidize  VC aerobi-
                          cally, however, with oxygen as an electron acceptor or
                          even under anoxic conditions with iron(lll) (17).  In addi-
                          tion, VC is readily and  very efficiently co-metabolized
                          aerobically by methane, phenol, or toluene oxidizers
                          (18, 19). Here, transformation yields of over 1 gram of
                          VC per gram of methane have been obtained. Thus, at
                          the aerobic  fringes of  plumes with methane and VC
                          present, or where sufficient iron(lll) is  present,  natural
                          attenuation of VC through oxidation can occur.

                          Case Studies

                          Major et al. (20)  reported field evidence for intrinsic
                          bioremediation of PCE to ethene and ethane  at a chemi-
                          cal transfer facility in North Toronto. In addition to high
                          concentrations of PCE  (4.4 milligrams  per liter [mg/L]),
                          high concentrations of methanol (810 mg/L) and  acetate
                          (430  mg/L)  were found as co-contaminants  in the
                          ground  water and served as  electron donors  for the
                          transforming organisms. Where high concentrations of
                          PCE were found, TCE  (1.7 mg/L), cis-DCE  (5.8 mg/L),
                          and VC (0.22 mg/L) were also found,  but little ethene
                          (0.01 mg/L). At one downgradient well, however, no PCE
                          or TCE were found, but cis-DCE (76 mg/L), VC  (9.7
                          mg/L) and ethene (0.42 mg/L) were present,  suggesting
                          that significant dehalogenation had occurred. Micro-
                          cosm studies also suggested that biotransformation was
                          occurring at the site, with complete disappearance of
                          PCE, TCE, and cis-DCE and production of both VC and
                          ethene. The conversions were accompanied by significant
                          methane production, indicating the presence of suitable
                          redox conditions for the transformation.
                          Fiorenza et  al. (21) reported on PCE,  TCE, TCA,  and
                          dichloromethane  (DCM) contamination of ground water
                          at a carpet backing manufacturing plant in Hawkesbury,
                          Ontario. The ground water contained 492 mg/L of volatile
Table 3.  Environmental Conditions for Reductive Transformations of Chlorinated Solvents

                                                         Redox Environment

Chlorinated Solvent
All
Denitrification
Sulfate Reduction
Methanogenesis
Carbon tetrachlonde CT -> CF
1,1,1-Trichloroethane TCA-> 1,1 -DCE
+ CH3COOH
Tetrachloroethene
Trichloroethene
CT -> CO2+Cr
TCA-»1,1-DCA
PCE->1,2-DCE
TCE->1,2-DCE

TCA -» CO2+Cr
PCE -» ethene
TCE -» ethene

-------
fatty acids and 4.2 mg/L of methanol,  organics that
appeared to serve as  electron donors for dehalogena-
tion. Sulfate was nondetected, but the concentration in
native  ground water was  about 15 to 18 mg/L. Total
dissolved iron was quite high (19.5 mg/L) and above the
upgradient concentration of 2.1 mg/L. Methane was pre-
sent. This supports conditions suitable for natural biode-
gradation  of the chlorinated solvents.   While some
chemical  transformation of TCA to 1,1-DCE  was  indi-
cated (0.4 mg/L) biotransformation was  extensive, as
indicated by a 1,1 -DCA concentration of 7.2 mg/L, com-
pared with the TCA concentration of 5.5 mg/L.  Some CA
was also present (0.19 mg/L). Transformation was also
indicated for PCE and TCE because the cis-DCE, VC,
and ethene concentrations  were  56,  4.2, and 0.076
mg/L, respectively. Only traces of ethane were found.
Downgradient from the lagoon, the dominant products
were cis-DCE (4.5 mg/L), VC (5.2  mg/L), and 1,1-DCA
(2.1 mg/L). While good evidence for natural attenuation
exists for this site, the ethene and ethane concentrations
were low compared with the VC concentration, suggest-
ing that biotransformation was not  eliminating the chlo-
rinated  solvent  hazard  at the  site,  although  it  was
producing compounds that may be more susceptible to
aerobic co-metabolism.

Evidence for intrinsic  biotransformation of chlorinated
solvents has also been provided from analyses of gas
from municipal  refuse landfills where  active methane
fermentation exists.  A summary by McCarty and Rein-
hard (22) of data from  Charnley  et al.  (23)  reported
average gaseous concentrations in parts per  million by
volume from eight refuse landfills as PCE, 7.15; TCE,
5.09; cis-DCE, not measured; trans-DCE, 0.02; and VC,
5.6. While these averages indicate that, in  general,
transformation was not complete, the high VC concen-
tration indicates the  transformation was significant. For
TCA, gaseous  concentrations were  TCA, 0.17;  1,1-
DCE, 0.10; 1,1-DCA,  2.5;  and CA,  0.37. These  data
indicate that TCA biotransformation was quite extensive,
with the transformation intermediate, 1,1-DCA,  present
at  quite significant  levels, as is  frequently  found  in
ground water.
Perhaps the most extensively studied and reported in-
trinsic chlorinated solvent biodegradation  is that at the
St.  Joseph,  Michigan,  Superfund  site  (7,  24-27).
Ground-water concentrations  of TCE as high as 100
mg/L were found, with extensive transformation to cis-
DCE, VC, and ethene. A high  but undefined COD (400
mg/L)  in  ground water,  resulting from waste leaching
from a disposal  lagoon, provided the energy source for
the  co-metabolic  reduction  of TCE.  Nearly  complete
conversion of the COD to methane provided evidence
of the  ideal conditions for intrinsic bioremediation (7).
Extensive analysis  near the source of contamination
indicated  that 8 to  25 percent of the TCE had been
converted to ethene and that up to 15 percent of the
reduction in COD in this zone was associated with re-
ductive dehalogenation (25).  Through  more extensive
analysis of ground water further downgradient from the
contaminating source, Wilson et al. (26) found a 24-fold
reduction in CAHs across the site. The great extent of
aerobic co-metabolic VC transformation in the methane
present suggests that aerobic oxidation at the  plume
fringes is likely to be occurring (18). A review of the data
at individual sampling points  indicated that conversion
of TCE to ethene was most complete  where methane
production was highest and removal of nitrate and sul-
fate by reduction was  most complete.

Since the above  early reports, many others have re-
ported on the natural biological attenuation of CAHs in
ground water, all showing conversion of PCE, TCE,  or
TCA to  nonchlorinated  end  points (28-31). Whether
complete dehalogenation is likely to occur over time at
these sites is still not clear. Review of this literature by the
reader interested in these processes is recommended.

References

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-------
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14. Sharma,  P., and RL  McCarty. 1996.  Isolation  and charac-
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15. Maymo-Gatell, X., Y.T.  Chien, T. Anguish, J.  Gossett, and  S.
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18. Dolan, M.E., and  P.L. McCarty. 1995. Small-column microcosm
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20. Major,  D.W., W.W.  Hodgins, and  B.J. Butler.  1991. Field and
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24. McCarty, PL., L. Semprini, M.E. Dolan, T.C. Harmon, C. Tiede-
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25. Semprini, L., P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson. 1995.
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    In: Hinchee, R.E., J.T. Wilson,  and D.C.  Downey, eds. Intrinsic
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29. Lee, M.D.,  P.P.  Mazierski,  R.J. Buchanan, D.E. Ellis, and L.S.
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30. Cox, E., E.  Edwards,  L. Lehmicke, and D. Major. 1995. Intrinsic
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    Battelle Press, pp. 223-231.

31. Buchanan, J.R.J., D.E. Ellis, J.M. Odom, P.F. Mazierski, and M.D.
    Lee. 1995. Intrinsic and accelerated anaerobic biodegradation of
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    Battelle Press, pp. 245-252.
                      0-t

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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) HC1     2[H] HC1

 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.
                                                  10

-------
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 H2 10 times greater than that of the methano-
gens in the culture (8). We do not know whether this 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
mu/tivorans
Strain TT4B
Enterobacter
agglomerans
Desulfitobacterium
sp. strain PCE1
Dehalococcus
ethenogenes
strain 195
a Maymo-Gatell X
tetrachloroethene
Dechlorination
Reactions
PCE, TCE -» c/s-DCE

PCE, TCE -H> c/s-DCE
PCE, TCE -> c/s-DCE
PCE, TCE -> c/s-DCE
PCE, TCE -> (c/s-DCE)
o-chlorophenols
PCE, others -> ethene
Y-T Chien, J.M. Gossett,
to ethene. Unpublished dat
Electron
Donors
H2

H2, formate,
pyruvate, etc.
Acetate
Nonfermentable
substrates
Lactate,
pyruvate,
butyrate,
ethanol, etc.
H2
and S.H. Zinder.
a.
Other Electron
Acceptors
None

Thiosulfate, nitrate,
fumarate, etc.
None
O2, nitrate, etc.
Sulfite, thiosulfate,
fumarate
None
1996. Isolation of a

Morphology
Rod

Spirillum
Rod
Rod
Curved rod
Irregular
coccus
novel bacterium

Phylogenetic
Position
Gram +
Desulfotomaculum
group
Epsilon
proteobacteria
?
Gamma
proteobacteria
Gram +
Desulfotomaculum
group
Novel eubacterium
capable of reductively

References
9-11

12
13
14
15
16"
dechlorinating

                                                   11

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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.
                                                     12

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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. Micro-bid. 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 ortho-chlorinated phenols.
    Arch. Microbiol. 165:132-140.

16.  Maymo-Gatell, X., V. Tandoi, J.M. Gossett, and S.H. Zinder. 1995.
    Characterization of an  Hg-utilizing enrichment culture that reduc-
    tively dechlorinates tetrachloroethene to vinyl chloride and ethene
    in the absence of methanogenesis and acetogenesis. Appl. En-
    viron. Microbiol. 61:3928-3933.

17.  DiStefano, T.D., J.M. Gossett, and S.H. Zinder.  1991. Reductive
    dechlorination  of  high concentrations of tetrachloroethene to
    ethene by an anaerobic enrichment  culture in the absence of
    methanogenesis. Appl. Environ. Microbiol. 57:2287-2292.

18.  Fennell, D.E., M.A. Stover, S.H. Zinder, and  J.M. Gossett. 1995.
    Comparison of alternative  electron donors to  sustain PCE an-
    aerobic reductive  dechlorination. In  Hinchee,  R.E., A. Leeson,
    and L.  Semprini,  eds. Bioremediation of chlorinated solvents.
    Columbus, OH: Battelle Press, pp. 9-16.
                                                               13

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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.
                                                  14

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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 product(s) 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 CI" 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
                                                   15

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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 (TEAP).  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 TEAP, 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) under the same conditions may undergo
further transformation, however,  if appropriate electron
acceptors such as O2 or Fe3+ are present.  Under these
conditions,  the oxidation  of vinyl chloride represents a
resource (energy) to the subsurface microbial populations.

Mechanisms of Adaptation
While an understanding of the ecological processes is
useful in predicting whether a transformation is likely to
occur, an understanding of the adaptation processes'
mechanisms is needed to predict the onset of the 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. Criddle, and P.L. McCarty. 1987. Transformations of
halogenated aliphatic compounds. Environ. Sci. Technol. 22:722-736.
                                                   16

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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 + SCI -^
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 microbia!
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
                                                 17

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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 for sulfate reduction the characteristic
range is between 1 and 4 nM. Methanogenesis, the least
energetically favorable anoxic process, is characterized
by H2 in the 5 to 15 nM range.

