f/EPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington, DC 20460
EPA/600/R-94/162
September 1994
Symposium on
Natural Attenuation of
Ground Water
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EPA/600/R-94/162
September 1994
Symposium on Natural Attenuation of Ground Water
U.S. Environmental Protection Agency
Office of Research and Development
Washington, DC 20460
Printed on paper that contains at least
50 percent recycled fiber.
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Notice
The information in this document has been reviewed in accordance with the EPA's peer and
administrative review policies and 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
Executive Summary 1
State and Federal Issues Impacting Natural Attenuation 5
State and Federal Issues Impacting Natural Attenuation: Meeting Overview 6
Panel Discussion on Regulatory Issues 9
Symposium on Natural Attenuation of Ground Water: Summaries and Abstracts 13
Regulatory Perspective on Natural Attenuation 14
Department of Defense Perspective 14
Jim Owendoff
U.S. EPA Regulatory Perspective: Natural Attenuation as a Remedial Option for 14
Contaminated Ground Water
Guy A. Tomassoni
State Perspective 15
John Shauver
Overview of Investigations Related to Natural Attenuation 16
USGS Perspective 16
Gail E. Mallard
U.S. EPA Perspective 16
Fran V. Kremer
Overview of Process and Procedures 19
Site Characterization: What Should We Measure, Where (When?), and How? 20
Michael J. Barcelona
Processes Controlling the Distribution of Oil, Air, and Water 26
John L Wilson
New Tools To Locate and Characterize Oil Spills in Aquifers 34
Bruce J. Nielsen
Microbiological and Geochemical Degradation Processes 40
E. Michael Godsy
Field and Laboratory Results: Getting the Whole Picture 44
Mary Jo Baedecker
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Contents (Continued)
Page
Case Studies: Public Service's Site 47
In S/fry Bioremediation of the Seventh Avenue Site in Denver: The Remediation of 48
Soils and Ground Waters
Christopher Nelson
The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through In Situ 50
Bioremediation: The Role of Mathematical Models
Tissa H. Illangasekare, David C. Szlag, and John T. Wilson
The Importance of Knowledge About Intrinsic Bioremediation for Cost-Effective 55
Site Closure: The Client's Perspective
Harry E. Moseley
A Regulators' Perspective of a Bioremediation Site 57
Lisa C. Weers
The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through In Situ 58
Bioremediation: The Regulator's Perspective
Mark E. Walker and Lisa C. Weers
Case Studies: Sites Where Natural Attenuation Has Been Documented 59
Intrinsic Bioremediation of JP-4 Jet Fuel 60
John T. Wilson, Fredrick M. Pfeffer, James W Weaver, Don H. Kampbell,
Todd H. Wiedemeier, Jerry E. Hansen, and Ross N. Miller
A Natural Gradient Tracer Experiment in a Heterogeneous Aquifer With Measured In Situ 68
Biodegradation Rates: A Case for Natural Attenuation
Thomas B. Stauffer, Christopher P. Antworth, J. Mark Boggs, and William G. Maclntyre
Traverse City: Distribution of the Avgas Spill 75
David W. Ostendorf
Traverse City: Geochemistry and Intrinsic Bioremediation of BTX Compounds 80
Barbara H. Wilson, John T. Wilson, Don H. Kampbell, Bert E. Bledsoe, and John M. Armstrong
Mathematical Modeling of Intrinsic Bioremediation at Field Sites 85
Hanadi S. Rifai
Biogeochemical Processes in an Aquifer Contaminated by Crude Oil: An Overview of 89
Studies at the Bemidji, Minnesota, Research Site
Robert P. Eganhouse, Mary Jo Baedecker, and Isabelle M. Cozzarelli
Simulation of Flow and Transport Processes at the Bemidji, Minnesota, Crude-Oil Spill Site 95
Hedeff I. Essaid
IV
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Contents (Continued)
Page
Natural Attenuation of Trichloroethylene and Similar Compounds: Case Studies 103
An Overview of Anaerobic Transformation of Chlorinated Solvents 104
Perry L McCarty
Contamination of Ground Water With Trichloroethylene at the Building 24 Site at 109
Picatinny Arsenal, New Jersey
Mary Martin and Thomas E. Imbrogiotta
Intrinsic Bioremediation of TCE in Ground Water at an NPL Site in St. Joseph, Michigan 116
John T. Wilson, James W. Weaver, and Don H. Kampbell
Posters 121
Technical Protocol for Implementing the Intrinsic Remediation with Long-Term Monitoring 122
Option for Natural Attenuation of Fuel Hydrocarbon Contamination in Ground Water
Todd H. Wiedemeier
Wisconsin's Guidance on Naturally Occurring Biodegradation as a Remedial Action Option 127
Michael J. Barden
Assessing the Efficiency of Intrinsic Bioremediation 128
Francis H. Chapelle
A Practical Approach To Evaluating Natural Attenuation of Contaminants in Ground Water 129
P.M. McAllister and C.Y. Chiang
The Use of Low Level Activities To Assist Intrinsic Bioremediation 130
Robert D. Norris, Jeffrey C. Dey, and Daniel P. Shine
Natural Attenuation of Jet Fuel in Ground Water 132
G. Doyle, D. Graves, and K. Brown
Evaluation of Intrinsic Bioremediation at an Underground Storage Tank Site in Northern Utah ... 133
R. Ryan Dupont, Darwin L Sorensen, and Marion Kemblowski
Case Studies of Field Sites To Demonstrate Natural Attenuation of BTEX Compounds in 135
Ground Water
C. Y. Chiang and P.M. McAllister
Demonstrating Intrinsic Bioremediation of BTEX at a Natural Gas Plant 136
Keith Piontek, Tom Sale, Jake Gallegos, Steve de Albuquerque, and John Cruze
Demonstrating the Feasibility of Intrinsic Bioremediation at a Former Manufactured Gas Plant ... 137
Ian D. MacFarlane, Edward J. Bouwer, and Patricia J.S. Colberg
Natural and Enhanced Bioremediation of Aromatic Hydrocarbons at 139
Seal Beach, California: Laboratory and Field Investigations
Harold A. Ball, Gary D. Hopkins, Eva Orwin, and Martin Rein hard
The Complete Dechlorination of Trichloroethene to Ethene Under Natural Conditions in a 142
Shallow Bedrock Aquifer Located in New York State
David Major, Evan Cox, Elizabeth Edwards, and Paul W. Hare
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Contents (Continued)
Page
Appendix A: Regulatory Meeting Participants 145
VI
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Executive Summary
Introduction
The U.S. Environmental Protection Agency's (EPA's)
Biosystems Technology Development Program and the
U.S. Geological Survey (USGS), with sponsorship from
the U.S. Air Force (USAF), coordinated a meeting on the
Natural Attenuation of Ground Water at the Hyatt Re-
gency in Denver, Colorado, on August 30 through Sep-
tember 1, 1994. The purpose of this meeting was to
bring together interested members of the bioremediation
community, including state and federal regulators, aca-
demicians, industry, and remedial contractors, to share
regulatory and field perspectives, present laboratory
and field research, and discuss issues related to the
successful implementation and evaluation of natural at-
tenuation. Fran V. Kremer, EPA, Cincinnati, Ohio; John
T Wilson, EPA, Ada, Oklahoma; and Gail E. Mallard,
USGS, Reston, Virginia, served as co-organizers of the
symposium, along with support from Lieutenant Colonel
Ross N. Miller, Headquarters, USAF, Center for Environ-
mental Excellence, Brooks Air Force Base, Texas.
In conjunction with this symposium, a smaller meeting
on State and Federal Issues Impacting Natural Attenu-
ation was held on Monday, August 29, at the same
location. The results of the regulatory meeting were
presented at the symposium and are reported in a sepa-
rate section of this proceedings.
More than 420 participants attended the 3-day sympo-
sium which included over 25 oral presentations, an over-
view of Monday's regulatory meeting and a followup
panel discussion, and a poster session with 12 presen-
tations. The oral presentations were divided into five
topic areas: a Regulatory Perspective on Natural Attenu-
ation, an Overview of Investigations Related to Natu-
ral Attenuation, an Overview of Processes and
Procedures, and Case Studies on the Public Service
Company's Site, Sites Where Natural Attenuation Has
Been Documented, and Natural Attenuation of Trichlo-
roethylene and Similar Compounds.
Regulatory Perspective on Natural
Attenuation
Day One of the symposium opened with the regulatory
perspective session, chaired by Fran Kremer, EPA's
Biosystems Technology Development Program. Colonel
Jim Owendoff, U.S. Department of Defense (DOD), of-
fered DOD's perspective as a facilities owner. With over
10,000 sites where biological and nonbiological cleanup
technologies are being considered or implemented,
DOD has attempted to adopt technologies that eliminate
contaminants from media rather than transferring them
between media. Bioremediation has emerged as an
attractive remedial option. Through several activities,
DOD has sought to engender greater participation and
communication among regulatory, public, and industrial
communities in carrying out remedial efforts.
Guy A. Tomassoni, EPA Office of Solid Waste, presented
an overview of the Superfund National Contingency
Plan's (NCP's) decision criteria for natural attenuation
and discussed some of the issues and obstacles con-
cerning the acceptance of natural attenuation as an
appropriate remediation technology, primarily the issue
of reasonable remedial time frame. A state perspective
was presented by John Shauver, Michigan Department
of Natural Resources. Mr. Shauver emphasized that
bioremediation, naturally occurring or otherwise, should
not be considered a "no action" remedial option. He
spoke of the increasing complexity and stringency of
regulations over the last 10 years and discussed the
emergence of bioremediation as a viable tool to meet
these challenges. He then outlined the steps that regu-
lators look for in evaluating the potential efficacy of a
bioremediation proposal, specifically performing ade-
quate site characterization, providing detailed descrip-
tion of the remedial process, and implementing an
extensive site monitoring program.
Overview of Investigations Related to
Natural Attenuation
Gail E. Mallard, USGS, presented her organization's
perspective in the context of its unique role as an earth
science information center without regulatory or man-
agement responsibilities. In relation to research con-
ducted at a petroleum spill in New Jersey, Ms. Mallard
illustrated three points that USGS stresses regarding
bioremediation: performing site characterization, pre-
senting multiple lines of evidence, and gathering as
much information as possible. By covering all three
issues thoroughly, USGS was able to evaluate and work
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towards developing an optimal engineered solution for
the site.
Fran Kremer's presentation on research within EPA cov-
ered the Office of Research and Development's (ORD's)
mission to provide technical assistance to states and
regions while fulfilling longer term environmental and
regulatory needs. Dr. Kremer described the reorganiza-
tion of ORD to emphasize health risk management and
outlined EPA Administrator Carol Browner's 5-year stra-
tegic plan for the Agency stressing collaboration, long-
term research, and expansion of competitive grants.
Finally, she reviewed some recent EPA natural attenu-
ation projects including collaborative efforts with the
U.S. Air Force on developing a protocol for evaluating
the fate of dissolved hydrocarbons and with the Public
Service Company and Colorado Department of Health
to establish a methodology for evaluating possible risks
associated with benzene, toluene, ethylbenzene, and
xylene (BTEX) in soils following remediation.
Overview of Process and Procedures
Session Two: Overview of Process and Procedures was
chaired by Ross N. Miller, USAF, Brooks Air Force Base.
The opening speaker, Michael J. Barcelona, University
of Michigan, Ann Arbor, presented an overview of site
characterization, which sets the stage for evaluating the
progress of the natural transformation of contaminants.
In addition to providing evidence of parent compound
disappearance, active microbial populations with
biotransformation capabilities, and the appearance or
disappearance of organic and inorganic constituents,
site remediators need to demonstrate with quantitative
evidence the net removal of toxic compounds solely by
biological processes. To meet this need, Mr. Barcelona
described a process of identifying the mass of contami-
nants and total reactive carbon and estimating the net
removal/transformation of reactive compounds over
time.
John L. Wilson, Department of Geoscience, New Mex-
ico Institute of Mining and Technology, Socorro, New
Mexico, described the implications of the processes
controlling the distribution of oil, air, and water on natural
attenuation. Capillary, viscous, and buoyancy forces
control the movement and distribution of nonaqueous
phase liquids (NAPLs), a major source of aquifer con-
tamination. Natural attenuation at a NAPL contaminated
site is most effective in the downgradient aqueous
phase of the dissolved plume, because of mass transfer
limitations, toxicity concerns, and limited nutrient avail-
ability within the NAPL-contaminated zone. Dr. Wilson
contended that to evaluate the effectiveness of bioreme-
diation within the plume, however, it is necessary to
determine the distribution of the plume's source or
whether that source is moving.
Bruce J. Nielsen, USAF, Tyndall Air Force Base (AFB),
Florida, described new tools for locating and charac-
terizing oil spills in aquifers. Tri-Services (the Air Force,
Army, and Navy) is executing a program to demonstrate,
test, and evaluate the application of cone penetrometers
to support natural attenuation demonstrations. This tool,
consisting of a laser spectrometer with fiber optics inte-
grated with a cone penetrometer, can take a "snapshot"
of subsurface conditions showing the type and amount
of contamination, thereby helping to determine the best
method of remediation, monitor the effectiveness of
treatment, and assist technicians and engineers in judg-
ing when a site is clean. This technology has been field
tested at numerous Air Force, Army, and Navy installa-
tions, and in the future will undergo refinements to ex-
tend its sensitivities and expand its capabilities to other
types of contaminants.