Patterns of electron-acceptor consumption, final  product
accumulation, and H2 concentrations can be combined to
logically identify redox processes (13). For example, if
sulfate concentrations are observed to decrease along an
aquifer flowpath, if sulfide concentrations are observed to
increase,  and if H2 concentrations are in the 1 to 4 nM
range characteristic of sulfate reduction, it may be 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). A standard gas-sampling
bulb is attached to a stream of water produced from a
well  and  purged for several minutes (at approximately
500 milliliters/minute) to eliminate all gas bubbles. Next,
20  milliliters  of nitrogen, made H2-free by passage
through a hopcalite column, is introduced to the  bulb
through a septum. As water continues to purge the bulb,
H2 and other slightly soluble gases partition to the 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.
    16
Explanation
  monitoring
  well location
  and number
     0    100 feet
     scate
                                                      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-
                                                   18

-------
  I
A.
A
\ rt 1
-•- Dissolved Oxygen
-»- Fe(ll)
A Methane
A . A

                                                 (A)
              100200300400500600700800
— ^

Ł40
I 30 H
I 20
8 10 -
                                                 (B)
         ~l	1	1	1	1	1	1	1—
          0   100200300400500600700600
|81
!6~
|2-
S 0 -
                            Methanogenesis


                   i Sulfato Reduction
       Fe(lll) Reduction
(C)
          0       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.

A cross section showing the  interpretation of these data
and including wells screened deeper in the flow system
is given in Figure 3. A methanogenic zone  is present
near the contaminant source, surrounded by sulfate-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
                                                          .2 20
                                                          2 15
                                                          | 10

                                                          J:
                                                         (A)
                  0   100  200  300  400  500  600  700  800
                                XAxis
                                                            700
         | 600 -r
         O 500 --
         .u
         5 300

         | 20°
         •s 100
                                                         (B)
                   i    r   r    i    i    i     i    i
                   0   100  200  300  400  500  600  700  800
                                                         (C)
                  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-
                                                     19

-------
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
NAS 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. Criddle, and P.L. 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., P.B. McMahon, N.M. Dubrovsky, R.F. Fujii, E.T.
    Oaksford, and D.A. Vroblesky. 1995. Deducing the distribution of
    terminal electron-accepting processes in  hydrologically diverse
    groundwater systems. Water Resour. Res. 31:359-371.

14.  Lovley, D.R.,  F.H. Chapelle, and J.C. Woodward. 1994. Use of
    dissolved H2  concentrations  to determine  distribution of micro-
    bially catalyzed redox reactions in anoxic groundwater. Environ.
    Sci. Technol. 28:1255-1210.
                                                            20

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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.

Applteation 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
                                                 21

-------
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 for daughter
products ethane  and  ethene.  A comparison of the
ground-water chemistry from the interval in which the
cores were acquired with that in neighboring monitoring
wells will demonstrate whether the collected cores are
representative of that section of the contaminant plume.

Reductive  dechlorination  of  chlorinated  solvents re-
quires an electron donor for the process to proceed. The
electron donor could be soil organic matter, low molecu-
lar weight 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 transfer the aquifer material
into a jar, exclude air by adding ground water, and seal
the jar without headspace.  The material should  be
cooled during transportation to the laboratory, then 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
                                                   22

-------
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 (SOT 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
                                                    23

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Table 2.  Contaminants and Daughter Products

Analysis             Interpretation
PCE

TCE


1,1,1 -Trichloroethane

c/s-DCE

trans-DCB

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
                                                    24

-------
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  jag/L and  1,100  |ig/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
|j,g/L, 450 u.g/L, and 330  (ig/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
     0   5   10   15   20   25   30   35   40   45
                    Time (Weeks)

Figure 1.  TCE concentrations in the Tibbetts Road microcosm
         study.
                                                                                                o Benzene Microcosm |
                                                                                                • Benzene Contro
                                               rocosml
                                               ntrol  |
     0   5    10  15   20   25   30   35  40   45
                   Time (Weeks)

Figure 2.  Benzene concentrations in the Tibbetts Road micro-
         cosm study.
                                          o Toluene Microcosm
                                          • 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.
                                                    25

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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
167 ± 130
228
304
19.9
184+147
Week 23
Controls
180
116
99.9
132 ±42.4
236
180
152
1 89 ± 42.8
254
185
157
1 99 + 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 for the three compounds are shown in Table 4. The
appropriate rate constant to be used in a model or a risk
assessment would be the first-order removal in the living
treatment  minus the first-order removal in the control, in
other words, the removal that is in excess of the removal
in the controls.

The first-order removal in  the living and control 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 frans-DCE,  1,1-DCE, vinyl
chloride, or ethene,  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
                                                   26

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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
Monitoring well
TCE
c/s-DCE
/rans-DCE
1, 1-DCE
Vinyl Chloride
Ethene
Benzene
Toluene
o-Xylene
m-Xylene
p-Xylene
Ethylbenzene
Methane
Iron
Segment A
SOS
Upgradient g/L
200
740
0.41
0.99
<1
<4
510
10,000
1,400
2,500
1,400
1,300
353

79S
Downgradient g/L
13.7
10.9
< 1
< 1
<1
<4
2.5
< 1
8.4
< 1
22
0.7
77

70S
Upgradient
710
220
0.8
<1
<1
7
493
3,850
240
360
1,100
760
8

Segment B
52S
g/L Downgradient g/L
67
270
0.3
1.6
< 1
<4
420
900
71
59
320
310
3

Segment C
70S 53S
Upgradient g/L Downgradient g/L
710 3.1
220 2.9
0.8 <1
< 1 <1
< 1 < 1
7 <4
493 <1
3,850 < 1
240 <1
360 <1
1,100 <1
760 <1
8 <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
                                                     27

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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.
    McGraw-Hill.
1992. Primer  of  biostatistics. New  York,  NY:
                                                              28

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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.
                                                 29

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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- orcis-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.-R
   Koschitzky,  eds. Groundwater and subsurface remediation:  Re-
   search strategies for in-situ remediation. Berlin/New York: Sprin-
   ger-Verlag. pp.102-127.
                                                   30

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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 pumping test, theflowdistribution entering the well
from the surrounding formation. If certain conditions are
met, the distributions provide sufficient information to
determinethehydraulicconductivityoftheaquiferzones
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
                                                31

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             PUMP-
    CAP ROCK
 (O-DISCHARGE RATE) !
                       I
                       I
      BOREHOLE FLOWjf
      METER  	•	nl
       ELEVATION'
                            . TO LOGGER (Q)
                               ^-LAND SURFACE
..CASING
                                      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.
                                                                                 .AmbtentHow
                                                             0    0.1   1.2   0.1   lut   «J
                                                                              Row(L/mM
                                                                                              a-'"
                                                                                                  ...o
                                                   N«t Pumping Flow
                                            114
                                               HewO/mtol
                        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
                        K distribution obtained in well P26-M1  is shown in Figure
                        4. Heterogeneity is evident, with K varying by an order of
                        magnitude at various locations in the aquifer (3).

                        A particularly illuminating  EM flowmeter application at
                        George AFB was reported by Wilson et al. (Figure 5) (6).
                        The concentration  data were obtained from core 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).
                                                   32

-------
                              Differential Ambracn Bow
   -0.4
                                               0.4
                       Benzene (pg/U)

                       2000      4000
                                                                                                      6000
                        O.f    O->     1
                        FlowIJ_/mln>
                                                                  836
                                                                 835
                                                                 834 .
                                                                 833 -
                                                                 832
                                                                 831
                                                         Elevation
                                                         (meters)
                                                r
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
  §-120
     130
     140
         830
                                                                 829
         828
                                                                 827  ..
                                                                 826
                                                                 825
                                                 r
                   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).
          0         5        10        15       20
             Hydraulic Conductivity  (m/day)
Figure 4.  Hydraulic conductivity as a function of depth in Well
         P26-M1 at the Savannah River site. The measurement
         interval (layer thickness) was 1 foot.
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
                                                     33

-------
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. Giiven.
   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.
                                                          34

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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 lor Natural Attenuation
of Fuel Contamination Dissolved in Groundwater, recently
published by the U.S. Air Force Center for Environmental
                                                35

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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.

For this reason, it is more difficult to predict the long-term
behavior of chlorinated  aliphatic hydrocarbon plumes
than fuel hydrocarbon plumes. Thus, it is important to
have a thorough understanding of the operant natural
attenuation mechanisms. In addition to having a better
understanding of the  processes  of advection, 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.
                                                   36

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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
                                                   37

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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 EPA's 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
                                                   38

-------
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
                                                   39

-------
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 sorptior:. 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
                                                   40

-------
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 water to 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-pathway analysis.

7. Evaluate the practicability and potential efficiency of
   supplemental source removal options.

8. If natural attenuation with or without source removal
   is acceptable, prepare a long-term monitoring plan.

9. Present findings to regulatory agencies, and obtain
   approval for remediation by natural attenuation.

Review Available Site Data, and Develop a
Preliminary Conceptual Model

Existing site characterization data should be reviewed
and used to develop a conceptual model for the site. The
preliminary  conceptual  model  will help identify any
shortcomings in the data and will  allow placement of
additional data collection points in the most scientifically
advantageous and cost-effective manner. A conceptual
model  is  a three-dimensional  representation  of  the
ground-water flow and solute transport system based on
available geological, biological, geochemical, 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.
                                                   41

-------
Engineered Remediation Required,
    Implement Other Protocols
                                                                             Perform Site Characterization
                                                                          to Support Remedy Decision Making
        Assess Potential For
        Natural Attenuation
         With Remediation
         System Installed
                                                                                        Refine Conceptual Model and
                                                                                           Complete Pre-Modeling
                                                                                                Calculations
                                                      Are
                                                  Sufficient Data
                                               Available to Properly
                                                 Screen the Site?
      Are
Screening Criteria
      Met?
               Does it
             Appear That
       Natural Attenuation Alone
         Will Meet Regulate
              Criteria?
                                            Evaluate Use of
                                          Selected Additional
                                           Remedial Options
                                              Along with
                                          Natural Attenuation
    Perform Site Characterization
    to Support Natural Attenuation
     Refine Conceptual Model and
       Complete Pre-Modeling
            Calculations
     Simulate Natural Attenuation
        Using Solute Fate and
           Transport Models
          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 ?
                                                                               Simulate Natural Attenuation
                                                                                Combined with Remedial
                                                                                 Option Selected Above
                                                                              Using Solute Transport Models
                                                                                   Initiate Verification of
                                                                                   Natural Attenuation
                                                                               using Long-Term Monitoring
Figure 1.  Natural attenuation of chlorinated solvents flow chart.
                                                          42

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Table 1A. Soil and Ground-Water Analytical Protocol3
Matrix Analysis Method/Reference6"6 Comments1'9
Soil





Soil





Soil
gas


Soil
gas



Water







Water









Water





Water




Water


Volatile
organic
compounds



Total
organic
carbon
(TOC)


O2, CO2



Fuel and
chlorinated
volatile
organic
compounds
Volatile
organic
compounds





Polycyclic
aromatic
hydro-
carbons
(PAHs)
(optional;
intended
for diesel
and other
heavy oils)
Oxygen





Nitrate




Ironfll)
(Fe+^)

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 SW8310




DO meter





Iron chromatography
Method E300; anion
method


Colorimetric HACH
Method 8146

Handbook
method
modified for
field extraction
of soil using
methanol
Procedure
must be
accurate over
the range of
0.5 to 15%
TOC









Handbook
method;
analysis may
be extended to
higher
molecular-
weight alkyl
benzenes
Analysis
needed only
when required
for regulatory
compliance





Refer to
Method A4500
for a
comparable
laboratory
procedure
Method E300
is a handbook
method; also
provides
chloride data
Filter if turbid


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
solvents/byproducts





PAHs are components
of fuel and are
typically analyzed for
regulatory compliance






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
Recommended
Frequency of
Analysis
Each soil
sampling round