E. Michael Godsy, USGS, discussed microbiological
and geochemical constraints to natural attenuation with
the focus on optimizing conditions. His presentation
covered the role of electron acceptors, factors affecting
bacterial-contaminant interactions in aquifer material
(e.g., bacterial nutrition, pH and redox potential, tem-
perature, and physical deterrents to bioremediation),
and aerobic versus anaerobic metabolism for a variety
of organic contaminants.
The final presentation in this session was given by Mary
Jo Baedecker, also of USGS. She discussed the inte-
gration of field and laboratory work in demonstrating the
efficacy of bioremediation strategies at a crude-oil spill
near Bemidji, Minnesota. At sites where the rates of
solubilization, volatilization, and biodegradation of hy-
drocarbons are such that the plume is either contained
or spreading at a slow rate, she concluded, natural
attenuation can be considered an important component
of a remediation program.
State and Federal Issues Impacting
Natural Attenuation
Michael Jawson, EPA Robert S. Kerr Environmental
Research Laboratory (RSKERL), presented a summary
of Monday's regulatory meeting at the end of Day One.
His presentation highlighted major themes and key is-
sues, with a focus on what regulators need to know to
assess and evaluate natural attenuation. His emphasis
was on data gaps and needs for natural attenuation,
including the key areas of degradation rates, site char-
acterization, monitoring, guidance materials, closure cri-
teria, and risk assessment/management issues. A
general conclusion was that often characterization and
monitoring needs are even greater for natural attenu-
ation than for other remedial alternatives due to the lack
of hydraulic controls on sites and the complexity of
establishing plume containment and degradation. The
more information that can be provided about site and
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contaminant characteristics and the more complete the
monitoring information, the more confident regulators
can be in evaluating remediation proposals and in clos-
ing sites where bioremediation was used.
Mr. Jawson's overview presentation was followed by a
panel discussion moderated by representatives of state
and federal regulatory agencies. Questions and an-
swers focused on criteria regulators use in evaluating
applications for proposals to implement natural attenu-
ation and for site closures. Day One closed with a poster
session, consisting of 12 posters.
Case Studies: Public Service's Site
Days Two and Three of the symposium were devoted to
several sessions covering case studies, the first of
which concerned the successful application of in situ
bioremediation at the Public Service Company site in
Denver, Colorado. This session, chaired by EPA's John
Wilson, consisted of five presentations made by the
various parties involved in the cleanup and subsequent
investigation, including the remedial contractor, a scien-
tist at the site, a Public Service Company repre-
sentative, and regulators from the state department of
health. Christopher Nelson, Groundwater Technology,
presented a short history of the site and explained the
treatment process from site assessment and treatability
studies to operation, monitoring, and results. At this site,
indigenous microorganisms, stimulated through the in-
troduction of oxygen and inorganic nutrients, achieved
removal of nearly 94 percent of the original contaminant
mass. The bioremediation system operated from July
1989 to March 1992.
Tissa Illangasekare, University of Colorado, Boulder,
Colorado, described the role of mathematical models as
tools for the design and evaluation of ground-water re-
mediation schemes and the application of a particular
model in developing a methodology for retrospective
evaluation. At the Public Service Company site, a meth-
odology was developed to help determine, after closure,
whether or not the contaminant could be expected to
reappear and, if so, how long it would take and what
concentration levels it would reach in the ground water.
Harry E. Moseley, Public Service Company, presented
the steps the company took to remediate the site as
efficiently and cost-effectively as possible. He described
the remediation, monitoring, and evaluation processes,
along with precautionary measures taken and lessons
learned. Mr. Moseley also provided a fairly detailed
breakdown of cost data from the project.
This session concluded with two presentations by Lisa
Weers and Mark Walker, Colorado Department of
Health, regarding closure and postclosure issues at the
Public Service Company site. Ms. Weers and Mr. Walker
reviewed the factors they consider in evaluating a site
for closure, including whether the point of compliance
wells are meeting the state maximum contaminant lev-
els, whether the existing monitoring program is ade-
quate for detecting contamination from the release, and
whether the extent of remaining contamination could
adversely affect human health or the environment. Ms.
Weers and Mr. Walker emphasized that after a site is
closed, the owner/operator still remains liable for any
residual subsurface contamination, and the Department
of Health needs to be informed of any developments that
could increase the potential for adverse affects to hu-
man health or the environment.
Case Studies: Sites Where Natural
Attenuation Has Been Documented
The first part of the second case studies session,
chaired by Jim Spain, USAF, Tyndall AFB, Florida, con-
sisted of five presentations focusing on sites where
natural attenuation has been documented. EPA's John
Wilson described studies and findings related to charac-
terization of JP-4 jet fuel spills at Eglin AFB, Florida, and
Hill AFB, Utah, and presented data on ground-water flow
and chemistry as well as measured and predicted con-
taminant concentrations. He also outlined the method-
ology behind distinguishing between contaminant
attenuation due to destruction from biological activity
and attenuation due to dilution or sorption and how
these methods were applied to correcting for dilution of
BTEX compounds at Eglin and Hill.
Thomas B. Stauffer, Tyndall AFB, described research at
Columbus AFB, Columbus, Michigan, that was con-
ducted to measure in situ biodegradation rates of or-
ganic compounds. He presented the methods and
parameters of the pulse injection experiments used in
determining the amount of 14C carbonate produced by
complete degradation of 14C p-xylene and the amount
of 14C in water-soluble organic intermediate products. In
addition, Mr. Stauffer introduced the leaky reactor model
used to describe solute concentrations as functions of
time. He concluded with results of the experiment and a
discussion of the applicability of the leaking reactor
model.
In the next presentation, David W Ostendorf, University
of Massachusetts, Amherst, summarized the theory de-
scribing capillary tension/saturation relations for immis-
cible continuous fluids. He discussed the theory's
applicability to characterization of distribution of sepa-
rate phase aviation gasoline (avgas) in solid core sam-
ples, specifically those taken from the U.S. Coast Guard
Air Station in Traverse City, Michigan. Mr. Ostendorf
also presented distribution profiles of total saturation,
free avgas, and residual avgas and reviewed the impor-
tance of such characterizations in the estimation of
separate phase contamination from monitoring-well ob-
servations.
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Barbara H. Wilson, EPA, Ada, Oklahoma, provided fur-
ther discussion of the bioremediation studies conducted
at the U.S. Coast Guard Air Station in Traverse City,
specifically research into the geochemistry of BTX com-
pounds. Ms. Wilson summarized the results of geo-
chemical characterization of water samples and
presented evidence of natural attenuation of BTX com-
pounds in aerobic and anaerobic aquifer material.
The next presentation, by Hanadi S. Rifai, Energy and
Environmental Systems Institute, Rice University, Hous-
ton, Texas, focused on mathematical modeling of natural
attenuation. Dr. Rifai presented four case studies that
combined field and laboratory investigations with mod-
eling studies: the Conroe Superfund Site, Texas; the
Traverse City Site; a gas plant facility in Michigan; and
the Cliffs-Dow Superfund Site.
Natural Attenuation of Trichloroethylene
and Similar Compounds: Case Studies
The final session of the symposium focused on case
studies of natural attenuation of trichloroethylene (TCE)
and similar compounds. This session, also chaired by
Gail Mallard, opened with a presentation by Perry L.
McCarty, Department of Civil Engineering, Stanford Uni-
versity, on the anaerobic transformation of chlorinated
solvents. In his presentation, Mr. McCarty provided
background information on major chlorinated solvents
and described the chemical transformation of 1,1,1-
trichloroethane (TCA) and the biological transformation
of chlorinated aliphatic hydrocarbons (CAHs). He con-
cluded his talk by presenting case studies that provide
evidence for natural attenuation of chlorinated solvents.
The second part of the session on documented case
studies, chaired by Gail Mallard, USGS, concluded on
Day Three with two presentations on studies conducted
at a crude-oil spill site in Bemidji, Minnesota. Robert P.
Eganhouse, USGS, Reston, opened with an overview
of the investigation at the Bemidji research site. Mr.
Eganhouse presented the methodology for water, sedi-
ment, and oil sampling; chemical analyses; and site
characterization. He further presented evidence of
change in ground-water geochemistry due to the inad-
vertent introduction of crude oil to the aquifer and pro-
vided data that supports the attenuation of oil-derived
contaminants due to biological, rather than physical,
processes.
Mary Martin Chepiga, USGS, West Trenton, New Jer-
sey, spoke next about TCE contamination of ground
water at the Building 24 Site at Picatinny Arsenal, New
Jersey. She described how the infiltration ofwastewater
from lagoons behind the building and of chlorinated
solvents from a dry well in front of the building had
created a plume of contaminated ground water down-
gradient of Building 24. Ms. Chipega further described
the conceptual model of the physical, chemical, and
biological processes that affect the transport and mass
balance of TCE within the plume. She also presented a
preliminary solute mass balance of TCE in the uncon-
fined aquifer at the site and compared those findings
with results of numerical solute-transport simulations.
Supplementing Mr. Eganhouse's talk, Hedeff I. Essaid,
USGS, Menlo Park, California, presented research into
the simulation and flow of transport processes at the
Bemidji crude-oil spill site. Mr. Essaid stressed the im-
portance of using numerical models for integrating infor-
mation collected in the field, for testing hypotheses, and
for studying the relative importance of simultaneously
occurring processes in complex, real-world systems. In
addition to outlining the procedure for collecting core
samples, Mr. Essaid summarized the assumptions and
results of the two-dimensional model used to simulate
multiphase flow through an oil pool at the site. He also
provided background and results of the two-dimensional
multispecies solute-transport model used to quantify the
field-scale degradation process and to identify the im-
portant factors affecting the distribution of solute species
in the field.
James W Weaver, EPA, Ada, Oklahoma, closed the
case studies session with a presentation of the natural
attenuation of TCE in ground water at the St. Joseph,
Michigan, National Priority List (NPL) site. Mr. Weaver
summarized the data collection, presentation, and in-
terpretation methods used to infer and analyze TCE
concentration in the western plume near Lake Michi-
gan. He emphasized that the contaminant mass flux
and apparent degradation constants were based on
visualization of the data, that presentation of the data
was free from artifacts of interpolation, and that ex-
trapolation of the data beyond the measurement loca-
tions was controlled.
Fran Kremer provided closing remarks, and the sympo-
sium was adjourned.
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State and Federal Issues
Impacting Natural Attenuation
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State and Federal Issues Impacting Natural Attenuation:
Meeting Overview
Introduction
A meeting to discuss state and federal regulatory issues
impacting natural attenuation was held on Monday, Au-
gust 29, 1994, in Denver, Colorado, in conjunction with
the U.S. Environmental Protection Agency's (EPA's)
Symposium on Natural Attenuation of Ground Water.
The objective of the meeting, which was attended by 28
state and federal regulators, was to discuss obstacles
and issues related to the use of natural attenuation as
an accepted cleanup option at waste sites. Critical top-
ics addressed included:
• What regulators need in order to assess and evaluate
natural attenuation proposals.
• Data gaps that need to be filled to increase the vi-
ability of natural attenuation as a cleanup option.
• Site closure criteria.
• Long-term monitoring requirements.
Fran Kremer and John Wilson of EPA's Office of Re-
search and Development co-chaired the meeting. (Ap-
pendix A contains a list of the meeting participants.)
Definitions
The meeting opened with a discussion of terminology
and an attempt to develop a definition for natural attenu-
ation that is accurate and avoids misconceptions and
negative connotations currently associated with related
terms such as natural attenuation. Participants agreed
that the perception of natural attenuation and related
processes as "no action" remedies represents a major
obstacle to the acceptance of natural attenuation.
A working definition was crafted for purposes of the
meeting discussions, but participants concluded that
formalizing a suitable definition of natural attenuation is
a high priority, both to change public perception and to
promote regulatory consistency.
Overview
The major issues raised in the meeting included the
following:
• Obtaining degradation rates
• Improving site characterization
• Cost-effective monitoring
• Developing guidance materials
• Developing closure criteria
• Addressing risk assessment/management
Conclusions and recommendations from discussion of
these issues follow.
Degradation Rates
There is a need for contaminant degradation rates,
which would be used to facilitate site characterization,
assess the viability of natural attenuation as a remedial
action, and predict how long cleanup would take. This
latter information would help regulators justify the use of
natural attenuation where it could be used for screening
purposes or where cleanup could be shown to be com-
pleted in a reasonable time frame. The need for regula-
tors to identify an endpoint for cleanup via natural
attenuation was a recurring theme. Degradation rates
are especially needed for contaminants other than pe-
troleum, for which a fair amount of data currently exist.
Participants suggested, however, that a published re-
cord of existing petroleum data would be very useful.
Data on regulated contaminants and mixtures are par-
ticularly needed, as most spills are not single com-
pounds but mixtures of several contaminants. Field data
would be most useful, and better methods for relating
laboratory data to the field would also be helpful. Ideally,
the data produced should be site specific and take into
account site geology, contaminant transport, electron
acceptors at the site, and other site characteristics.
Degradation rates also should be tied to mass balances,
such that the mass balances coincide with projected
rates.
Site Characterization
Detailed site characterizations are needed for all reme-
dial efforts, but natural attenuation requires particularly
intensive characterizations. To justify natural attenu-
ation, regulators need to be assured that contaminants
will not migrate to sensitive receptors or pose health
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threats during cleanup. Thus, cost-effective methods for
characterizing the location and possible migration of the
plume, particularly in relation to sensitive receptors and
other health exposure points (e.g., utilities, drinking
water wells, and discharge points at streams, ponds,
and lakes), are needed. Regulators want full three-di-
mensional plume characterization, including vertical dis-
tribution, along with information on physical and
hydrogeological parameters to determine the fate and
transport of contaminants. Also needed are visualization
systems that synthesize and display all of these charac-
terization data in an easily understandable manner.
Site characterization is considered one of the more
costly components of natural attenuation. For small sites
(e.g., those in the leaking underground storage tank
(LIST) program), assessment costs can run as high as
$40,000 or $50,000, a burden that many small busi-
nesses cannot afford. In Utah, for example, only about
10 percent of LIST sites can afford to hire contractors to
perform thorough assessments. In addition, in some
states, such as Utah and Nevada, there are very few
environmental consultants available, even if cost were
not an issue. One response was that states could offer
reimbursement programs, similar to that in Florida. An-
other suggestion was to prepare guidance that specifically
addresses how to characterize sites with limited funding.
Monitoring
Bioremediation is not a "walkaway" remedy; on the con-
trary, it requires intensive, often long-term, monitoring.
Natural attenuation, in fact, can require more monitoring
and analysis than many other types of remediation.
Issues of concern to state and federal regulators in-
cluded how frequently to monitor and at how many
points, exactly which parameters to measure to assess
the progress of biodegradation processes and cleanup,
and whether surrogate parameters would be adequate
if a project was constrained by budget. In addition to
establishing points of compliance, a primary goal of
monitoring is to ensure that the plume is not migrating
beyond the points of compliance. Monitoring should oc-
cur more frequently initially, then less frequently when
remediation rates, migration patterns, and seasonal in-
fluences are understood.
Potential monitoring parameters mentioned by partici-
pants included dissolved oxygen (DO), temperature gra-
dients, and electron acceptor levels. Analysis of volatile
organic chemicals (VOCs) was cited as a requirement
that could greatly increase costs. Information on tech-
niques and methods for analyzing different classes of
compounds would be useful. Participants also expressed
a need for validation of data from conventional wells,
and for mass balance calculations. Finally, participants
discussed the role of models versus field data, and the
extent to which models could be used to complement or
even replace field data for the purposes of predicting
plume behavior and contaminant degradation.
Type of Guidance Materials
The need for guidance materials was stressed. Informa-
tion on natural attenuation needs to be presented in
simple terms so that it can be offered clearly and per-
suasively as a remedial alternative. More information is
needed on the types of situations for which natural
attenuation is suitable, considering such factors as con-
taminant types and levels and site characteristics. State
and federal participants called for protocols, in the form
of guidelines, on what parameters to measure and pro-
cedures to use. It was also suggested that user-friendly
computer models and decision support systems would
be invaluable to assist site coordinators with implemen-
tation.
Closure Criteria
A key need is the establishment of closure criteria to
determine at what point a site can be considered "clean."
There was discussion as to which parameters should
serve as "endpoint parameters," or parameters that
must be measured to prove that cleanup is complete.
Participants questioned the adequacy of current criteria,
particularly with regard to toxicological tests. Toxicologi-
cal testing should be made on the weathered rather than
nonweathered material, and the material should come
from the site itself.
Participants also felt that criteria other than total petro-
leum hydrocarbons (TPHs) should be used and that
daughter products and by-products for the degradation
process should be carefully measured and assessed.
Resolving these issues would provide regulators with
more confidence in closing sites where natural attenu-
ation was used.
Summary: Needs for Effective Risk
A ssessment/Management
The final critical theme, drawing on all of these issues,
was what information regulators would need in order
to perform risk assessment and management of natu-
ral attenuation sites. Regulators at the meeting
pointed out that they frequently receive proposals for
intrinsic remediation that do not provide adequate
data on physical parameters. Participants concluded
that they would need information related to plume
containment, including whether or not degradation
was occurring. For petroleum sites, it was felt that
bioremediation usually is occurring, but more data on
retardation and degradation coefficients would be
helpful. For nonpetroleum sites, assessment is more
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difficult, and would require databases based on actual
sites showing the range of rates and parameters. A
database of case studies documenting both successes
and failures of natural attenuation in the field would be
of enormous value. Such a database is under develop-
ment; it currently has information on over 450 bioreme-
diation sites, some of which have intrinsic components.1
Regulators would also find useful specific information on
what distinguishes intrinsic from engineered bioreme-
diation. All of this information should be presented in
clear, simple terms in guidance materials geared for
decision-makers and regulators.
Cost-effectiveness and timeliness are other key issues
in assessing the natural attenuation option. Sites need
to document that natural attenuation can be achieved in
a "reasonable time frame," a phrase that carries some
flexibility depending on the relative containment of the
contamination and its threat to receptors. Some partici-
pants suggested that natural attenuation is most appropri-
ate for sites where contamination is relatively low, little risk
is posed to drinking water supplies, and the contaminant
source has been removed. In many cases, natural attenu-
ation can be shown to be more cost effective and speedier
than other potential remedies.
Key data needs are summarized in Table 1.
Table 1. Checklist of Data Gaps and Needs
A. Degradation Rates
• For more than petroleum and regulated products/mixtures
• Field data (relation to lab)
• Tied into site characteristics (geology, transport, etc.)
• Multiple electron acceptors
• Tied into mass balances
B. Site Characterization
• More intensive for natural attenuation
• Three-dimensional characterization
• Plume location in relation to sensitive receptors and health
exposure points
• Data (physical, hydrogeological, etc.) to determine fate and
transport
• Visualization systems
C. Monitoring
• Frequency
• Parameters to measure (use of surrogate parameters)
• Validity of data from conventional wells, need for mass balances
• Role of models vs. field data
D. Guidance Materials
• Information in simple terms
• Situations (sites, levels, and types of contamination) where
natural attenuation is applicable
• Protocols (parameters to measure, procedures to use)
• Models (user friendly) and decision support systems
E. Closure Criteria
• Endpoint parameter data
• Adequacy of current criteria/toxicological tests
• Toxicological testing of "weathered" object
• Daughter and/or by-products
F. Risk Assessment/Management
• Plume containment, contaminant degradation
• Time frame
• Cost effectiveness
• Database of case studies
• Guidelines for decision-makers
The Bioremediation in the Field Search System (BFSS) 2.0 is a
PC-based bioremediation site database that allows users to search
for data electronically, view the data, and print reports. BFSS is
available on EPA's Alternative Treatment Technology Information
Clearinghouse (ATTIC, 703-908-2138), Cleanup Information (CLU-IN,
301-589-8366), and Office of Research and Development (ORD,
513-569-7610) electronic bulletin board systems. BFSS is also avail-
able on diskette from EPA's National Risk Management Research
Laboratory. To request a copy, call 513-569-7562.
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Panel Discussion on Regulatory Issues
Panel Members:
Michael Barden, Wisconsin Department of Natural Re-
sources, Madison, Wl
Matthew Charsky, Hazardous Site Control Division, Of-
fice of Solid Waste and Emergency Response, U.S.
EPA, Washington, DC
Daniel Cozza, Region 5, Superfund Program Manage-
ment Branch, U.S. EPA, Chicago, IL
Fran Kremer, Office of Research and Development, U.S.
EPA, Cincinnati, OH
Gail Mallard, U.S. Geological Survey, Reston, VA
John Shauver, Michigan Department of Natural Re-
sources, Lansing, Ml
Guy Tomassoni, Office of Solid Waste, U.S. EPA, Wash-
ington, DC
John Wilson, Office of Research and Development, U.S.
EPA, Ada, OK
Question: The list of data gaps and needs presented
by Mike Jawson is very good. It would be helpful if the
list were written up and distributed so that we could work
from it.
Answer (Fran Kremer): Symposium participants will
receive a final report from this meeting, which will in-
clude the results of the meeting and panel discussion.
Question: The list presented was very comprehensive,
sort of a "wish list." Is it necessary to address all of it
before regulatory agencies begin to accept applications
of bioremediation? Which two or three critical needs
should be met now, as opposed to all 20 or 25 men-
tioned?
Answer (Dan Cozza): For the most part, nothing on
the list is new. For any technology, you need essentially
the same list of criteria to close out a site or choose a
remedy. You can however, have flexibility, for example
with respect to monitoring and site characterization. But
since we don't want natural attenuation to be labeled as
"walkaway" or "no action," we may require even more
monitoring than for an active pump-and-treat system. So
for the most part, you will need nearly everything on the
list, but there will be flexibility. Every site is different, and
sometimes the information will already be available.
Closure criteria are the same regardless of the type of
technology.
Answer (John Shauver): I'd like to say "all of the
above," but we live in the real world. Remember that
bioremediation is still site-specific; therefore, we can't
paint with a broad brush. If you know what kinds of
systems operate best on the contaminant at the site, you
can limit the monitoring and the data that need to be
generated. Often responsible parties propose bioreme-
diation without analyzing forsulfate, nitrate, iron, etc. or
knowing anything about the geochemistry, biochemistry,
or microbiology. It is important to think about what the
processes need and develop analytical data based on
what is likely to happen before submitting a proposal.
Because intrinsic and any other kind of bioremediation
is site specific, the state regulatory agency can't provide
a cookbook.
Question (Fran Kremer): South Carolina has devel-
oped some guidelines for what they would like to see in
proposals that come to them. I would like to ask
Read Miner from South Carolina to comment on those
guidelines.
Answer (Read Miner): In 1992, we published a guid-
ance document for completion of comprehensive as-
sessments. We have had a lot of success with
contractors completing more comprehensive assess-
ments to more fully characterize sites so that the best
remediation options could be determined. We are cur-
rently developing criteria for natural attenuation; we are
gathering information at this symposium to help com-
plete the development of these criteria.
Question: What criteria will EPA use to evaluate a plan
specifically for a Superfund site (how many samples
should be collected and where, what parameters should
be tested for, etc.)?
Answer (Matthew Charsky): In the past, we have
looked at four factors for Superfund sites: practicability,
cost-effectiveness, the unlikely future use of drinking
water, and the time frame to reach the cleanup level.
Historically, we have used the term natural attenu-
ation rather than natural attenuation. Natural attenu-
ation has been used at approximately 57 sites almost
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nationwide between 1986 and 1993. In order to use
natural attenuation, these sites have had to meet the
following criteria: contaminant concentrations are rela-
tively low, drinking water is not threatened or can be
easily replaced, the source of contamination is no longer
present, source removal will result in a predictable de-
crease in ground-water contamination, and the active
restoration time frame is not radically different from that of
active remediation.
Question: In a case where EPA has already agreed
that we can evaluate the process, how will the adequacy
of the plan be evaluated?
Answer (Dan Cozza): It is hard to specify what we
would accept (for example, a certain number of wells)
because every site is different. The bottom line is that
you have to show that the plan will not affect human
health and the environment and you need to do the site
characterization. Another key element in a plan is part-
nership among EPA, state and local agencies, and the
responsible party to work together on a proposal before
it comes in. Such a partnership among everyone in-
volved is critical if natural attenuation is going to become
a reliable alternative.
Question: There seems to have been a somewhat con-
flicting message today: On the one hand, there has been
discussion of the need for site-specific evaluations; on
the other hand, it has been implied that one must do
mass balances, microcosm studies, and so on. I hope
we don't lose sight of the fact that the proof that must
be provided on natural attenuation projects is a function
of the objectives and criteria you are trying to meet and
the site conditions. For example, at one site you might
have to show that a plume is not growing or is shrinking;
at another site you might have to show that a plume will
not reach a receptor location a certain distance from its
current extent. In the latter case, you need a more
rigorous proof than in the former case. We need to
recognize the need for flexibility in the level of proof
required.
Answer (Guy Tomassoni): Particularly in the RCRA
program, we often have years of data from a site. If we
have a good monitoring program over the years that
shows the plume reducing in size and concentration,
then we shouldn't always require the regulated entity to
expend a tremendous amount of time and resources to
prove why good things have happened. There might be
relatively inexpensive ways to show that bioremediation
is taking place, but at some sites it may be very difficult
and costly to figure out why levels are going down and
why the plume is not growing. Resource Conservation
and Recovery Act (RCRA) regulations require remedia-
tion of ground water to achieve designated standards;
the regulation does not require proof of why levels are
being achieved.
Answer (Fran Kremer): There is a difference as well
between RCRA/Superfund sites and LIST sites. There
are different levels of resources available to do the
necessary work, and state LIST programs have varying
opinions about what is considered acceptable.