At initial
sampling




At initial
sampling and
respiration
testing
At initial
sampling



Each sampling
round






As required by
regulations








Each sampling
round




Each sampling
round



Each sampling
round

Sample Volume,
Sample Container,
Sample Preservation
Collect 1 00 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







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 or
Fixed-Base
Laboratory
Fixed-base





Fixed-base





Field



Fixed-base




Fixed-base







Fixed-base









Field





Fixed-base




Field


depletion of oxygen,
nitrate, and
manganese
      43

-------
Table 1A. Soil and
Matrix
Water





Water

Water
Analysis
Sulfate
(S04-2)





Methane,
ethane,
and ethene

Alkalinity
Ground-Water Analytical Protocol9 (Continued)
Method/Reference1"
Iron chromatography
Method E300 or
HACH Method 8051





Kampbell et al. (35)
or SW3810, modified

HACH alkalinity test
kit Model AL AP MG-L
Comments*'9
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
Data Use
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
Recommended
Frequency of
Analysis
Each sampling
round





Each sampling
round

Each sampling
round
Sample Volume,
Sample Container,
Sample Preservation
Collect up to 40 mL
of water in a glass or
plastic container; cool
to4°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 mLof
water in glass
container
Field or
Fixed-Base
Laboratory
Ł300 =
Fixed-base
HACH
Method
8051 = Field



Fixed-base

Field
Water  Oxidation-
       reduction
       potential
A2580B
Water  pH
Field probe with
direct reading meter
Water  Temperature Field probe with
                   direct reading meter

Water  Conductivity E120.1/SW9050,
                   direct reading meter
Water  Chloride
Mercuric nitrate
titration A4500-CI" C
               capacity of ground
               water; can be used to
               estimate the amount
               of CO2  produced
               during biodegradation

Measurements  The oxidation-         Each sampling
made with      reduction  potential     round
electrodes,      of ground water
results are      influences and is
displayed on a  influenced by the
meter, protect   nature of the
samples from   biologically mediated
exposure to     degradation of
oxygen; report  contaminants; the
results against  oxidation-reduction
a silver/silver   potential of ground
chloride        water may range from
reference       more than 800 mV to
electrode       less than  -400 mV

Field           Aerobic and           Each sampling
               anaerobic processes    round
               are pH-sensitive
                                                           Each sampling
                                                           round

                                                           Each sampling
                                                           round
Field only      Well development


Protocols/      Water quality
Handbook      parameter used as a
methods       marker to verify that
               site samples are
               obtained from the
               same ground-water
               system

Ion            Final product of        Each sampling
chromatography chlorinated solvent     round
Method E300;  reduction; can be
Method        used to estimate
SW9050 may   dilution in calculation
also be used   of rate constant
Collect 100 to
250 mL of water
in a glass container
                                                                                                                   Field
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
                                                                                                                   Field
                                                                                                                   Field
Field
                                                                                                                   Fixed-base
                                                               44

-------
Table 1A.  Soil and Ground-Water Analytical Protocol3 (Continued)
Matrix
Water
Analysis
Chloride
(optional;
see data
use)
Method/Reference1"
HACH chloride test
kit Model 8-P
Comments''9
Silver nitrate
titration
Data Use

As above, and to
guide selection of
additional data points
in real time while in,
the field
Recommended
Frequency of
Analysis
Each sampling
round
Sample Volume,
Sample Container,
Sample Preservation
Collect 100 mLof
water in a glass
container
Field or
Fixed-Base
Laboratory
Field
Water  Total
       organic
       carbon
SW9060
Laboratory
Used to classify
plumes and to
determine whether
anaerobic metabolism
of chlorinated solvents
is possible in the
absence of
anthropogenic carbon
Each sampling
round
Collect 100 mLof
water in a glass
container; cool
Laboratory
 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).
0 "E" refers to Methods for Chemical Analysis of Water and Wastes (30).
d "HACH" refers to the Hach Company catalog (31).
6 "A" refers to Standard Methods for the Examination of Water and Wastewater (32).
f "Handbook" refers to the AFCEE Handbook to Support the Installation Restoration Program (IRP) Remedial Investigations and Feasibility
 Studies (RI/FS) (33).
9 "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 Consideration3'15

                                                                               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-fert-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


SW8260/80150






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
dechlonnation
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.
 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).
                                                             45

-------
Screen the Site, and Assess the Potential for
Natural Attenuation

After reviewing available site  data and developing a
preliminary conceptual model, an  assessment of the
potential for natural attenuation must be made. As stated
previously,  existing data can be  useful in determining
whether natural attenuation will be  sufficient to prevent
a dissolved contaminant plume from completing 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 Ta-
  ble 2  for  a minimum  of six samples. Figure  3 shows
  the approximate  location of these data collection
  points. If other contaminants  are suspected, then
  data on the concentration and  distribution  of these
  compounds also should be obtained.

• Locations of source(s) and  receptor(s).

• An estimate of the contaminant transport velocity and
  direction  of ground-water flow.

Once these data have been  collected, the screening
process can be undertaken. The following steps 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.
                                                   46

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       Analyze Available Site Data
     to Determine if Biodegradation
              is Occurring
                                Collect More Screening Data
     Engineered
Remediation Required
   Implement Other
      Protocols
                                                         Are
                                                    Sufficient Data
                                                      Available ?
Biodegradation
  Occurring?
Insufficient
Data
    Determine Groundwater Flow and
    Solute Transport Parameters using
     Site-Specific Data; Porosity and
      Dispersivity May be Estimated
            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?
      Perform Site Characterization
      to Support Natural Attenuation
                    Proceed to
                    Figure 1
Figure 2.  Initial screening process flow chart.
                                                        47

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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
pHa
DOC
Temperature3
Carbon dioxide
Alkalinity
Chloride3
Hydrogen
Hydrogen
Volatile fatty acids
BTEXa
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 < pH < 9
> 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-Trichloroethanea                                    Material released
1,1-dichloroethene3                                      Daughter product of trichloroethene or chemical reaction
                                                        of 1,1,1-trichloroethane
a Required analysis.
  Points awarded only if it can be shown that the compound is a daughter product (i.e., not a constituent of the source NAPL).
                                                                48

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                         Helps Define
                         Lateral Extent
                     F   of Contamination
                                      Helps Define
                                      Downgradlent Extent
                                      of Contamination
              Plume Migration
       LEGEND
 8 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
Oto 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
-190 mV
3x background
1 ,000 ug/L
1 ,200 ug/L
500 ug/L
50 ug/L
Total points awarded
Points
Awarded
3
2
3
2
3
2
2
0
2
2
2
23
                                                    49

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In this example, the investigator can infer that biodegra-
datioFMS 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
10 mg/L
ND
100mV
Background
1,200 ug/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
                                                   50

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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 whether the 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 hydraulic
  conductivity and hydraulic gradient, as well as estimates
  of effective porosity and contaminant retardation?
• Is it likely that  the contaminant mass is attenuating
  at rates sufficient to be  protective of human health
  and  the environment at a point of  discharge to  a
  sensitive environmental receptor?
• Is the plume going to attenuate to concentrations less
  than risk-based corrective action  guidelines before
  reaching potential receptors?

Collect Additional Site Characterization  Data
To Support Natural Attenuation, As Required
Detailed site characterization is necessary to document
the potential for natural attenuation. Review of existing
site characterization data is particularly  useful before
initiating site characterization activities.  Such review
should allow identification of data gaps and guide the most
effective placement of additional data collection points.
There are two goals during the site characterization
phase of a natural attenuation investigation. The first is
to collect the data needed to determine whether natural
mechanisms of contaminant attenuation  are occurring
at rates sufficient to protect human health and the 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 aqui-
  fers 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 ar-
  eas 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.
The analytical protocol to be used for soil sample analy-
sis  is presented in Table  1. This analytical protocol
                                                   51

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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-
tassium, or ammonia. These species  result from the
dissolution of  rock (especially  carbonate rocks), the
transfer of carbon dioxide from the atmosphere, and the
                                                   52

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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
acquisition  using a direct-reading meter. Because most
well purging techniques can allow aeration of collected
ground-water samples (which can affect oxidation-reduction
                                          Redox Potential (Eh°)
                                          in Millivolts @ pH = 7
                                             and T=25rC
Decreasing Amount of Energy Released During Electron Transfer
 HS"
CO, + 8rt' + 8e 	 > CH4
^
                                                                                 !k° = +
                                                                                       (Eh' = +740)
                                                                                   MnCOs(s) + 2H3O
                                                                                  •  FeCO,+2H1O


                                                                                  1   (Eh1 = -220)
                                                                         CH4+2H,0   (Eh" = -240)
         Modified From Bouwer (1994)

Figure 4.  Redox potentials for various electron acceptors.
                                                   53

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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
dissolved hydrogen content is  not yet commonplace or
standardized, however, and requires a relatively expen-
sive field laboratory setup.
Table 4.  Range of Hydrogen Concentrations for a Given
        Terminal Electron-Accepting Process
Terminal
Electron-Accepting Process
Hydrogen Concentration
(nanomoles per liter)
 Denitrification

 Iron(lll) reduction

 Sulfate reduction

 Methanogenesis
0.2 to 0.8

1 to 4

>5
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"io 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 CI" 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-
ity.  Estimating hydraulic conductivity and gradient in the
field is fairly straightforward,  but obtaining field-scale
information on porosity and dispersivity can be difficult.
                                                    54

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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).

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
                                                   55

-------
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
manage the risk associated with such expansion. Figure
5 depicts 1) an upgradient well  in unaffected ground
water, 2) a  well in the NAPL source  area, 3) a well
downgradient of the NAPL source area in a zone of
anaerobic treatment, 4) a well in the zone of aerobic
treatment, along the periphery of the plume, 5)  a well
                                                  56

-------
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.
                            Anaerobic Treatment Zone
                 Plume Migration
                                    Extent of Dissolved
                                    BTEX Plume
                                    Aerobic Treatment
                                    Zone
 • Poinl-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 m conjunction
with the appropriate regulators


Figure 5.  Hypothetical long-term monitoring strategy.
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 fora 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.

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34. AFCEE. 1992. Environmental chemistry function Installation Res-
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35. Kampbell, D.H., J.T. Wilson, and S.A. Vandegrift. 1989. Dissolved
    oxygen  and methane in water  by  a GC headspace equilibrium
    technique. Int.  J. Environ. Anal. Chem. 36:249-257.
                                                               58

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36. Wiedemeier, T.H., M.A. Swanson, J.T. Wilson,  D.H.  Kampbell,
    R.N. Miller, and J.E. Hansen. 1996. Approximation of biodegra-
    dation rate constants for monoaromatic hydrocarbons (BTEX) in
    groundwater.  Ground  Water Monitoring and  Remediation.  In
    press.

37. Wiedemeier, T.H., M.A. Swanson,  D.E. Moutoux, J.T. Wilson,
    D.H. Kampbell, J.E.  Hansen, P. Haas, and F.H. Chapelle. 1996.
    Technical protocol for natural attenuation of chlorinated solvents
    in groundwater. San Antonio, TX: U.S. Air Force Center  for En-
    vironmental Excellence. In preparation.

38. Domenico,  P.A. 1987.  An analytical model for multidimensional
    transport of a decaying contaminant species. J. Hydrol. 91:49-58.

39. Domenico, P.A., and F.W. Schwartz. 1990. Physical and chemical
    hydrogeology. New York, NY: John Wiley and Sons.

40. Morel, F.M.M., and J.G. Hering. 1993. Principles and applications
    of aquatic chemistry. New York, NY: John Wiley & Sons.