Answer (Michael Barden): At all the sites ranging
from Superfund to leaking USTs, we look for adequate
site characterization (often including vertical gradients)
so that we know where the plume is and where it is
going, and we look for adequate monitoring to show that
concentrations are decreasing overtime. If we can get
a better understanding of how these processes work,
that will help us consider which sites are suitable for
natural attenuation. From the regulatory standpoint,
however, the fundamental point is that concentrations
are decreasing, not necessarily why Ms is happening
so long as the plume isn't migrating. Mass balance is
the key indicating that we are losing the contaminant
mass over time, but cleanup standards are concentra-
tion-based, not mass-based, so we are looking at con-
centration changes.
Question: Regarding closure criteria, many sites with
natural attenuation will not reach maximum contaminant
levels (MCLs) or other levels in the entire plume, and
decisions must be made to leave some contamination
in place. How do you make those decisions, and how
do you use the concept of alternate compliance points,
which fit in well with the natural attenuation approach?
Answer (John Shauver): One option is to clean up to
background levels, get a clean bill of health, and have
no future liability. But remember, risk-based numbers
are generated by an algorithm using conservative as-
sumptions, based on complete freedom of use of the
site. With a risk-based number, the potential exists for
the regulatory community to say that the number is too
high and the site must be cleaned up again. Therefore,
in the case of natural attenuation, you must look at
cleanup not only from the standpoint of environmental
protection but also long-term liability. These sites may
be looked at for 25 or 100 years. As long as you can
show that public health and the environment are pro-
tected, we don't care how long the site is monitored, but
you will be held to a cleanup standard based on future
use of that resource. We may be willing to write off the
site for 50 or 100 years while we watch the microorgan-
isms degrade the contaminants, but there must be some
activity on the site to make sure that the microorganisms
are not being affected in some other way and that deg-
radation is really taking place. We like bioremediation
because it is not temporary storage and does not trans-
fer contaminants from one medium to another, but there
are liability risks and social risks, and the regulator must
take those risks into account in determining how much
contamination can be left on a site. Much of that deter-
mination is not based on science.
10
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Answer (Michael Barden): The enabling statutes in
the states vary widely. For example, in Wisconsin,
ground-water standards must be met for all ground
water in the state, whether you drink it or not. We do
have flexibility with respect to how long you have to meet
those standards. That is where we can consider risk
management, plume migration, and the potential impact
on receptors. If there is no potential impact, we don't
care how long it takes, but the standards must ultimately
be met.
Question: Has there been much movement toward
combining the technology of natural attenuation with
other methods of attenuation? For example, natural at-
tenuation might be effective at very small sites, but not
at large sites, where additional types of remediation
might be necessary. But if you resort to pump-and-treat,
you may upset the microbiological environment and no
longer have the natural attenuation. Has there been
much work on combining natural attenuation with other
passive forms of treatment, such as a funnel and gate
system?
Answer (John Shauver): From our perspective, it
doesn't matter what suite of technologies a responsible
party uses to address a site. When you have reasonably
low concentrations (parts per billion or low parts per
million range), water supplies are not at risk, and you
can show that degradation is taking place, then you can
use something more passive. The problem in areas with
increasing populations is that it is difficult to predict how
long that cleanup will take. Some communities don't
want to wait 25 years or more because of lost economic
development opportunities. Therefore, you have to look
at community concerns and long-term land use as well
as regulatory needs.
Answer (Guy Tomassoni): States or areas of the
country might have varying ideas about what constitutes
a reasonable time frame. We are seeing developments
in more passive technologies, such as treatment walls,
that could possibly supplement completely natural sys-
tems. Combinations of a conventional or more active
remedy for a part of a plume with an intrinsic-based
remedy for the remaining plume may prove in many
settings to yield a cost-effective and protective remedial
strategy. Clearly, what an entity views as a reasonable
time frame is crucial. It is also important to realize that
time frames for cleanups using a conventional pump-
and-treat remedy can be very long as well.
Question: Wisconsin has had clear and usable guid-
ance for a year and a half on how to evaluate a site for
natural attenuation. At what percentage of the sites has
natural attenuation been found to be a viable option?
Answer (Michael Barden): We don't have any hard
numbers because we don't have the personnel to put all
the information into a database. We recently surveyed
site owners to determine what remediation technologies
are being used. In the past year, about 80 sites have
applied natural attenuation at least to soil contamination.
Fifteen or 20 are looking at this technology for ground-
water contamination. Larger stable entities such as ma-
jor oil companies and active industrial sites are looking
at this as a viable option. We are getting a lot of interest
in natural attenuation, but more as a followup to more
active remediation, because many responsible parties
are not willing to wait long enough.
Question: What is the real objective of natural attenu-
ation—is it to stop the plume, to shrink the plume, or
even to attack the residual nonaqueous phase liquid
(NAPL)? How are objectives viewed by the regulators?
Answer (John Shauver): In Michigan, we have a mul-
tiple use ethic, looking at resource management and
resource use. If you are going to let a plume get bigger,
we are concerned about future users as well as existing
users. As a regulator, I want control so that future users
won't be affected and don't have to worry about a public
water supply, and I want to know that degradation is
taking place. When I say stop the plume, I want some
mechanism in place that can achieve that control. If
microorganisms are achieving that goal, then put the
monitoring system in and demonstrate that it is working.
You don't have to do all the scientific studies, but we
want to know that degradation is taking place and that
the plume is not getting bigger in three dimensions.
We're willing to watch it 5, 30, or 100 years while the
microorganisms do the job. But at a fresh spill where
you've removed the source and the plume is likely to get
much bigger (and there are homeowners nearby using
the water), we will probably say that you must keep the
plume that size, using some mechanical control if need
be. We will evaluate every site, but our basic concept is
not to contaminate more of the resource if there is a
reasonable, feasible way to maintain the size of the
plume. Therefore, we put limits on when natural attenu-
ation is appropriate.
11
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Symposium on Natural Attenuation of Ground Water:
Summaries and Abstracts
13
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Regulatory Perspective on Natural Attenuation
Chairperson: Fran V. Kremer, U.S. Environmental Protection Agency
Department of Defense Perspective
Colonel Jim Owendoff, U.S. Department of Defense,
Washington, DC
The U.S. Department of Defense (DOD) consists of over
1,000 installations across the United States and over-
seas, with over 10,000 sites where biological and non-
biological cleanup technologies are being considered or
implemented. DOD has an annual budget of about $2.3
billion for cleanup at closing and nonclosing bases.
Funds for the overall defense budget have decreased
while funding for environmental programs has increased
or remained constant over the past several years.
DOD hopes to move toward using technologies that
eliminate contaminants from media rather than toward
those that transfer contaminants between media. DOD
also must address the issues of prioritization and se-
quencing of work, timeliness of cleanup, ease of tech-
nology implementation, and affordability of treatment. In
assessing these factors, DOD hopes to engender a
greater participation among the regulatory, public, and
industrial communities. The drafting of management ac-
tion plans constitutes a step in the direction of greater
collaboration. These plans provide comprehensive de-
scriptions of each site along with detailed information
about its contaminants and the strategy, schedule, and
initiatives for site cleanup. The establishment of a Res-
toration Advisory Board, a body of regulators and com-
munity members that would evaluate the relative risk of
site contamination, is a further move towards greater
communication of problems and solutions among the
regulatory and public communities.
In view of the large number of DOD sites which have
hydrocarbon contaminants, intrinsic bioremediation
holds substantial promise for application. For intrinsic
bioremediation to be accepted, the remedial project
manager (RPM) must perform site evaluations and pre-
sent evidence that bioremediation has occurred as well
as implement monitoring systems to confirm the con-
tinuance of natural biodegradation. Utilizing these prin-
ciples and guidelines, DOD will be able to implement
intrinsic bioremediation while protecting and preserving
human health and the environment.
U.S. EPA Regulatory Perspective: Natural
Attenuation as a Remedial Option for
Contaminated Ground Water
Guy A. Tomassoni, U.S. Environmental Protection
Agency, Office of Solid Waste, Washington, DC
The two primary goals of both the RCRA and CERCLA
(Superfund) programs for contaminated ground water
are (1) protecting human health and the environment,
and (2) restoring contaminated ground water to benefi-
cial uses, where technically practicable. EPA has recog-
nized in policy statements and guidance that one
approach to achieve these goals in certain settings can
involve natural attenuation. Thus, EPA policy allows for
natural attenuation to be used as a component of pro-
tective and cost-effective remedies.
Natural attenuation is addressed in the preambles of
both the National Contingency Plan (NCP) (55 FR 46
March 8,1990) and the proposed Subpart S rule (55 FR
145 July 27, 1990), which provide the regulatory frame-
work for implementing the Superfund and RCRA Correc-
tive Action programs, respectively. The NCP defines a
remedy based on natural attenuation as one where
biodegradation, dispersion, dilution, and adsorption will
effectively achieve remedial goals. The NCP preamble
also states that natural attenuation is generally recom-
mended only when active restoration is not practicable,
cost effective, or warranted based on site-specific con-
ditions (e.g., where ground water is not a likely source
of drinking water), or where natural attenuation could
achieve remedial goals in a reasonable time frame.
Similar to the NCP, the Subpart S preamble discusses
natural attenuation in the context of reasonable cleanup
times, and where the likelihood of exposure is minimal.
Furthermore, the Subpart S preamble states that natural
attenuation could play a major role in a remedy that can
be carried out over an extended period of time.
14
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An example of EPA guidance that recognizes natural
attenuation was issued very recently by the Office of
Underground Storage Tanks. The guidance document
titled, Evaluating Alternative Cleanup Technologies for
Underground Storage Tanks: A Guide for Corrective
Action Plan Reviewers, covers eight technologies
including natural attenuation. The chapter on natural
attenuation includes guidance on initial screening, de-
tailed evaluation, and monitoring. While EPAs recogni-
tion in policy and guidance helps to determine when to
select natural attenuation as part of remedial action,
significant but not insurmountable obstacles still exist.
Two major obstacles include (1) the association of natu-
ral attenuation with a no action remedy; and (2) the
preference for rapid remediation.
Associating natural attenuation with no action often re-
sults in the conclusion that the regulators would be
letting the responsible party "off-the-hook." Overcoming
this obstacle will involve educating all interested parties
that a remedy based on natural attenuation is intended
to achieve remedial goals and will still involve a signifi-
cant cost to the responsible party. For example, a natu-
ral attenuation remedy will generally require thorough
site characterization, source control or removal, docu-
mentation (where appropriate) of attenuation processes,
and comprehensive long-term monitoring to ensure
protectiveness and contaminant reduction. Natural at-
tenuation remedies should also generally include contin-
gencies for more "active" remedial measures in the event
that natural processes will not achieve the goals, or where
such measures are deemed necessary for part of the
contaminant plume. Activities such as these need to be
conveyed to distinguish a remedy involving natural attenu-
ation from a no action or walkaway decision.
The preference for rapid remediation can be an obstacle
because natural attenuation is very often associated
with prolonged remedial time frames. Overcoming this
obstacle can be achieved by (1) recognizing that reme-
dial time frames associated with conventional pump-
and-treat systems also may be prolonged; and (2)
recognizing that longer remediation time frames may be
appropriate in certain circumstances where there is no
urgent need for the resource, and human and environ-
mental exposures can be prevented.
Scientific evidence has documented nature's capacity to
rectify environmental contamination in certain settings.
Where natural processes are determined capable of
achieving remedial goals for ground water, EPA affords
the flexibility to evaluate and select natural attenuation
as a component of a protective remedy. The further
acceptance of remedies involving natural attenuation will
likely be facilitated by continued advances in the scientific
understanding of the processes involved, and the ability to
address public and regulatory concern that such remedies
be protective and will achieve the intended cleanup goals.
State Perspective
John Shauver, Michigan Department of Natural
Resources, Lansing, Ml
Bioremediation is not a no action, walkaway alternative,
whether it occurs naturally or with assistance. A presen-
tation given 10 years ago at the 7th Ground Water
Conference at Michigan State University allows one to
examine the Department of Natural Resources perspec-
tive on ground water in Michigan and assess the
changes that have occurred in the past 10 years. Al-
though the approach to seeking legal relief at any
ground-water contamination site has remained the
same, regulations have become more complex and
stringent. Consequently, regulation prompted the re-
search and development of cleanup technologies that
are quicker and less costly, and that provide greater
long-term liability. Bioremediation has emerged as a tool
that can meet those needs.
Bioremediation has significant advantages over other
treatment technologies. True degradation of contami-
nants will limit long-term potentially responsible party
(PRP) liability. In addition, although initial startup costs
and long-term monitoring costs may be somewhat
higher than for other remedies, bioremediation allows
for in-place treatment, thereby eliminating the need for
repeated excavations over time. Bioremediation also
offers an element of environmental sanity in that it does
indeed eliminate waste by eliminating toxicity and the
need for continued storage and disposal.
Whether natural or enhanced, bioremediation is a treat-
ment technology that must be described to regulators in
detail. Being able to describe the requirements of the
remedial process to regulators helps them to under-
stand the treatment process, to settle any permitting
issues that may arise, and to be assured that the tech-
nology has a reasonable chance of remediating the
contaminants so that the resource is protective of public
health and the environment.