41. Willey, L.M., Y.K. Kharaka, T.S. Presser, J.B. Rapp, and I. Barnes.
    1975. Short chain aliphatic acid anions in oil field waters and their
    contribution to the measured alkalinity.  Geochim.  Cosmochim.
    Acta 39:1707-1711.

42. Lovley, D.R., and S. Goodwin.  1988. Hydrogen  concentrations
    as an indicator of the predominant terminal electron-accepting
    reaction in aquatic sediments.  Geochim.  Cosmochim.  Acta
    52:2993-3003.
43. Lovley, D.R., F.H. Chapelle, and J.C. Woodward.  1994. Use of
    dissolved Ha concentrations to determine distribution of micro-
    bially catalyzed redox reactions in anoxic groundwater. Environ.
    Sci. Technol. 28(7):1205-1210.

44. Chapelle, F.H., P.B.  McMahon, N.M. Dubrovsky, R.F. Fujii, E.T.
    Oaksford, and D.A. Vroblesky. 1995. Deducing the  distribution of
    terminal electron-accepting processes in hydrologically diverse
    groundwater systems. Water Resour. Res. 31:359-371.

45. Hem, J.D. 1985.  Study and interpretation of the chemical char-
    acteristics of natural water. U.S. Geological Survey  Water Supply
    Paper 2254.

46. Kaufman, W.J., and  G.T. Orlob. 1956. Measuring  ground  water
    movement with  radioactive and chemical  tracers. Am. Water
    Works Assn. J. 48:559-572.

47. Buscheck, T.E., and C.M. Alcantar. 1995. Regression techniques
    and analytical solutions to demonstrate intrinsic bioremediation.
    In: Proceedings of the 1995 Battelle  International Conference on
    In-Situ and On Site Bioreclamation. April.

48. Bear,  J.  1979.  Hydraulics of groundwater.  New York, NY:
    McGraw-Hill.

49. Rice, D.W., R.D. Grose, J.C. Michaelsen, B.P. Dooher, D.H. Mac-
    Queen, S.J. Cullen, W.E.  Kastenberg, L.G. Everett,  and M.A.
    Marino. 1995. California leaking underground fuel tank (LUFT)
    historical case analyses. California State Water Resources Con-
    trol Board.
                                                              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
                                                 60

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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 EPA's Center for
Subsurface Modeling Support (CSMoS), NRMRL/SPRD,
P.O. Box 1198, Ada, OK 74821-1198, telephone 405-436-
8594, fax 405-436-8718, bulletin board  405-436-8506
                                       CENTERLINE
                                        ••••••i
                                        View Output
                            Restore Formulas lor Vs
                          DispersMties, R, lambda
Figure 1. BIOSCREEN input screen.
                                                 61

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 Figure 2.  BIOSCREEN centerline output screen.
                                                                    *'
~4
— M
5s
-ISO
•M';.-Q tj •Ł;

0.100
1 500
0.100
0.000
- <f
0,000
0.000
0000
0000
0.000

0000
0000
0.000
0.000
0.000

0000
0000
, 0,000
0000
0000
' -. • '•..,'•• '•'••<•.;• :••
0000 0000
0.000 0000
0,000 0.000
0000 0.000
0000 0000
                                                                                                            Wo Degradation
                                                                                                                Model
                                                                                                           -^-';	'	'1	''  '- '*'',•
                                                                                                            1st Order Decay
                                                                                                                Modal
                                       Tafgeti0Vi9»; i  OOPS   rno/L
Figure 3.  BIOSCREEN concentration array output screen.
                                                            62

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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.
                                                          63

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                     Case Study: Naval Air Station Cecil Field, Florida
                              Francis H. Chapelle and Paul M. Bradley
                         U.S. Geological Survey, Columbia, South Carolina
Redox processes at a fire-training area at Naval Air
Station Cecil Field in Florida are segregated into distinct
and clearly definable zones. Near the source of contami-
nation, methanogenesis predominates. As ground water
flows downgradient, distinct sulfate-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.
                                                   64

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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.
                                                65

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                      587 588  >589 590  591 592
    589 around Water Elevation Control
    '       (Data from 9-23-87)
Figure 1.  St. Joseph Superfund site plan.
0
03
•g
CU
O
cr
o
O

-------
         6
                                    1,23

                                    lake
              0     100    200   300    400
               Micromolar Oxygen Concentration

Figure 4.  Composited chlorine number plotted against oxygen
         concentration.


 . F ** D4.5
A » A A*" * ° lakC

* n* 1'*^.*^ jr" * *•
i*i* • * A A
A A S A A
11 A A * A
* " .A "* ' *
A * A*"
"A A • A » A
                                                        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 ^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, NC|, that is  defined by
                   NCi=-
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),
        o
        O
             104
             10'
10'
             10°
                                  -A SCH
                                  -* OXYGEN
                                  -• METHANE
               40    50    60    70   80
                        Depth (teet)
                                        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
   600


   500

§"
H. 400

6
1° 300

(U
                                                              o
                                                              O
                                                                 200
                                                                 100
                                                              -aTCE
                                                              -o c-DCE
                                                              -* VC
                                                              -i ETHENE
                                                      40    50   50    70    80
                                                                Depth (feet)
                                                                                  90
                                         Figure 7.  Distribution of ethenes with depth at Borehole T23.
                                                   67

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             10'
             103
             102
             10'
             10°
               40   50   50   70   80   SO   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.
                                                                o
           600

           500

           400

           300

           200

           100
              40   50  60   70   80
                       Depth (feet)
                                    90   100
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, P.L., and J.T. Wilson. 1992. Natural anaerobic treatment
  of a TCE plume St. Joseph, Michigan, NPL sites. In: U.S. EPA.
  Bioremediation of Hazardous Wastes. EPA/600/R-92/126. pp. 47-50.
3. U.S. EPA. 1995. Natural bioattenuation  of trichloroethene at the
  St. Joseph, Michigan, Superfund site. EPA/600/V-95/001.
4. Semprini, L., P.K. Kitanidis, D. Kampbell, and J.T. Wilson. 1995.
  Anaerobic transformation of chlorinated aliphatic hydrocarbons in
  a sand aquifer based on spatial chemical distributions. Water 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.
                                                     68

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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 tjjg/L)
                        Highest Concentration
Distance
From    Transport Transect
Source     Time    Width
(m)         (y)      (m)     TCE
          Vinyl
cis-DCE   Chloride
130
390
550
855
3.2
9.7
12.5
17.9
108
150
192
395
6,500
68,000
520
8,700
15
56
<1
0.4
8,100
128,000
830
9,800
18
870
<1
0.8
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
                                                69

-------
by summing over the blocks and dividing by the area of
the transects.
In Table 2, concentration estimates are presented for the
perpendicular transects ordered from furthest upgradient
(Transect 2) to furthest downgradient (Transect 5). The
concentration estimates are based only on blocks from
the anaerobic portion of the aquifer (and thus differ from
the averages in Table 1). All of the chlorinated ethenes
show decreasing concentration with distance 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 (ng/L) From the
        Anaerobic Zone
Chemical
                                  Lake
Transect 2  Transect 4  Transect 5   Transect
TCE
cis-DCE
t-DCE
1,1 -DCE
VC
Ethene
Sum of the
ethenes
7,411
9,117
716
339
998
480
19,100
864
1,453
34.4
24.3
473
297
3,150
30.1
281
5.39
2.99
97.7
24.2
442
(1.4)
(0.80)
(1.1)
blq
(0.16)
No data
(3.5)
Chloride
            65,073
          78,505
92,023
44,418
Note: Values in parentheses were based on one or more estimated
values; blq indicates no detection above the limit of quantitation.
                                            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
                                                                              dc
                                             (Eq. 1)

where R is the retardation coefficient; c is the concen-
tration; t is time;  D^ 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 for the
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.
                                                    70

-------
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) = 0
                    c(0,y,f) = c0exp(
         c (°°,y,0 = c (x,-°°,f) = c (x.00,9 = 0
                                           (Eq. 2)
                                                  71

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has the approximate steady state solution (9)
                               Net and Gross Decay Rates
                      -j,^
                                     2 +•
c(x,y)
                                          1 +-
                                              (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, a, 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 = 6a. A mass balance can then be solved for the
peak concentration of the Gaussian distribution, c0, from
r A     f
jncdy = J
                  nc0
-y2
-^) dy= nc0 a
                                              (Eq. 4)
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
     1 do2
     2dt
                                              (Eq. 5)
where a^ 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 (A x), and change in  variance of the  Gaussian
distributions for the transect  concentrations (A a2):
                           1 Aa2
                     (Eq. 6)
The only remaining unknown in Equation 3 is the decay
constant X*, which  is determined through a bisection
search. Table 5 gives the rate constants from the two-
dimensional model.
                               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:
                                                                                          "-M
                                                                                                      (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 Xj(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).


References

1. McCarty, PL., and L. Semprini. 1994. Ground-water treatment for
   chlorinated solvents. In: Norris R.D., R.E. Hinchee, R. Brown, PL.
   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.
                                                      72

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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.
                                                              73

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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
                                                 74

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Table 1.  Analytical Data, Pittsburgh 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 TMB
(feet) (ug/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
(Ug/L) (ug/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
(ug/L) (ug/L)
51,412
12,626
14,968
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
(H9/L)
1,420
1,600
305
339
1,010
714
NA
617
3,530
1,800
115
44
Ethene Chloride
(ug/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
a 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
                                                  75

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                     Extent of
                     Dissolved
                     Contaminant
                     Plume (1996)
                        1L800
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.
                                                  76

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       A
   Northwest
                                  Southeast
                                                                                      Discharge
                                                                                      to
                                                                                      Wetlands
                                                   Well-Sorted, Fine-to
                                                   Medium-Grained Sand
                          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
resQIts 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.
                                                  77

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    TOTAL BTEX
   —4
            TRICHLOROETHENE

  > 4,000 pg/L     /"VV  \ v-s-gJu^     B>10.000pg/L
  2,000 - 4,000 pg/L  jg^ndSiilil^.  • 1'000 '10'°00 P8"-
5 ND - 2,000 pg/L   AM"WiBiii!Mk«, II ND -1,000 pg/L
             DICHLOROETHENE
   VINYL CHLORIDE
    'h
    c
• >1,000 pg/L
• 500 -1,000 pg/L
1 ND- 500 pg/L
>600pg/L
100- 500 pg/L
ND -100 pg/L
                                                                                  >S,000 \iglL
                                                                                  1,000 - 5,000 pg/L
                                                                                  ND-1,000 pg/L
>100mg/L
50 -100 mg/L
ND - 50 mg/L
Figure 3.  Chlorinated solvents and byproducts (1995).
 TOTAL BTEX
 NITRATE
                           DISSOLVED OXYGEN
                             > 4,000 ug/L

                             2,000- 4,000 MS/L

                             ND.2,000ug/L
                            2 -4 mg/L

                            0.05 - 2 mg/L
                            < 0.05 mg/L
                                                                  5-10 mg/L

                                                                  1 -5 mg/L

                                                                  <1mg/L
                                                                 10 -20 mg/L

                                                                 0.05 -10 mg/L

                                                                 < 0.05 mg/L
Figure 4.  BTEX and electron acceptors (1995).
                                              78

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Table 3. Approximate First-Order Biodegradation Rate Constants
A- B B -C C - E
Oto 970 to 1,240 to
Correction 970 feet 1,240 feet 2,560 feet
Compound Method (1/year) (1/year) (1/year)
TCE Chloride 1.27 0.23 -0.30
TMB 1.20 0.52 NA
Average 1.24 0.38 -0.30
DCE Chloride 0.06 0.60 0.07
TMB 0.00 0.90 NA
Average 0.03 0.75 0.07
Vinyl chloride Chloride 0.00 0.14 0.47
TMB 0.00 0.43 NA
Average 0.00 0.29 0.47
BTEX Chloride 0.13 0.30 0.39
TMB 0.06 0.60 NA
Average 0.10 0.45 0.39
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 + 2H2O + 2.5C-CI -> CO2 + 2.5H+ +
2.5CI- + 2.5C-H (Eq. 1)
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.o5iOo.6i9 + 1.38H2O + 1.91C-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.
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
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
                                                 79

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      TOTAL BTEX
         IRON(II)
                                   > 4,000 |ig/L

                                   2,000-4,000 ug/L
                                    >10mg/L

                                    5 -10 mg/L

                                    O.OS - 5 mg/L
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(ll) 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-
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
                                                  80

-------
 0    450   900
           ^^
            FEET
                                                        e oi^/ype
                                                        avioV
             T
                        Zone of
                        Type3
                        Behavior
1,800
                                                                                         47PITK2
                                                                                         4W1.1W22*
                                                                                         4*11*22
Figure 6. Zonation of CAM plume.
    60000
 J  50000
        0    500  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).
                                                  81

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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 Ha 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.
                                                              82

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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 (ng/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 |j,g/L are found in the source
area, the TCE concentrations are highest (greater than
10,000 |ig/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.
                                                 83

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                                                              A.
                                                                                     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	A'   Line of section

                                                                 • 41-9      Ground-water sampling site location
                                                                            and local identifier
                                                              B.
 „___-
 METERS
    200    100     0     100     200    300    400    SOO    800

             DISTANCE FROM BUILDING 24 (B-24). IN METERS
                                                                                    EXPLANATION
                                                      -210	LINE OF EQUAL TRICHLOROETHENE
                                                                          CONCENTRATION-Shows trichloroethene
                                                                          concentration, in micrograms per liter.
                                                                          Dashed where approximate

                                                                         Well screen and trichloroethene concentration,
                                                                    10     hi 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) area! 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
ground water in the plume is near neutral (6.5 to 7.5),
and concentrations of both dissolved oxygen (less than
0.5 milligrams per liter [mg/l]) and nitrate (less than  1
mg/L) are very low. Concentrations of iron(ll) are greater
than 1 mg/L in some areas of the plume, whereas sulfate
and carbon dioxide are consistently plentiful (greater than
40 mg/L and 100  mg/L as bicarbonate, respectively) as
                                                       84

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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 [ig/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
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  (ig/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-
                                                  85

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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.
                                                  86

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                                 ADVECTION-DRIVEN
                                  VOLATIUTZATION
                                     (50 kg/yr)
                          Wtt*rt»ble
                    DIFFUSION-DRIVEN
                     VOLATILIZATION
                          kg/yr)
               ADVECTIVE
              TRANSPORT TO
            GREEN POND BROOK
                (50 kg/yr)
                            .•-•••- -.•.•••-.•.•••-.  INFILTRATION ••'-.-.•.••.•
                            •"••••".•.••    <1 fc      "•'•"••
                             TRICHLOROETHENE  PLUME
                                                                            ANAEROBIC
                                                                        BIOTRANSFORMATION
                                                                            (360 kg/yr)
                                DISSOLUTION
                                 OF DNAPL
                                (not ostlmatad)
                                  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
                        SORPTION	<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 |j,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
                                                87

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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 ng/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.
                                                     88

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 8. Koller,  D., T.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, T.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 T.E. Imbrigiotta. 1994. Contamination of ground
    water with trichloroethylene at the Building 24 site at Picatinny
    Arsenal, New Jersey. In: U.S. EPA Symposium on Intrinsic 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, T.E. Imbrigiotta, and J.T. Wilson. 1991.
    Reductive dechlorination of trichloroethylene in anoxic aquifer
    material from Picatinny Arsenal, New Jersey.  In:  Morganwalp,
    D.W., and D.A. Aronson, eds. U.S. Geological Survey Toxic 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., T.E. Imbrigiotta,  B.H. Wilson, and J.T. Wilson. 1991.
    Biotransformation of cis-1,2-dichloroethylene in aquifer material
    from Picatinny Arsenal,  Morris County, New Jersey. In: 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., P.R. 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.
                                                              89

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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
                                                 90

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     E-4 •  Extraction wells

     R-SA  Recharge wells
 A
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
                                                       ^M Site II


                                                       • Site XV

                                                       I Site I
                          Site IV
     Note: Sites II, III, VII, and VIII were reportedly used for disposal of DNAPL 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.
                                                   91

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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     ^ Hargis and Montgomery, Inc. 1982. Phase II 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. I. 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.
                                                       92

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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.

A wide 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.
                                              93

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Background

The  RTDF  Bioremediation Consortium  initiated  this
study in February 1995. Dover AFB was chosen over the
many other sites evaluated for the study because:

• The  plume is well-characterized.

• Analyses  of ground-water chemistry provided clear
  evidence  that chlorinated solvent contaminants are
  being biodegraded.

• The  deep zone of the aquifer has  relatively simple
  geology and is underlain by a thick confining layer.

• Access for sampling and testing is good, and the site
  is easily  reached  by offsite  personnel and visitors.

• The  base has a proactive environmental program.

The  plume  contains primarily TCE and dichloroethene
(DCE), with smaller amounts  of vinyl  chloride  (VC). It
occupies an area north and south of U.S. Highway 113
approximately  9,000 feet long  and 3,000 feet wide.
There are multiple sources of solvent  contamination in
the area north of the highway, as well  as several minor
sources of petroleum hydrocarbons. There appear to be
at least three sources of TCE.

The water-bearing unit in the study area is composed of
fine- to coarse-grained sands ranging in thickness from
30 to 60 feet. The ground-water elevation ranges from
approximately 13 feet mean sea level (MSL) at the north
end  of the  plume to less than 3 feet MSL near the
southern end.  Ground water flows  to the south. The
plume velocity ranges from about 150 feet per year in
the northern portion of the study area to over 200 feet
per year beneath the southern area. The consortium
believes that the aquifer contains aerobic and anaerobic
microzones. This simple sand aquifer exhibits complex
metabolic activity that might not be apparent  from  a
cursory examination of geochemical information.

This paper focuses on the  lower third of the aquifer,
which has  the highest permeability and  contains the
majority of the contaminants.

Current Findings

Intrinsic remediation is clearly occurring in the  ground
water, and results suggest that multiple biodegradation
pathways are operating. These findings are based on data
on the plume profile, contaminant concentrations and 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.
                                                  94

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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 (|o.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.
    //?

  1 Monitoring well
                                                                         500
                                                                                             1000 FEET
Figure 2. Total chlorinated compounds (mg/L) in the deep zone at Dover AFB.
                                                 95

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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  SOO^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
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
                                                  96

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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.
                                                      97

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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
                                                 98

-------
 EXPLANATION

        IRPSrte
       Gfoundwater Plume

       Direction of Groundwater Flow
• - - — Base Boundary


   ^ EOD Range Safety Zone
0  750  1500   3000 Feet
                     o
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.
                                                    99

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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 oxic to transitional redox
conditions exist in the  dissolved plume. The extent of
contamination  of aquifer materials remains unknown.

Facilities

EPA (Region 5),  Michigan Department  of Natural 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 hte measurements for delineat-
ing redox processes in a contaminated aquifer. In preparation.
                                                   100

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                                                     v
                                      \  lw.-.'l«jilM        ^
                                      *., v..^,™    V    V— HA-M
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)8
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
Water
(mg/L)
0.08





0.003





Aquifer
Solids
(mg/L)
10.5





6.3





% of Total
Associated
With
Solids
99%





99%





Volume
% of
Oily
Phase"
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.
5 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).
                                                     101

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                           TCE

                        Deep Wells

                 (At Midfield and Farfield)
                                                           25
                                                           15 -
      9 -
      *
                                                         S 10
                                                         c
                                                         o
                                                         u
      o =_
          8/87 4/88 8/90 12/93 4/94 6/94 9/94  1/95 4/95 7/95 10/95 1/96

                             Tim*
                                                                                TCE

                                                                            Shallow Wells

                                                                       (At Midfield, and Farfield)
                                                           0 -
                                                                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
                     cis-1,2

                    Shallow Wells

                     (At Midfield)
-..— PTJS

—=— FT12S

—«---FT13S
                                                                            cis-1,2-DCE

                                                                            Deep Wells

                                                                     (At Nearfield and Midfield)
   1600
  1400


  1200 -

*
* 1000 7-


5 800 ^
c
o
c 600 -

         12/93  4/94  6/94  9/94   1/95  4/95   7/95  10/95  1/96

                           Tim.
                                                         350 -


                                                         300


                                                         250


                                                         200
                                                       c 150
                                                       o
                                                       U

                                                         100


                                                          50


                                                          0
                                                              12/93  4/94  6/94  9/94  1/95  4/95  7/95 10/95  1/96

                                                                                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
                                                     102

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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 contaminati°n remediation efforts, in: Gamboiati, G. and G.
                                                                   Verri, eds. Advanced methods for ground water pollution central.
                                                                   International Center for Mechanical Sciences, University of Udine,
                                                                   University of Padua, May 5-6, 1994, Udine,  Italy.  Courses and
                                                                   Series No. 364. Wien/New York: Springer-Verlag. pp.  161-175.

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.
                                                            103

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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 Nume .
Condflfons?
2. Estimate
Contaminant
Degradation Rate
3. Estimate Source
Mass
4. Estimate Source
lifetime
- 5. Long-Term Behavior
4. Intrinsic
Remediation for Site?
7. Long-Term
Monitoring for Site

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
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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
 for the site. Center of mass (CoM) and total mass results
 for Site 45/57 were generated from  ground-water
 concentration data collected in this field  study using a
 Thiessen area approach (1-3). Both TCE centerline
concentrations and dissolved plume mass estimates
using a consistent set of sampling locations over time
indicated a decreasing plume mass, with CoM locations
indicating no net plume migration  over the  sampling
interval. The data indicated a finite source producing a
stable TCE plume at Site 45/57 (2, 3, 6).

Estimation of Contaminant Degradation Rate

Estimation of contaminant degradation  rates can  be
carried out using dissolved contaminant mass data if a
declining mass of contaminant is observed over time in
the plume. With estimated dissolved TCE 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
                                                                 EIELSON AFB SITE 45/57
                                                                  Overlay plot sh owing TCE
                                                                 and reduction by-products
                                                                        JuV 1995
Figure 2. Overlay plot of TCE and its degradation products measured in July 1995 at Site 45/57, Eielson AFB, Alaska.
                                                 105

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                                       40      60      80      100      120

                                       DUtonc* Downgradant from Source Araa at 45MWM (m)
                                                                            UO
                                                                                    160
                               20
                                      40      60       80      100      120
                                      Dfstanc* DowngradlMit hom Souic. Ama at 45MWM (m)
                                                                            140
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 over time
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) (6)     (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 0 = 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 0 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/L and a ground-water
                                                  106

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impact concentration (maximum contaminant level [MCL])
of 5 micrograms per liter (ng/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 1Cr5)/(-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 l_s). 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/MWResldua| 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:
                      E Residual =
     Ys (Ls) (b) (0) (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) (6) (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 =
          MassTCE Residual/Mass flux =
   (1,677 kg)/(0.15 kg/d) = 10,897 d = 29.9 yy (Eq. 6)
As this example illustrates, if residual mass does exist,
the lifetime of the  plume is  extended significantly,  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,
                                                  107

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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 .
                                      SP)0«
SP41 •
                SP38
                                                    SP12
                                                                        SP32
                                                                             SP29
                                SP42
        Approximate Scale
            100 meters
                                              SP40
                               45MW02
                                                                   GP02
                                          SP8
                                             MW08
                                                                            45MW07 *
   %  Existing Long-Term Monitoring Wells
   O  Proposed Long-Term Monitoring Wells
   •§•  Proposed Point-of-Compliance Wells
                               | | t = 0 years after source removal/depletion

                                   t = 3 years after source removal/depletion

                               ^1 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 over time. 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 |ig/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 over time to verify plume stability and source area
depletion, at a reasonable frequency  based  on cost
considerations without compromising human health or
environmental quality.
                  Conclusion

                  This paper highlights the application of an intrinsic 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 over time, and calibration of field data to a
                                                   108

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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, P.A.  1987. An analytical model  for multidimensional
   transport of decaying contaminant species. J. Hydrol. 91:49-58.