The first step in any treatment process should consist of
an adequate site characterization. Next, it is necessary
to understand microorganism behavior and response to
the contamination so that plume shrinkage or expansion
can be described and predicted. Monitoring systems
can assist in addressing plume migration in current and
long-term site characterizations.
The biggest barriers to the implementation and success
of bioremediation in Michigan are the numerous unreal-
istic and uninformed proposals that are brought before
state environmental agencies. An adequate site charac-
terization must be performed, and initial site investiga-
tions are not necessarily less costly for bioremediation
technologies than for other processes.
15
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Overview of Investigations Related to Natural Attenuation
Chairperson: Fran V. Kremer, U.S. Environmental Protection Agency
USGS Perspective
Gail E. Mallard, U.S. Geological Survey, Reston, VA
Many of the investigations that the U.S. Geological Sur-
vey (USGS) conducts on the processes that affect the
fate and transport of contaminants directly relate to the
topic of intrinsic remediation. A division of the Depart-
ment of Interior, USGS is an earth science information
agency that has no regulatory or management respon-
sibilities. Thus, USGS is able to provide information in
an unbiased manner to both the regulatory and regu-
lated communities. The organization has offices in every
state and conducts studies in close collaboration with
local, state, and other federal agencies as well as uni-
versities. In choosing field laboratories, USGS selects
sites that are known to be contaminated with what the
agency considers to be the most important contami-
nants in the environment. USGS's approach in perform-
ing field investigation is to integrate disciplines to
conduct research that produces thorough site analyses
and reports.
In addition to field-based investigations, USGS research
includes a significant laboratory component. Laboratory
studies help to verify field observations and to under-
stand controlling processes. The ultimate goal of USGS
research is to understand processes so that information
transfer can occur between sites about the fate and
transport of contaminants.
Research conducted at the site of a petroleum spill in
Galloway Township, New Jersey, illustrates the three
steps that USGS stresses regarding the use of bioreme-
diation as a treatment technology:
• Performing site characterization
• Presenting multiple lines of evidence
• Gathering as much information as possible
Understanding site, system, and process helps to evalu-
ate cleanup alternatives. At the Galloway site, a gasoline
tank on a fuel farm had leaked, but the extent and
duration of the leak were unknown. USGS assisted in
performing a site characterization that enabled the
agency to gain approval from New Jersey regulatory
authorities to conduct ongoing research at the site.
Multiple lines of evidence, such as information based on
both field and laboratory observations, data on hydro-
geologic parameters, and information on accumulation
of by-products, further strengthen the case for bioreme-
diation. Periodic reappearances of petroleum by-prod-
ucts combined with information on hydraulic properties
of the unsaturated zone soil and other hydrogeologic
data helped USGS and its collaborators to infer the
overall degradation rate of hydrocarbons at the Gallo-
way site. Comparing rates inferred from field studies
with those from laboratory experiments gave re-
searchers confidence that the remedial approach was
effective.
Having information such as degradation rates gives re-
searchers a better understanding of the natural system
so that they can arrive at optimal engineered solutions.
For instance, degradation rates can be used in a passive
remediation system to predict how long hydrocarbons
are likely to persist in the unsaturated zone, or they can
be used in an engineered solution to design models that
optimize hydrocarbon removal and vapor extraction
(modeling currently being developed at Drexel Univer-
sity). Thus, the more that is known about a natural
system, the better the engineered solution will be.
U.S. EPA Perspective
Fran V. Kremer, U.S. EPA, Office of Research and
Development, Cincinnati, OH
The primary mission of the U.S. EPAs Office of Re-
search and Development (ORD) consists of providing
technical assistance to regional offices and the states
while carrying out a longer term mission to conduct
research to fulfill needs in the environmental area. ORD
also supports activities for the regulatory offices to sup-
ply them with necessary data to issue regulations.
ORD has a staff of approximately 1,200 employees and
had a budget of about $535 million for the 1993 fiscal
16
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year. The Agency had an overall increase in its budget
for 1994 and is currently undergoing evaluations for the
1995 budget. ORD's projected total budget of $564
million includes $9.2 million tentatively reserved for
bioremediation activities and $11.7 million dedicated to
ground-water research.
ORD consists of offices and laboratories across the
country as well as 8 offices at headquarters in Washing-
ton, DC, 12 other laboratories, and 5 additional field
facilities. ORD is currently undergoing a major reorgani-
zation around four "mega" laboratories. Although the
consolidation will not force any labs into closing, at this
time, management will be realigned to reflect the
changes in office structure. The new Risk Management
Laboratory will encompass most of the engineering op-
erations including the air engineering research being
conducted out of Research Triangle Park, North Caro-
lina. The Environmental Criteria and Assessment Office
will move under the Risk Assessment Laboratory, and
the Monitoring Support Laboratory as well as the Ada
and Athens labs will become a part of the newly
structured Exposure Research Laboratory. Finally, the
Health Effects Research Laboratory and the Gulf
Breeze Laboratory will move under the Health and En-
vironmental Effects Research Laboratory.
Some projects on natural attenuation include a collabo-
rative effort with the U.S. Air Force Center for Environ-
mental Excellence in developing a protocol for
evaluating the fate of dissolved-phase hydrocarbons in
ground water. Another project with the Public Service
Company in Colorado involves the development of a
model to establish a methodology to evaluate possible
risks associated with BTEX in soils after remediation.
Research also has been conducted toward distinguish-
ing between attenuation due to destruction from biologi-
cal treatment and attenuation due to dilution orsorption.
Evaluations also have been conducted on the geochem-
istry of ground water to confirm the existence of biologi-
cal activity through examinations of product and
consumed compounds. Through these investigations,
EPA hopes to acquire a better understanding of biore-
mediation and its potential for environmental cleanup.
17
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Overview of Process and Procedures
19
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Site Characterization: What Should We Measure, Where (When?), and Why?
Michael J. Barcelona
Department of Civil and Environmental Engineering, University of Michigan, Ann Arbor,
Abstract
Site characterization represents the initial phase of the
active monitoring process that occurs as part of intrinsic
organic contaminant bioremediation efforts. Initial char-
acterization work sets the stage for evaluating the pro-
gress of the natural transformation of contaminants. The
following have frequently been observed: parent com-
pound disappearance, active microbial populations with
biotransformation capabilities, and the appearance or
disappearance of organic and inorganic constituents
that provide evidence of bioremediation at contaminated
sites. Quantitative evidence is lacking, however, for net
removal of toxic compounds from complex mixtures
solely by biological processes. This is due largely to the
reliance on monitoring well samples for evidence of
biological activity, rather than on identifying the mass of
contaminants (and total reactive organic carbon) and
estimating the net removal/transformation of reactive
compounds overtime.
A dynamic approach to quantitative site characterization
is needed that recognizes intrinsic bioremediation as an
active cleanup approach. Careful attention must be paid
to the identification of the three-dimensional distribution
of contaminant mass. Then the correspondence be-
tween contaminant distribution and favorable physical,
geochemical, and microbial conditions in the subsurface
over time provides a basis for net contaminant-removal
estimates. Mere adaptations of detective ground-water
monitoring networks are insufficient for quantitative
evaluation of intrinsic bioremediation technologies.
Introduction
The practice of site characterization for remediation of
subsurface organic contaminants has evolved slowly in
the past decade. Early guidelines (1-3) for minimal
ground-water contamination detection monitoring (i.e.,
monitoring wells upgradient and downgradient) have
been applied to many sites of potential concern from
detection through remedial action selection phases.
This minimal approach has been applied widely, regard-
less of the physicochemical characteristics of contami-
nant mixtures or the complexity of hydrogeologic
settings. With soluble inorganic constituents, this ap-
proach may be adequate for detection purposes, but
assessment efforts require substantially more compre-
hensive approaches. For organic contaminant assess-
ment efforts (i.e., determinations of the nature and
extent of contamination), wells alone have been found
to be inadequate monitoring tools. Recognition of the
value of subsurface soil vapor surveys for volatile
organic components of fuel and solvent mixtures has
generated a flurry of modified site characterization
approaches based on monitoring wells (4). These
approaches to site characterization and monitoring net-
work design suffer also from a failure to identify the total
mass of contaminant in the subsurface.
This failure occurs for three main reasons. First, al-
though volatile organic compounds (VOCs) are mobile
in ground water and are frequently early indicators of
plume movement (5), their detection in vapor or well
samples and their apparent aqueous concentration dis-
tribution do not identify the total mass distribution of
organic contaminant (6). Second, efforts to correlate
observed soil vapor or ground-water VOC concentra-
tions with those in subsurface solid cores have often
been unsuccessful, because current bulk jar collec-
tion/refrigeration at 4°C guidelines for solid core sam-
ples for VOC analyses lead to gross negative errors (7).
Third, "snapshots" (i.e., one-time surveys) of back-
ground and disturbed ground-water chemistry condi-
tions have been interpreted as "constant," ignoring
temporal variability in subsurface geochemistry.
The unhappy result of the slow improvement in site
characterization and monitoring practices has often
been the very low probability of detecting the source of
mobile organic contaminants. This outcome may be
followed by the misapplication of risk assessment or
remediation models and fiscal resources. Nonetheless,
good reasons exist for a more optimistic view of the
20
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future reliability of site-characterization and monitoring
efforts.
The shortcomings of previous contaminant detection
and assessment efforts have been recognized. New
guidelines and recommendations for network design
and operations will lead to more comprehensive, cost-
effective site characterization (7,8) in general. Also, ex-
cellent reviews of characterization and long-term
monitoring needs and approaches in support of in situ
remediation efforts should guide us in this regard (9,10).
Site characterization efforts provide a basis for long-
term monitoring design and actually continue throughout
the life of a remediation project.
Advanced Site Characterization and
Monitoring
How do we estimate the potential for subsurface intrinsic
bioremediation success and track its performance into
the future? Clearly, we should seek to design technically
defensible characterization and monitoring networks
that will provide reasonable estimates of in-place con-
taminant distributions over time. Therefore, a dynamic,
ongoing site-characterization effort includes the follow-
ing objectives:
• Identify the spatial distribution of contaminants, particu-
larly their relative fractionation in subsurface solids,
water, and vapor, along potential exposure pathways,
recognizing that the mass of contaminants frequently
resides in the solids.
• Determine the corresponding spatial distribution of total
reactive organic matter (e.g., degradable normal, ali-
phatic, and aromatic hydrocarbon compounds), be-
cause overall microbial activity and disruptions in
subsurface geochemical conditions (and bioremedia-
tion indicators) are due to the total mass of reactive
organic carbon.
• Estimate the temporal stability of hydrogeologic and
geochemical conditions that may favor microbial
transformations in background, source, and down-
gradient zones during the first year of charac-
terization and monitoring.
• Derive initial estimates of net microbial transforma-
tions of contaminant-related organic matter overtime
that may be built into an efficient long-term monitoring
network design.
The first three objectives establish the environment of
major contamination and the conditions under which
bioremediation may occur. The latter two objectives are
vitally important, because evaluating the progress of
intrinsic bioremediation processes depends on distin-
guishing compound "losses" due to dilution, sorption,
and chemical reactions from microbial transformations.
This approach has been suggested emphatically by
Wilson (9) and was recently developed into a draft tech-
nical U.S. Air Force (USAF) protocol by Wiedemeier et
al. (10).
The latter reference focuses directly on the implemen-
tation of intrinsic remediation for dissolved fuel contami-
nation in ground water. The general approach is shown
in Figure 1, which has been modified from the original
work. The draft USAF protocol (10) has as its goals the
collection of data necessary to support:
• Documented loss of contaminants at the field scale.
• The use of chemical analytical data in mass balance
calculations.
• Laboratory microcosm studies using aquifer samples
collected from the site.
These data, if collected in three dimensions for an ex-
tended period, should be sufficient to implement intrinsic
remediation successfully (11). The data collected in the
initial site characterization effort (Figure 1) support the
development of a site-specific conceptual model. This
model is a three-dimensional representation of the
ground-water flow and transport fields based on geo-
logic, hydrologic, climatologic, and geochemical data for
a site. The conceptual model, in turn, can be tested,
refined, and used to determine the suitability of intrinsic
remediation as a risk-management strategy. The validity
of the conceptual model as a decision tool depends on
the complexity of the actual hydrogeologic setting and
contaminant distributions relative to the completeness
of the characterization database. The draft USAF proto-
col is quite comprehensive in identifying important pa-
rameters, inputs, and procedures for data collection and
analysis. The major categories of necessary data are
listed in Table 1 from the draft protocol (10). Ongoing
work on the protocol has revised some of the detailed
guidance it provides on sampling and analytical proto-
cols for these critical parameters; thus, recent drafts of
the protocol should be even more useful to practitioners.
Typical detective monitoring data sets available prior
to in-depth site characterization are more likely to
contain contaminant-related information rather than
the three-dimensional aquifer property, hydro-
geologic, or geochemical data needed to formulate a
conceptual model. A recognition of the variability
inherent in these parameter distributions is critical to
site characterization efforts.