8.  Gorder, K., R.R. Dupont, D.L. Sorensen, M.W. Kemblowski, and
   J.E. McLean. 1996. Application of a simple ground-water model to
   assess the potential for intrinsic remediation of  contaminated
   ground-water. Presented to the First IBC International Conference
   on Intrinsic Remediation, London, UK. March 18-19.

9.  Parker, J.C., R.J. Lenhard, and T. Kuppusamy. 1987. A parametric
   model  for constitutive properties governing multiphase flow in 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 (EPA's) Ground  Water Protection
Strategy. This strategy, which is outlined in the preamble
to the National Contingency Plan (NCP), includes  a goal
to return usable ground waters to their beneficial uses
within a timeframe that is reasonable  given the particular
circumstances of the site. The preamble  to the NCP
further states  that ground-water  remediation  levels
should generally be attained throughout the 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
/Ffl/8711), which states that institutional controls, while
not actively cleaning up  the contamination at the site,
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can control exposure and, therefore, are considered to
be limited-action alternatives.

The selected remedy could be one of the alternatives or
a combination, depending on the degree of active res-
toration required. Considerations for remedy selection
were the amount of contamination remaining in the
unsaturated and saturated zones and the availabil-
ity of the contamination for removal and treatment.
The technical evaluation is largely  an issue of a
balance between the need for and feasibility of con-
taminant removal in the unsaturated and/or saturated
zones  and the efficiency  of natural attenuation.  The
NCP (55 FR8734) 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 (NCP §300.430(f)(4)(ii)).
At Eielson AFB, the regulatory options considered  are
described  below.

Alternate Concentration Limits
Alternate concentration limits (ACLs, 55 FR 8732)  are
considered when the ground water has  a  known  or
projected point of entry to  surface water with no 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 FR 8754).
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 FR8753).

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 water will protect
human health.

Selection of Natural Attenuation With or
Without Institutional Controls

Natural attenuation  is  generally  recommended  only
when more active restoration is not practicable, cost-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 FR8734).

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.
                                                  113

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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 or stakeholders
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.
                                                 114

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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. A site
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.
                                                 115

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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?
                                                116

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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      _ To examine plume flow velocity?
pume  e avior.                                      _ TQ examjne proximity to sensitive  receptors?
What should  be  the relative investment  in sample
acquisition,  sample analysis, data reduction, mathe-
matical modeling, and report preparation?
                                              117

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       Introductory Remarks: Appropriate Opportunities for Application-
                      Civilian Sector (RCRA and CERCLA)
                                   Fran Kremer
        U.S. Environmental Protection Agency, Office of Research and Development,
                                 Cincinnati, Ohio
(Paper unavailable at press time.)
                                       118

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       Introductory Remarks: Appropriate Opportunities for Application—
                    U.S. Air Force and Department of Defense
                                   Patrick Haas
      U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, Texas
(Paper unavailable at press time.)
                                       119

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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 treat-
  ment.

• 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 greater the analytical
cost. Therefore, there is often  a reluctance to evaluate
a large number of remedial alternatives. DuPont has
found,  however, that the incremental cost of evaluating
intrinsic bioremediation along with other options is rela-
tively small. This incremental  cost may  be  more  than
offset if intrinsic remediation is chosen over a technology
                                                 120

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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
                                                    121

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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 bio-de-
  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 TOG 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 already
   occurring  in the aquifer.

•  It is possible to predict how far the plume might ex-
  tend if the pump-and-treat system was not operating,
   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. http7Awww.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.
                                                      123

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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 for microbial 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
                                                 124

-------
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
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


Not identified

Cecil Field
NAS, FL
     1-3




    4, 5



     6

     7
Weidemeider,
 this volume
 B. Wilson,
 this volume
     9

    10


    11

 Chapelle,
 this volume
  (meters)


 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

   Oto24
   0 to 40
   Oto55
   0 to 600
   0 to 140
                                                  (years)
 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
 0.0 to 1.2
 (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
   1.2
  0.8

3.3 to 7.3
                                                                        5.11
                                                         0.32

                                                         0.73
                                                         2.3

                                                         0.8
                          3.1
Produced
0.6
0.07
0.21
0.42
0.73
Produced
1.16
0.47
Produced
0.68
>0.73
             0.8

           3.3 to 7.3
                                                    125

-------
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) are similar to the rates
of removal of the chlorinated alkenes.
                                                       TCE Removal in Field

                                       5

                                      I-
                                       O
                                       S
                                       S.
                                       B^
                                      "- <=
                                                                   10  11  12  13  U 15  18  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
            Vinyl
cis-DCE      Chloride    1,1,1-TCA
                                                                         Apparent First-Order Loss (1/year)
                                (meters)     (years)       (years)           Apparent Zero OrderLoss (/ifir/t* day)
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
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
1.62
1.20
1.65
3.6
0.012

1.75
1.42


                                                     126

-------
References

 1. Semprini, L, P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson. An-
    aerobic Transformation of chlorinated aliphatic hydrocarbons in
    a sand aquifer  based on spatial chemical distributions. Water
    Resour. Res. 31(4):1051-1062.

 2. Weaver, J.W., J.T. Wilson, D.H. Kampbell, and M.E.  Randolph.
    1995. Field derived transformation rates for modeling natural
    bioattenuation of trichloroethene and its degradation products. In:
    Proceedings: Next Generation Environmental Models and Com-
    putational Methods, August 7-9, Bay City, Ml.

 3. Wilson, J.T, J.W. Weaver, D.H. Kampbell. 1994. Intrinsic biore-
    mediation of TCE in ground water at an NPL site in St. Joseph,
    Michigan. In: U.S. EPA. Symposium on  Natural  Attenuation of
    Ground Water, Denver, CO, August 30-September 1. EPA/600/R-
    94/162. pp. 116-119.

 4. Ehlke, T.A., B.H. Wilson, J.T. Wilson, and T.E. Imbrigiotta. 1994.
    In-situ biotransformation of trichloroethylene and cis-1,2-dichlo-
    roethylene at  Picatinny Arsenal, New Jersey.  In:  Morganwalp,
    D.W., and D.A. Aronson, eds. Proceedings of the U.S. Geological
    Survey Toxic Substances Hydrology Program, Colorado Springs,
    Colorado, September 20-24, 1993. Water Resources Investiga-
    tions Report 94-4014. In press.

 5. Martin, M., and  T.E. Imbrigiotta. 1994. Contamination of ground
    water with trichloroethylene  at  the Building 24 site at Picatinny
    Arsenal, New Jersey. In:  U.S.  EPA. Symposium  on Natural At-
    tenuation of Ground Water.  Denver, CO, August 30-September
    1. EPA/600/R-94/162. pp. 109-115.

 6. Cox, E., E.  Edwards, L. Lehmicke, and D. Major. 1995. Intrinsic
    biodegradation of trichloroethylene and trichloroethane in a se-
    quential anaerobic-aerobic aquifer. In: Hinchee,  R.E., J.T. Wilson,
    and D.C. Downey, eds. Intrinsic bioremediation. Columbus, OH:
    Battelle Press, pp. 223-231.

 7. Lee, M.D., P.P. Mazierski,  R.J. Buchanan, Jr., D.E. Ellis, and L.S.
    Sehayek. 1995.  Intrinsic and in situ anaerobic biodegradation of
    chlorinated solvents at an industrial landfill. In: Hinchee, R.E., J.T.
    Wilson, and D.C. Downey, eds. Intrinsic bioremediation. Colum-
    bus, OH: Battelle Press, pp. 205-222.

 8. Buscheck, T, and K. O'Reilly. 1996. Intrinsic anaerobic biodegra-
    dation of chlorinated solvents at a manufacturing plant. Abstract
    presented at the Conference on Intrinsic Remediation  of Chlorin-
    ated Solvents, Salt Lake City, UT, April 2. Columbus, OH: Battelle
    Memorial Institute.

 9. Benker, E., G.B. Davis, S.  Appleyard, D.A. Berry, and T.R. Power.
    1994. Groundwater contamination by trichloroethene (TCE) in a
    residential area of Perth: Distribution, mobility, and implications for
    management. In: Proceedings of the Water Down Under 94, 25th
    Congress of IAH, Adelaide, South Australia, November 21-25.
10. Gorder,  K.A., R.R.  Dupont,  D.L.  Sorensen, and M.W.  Kem-
    blowski.  1996. Intrinsic remediation of TCE in cold regions. Ab-
    stract presented at the Conference on Intrinsic Remediation of
    Chlorinated Solvents, Salt Lake City, UT, April 2. Columbus, OH:
    Battelle Memorial Institute.

11. De, A., and D. Graves. 1996. Intrinsic bioremediation of chlorin-
    ated  aliphatics  and aromatics at a complex industrial site. Ab-
    stract presented at the Conference on Intrinsic Remediation of
    Chlorinated Solvents, Salt Lake City, UT, April 2. Columbus, OH:
    Battelle Memorial Institute.

12. Ehlke, T.A., T.E. Imbrigiotta, B.H. Wilson, and J.T. Wilson.  1991.
    Biotransformation of cis-1,2-dichloroethylene in aquifer material
    from  Picatinny Arsenal, Morris County, New Jersey. In: U.S. Geo-
    logical Survey Toxic Substances Hydrology Program—Proceed-
    ings of the Technical Meeting, Monterey, CA, March 11-15.  Water
    Resources Investigations Report 91 -4034. pp. 689-697.

13. Wilson, B.H., T.A.  Ehlke, T.E. Imbigiotta, and J.T. Wilson.  1991.
    Reductive dechlorination of trichloroethylene in anoxic aquifer
    material from Picatinny Arsenal, New Jersey. In: U.S. Geological
    Survey Toxic Substances Hydrology Program—Proceedings of
    the Technical Meeting, Monterey, CA, March 11-15. Water Re-
    sources Investigations Report 91 -4034. pp. 704-707.

14. Haston, Z.C., P.K. Sharma, J.N.P. Black, and P.L. McCarty.  1994.
    Enhanced reductive dechlorination of chlorinated  ethenes.  In:
    U.S.  EPA. Proceedings of the EPA Symposium on Bioremediation
    of Hazardous Wastes: Research, Development, and Field Evalu-
    ations. EPA/600/R-94/075. pp. 11-14.

15. Wilson, B.H., J.T. Wilson, D.H. Kampbell, B.E. Bledsoe, and J.M.
    Armstrong.  1990. Biotransformation of monoaromatic and chlo-
    rinated hydrocarbons at an aviation gasoline spill site. Geomicro-
    biol. J. 8:225-240.

16. Parsons,  F., G. Barrio Lage, and R. Rice. 1985. Biotransformation
    of  chlorinated organic solvents in static microcosms. Environ.
    Toxicol. Chem. 4:739-742.