Sampling in Space
The initial site characterization phase should be de-
signed to provide spatially dense coverage of critical
data over volumes corresponding to 10-yr to 100-yr
travel times along ground-water flow paths. If the flow
path intersects a discharge zone in less than 100 yr,
then the volume should be scaled accordingly. For
21
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Review Available
Site Data
Develop Preliminary
Conceptual Model
Make Preliminary
Assessment of Potential
for Intrinsic Remediation
Based on Existing Site
Characterization Data
- Contaminant Type
and Distribution
- Hydrogeology
- Location of Receptors
Perform Site Characterization
in Support of Intrinsic Remediation
Refine Conceptual Model and
Complete Premodeling
Calculations
Document Occurrence of
Intrinsic Remediation and
Model Intrinsic Remediation
Using Numerical Models
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
Evaluate Use of
Other Remedial
Options in
Conjunction With
Intrinsic Remediation
Assess Potential for
Intrinsic Remediation
With Remediation
System Installed
Refine Conceptual Model and
Complete Premodeling
Calculations
Model Intrinsic Remediation
Combined With Remedial
Option Selected Above
Using Numerical Models
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
Yes
Is There
< Unacceptable
Risk to Potential
,. Receptors
No ?.
Examine Temporal/
Spatial Distributions
Site Point-of-Compliance
Monitoring Wells and
Prepare Long-Term
Monitoring Plan
A
V
Prepare Refined
Monitoring Plan
Present Findings
and Long-Term
Monitoring Plan to
Regulatory Agencies
and Reach Agreement
on Monitoring Strategy
Figure 1. Intrinsic remediation flow chart.
22
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Table 1. Site-Specific Parameters To Be Determined During Site Characterization (10)
Fractionation and Spatial Extent
of Contamination
Hydrogeologic and Geochemical
Framework
Extent and type of soil and ground-water contamination
• Location and extent of contaminant source area(s) (i.e., areas containing free- or
residual-phase product)
• Potential for a continuing source due to leaking tanks or pipelines
Ground-water geochemical parameter distributions (Table 2)
• Regional hydrogeology, including:
- Drinking water aquifers
- Regional confining units
• Local and site-specific hydrogeology, including:
- Local drinking water aquifers
- Location of industrial, agricultural, and domestic water wells
- Patterns of aquifer use
- Lithology
- Site stratigraphy, including identification of transmissive and nontransmissive units
- Grain-size distribution (sand versus silt versus clay)
- Aquifer hydraulic conductivity determination and estimates from grain-size distributions
- Ground-water hydraulic information
- Preferential flow paths
- Location and type of surface water bodies
- Areas of local ground-water recharge and discharge
• Definition of potential exposure pathways and receptors
Table 2. Target Constituents for Site Characterization in Support of Intrinsic Bioremediation
Contaminant Mixture Inorganic Constituents
Contamination
Area
Source
Downgradient
Apparent/Geochem-
ical Redox Zone
Reducing
Transitional/Suboxic
Fuels
Chlorinated solvents
Fuels
Chlorinated solvents
02, C02, H2S; pH, Fe2+
HSVS=, N02-, NH3,
alkalinity
02, C02, H2S; pH, Fe2+
alkalinity, NO2-, NO3-,
NH3, HSVS=
Intrinsic Constituents
Organic carbons, CH4,
organic acids, phenols
As above and:
chlorinated metabolites,
ethylene, ethane
Organic carbon, CH4,
organic acids, phenols
As above and:
chlorinated metabolites,
ethylene, ethane
Upgradient/Far-field Oxic
downgradient
Fuels
Chlorinated solvents
02, C02, H2S;
alkalinity, Fe2+, NO3-,
NO2-, NH3
Organic carbon, CH4,
organic acids, phenols
As above and:
chlorinated metabolites,
ethylene, ethane
example, if the flow path discharges after 10 yr, the
critical volume would be 1 yr of travel time. The "volume-
averaged" values of the contaminants, hydrogeological
and geochemical parameters within zones along the
flow path(s), should be derived from data sets that are
large enough to permit estimation of statistical proper-
ties (e.g., mean, median, correlation distance, and vari-
ance). In general, this means that the data sets for
derived mass loadings of contaminants, aquifer proper-
ties, and geochemical constituents (Table 2) derived
from spatial averages of data points must include
approximately 30 or more data points (12-14). Indeed,
this minimum data-set size strictly applies to points in
a plane.
23
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Two major decisions must be made with regard to how
spatially averaged masses of contaminants, electron
donors (e.g., organic carbon, Fe2+, S=, and NH3), and
electron acceptors (e.g., O2, NO3-, NO2-, Fe and Mn
oxides, and SO4=) are to be estimated.
The first question deals with identification of the media
in which the bulk of the constituent's mass resides. For
aquifer properties (e.g., grain size and laboratory esti-
mates of hydraulic conductivity), the answer is simple.
In this case, the solids are clearly the media of interest.
For constituents, particularly VOCs, which are sparingly
water soluble, the bulk of the mass may in fact reside in
the solids, though both solids and water samples must
be collected carefully.
The second question pertains to the depth interval over
which "planar" data points might be averaged. With
fuel-related aromatic contaminants, the depth interval
above and below the capillary fringe/water table inter-
face typically exhibits order-of-magnitude differences in
solid-associated concentrations. In this situation, aver-
aging data points over depths greater than 0.5 m could
easily lead to order-of-magnitude errors in estimated
masses for a site. Continuous coring of subsurface sol-
ids and close interval (i.e., <1 m) sampling of water
should be considered in many VOC investigations. To
approach this level of depth detail in sampling, "push"
technologies and/or multilevel sampling devices present
very useful tools for site characterization. Push tech-
nologies rely on hydraulic or hammer-driven, narrow
diameter (i.e., <2 in.) probes for solid or water sampling.
These technologies have the potential to provide greater
spatial coverage of the subsurface at less cost than
drilling techniques.
The approach to site characterization for chlorinated
hydrocarbons is significantly more difficult. Very few
models of site characterization for these contaminants
have estimated mass loadings in specific media. Many
of the previously referenced methods may work satis-
factorily. Free-phase detection, assessment, and quan-
titation, however, may be more a matter of luck and
exhaustive sampling than intuition based on experience.
Sampling Over Time
VOC compounds (e.g., aromatic hydrocarbons and
chlorinated solvents) are among the target contami-
nants that are considered constituents of concern in
remedial investigations. Their aqueous solubility and
demonstrated association with aquifer solids require
sampling of these media during the site characterization
phase. This suggestion also applies to organic metabo-
lites of complex organic mixtures (e.g., ethylene, vinyl
chloride, aromatic acids, and phenols). Aqueous plumes
that develop subsequent to the release of these organic
mixtures and byproduct compounds have received the
most attention in the past. The fact that the mass of
these contaminants frequently resides in the solids
strongly suggests that the solids should receive the
most attention in the initial site characterization effort.
This should also be the case forthe physical, geochemi-
cal, and microbial determinations.
Initially, conventional nested monitoring wells with
screened lengths of 1 m or more will be useful for
estimating the spatial extent of the dissolved plume, for
delineating apparent geochemical zones, and for provid-
ing data on water level and aquifer property (e.g., slug-
and pump-test derived hydraulic conductivity esti-
mates). Semiannual or annual sampling of wells,
particularly multilevels appropriately designed and com-
pleted, should be quite useful over the course of the
long-term monitoring program. Sampling should track
the downgradient progress of risk-associated target
compounds and permit testing predictions of intrinsic
bioremediation effects on risk reduction.
Proof of the effects of the net removal of specific solid-
associated contaminants due to intrinsic bioremediation,
however, will depend on solid sampling and analysis at
annual or greater intervals, because solid-associated
concentrations may be expected to change slowly. Un-
less biotransformation can be shown to be a major loss
mechanism for contaminants mainly in solids over ex-
tended periods, it will remain an area of research rather
than practice.
Because very few contamination situations have been
monitored intensively for periods exceeding several
years, it is difficult to define specific sampling frequen-
cies for the range of hydrogeologic and contaminant
combinations that may be encountered. The adoption
and future refinement of recently developed, technically
defensible protocols will improve intrinsic remediation
approaches to risk management in subsurface contami-
nation situations.
Acknowledgments
The author would like to express his gratitude to the
following individuals who aided in the preparation of the
manuscript: Dr. Gary Robbins, Mr. Todd H. Wiedemeier,
Dr. John T. Wilson, Dr. Fran Kramer, and Ms. Rebecca
Mullin.
References
1. Scalf, M.R., J.F. McNabb, WJ. Dunlop, R.L. Cosby,
and J.S. Fryberger. 1981. Manual of ground-water
sampling procedures. National Water Well Associa-
tion.
2. Barcelona, J.J., J.P Gibb, J.A. Helfrich, and E.E.
Garske. 1985. Practical guide for ground-water
sampling. Illinois State Water Survey, SWS Con-
tract Report 374 U.S. Environmental Protection
Agency, Ada, OK.
24
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3. U.S. EPA. 1986. RCRAtechnical enforcement guid-
ance document, OSWER-9950.1. Washington, DC.
4. Eklund, B. 1985. Detection of hydrocarbons in
ground water by analysis of shallow soil gas/vapor.
API Publication No. 4394. Washington, DC.
5. Plumb, R.H. 1987. A comparison of ground-water
monitoring data from CERCLA and RCRA sites.
Ground Water Monitor. Rev. 7:94-100.
6. Robbins, G.A. 1989. Influence of using purged and
partially penetrating wells on contaminant detec-
tion, mapping, and modeling. Ground Water 2:155-
162.
7. U.S. EPA. 1992. RCRA ground-water monitoring:
Draft technical guidance document. EPA/530/R-
93/001. Washington, DC.
8. U.S. EPA. 1994. Proceedings of the Ground Water
Sampling Workshop, Dallas, TX, December 8-10,
1993. U.S. Environmental Protection Agency, Ada,
OK.
9. Wilson, J.T. 1993. Testing bioremediation in the field
In: National Research Council. In situ bioremediation—
when does it work? Washington, DC:
Academy Press, pp. 160-184.
National
10. Wiedemeier T.H., D.C. Downey, J.T. Wilson, D.H.
Kampbell, R.N. Miller, and J.E. Hansen. 1994. Draft
technical protocol for implementing the intrinsic re-
mediation (natural attenuation) with long-term moni-
toring option fordissolved-phase fuel contamination
in ground water. Air Force Center for Environmental
Excellence, Brooks AFB, San Antonio, TX (March).
11. National Research Council. 1993. In situ bioreme-
diation—when does it work? Washington, DC: Na-
tional Academy Press.
12. Journel, A.G. 1986. Geostatistics: Models and tools
for the earth sciences. Math. Geol. 18:119-140.
13. Hoeksema, R.J., and PK. Kitanidis. 1985. Analysis
of the spatial structure of properties of selected
aquifers. Water Resour. Res. 21:563-572.
14. Gilbert, R.O., and J.C. Simpson. 1985. Kriging for
estimating spatial patterns of contaminants: Poten-
tial and problems. Environ. Monitor. Assess. 5:113-
135.
25
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Processes Controlling the Distribution of Oil, Air, and Water
John L. Wilson
Department of Geoscience, New Mexico Institute of Mining and Technology,
Socorro, NM
Abstract
Oils and other nonaqueous-phase liquids (NAPLs) are
a major source of dissolved contamination in aquifer
systems. Three major forces control the movement and
distribution of NAPLs, as well as air and water, in both
porous and fractured media. These are capillary, vis-
cous, and buoyancy forces. These forces interact with
the complex pattern of geological features, making fluid
behavior difficult to predict and fluid distribution difficult
to characterize. Intrinsic bioremediation within a NAPL-
contaminated zone is probably of limited effectiveness
because of mass transfer limitations, toxicity concerns,
and the limited availability of other nutrients. Thus, in-
trinsic bioremediation is aimed at the contamination in
the downgradient aqueous phase of the dissolved
plume. Evaluating the effectiveness of bioremediation is
impossible within the plume without knowing the distri-
bution of the NAPL—the plume's source—or determin-
ing whether that source is still moving.
Introduction
Oil or another NAPL, such as gasoline or trichlo-
roethylene (TCE), may be released at or near the
ground surface. These liquids are primary sources for
dissolved contaminant plumes in ground-water systems.
Even if the NAPL has ceased to move, trapped by
capillary forces as we describe below, it remains a long-
term source of dissolved contamination. The limited
aqueous solubility of the chemicals composing these
oily liquids implies that even in small volumes they can
lead to ground-water plumes of enormous dimensions.
Intrinsic bioremediation is aimed at the contamination in
the aqueous-phase plume, for reasons that will become
clear. The plume location and concentrations depend on
understanding the spatial pattern of its source, the
NAPL. In the vadose zone, the problem is complicated
by the presence of a third phase, air, and the propensity
of the NAPL to spread at the air-water interface and to
volatilize.
In this review, we first describe the three major forces
controlling the movement and distribution of fluids in the
subsurface, using natural processes such as infiltration
to illustrate. We then add NAPLs, to relate the discus-
sion to the contamination issue, and aquifer heteroge-
neity, to relate the discussion to natural hydrogeological
systems. Finally, we discuss the implications for intrinsic
bioremediation. Many of these issues are illustrated with
photomicrographs taken of appropriate processes in situ.
Three Major Forces That Control
Processes
Ground-water systems are composed of porous and/or
fractured aquifer material containing water in the spaces
between the solids (see Figure 1a). Above the water
table, in the vadose zone (see Figure 2), air is also
present within this pore space (see Figure 1c). Three
major forces control both the movement and distribution
of each of the fluid phases: capillary forces, viscous
forces, and gravity or buoyancy forces (1).
Capillarity is the result of the cohesive forces within each
fluid phase and the adhesive forces between the solid
phase and each of the fluids. A capillary force is propor-
tional to the interfacial tension at the fluid-fluid interface
Water
Air
Solid
(a) Water Saturated (b) Residual Gas (c) Continuous Gas
Figure 1. Diagram of fluid saturation in a porous media (4).
26
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and the strength of fluid wetting to the solid surface, and
inversely proportional to the pore size. In the vadose
zone, with air and water present, interfacial tension is
the same as air-water surface tension. In the saturated
zone, beneath the water table (see Figure 2), the only
fluid is water, and capillary forces are absent (see Figure
1a). The exception occurs beneath a fluctuating water
table, where gas bubbles may become entrapped by
capillary forces (see Figure 1b and Figure 3; Figure 3 is
a photomicrograph from a visualization experiment [see
also the appendix to this paper]). For most aquifer ma-
terials, water is the wetting fluid; that is, the solid aquifer
material has a greater affinity for water than for air, and
the air or gas is almost always the nonwetting phase
(see Figures 1b and 1c). Water occurs as an intercon-
nected film or layer of wetting liquid covering and con-
necting the solids in the vadose zone. The nonwetting
gas phase occupies the larger pores.
Viscous or flowing forces within a fluid phase require an
expenditure of potential energy. For example, in the
saturated zone the ground-water flow rate is propor-
tional to the hydraulic gradient. The flow rate is also a
function of the aquifer material and the structure of its
pore space, as represented by the effective permeabil-
ity. If more than one fluid phase is present, as in Figures
1b and 1c, the interconnected paths in each phase are
more tortuous. The permeability of each phase is re-
duced, with a relative permeability closer to zero than to
one, the value that applies when only a single fluid
occupies the pore space.
Buoyancy is a gravitational force proportional to the
density difference between two fluids. The gas phase
has a much lower density than water, so that gravity
(buoyancy) forces play a significant role in the vadose
zone. Water infiltrates downward, toward the water ta-
ble, primarily under the influence of gravity. As it moves
downward, water easily displaces the less dense and
less viscous air phase. In the saturated zone, with only
water present, gravity usually plays no direct role
(although, of course, it ultimately drives the hydraulic
gradient). In the saturated zone, if the chemical concen-
tration varies enough to influence water density, gravity
can again play a direct role, causing the water and its
chemicals to move.
Oil and Other NAPLs
A third fluid is often present at many hazardous waste
sites and most leaking underground storage tanks.
NAPLs often share the pore space with both gas and
Hazardous Waste Site
Ground
Surface
Water Table
KxraisraiOTKttBttirarasiOTBaHttarora
^•HWWBU UM ouu «»j i-x MA m w*BHBBSOTS*< w«odSHRfllfiflR!HHWfl Vn4Rffl49M])W!ffln]K!ffl1!nBiK4W1fflKHUSHH0!fW KOCK Sfllfif
..mn^i?s^ Dense NAPL
j mat mss ass saa vast BBS SEE BHU ssa am OTS sas a«a« saa imittttitt^ttttttttttimtttttttt^ttis&WRVissivs&sissiv
Figure 2. Diagram of the saturated and vadose zones, showing the migration pattern for a NAPL more dense than water (left), and
less dense than water (right) (3).
27
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Figure 3. Photograph of a micromodel with entrapped air bub-
bles. The bubbles are trapped in the pore bodies,
which are roughly 200 urn in diameter, while the pore
throats are 50 urn in diameter (4).
water, in the vadose zone, or water alone, in the satu-
rated zone. The NAPL may be more dense than water
(a dense NAPL, or DNAPL), such as TCE, or it may be
lighter than water, such as gasoline. In either case, a
NAPL is heavier than air, so it easily moves downward
through the vadose zone. Once it reaches the vicinity of
the water table, it may continue to move deeper if it is
denser than water (DNAPL), or it may remain in the
vicinity of the water table if it is lighter, as illustrated in
Figure 2 (2,3); in the latter case, the term "water table"
begins to lose its meaning. In either case, the NAPL
leaves behind a trail of residual as it moves.
Most NAPLs appear to be of intermediate wettability in
typical aquifer materials; that is, they are nonwetting
relative to water but wetting relative to gas. Thus, de-
pending on whether it encounters gas or water, a NAPL
can then display either wetting or nonwetting behavior
or both. Many NAPLs, such as gasoline, also have low
internal cohesion and will spread at a gas-water inter-
face, presumably forming a film between the water
phase, which because of capillarity preferentially occu-
pies the smallest pores, and the gas phase, which pref-
erentially fills the largest pores. In the vadose zone, this
film interconnects the pockets of NAPL, which even at
residual saturation should be largely continuous, as
shown in Figure 4. Some NAPLs, such as TCE, have
more internal cohesion and do not spread. In this case,
the sum of the interfacial tensions between the NAPL
and the water and air, oow + oao, exceed the surface
tension between the gas and water, oaw (see Figure 5,
theta>0). These NAPLs will not spread as films. On a
flat water surface (the gas-water interface), nonspread-
ing liquids tend to coalesce into lenses that float on the
surface (much as depicted in Figure 5), even though
many of these liquids are denser than water. In porous
media, this leads to complications, as shown in Figure 6.
In the saturated zone, the NAPL fills the larger pores,
while water occupies the smaller pores and lies as a film
Figure 4. A micromodel after a spreading NAPL (Soltrol 130) is
drained with air. Air fills the pore bodies and is inter-
connected through some of the pore throats. Water
fills the other pore throats and some of the pore
wedges and is a wetting film everywhere else. The
NAPL forms a thick film filling some pore wedges and
surrounding the air everywhere. This ubiquitous and
interconnected film is particularly thick near the
water-filled pore throats (3).
Gas (air
NAPL
Lens
Figure 5. Cross-sectional diagram of spreading potential for a
drop of NAPL floating on the air (gas)-water interface,
after Adamson (10) and others. The water is wetting,
the gas is nonwetting, and the NAPL is intermediate
wetting.
between the NAPL and the solid surfaces. After the
NAPL moves on, it leaves behind a residual saturation
trapped by capillary forces, which occupies the larger
pores. NAPL blobs occupying one or two pore bodies
are shown in Figure 7. A similar process occurs in the
vadose zone, when a rising water table entraps air
bubbles, as shown in Figures 2b and 3. Because of
lower interfacial tension, NAPL blobs are often much
more complex than air bubbles, forming interconnected
groups of larger pore bodies and pore throats, as shown
in Figure 8.
So far we have looked at water-wet porous media.
Wetting, however, may change for a variety of reasons,
including the adsorption at hazardous waste sites of
(polar) organic compounds onto the solid. Figure 9
shows the saturated zone residual in a pore space with
28
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NAPL
teese-i
V
Figure 6. A micromodel after a nonspreading DNAPL (mineral
oil) is drained with air. The DNAPL is not connected.
As it leaves the photo at the bottom, the air is sur-
rounded by a film of oil, similar to that seen with the
spreading oil. As the air leaves to the left, the film no
longer exists; it truncates somewhere in between. In
this area, the air-water interface is dimpled, with small
lenses of mineral oil. These lenses are not intercon-
nected. The various pockets of oil, such as on the
right side of the figure, are also generally discontinu-
ous and not connected (11).
Figure 7. A micromodel with a singlet and doublet NAPL blob
(dark gray) in a water-filled pore space (lighter gray)
of an irregular pore network. The pore bodies are up
to 1,000 urn across, and the model was 10 cm x 15 cm
in area (8).
altered wetting and significant wetting hysteresis. The
residual NAPL is not trapped in the pore bodies, the
characteristic location for a nonwetting fluid. Some of it
is in the pore throats, some in the pore bodies, some
has an end in both. Because of the large wetting hyster-
esis of this surface, water wet conditions prevailed dur-
ing the invasion of the NAPL, while intermediate wet
conditions prevailed when the water reentered to trap
the NAPL. The NAPL-water interfaces in the figure
clearly show evidence of both histories. In some loca-
tions, clearly water wet, we can infer that the interface
was not disturbed as the water reentered. In other loca-
tions, the contact angle shows the recent displacement
of the NAPL by water.
Heterogeneity
All of this is complicated by the geology of the ground-
water system. The material composing an aquifer is
always heterogeneous. For example, fluvial-deposited
materials contain sand, gravel, and clay in a complex
geometric pattern of geologic fades. Many formations,
even those containing clay, are fractured and sometimes
faulted. Heterogeneities provide preferential paths for
fluid and chemical migration in an interplay of the geol-
ogy with the forces of capillarity, viscosity, and gravity.
For example, a DNAPL reaching the saturated zone will
tend to seek out the heterogeneities with larger pore
spaces such as the coarser sands or, more insidiously,
fractures where its movement will be hard to trace.
Heavier than water, it tends to move farther downward
into the aquifer, fingering out the bottom of a sand lens
or moving snakelike down the inside of a fracture.
Figure 8. A micromodel with a large, branched, nonaqueous,
nonwetting phase blob (dark gray) in an irregular pore
network (3).
Residual nonaqueous-phase saturation in a micro-
model treated with an alkoxysilane of proprietary
composition, GlassClad 18 (Huls Americal Petrarch,
Bristol, PA). This treated surface has significant con-
tact angle hysteresis. The water is light gray, and the
NAPL is darker gray in this image (9).
29
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Figure 10 shows a pore system with embedded zones
of larger pores surrounded by a network of smaller
pores. This simulates lenses of coarse porous materials
(e.g., coarse sand) embedded in a matrix of finer mate-
rial (e.g., fine sand). The NAPL invaded from the side,
mimicking the lateral migration of a NAPL plume along
the bottom of an aquifer. Afterwards, the NAPL was
displaced by water, leaving behind a residual saturation.
The photograph in Figure 10 was taken at this time. The
residual is dominated by the coarse lenses. In the fine
material, we see the typical single- and multiple-pore body
blobs, but the coarse lenses have been completely by-
passed. Capillary forces are too strong, and the viscous
forces in the moving water too weak, to displace the NAPL
from these isolated coarser zones. Note that the zones
with higher saturated permeability trap the NAPL.
This experiment was repeated at a much higher water
flow rate. A photograph taken while this was occurring
is shown in Figure 11. A significant amount of NAPL has
been swept from the coarse lenses, even though the
initial condition was essentially the same. In the swept
area, on the right side, there is still some bypassed
NAPL, but it is now located on the downstream end of
the lenses. Sufficient viscous forces were generated to
overcome the capillary forces that held NAPL in the
coarse lenses. The displacement is incomplete be-
cause, as the wetting front reaches the end of a lens, it
closes together and surrounds the lens. At that point, the
nonwetting NAPL in the lens is no longer connected with
the downstream NAPL. These two photographs suggest
that in heterogeneous systems the residual nonwetting-
phase saturation is a function of the structure of the
heterogeneities and the fluid history. Because the resid-
ual is the source of the dissolved plume, we would like
to be able to predict its location and magnitude from
knowledge of the geology and the fluid flow history. We
use mathematical models to assist with this kind of
prediction. There is a problem, however. The current
generation of multiphase flow models assumes that
there is a single value of residual and, worse, that we
know what it is. These models offer little assistance with
the important issue of predicting the amount and distri-
bution of residual.
These photographs were taken from experiments that
simulated the movement of DNAPL parallel to strati-
graphic bedding planes—for instance, as it migrates
along the bottom of an aquifer. What happens as it
moves perpendicular to the bedding, for example, dur-
ing its movement downward from the water table toward
the bottom of the aquifer? Figure 12 shows the same
kind of heterogeneous lenses, except for this change in
geometry. The flow is downward, across the lenses, with
both gravity and viscous components. As the DNAPL
encounters its first coarse lens, it tends to spread hori-
zontally because of capillary forces. As soon as it fills a
lens, it begins to spill downward, but because of the
Figure 10. Water has swept through a heterogeneous micro-
model containing a NAPL as a nonwetting fluid and
water as a residual wetting fluid. The nonwetting
NAPL in the isolated coarse-grained heterogeneities
has been bypassed (3).
Figure 11. Water is displacing the NAPL at a higher flow rate
through the same heterogeneous micromodel. The
flow is from right to left. With larger viscous forces,
the displacement is more efficient, and less NAPL is
bypassed. This photograph was taken while the dis-
placement was still under way. The bypassing is
complete on the right side of the model and has not
yet occurred at the left side (3).
gravity instabilities—DNAPL is heavier than water—this
movement occurs as fingers on the scale of a few pores.
When one of these fingers encounters another, deeper
coarse lens, the downward migration of the DNAPL is
again arrested, until the lens is filled. In essence, the
lenses promote horizontal spreading, here confined by
the lateral boundaries of the model, and vertical finger-
ing. The process leads to a very inefficient displace-
ment, with lots of water bypassed by the DNAPL, and
illustrates some of the reasons why it is difficult to di-
rectly measure DNAPLs in the field, predict their behav-
ior, or design effective remediation measures.