17. Davis, J.W., and C.L. Carpenter.  1990.  Aerobic biodegradation
    of vinyl chloride in groundwater samples. Appl. Environ. Microbiol.
    56(12):3878-3880.

18. Klecka, G.M., S.J.  Gonsior, and D.A. Markham. 1990. Biological
    transformations of 1,1,1-trichloroethane in subsurface  soils and
    ground water. Environ. Toxicol. Chem. 9:1437-1451.

19. Barrio-Lage, G.A., F.Z. Parsons, R.M. Narbaitz, and PA. Lorenzo.
    1990. Enhanced anaerobic biodegradation of vinyl chloride in
    ground water. Environ. Toxicol. Chem. 9:403-415.
                                                              127

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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 as trichloroethylene, are co-metabolic and
yield no  advantage to  the  bacteria that catalyze the
reactions. In fact, co-metabolism can select against the
organism because of the wasting of energy and production
of toxic metabolites.

Between the extremes of readily degradable 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
                                                  128

-------
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
                                                  129

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community or enrichment, which supports the premise
that dehalogenation provides a  selective advantage to
the organisms involved. The results also suggest that a
wide range of different bacteria have the ability to deha-
logenate PCB. The electron donors for the dehalogena-
tion  in sediment  are   unknown.  The  addition  of
exogenous carbon sources does not stimulate the reac-
tion. In contrast, "priming" the mixtures with low levels
of bromobiphenyl or  specific isomers of tetrachloro-
biphenyl (15) seems to selectively enrich a population
of PCB-dechlorinating bacteria and dramatically stimu-
late the dechlorination of the other congeners.

The lower chlorinated  PCB congeners, whether part of
the original Arochlor mixture or  derived from reductive
dehalogenation, are biodegraded by aerobic bacteria
(16). The initial attack is catalyzed by a 2,3- or 3,4-di-
oxygenase, followed by a sequence of reactions  that
lead to ring cleavage and the  accumulation of chlo-
robenzoates which are readily degraded by a variety of
bacteria. The enzymes that oxidize PCB are produced
by bacteria grown on biphenyl, and adding biphenyl to
slurry-phase reactors stimulates the growth and activity
of PCB degraders. Such stimulation has been shown to
be effective in the field (17). There is also good evidence
that aerobic PCB degradation is taking place in contami-
nated river sediments (18).

Clearly, reductive dechlorination is ongoing at a wide
range of PCB contaminated sites. The strategy of an-
aerobic dehalogenation followed by aerobic degradation
seems to be particularly effective with PCB whether in
an engineered system or in natural systems (e.g., during
resuspension of anaerobic sediments). To date the com-
plete  biodegradation  of  PCB is  slow and difficult to
predict or control in the field. Several new strategies,
including construction of novel strains, may increase the
potential for effective PCB biodegradation.

Chloroaliphatic Compounds

Several good reviews have recently appeared on the
biodegradation  of small  (one-  and two-carbon)  chlo-
roaliphatic  compounds (19-21); therefore,  this paper
briefly mentions only some aspects that might otherwise
be overlooked. Among the one- and two-carbon chlorin-
ated compounds, the more highly chlorinated molecules
are subject to reductive dehalogenation under a variety
of conditions. Thus, carbon tetrachloride can be se-
quentially reduced to chloroform and dichloromethane.
Similarly, perchloroethylene can be reduced to ethylene
via trichloroethylene, dichloroethylene, and vinyl chlo-
ride. The degradation of chloroethylenes is discussed in
considerable detail by Gossett and Zinder (this volume).
Most of the work to date has focused on mixed microbial
cultures that use chlorinated solvents fortuitously as
electron acceptors. Such  activity is very  widely distrib-
uted in anaerobic ecosystems and catalyzes the slow and
often partial reduction of chlorinated contaminants. In
contrast, some microbial communities and a few iso-
lated strains can derive energy from the use of chlorin-
ated  compounds  as   terminal  electron  acceptors
(Gossett and Zinder, this volume). Such processes are
much faster than the co-metabolic processes because
they provide a selective advantage for the bacteria and
are self-sustaining.

Several Chloroaliphatic compounds can serve as growth
substrates for aerobic bacteria. The more chlorinated
compounds such as trichloroethylene and chloroform do
not provide energy and  carbon for aerobic growth, al-
though they can be degraded co-metabolically.  In con-
trast, methylene chloride can support the growth of both
anaerobes  (22) and aerobes  (20). 1,2-Dichloroethane
(23) and vinyl  chloride  (20)  similarly can be  readily
degraded by aerobic bacteria. Any of these compounds
that serve as growth substrates would be excellent can-
didates for natural attenuation where oxygen is present.
Aerobic mineralization of the related molecule, ethylene
dibromide, has been reported  in soil,     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),
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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-fert-butyl Ether

Gasoline oxygenates such as ethanol, methyl-terf-butyl
ether (MTBE), and ferf-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  fe/t-butyl alcohol is
much more readily degradable than MTBE under aerobic
or anaerobic conditions.

Nitrate Esters

A variety of nitrate esters, including glycerol trinitrate,
pentaerythritol  tetranitrate, and nitrocellulose, have
been used extensively as explosives. Recent studies
indicate that the  nitrate esters can be  degraded  by
bacteria from a variety of sources  (29, 30). Bacterial
metabolism releases nitrite, which can serve as a nitro-
gen  source and  yield  a  selective  advantage for the
organisms. The  biodegradation of  nitrate  esters  has
only recently  been  studied extensively, and  little  is
known about degradation  in the environment. Recent
laboratory results strongly suggest that natural attenu-
ation is possible,  but more  information  is needed on the
bioavailability, toxicity, and kinetics  of the process.
1  Weaver, J. 1996. Personal communication with the author.

2  Cowan, R. 1996. Personal communication with the author.
Pesticides

Most pesticides used in the past 20 years in the United
States have been formulated to degrade in the environ-
ment, and a considerable amount of information is avail-
able on degradation kinetics in soil and water. The U.S.
Environmental  Protection Agency  Risk Reduction Engi-
neering Laboratory in Cincinnati, Ohio, has developed
an extensive Pesticide Treatability Database  that con-
tains information on a variety of compounds. Many pes-
ticides hydrolyze and yield compounds that  serve as
growth  substrates  for  bacteria.  For example,   car-
bamates, chlorophenoxyacetates, dinitrocresol,  cou-
maphos, atrazines, and some organophosphates serve
as growth substrates for bacteria and would be good
candidates  for natural attenuation.  A variety of other
pesticides are hydrolyzed by extracellular enzymes de-
rived from soil bacteria but provide no advantage to the
organisms that produce the enzymes. Similarly, some of
the organohalogen insecticides can be reductively de-
halogenated but provide no advantage to specific organ-
isms. Their biodegradation rates are proportional to the
biomass and activity  in the soil.

Conclusion

To date, the focus of natural  attenuation has been on
hydrocarbon fuels and chlorinated aliphatic solvents. A
wide range of synthetic chemicals released in the envi-
ronment are known to be biodegradable by bacteria, and
much is known about the processes and their require-
ments.  The potential for natural attenuation  of biode-
gradable contaminants  should be considered  before
more costly and disruptive treatment  options.

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              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
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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.
                                                    134

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Poster Session

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             Degradation of Chloroform Under Anaerobic Soil Conditions
                                       Frances Y. Saunders
   National Council of the Paper Industry for Air and Stream Improvement, Gainesville, Florida

                                            Van Maltby
   National Council of the Paper Industry for Air and Stream Improvement, Western Michigan
                                 University, Kalamazoo, Michigan
This study was designed to determine the rate of chlo-
roform biodegradation under anaerobic conditions. Sub-
surface soil samples were taken from leachate plumes
downgradient from two different bleached kraft mill land-
fills chosen to represent a northern climate soil  and a
southern climate soil. Anaerobic subsurface soil  condi-
tions were modeled by assembly of the sample soils into
microcosms, with  care taken throughout the study to
ensure that the soils and microcosms were maintained
and handled anaerobically. Following assembly, the mi-
crocosms were spiked with chloroform at either a 10,60,
or 160 micrograms per liter spike level. (Spiked steril-
ized soils and unspiked soil blanks  were included as
controls.) The microcosms were incubated at the year-
round  average ambient soil temperature and were ana-
lyzed  for chloroform  over a  period up to  64 weeks
following their preparation.
Data from these experiments showed  that  in accord-
ance with the literature, chloroform degraded at a rapid
rate under anaerobic conditions. For the southern site
microcosms, an 8-week adaptation period was noted,
followed by rapid degradation (ty2 = 4-16 weeks). For the
northern site soil microcosms, the chloroform concen-
tration was reduced to 5 percent of the initial concentra-
tion in a total of 4 weeks or less (ty2 = 0.4 to 3 weeks),
with no adaptation period noted. The absence of chlo-
roform degradation in sterilized control microcosms and
the absence of degradation intermediates (methylene
chloride and chloromethane) suggest  that chloroform
was degraded completely by a microbial pathway. The
data generated in these experiments were incorporated
into an attenuation fate-and-transport model for organic
substrates in subsurface  soils.  This  model demon-
strated that at the rates determined in this study and at
the most conservative rate estimates, several orders of
magnitude higher, the biodegradation process is a sig-
nificant factor in  the  rapid removal of organics from
subsurface soils. Modeling runs resulted in receptor well
concentrations for chloroform that were predominantly
orders of magnitude below current analytical capabilities.
                                                137

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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.
                                                 138

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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
                              McMaster University, Hamilton, Ontario

                                        Richard A. Mechaber
                         GEI Consultants, Inc., Concord, New Hampshire

                                           Benjamin Y. Su
                        GEI Consultants, Inc., Winchester, Massachusetts
Tetrachloroethene (PCE) and trichloroethene (TCE) are
being biodegraded under naturally occurring sequential
anaerobic/co-metabolic conditions in ground water at an
inactive landfill in New Hampshire. Ground water in the
vicinity of the landfill is predominantly aerobic, with the
exception of an anaerobic zone that has developed at
the landfill source  area  where significant  historical
biodegradation of dichloromethane, ketones, and 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.
                                                 139

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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 per liter (mg/L), respectively.
Determination of  precision  and accuracy for a 19.8
mg/L methane prepared  sample using  six replicates
was a  standard  deviation  of 0.6 mg/L, risk-specific
dose (RSD)  = 3.2 percent, and  average recovery  of
87 percent. Similar statistics for 118 mg/L ethylene 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.
                                                140

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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 -0.05 wk"1. Biodegradation of cis-DCE was
most  rapid in  soils underlying a peat  layer near the
plume discharge  point.

Degradation of TCE in situ also was estimated using the
concentrations of chlorinated ethenes determined for a
series of monitoring wells along the major ground-water
flow path within the plume. Chlorinated ethene 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.
                                                 141

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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.
                                                  142

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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.
                                               143

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 Evidence of Natural Attenuation of Chlorinated Organics at Ft. McCoy, Wisconsin
                                            Jason Martin
                   Rust Environment and Infrastructure, Sheboygan, Wisconsin
Ft. McCoy is a Resource Conservation and Recovery
Act regulated  U.S. Army facility located  in western
Wisconsin. Fire  Training Burn Pit 1  (FTBP1) on the
site was operated from approximately 1973 to 1987.
Operations at the pit consisted of filling the pit with a
layer of water and fuel, then repeatedly igniting and
extinguishing the  contents  until the fuel  was  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.
                                                  144

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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.
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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 near the
source area at  maximum concentrations of 130 and 12
micrograms per liter, respectively, then decreased to near
or below the detection level at the site boundary.  Low
levels of the nonchlorinated product,  ethylene, were 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.
                                                 146

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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.
                                               147

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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.
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                   Intrinsic Remediation of Chlorinated Solvents as an
                                Effective Containment Strategy
                                 Ronald Hughen and Randall Hicks
                          BDM Environmental, Albuquerque, New Mexico

                                              Leon Grain
                             BDM Environmental, Fair Oaks, California
At a site in California, a tetrachloroethene (PCE) plume
over 1,300 feet long was discovered within the capture
zone  of two  municipal supply wells. Fortunately,  the
plume was restricted to an upper water-bearing unit that
was separated from the  drinking water aquifer by a
continuous clay zone. Nevertheless, the concentrations
of PCE in this shallow zone exceeded 1,000 parts per
billion (ppb) and represented a potential threat to human
health and the environment. After 3 years of investiga-
tion, pilot-scale air sparging in the release areas, and
significant regulatory negotiation, a full-scale air sparging
system was installed and operated for 2 years. PCE con-
centrations within the source areas declined by an order
of magnitude  during this time. Despite operation of the
remedial system, PCE concentrations in the source areas
stabilized at 100 ppb, 20 times higher than the closure
criteria specified in the state  cleanup order.  Remedial
efforts outside the source areas were not required by the
approved plan unless  PCE concentrations rose above
unacceptable levels, signaling plume migration.
Outside the source area "hot spots," PCE concentrations
remained constant over the 5-year period at concentration
levels ranging from 20 to 50 ppb;  these monitoring data
demonstrated  that the plume  was not expanding. The
stability of the plume and the documented inefficiency of
the pump-and-treat/air-sparging remedial system permit-
ted establishment of a risk-based plume management plan
that called upon institutional controls rather than hydraulic
manipulation and ground-water treatment. The contain-
ment  strategy  was  permitted  based on  the  empirical
evidence of 5 years  of ground-water monitoring and the
acceptance by the regulatory agency that intrinsic reme-
diation was active at the site.

Based on 6 years of data monitoring concentrations of
halogenated solvents  (HVOC), pH, conductivity, tem-
perature, turbidity, dissolved oxygen, and salinity, intrin-
sic  remediation now appears to be occurring. In June
1995, the first chemical samples were obtained to spe-
cifically document intrinsic remediation processes  oc-
curring at this  site (HVOC, total hydrocarbons, volatile
hydrocarbons,  dissolved oxygen, nitrate, sulfate, meth-
ane, ethane, ethene, redox potential, pH, temperature,
conductivity, and chloride). These data were used to
validate our hypothesis of effective intrinsic remediation
at this site, using the protocol developed by the U.S. Air
Force Center for Environmental Excellence for fuel  hy-
drocarbon intrinsic remediation as a guide but  modeling
for the important parameters of chlorinated solvent in-
trinsic attenuation/remediation.

The purpose of this poster is to present another case
history of intrinsic remediation of chlorinated  solvents.
We maintain that case histories such  as this will en-
hance the acceptance of intrinsic remediation as an
effective containment  strategy, obviating  the  need for
extensive regulatory negotiations and, possibly, opera-
tion of mechanical remedial systems at source areas.
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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.
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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 -DCA and 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, CP, 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.
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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
(ng/L); by 1994, the concentration ranged from 11,000 |o.g/L
to BDL. The RDX concentration ranged from 7,600 to 33.8
Hg/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 (QMS)—which is used in the modeling  of the
site, also will be discussed and  illustrated.
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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 for treatment
of ground  water was signed by the state of California
and the Army on September 8,1995—the first approved
ROD in the United States selecting natural attenuation as
a primary  remedial alternative for trichloroethene (TCE)
and explosives in ground water. The natural attenuation
alternative  consists of institutional controls to eliminate
future use of ground water in the area surrounding the
site, long-term ground-water monitoring, and evaluation
of contaminant migration and degradation rates.
Explosives and volatile organic compounds (VOCs) are
present  in  shallow ground water over a 26-acre area of
Sierra Army Depot, which is located approximately 50
miles  northwest of Reno, Nevada. No surface water
features or water supply wells exist within the area of
the site.  A  ground-water plume of explosive compounds
originates  from the TNT Leaching Beds, a facility used
during the  1940s for percolation of  waste  water from a
shell washout facility. Dissolved explosive compounds
in the ground water include RDX, 1,3,5-trinitrobenzene,
HMX, and  minor concentrations of numerous other 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.
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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).
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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.
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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 «d (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 K^ 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.
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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.
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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.
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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.
                                                 159

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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  [|J.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 ng/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 NOa/N, total and dissolved iron, and SO^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.
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           Using Evidence of Natural Attenuation To Locate the Source of a
                      Chlorinated Volatile Organic Compound Plume
John M. Armstrong, John J. D'Addona, Charles W. Dittmar II, Greg M. Tatara, and Joel W. Parker
                          The Traverse Group, Inc., Ann Arbor, Michigan
An upper Midwest manufacturing plant has been the site
of recent subsurface investigations because of past mis-
handling practices associated with degreasing solvents,
namely trichloroethene (TCE) and 1,1,1-trichloroethane
(TCA), during the early 1970s. The site is situated on a
well-graded sandy silt aquifer, with limestone bedrock
located from 30 to 45 feet below grade.

Initial  investigations focused on a solvent storage area
adjacent to the building. Results of these investigations
revealed significant concentrations of TCE and TCA
breakdown  products—cis-1,2-dichloroethene   (cis-
DCE), 1,1 -dicloroethane (1,1 -DCA), and vinyl chloride—
ranging in concentration from 100 to 12,000 micrograms
per liter. Since this was the only known source area for
these  chemicals, the absence of parent compounds was
puzzling, especially given the low hydraulic conductivity
of the overburden aquifer (10~5 centimeters per second
range). Ground water was tested for the general water
quality parameters of hardness, sulfate, chemical oxy-
gen demand, phosphate, and nitrates. In addition, meth-
ane, ethane, and ethene were analyzed in the ground
water. These data revealed that in areas of high break-
down product concentrations, there were corresponding
decreases in sulfate, nitrate, and phosphate concentra-
tions and increases in the formation of byproduct gases.
Conversely, in  uncontaminated  areas,  sulfate (greater
than or  equal  to 200 milligrams per liter) and  nitrate
(greater than or equal to 10 milligrams per liter) were
present  and the gases were absent. This evidence of
natural attenuation did not explain the absence of parent
compounds. Contouring concentrations of electron do-
nors and acceptors, nutrients, and breakdown products,
combined  with ground-water  contour overlays and
plume  prediction  models,  indicated that the  actual
source of  contamination may  be under  the building.
Latest investigations resulted in isolating  source areas
from unknown  solvent disposal areas through a series
of borings inside the plant.
To the best of our knowledge, this represents one of the
first instances in which evidence of natural attenuation,
instead of historical information, has been used to locate
the source of contamination.
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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-water sampling, in situ
measurement of ground-water field parameters, a well
search, and an evaluation of potential receptor impacts
through modeling. The  results of these investigations
suggested that "steady state" conditions of contaminant
influx and attenuation were in effect, and that, even in
the absence of the source remediation conducted,  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.
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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.
                                                  164

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                  Natural Attenuation as Remedial Action: A Case Study
                                          Andrea Putscher
                           Camp Dresser & McKee, Wood bury, 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.

A lighting 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.
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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.
                                               166

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                 Applying Natural Attenuation of Chlorinated Organics
        in Conjunction With Ground-Water Extraction for Aquifer Restoration
                              W. Lance Turley and Andrew Rawnsley
                        Hull & Associates Engineering, Inc., Austin, Texas
Natural attenuation of dissolved chlorinated organics
(primarily tetrachloroethene, trichloroethene, and 1,1,1-
trichloroethane) via dilution is being successfully 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 near the end of the plume
was not practical, however, because of property access
difficulties, the presence of a flood plain, and anticipated
problems in conveying extracted water via a forcemain
due to expected low flows and a large head differential.

Field measurements were supported by finite-difference,
three-dimensional flow modeling and indicated that the
aquifer at the end of the plume discharges into the 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.
                                                167

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      Natural Attenuation of Chlorinated Organics in Ground Water Based on
     Studies Conducted at Naval Amphibious Base Little Creek Sites 12 and 13
                                           Scott Park
     U.S. Navy, Atlantic Division of Naval Facilities Engineering Command, Norfolk, Virginia

                                            Nitin Apte
               Foster Wheeler Environmental Corporation, Lyndhurst, New Jersey
The Department of Defense (DOD) designed the U.S.
Navy's Installation Restoration Program (IRP) to inves-
tigate  past disposal sites according to the Comprehen-
sive  Environmental  Response, Compensation,  and
Liability Act of 1983  and  the Superfund Amendments
and Reauthorization  Act  of 1986.  The IRP has been
underway at Naval Amphibious Base (NAB) Little Creek
since  1984. As part  of the program, multiple ground-
water  investigations  have been conducted at several
sites between 1986  and  1996. Chlorinated  organics,
namely  trichloroethylene (TCE),   perchloroethylene
(PCE), and pentachlorophenol (PCP),  have  been the
major  constituents of concern at the following sites:

• Site 12—Exchange Laundry Waste  Disposal Area:
  This site  consists  of an area surrounding a former
  storm drain used for disposal of soaps, sizing agents,
  dyes, and PCE sludges from a laundry  operation
  between 1973 and  1978. A sewer line, which received
  dry  cleaning waste from the former  laundry facility,
  drained to a canal that eventually  flows into Little
  Creek Cove. Remains of the laundry facility and the
  sewer line were removed by 1992. A new commissary
  building,  covering  a portion of  the  site, was con-
  structed in 1993.  Ground-water studies conducted
  from 1992 have indicated volatile organic compound
  levels as high as 18,200 parts per billion (ppb), mainly
  consisting of 1,2-dichloroethene, TCE, and PCE.

• Site 13—Public Works Dip  Tank and Wash Rack:
  This site consists of an area surrounding a former dip
  tank used to treat wood with PCP. The tank reportedly
  contained 300 to 400  gallons of PCP during its use
  from  early 1960s to 1974. All wood-treating opera-
  tions were discontinued, and the equipment was dis-
  mantled by 1982.  Ground-water and soil sampling
  has indicated PCP at levels as high as 890,000 ppb
  in subsurface soil and 1,700 ppb in ground water.

Slug tests have been conducted at both sites to charac-
terize the hydrogeology, and the  plumes  have  been
delineated by sufficient perimeter sampling. In addition,
step tests and an 8-hour pump test have been con-
ducted  at Site 12. The depth of the water table aquifer
at these sites is between 20 and 24 feet below ground
surface (bgs).

Remedial alternatives at these  sites, including natural
attenuation, will be evaluated to mitigate  the human
health and ecological risks as  well as the impact on
nearby surface water. The remedy selection process is
expected to be complete by the end of 1996. Although
no specific remedy has  been selected, data collected
over a  9-year span allow evaluation of natural attenu-
ation occurring at these  sites. This presentation identi-
fies trends and makes projections for future attenuation
periods. The two sites present an  opportunity to com-
pare natural attenuation of relatively mobile and vola-
tile compounds (TCE and PCE) with that of immobile
and semivolatile compounds (PCP) in almost identical
hydrogeological settings.
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           A Modular Computer Model for Simulating Natural Attenuation of
               Chlorinated Organics in Saturated Ground-Water Aquifers
                               Yunwei Sun and James N. Petersen
     Chemical Engineering Department, Washington State University, Pullman, Washington

                            T. Prabhakar Clement and Brian S. Hooker
                  Pacific Northwest National Laboratory, Richland, Washington
Although several field-scale natural attenuation projects
have already been considered for managing benzene,
toluene, ethylbenzene, and xylene (BTEX) plumes, a
rational basis for implementing natural attenuation 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.
                                                169

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