Influence on Intrinsic Bioremediation
The distribution of fluid and solid phases controls the
mass transfer and transport of chemicals and their
30
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Figure 12. DNAPL is moving downward through a water-
saturated heterogeneous micromodel (9).
availability for biotransformation by the microbial popu-
lation. For all practical purposes, biotransformation
takes place in the aqueous phase or in contact with the
aqueous phase. Bacteria colonize surfaces. In water-
saturated media, the only surfaces available are the
solids, but in the vadose zone and in NAPL-contami-
nated areas there are also fluid-fluid (e.g., air-water,
oil-water) interfaces that some bacteria will colonize.
These colonies develop in part due to the same inter-
facial forces that attract abiotic colloids to solid-liquid or
fluid-fluid interfaces. Figure 13 shows fluorescing col-
loids that are repulsed from the solid surface by electro-
static forces but that are attracted to an air-water
interface by hydrophobic and capillary forces. Similar
observations have been made for microorganisms (4,
5). Other important processes also lead to bacterial
attachment and colonization at interfaces that are
unique to living bacteria, with their complex surface
chemistry and ability to produce extracellular polymers.
Where NAPLs are concerned, there may also be toxicity
effects repelling the bacteria, or, if the NAPL is an im-
portant nutrient source, it may be an attractant. Figure
14 shows a bacteria colony in the vicinity of a blob of
iso-octane. The bacteria have colonized the solid surface
and the surface of the blob, and a few were even observed
to be living in the iso-octane. In a similar experiment with
toluene and a toluene-degrading bacteria, the bacteria
found high concentrations of toluene toxic and set up
housekeeping at a distance that allowed ambient flow and
mixing to dilute the concentrations many times.
The geometry of the fluid distributions have other indi-
rect influences on intrinsic bioremediation. Consider
Figure 13. Negatively charged hydrophilic polystyrene parti-
cles attached to the surface of gas bubbles en-
trapped in a micromodel. There is almost no
sorption on the glass surface at this ionic strength
(4).
Figure 14. A colony of bacteria (Arthrobacter sp. ZAL001) in
the vicinity of an iso-octane blob after 48 hr of
growth. The blob is trapped in a pore body. The
bacteria are light gray (5).
mass transfer of chemical components between a multi-
component residual NAPL and the water phase in the
saturated zone. Many spilled NAPLs—from gasoline
and other fossil fuels to transformer oils containing poly-
chlorinated biphenyls (PCBs), as well as the mixed bag
of organics that are sometimes found in industrial waste
pits—are mixtures having many (sometimes hundreds
of) organic chemical components. The more soluble
components of these mixtures can dissolve compara-
tively quickly, leaving behind less soluble components
to leach out more slowly.
There are also cosolvency effects to consider. Interac-
tion between components can either enhance the solu-
bility of a given component (cosolvency) or reduce the
solubility of that component (by a kind of salting-out
process). Capillary-trapped residual blob size and shape
influence the partitioning of NAPL components to the
aqueous phase. Mass transfer coefficients used in the
mathematical models of this partitioning often employ
31
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the analogy of an equivalent spherical blob. Certainly
singlet blobs can be represented by a spherical model,
but it is less clear that this model works for the tortuous
multiple pore-body elongated or branched blobs shown
in Figure 8, or in the presence of heterogeneity shown
in Figure 10. Large complex blobs hold the majority of
the NAPL volume, and mass transfer from them to the
aqueous phase is clearly rate limited (6-8); it can take
decades under natural conditions. Intrinsic bioremedia-
tion can do little to speed up this mass-transfer process,
except to possibly increase concentration gradients be-
tween the NAPL and the aqueous phase, or perhaps
increase the solubility of chemical components through
the production of biosurfactants.
Even in the absence of toxicity effects and mass-transfer
limitations, there may be little intrinsic bioremediation
within the NAPL zone. The aqueous-phase flow field is
complicated by the presence of the NAPL. Among other
things, the effective permeability to water is lower. Con-
sequently, it is difficult to supply the nutrients necessary
for growth (3,8).
For these reasons, intrinsic bioremediation of the NAPL
itself may be of limited value. If instead we consider the
downgradient plume of dissolved contamination, the ba-
sic concern with the NAPL is as a source of that con-
tamination.
To understand the plume, we must know something
about the spacial distribution of the source and deter-
mine whether or not it is still moving. This both locates
the source and controls the effective mass-transfer rate
between the source and the ground water. If the source
is a NAPL, it is obvious from some of the earlier photo-
graphs that even a small amount of liquid may bespread
throughout a relatively large volume of aquifer material,
and in a most complex pattern (see Figures 11 and 12).
Characterization of this distribution and movement is
difficult and may be beyond the current state of the art.
Summary
Three major forces control the movement and distribution
of fluids in the vadose and saturated zones: capillary,
viscous, and buoyancy forces. These forces interact
with the complex pattern of geological features, making
fluid behavior difficult to predict and fluid distribution
difficult to characterize. Wetting fluids have an affinity for
the solid surfaces of aquifer materials. In general,
NAPLs are wetting relative to air, but nonwetting relative
to water. Consequently, they tend to fill the larger pores
of the saturated zone, leading to entrapment in what are
usually regarded as high-permeability materials: the
coarse sand but not the fine silt. This tendency, together
with the propensity for gravity fingering through the finer
materials, causes a complex pattern of migration and
distribution in the saturated zone. In the vadose zone,
behavior is further complicated by the presence of air
and the question of whether or not the NAPL tends to
spread at the air-water interface. Wettability can be
altered by sorption processes, changing the relative
roles of water and the NAPLs and leading to new path-
ways and fluid distributions.
Intrinsic bioremediation within a NAPL-contaminated
zone is probably of limited effectiveness. Mass-transfer
limitations constrain the rate at which dissolved compo-
nents reach the aqueous phase. There are toxicity con-
cerns with the high concentrations that are found nearer
the NAPL blobs or films. Some other nutrients are only
limited in availability because of the more complex flow
patterns that water must take in this region.
Consequently, intrinsic bioremediation should probably
be aimed at the contamination in the downgradient
aqueous phase of the dissolved plume. It is impossible
to evaluate the effectiveness of bioremediation within
the plume without knowing the answer to two questions:
Where is the source? and, How much contamination is
entering the flowing ground water from the source? The
answer to both questions requires that we know the
distribution of the NAPL—the plume's source—and de-
termine whether that source is still moving. It is a direct
answer to the "where" question. The "how much" ques-
tion depends on the effective mass-transfer rate be-
tween the source and the passing ground water. The
trouble is that, even in a geologically simple aquifer, a
small amount of nonaqueous contamination may be
spread throughout a relatively large volume of aquifer
material, and in the most complex pattern. Charac-
terization of this distribution and movement is difficult
and may be beyond the current state of the art.
Appendix: Visualization of Processes in
Micromodels
Micromodels are transparent physical models of a pore
space network, created by etching a pattern onto two
glass plates which are then fused together (3,8,9). The
resulting pores have complex three-dimensional struc-
tures, although the network is only two dimensional. The
micromodel in Figure 3 shows pore bodies connected
together by narrower pore throats. The fused glass lo-
cated in between these connected channels represents
the solid material in this model of a porous media. Glass
is also at the top of the channels, toward the viewer, and
below the channels. When the pore body contains a
nonwetting fluid, a thin water film is attached to the top
and bottom of the pore body. The film is thin enough to
partially exclude the bacteria from colonizing the top
surface of the iso-octane blob in Figure 14. All of the
photographs in this paper can be found collected to-
gether and described in more detail in Wilson (9).
Greater detail is given in the original references.
32
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Acknowledgments
The photographs presented in this paper were taken
from work sponsored by the Subsurface Science Pro-
gram of the Department of Energy, or from work spon-
sored by the Robert S. Kerr Environmental Research
Laboratory of the U.S. Environmental Protection
Agency. M. Flinsch of New Mexico Institute of Mining
and Technology assisted with the preparation of this
manuscript.
References
1. Bear, J. 1972. Fluid flow in porous media. New York:
Elsevier.
2. Mercer, J.W., and R.M. Cohen. 1990. A review of
immiscible fluids in the subsurface: Properties,
models, characterization and remediation. J. Con-
tarn. Hydrol. 6:107-163.
3. Wilson, J.L., S.H. Conrad, E. Hagan, W.R. Mason,
W. Peplinski, and E. Hagan. 1990. Laboratory in-
vestigation of residual liquid organics. Report CR-
813571. U.S. Environmental Protection Agency,
Ada, OK.
4. Wan, J., and J.L. Wilson. 1994. Visualization of the
role of the gas-water interface on the fate and trans-
port of colloids in porous media. Water Resour. Res.
30:11-23.
5. Wan, J., J.L. Wilson, and T.L Kieft. 1994. The gas-
water interface as an influence on the transport of
microorganisms through unsaturated porous me-
dia. Appl. Environ. Microbiol. 60:509-516.
6. Miller, C.T., M.M. Poirier-McNeill, and A.S. Mayer.
1990. Dissolution of trapped nonaqueous phase
liquids: Mass transfer characteristics. Water Re-
sour. Res. 26:2783-2796.
7. Powers, S.E., C.O. Louriero, L.M. Abriola, and WJ.
Weber. 1991. Theoretical study of the significance
of nonequilibrium dissolution of NAPLs in subsur-
face systems. Water Resour. Res. 27:463-478.
8. Conrad, S.H., J.L. Wilson, W.R. Mason, and W.
Peplinski. 1992. Visualization of residual organic
liquid trapped in aquifers. Water Resour. Res.
28:467-478.
9. Wilson, J.L. 1994. Visualization of flow and trans-
port at the pore level. In: Dracos, T.H., and F. Stauf-
fer, eds. Transport and reactive processes in
aquifers. Rotterdam: Balkema. pp. 19-36.
10. Adamson, A.W 1982. Physical chemistry of sur-
faces, 4th ed. New York: Wiley and Sons.
11. Wilson, J.L. 1992. Pore scale behavior of spreading
and nonspreading organic liquids in the vadose
zone. In: Weye, U., ed. Subsurface contamination
by immiscible fluids. Rotterdam: Balkema. pp. 107-
114.
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New Tools To Locate and Characterize Oil Spills in Aquifers
Bruce J. Nielsen
Environics Directorate, Armstrong Laboratory, Tyndall AFB, FL
Abstract
The vision of the Tri-Services (Air Force, Army, and
Navy) scientists is about to become reality, as a partner-
ship between the Department of Defense (DOD), aca-
demia, and private industry evolves into a combined
technology that can save millions of dollars in long-term
hazardous waste site cleanup costs.
DOD has about 20,000 contaminated sites, many of
which require characterization. These sites also may
require monitoring for more than 30 years, at a cost of
millions of dollars per year. The cost of site charac-
terization and monitoring can be 25 percent to 30 per-
cent of the total remediation costs.
A quick fix is needed—one that would take a "snapshot"
of the type and amount of contamination, help determine
the best method of remediation, monitor the effective-
ness of treatment, and let technicians and engineers
know when the site is clean. The vision was that of a
cost-effective method that would replace the current
time- and labor-intensive processes of site charac-
terization and monitoring.
To solve the problem, the Tri-Services integrated a laser
spectrometer with fiberoptics with a cone penetrometer
to provide the required "snapshot" of subsurface condi-
tions. To date, this technology has been field-tested at
numerous Air Force, Army, and Navy installations. On-
going studies will extend its sensitivities; expand its
capability to other contaminants such as solvents, met-
als, and explosives; and make it more user-friendly for
operating technicians. The end result is a technology
that can significantly reduce the cost of site charac-
terization and monitoring.
Introduction
The Air Force's Armstrong Laboratory is executing a
program to demonstrate, test, and evaluate the applica-
tion of cone penetrometers in support of intrinsic biore-
mediation (natural attenuation) demonstrations. The Air
Force and its contractors, in cooperation with the U.S.
Environmental Protection Agency's (EPA's) Robert S.
Kerr Environmental Research Laboratory, have con-
ducted investigations at Tinker Air Force Base (AFB),
Oklahoma; Plattsburgh AFB, New York; Patrick AFB,
Florida; and Dover AFB, Delaware. One objective of the
investigations was to assess the subsurface conditions,
obtaining parameters for input into BIOPLUME® II, a
numerical model useful for determining the potential for
intrinsic remediation. A second objective was to validate
the cone penetrometer/laser spectrometer system by
soil sampling and analysis. The demonstrations proved
the cone penetrometer's capabilities to quickly locate
and define the areal and vertical extent of the liquid
oily-phase plume using laser-induced fluorescence
(LIF), and to rapidly install monitoring points and collect
soil samples to provide additional data necessary to
define the dissolved-phase plume.
This paper briefly describes the Tri-Services program for
developing this cone penetrometer technology and re-
cent results of the Air Force program.
Methods
Cone Penetrometry
The Site Characterization and Analysis Penetrometer
System (SCAPS), developed jointly by the Tri-Services,
has proven to be an effective technology for charac-
terizing contaminated DOD sites. The Army has pro-
vided leadership on developing SCAPS, including the
concept of using a sapphire window in the cone rod and
of using fiberoptics and spectroscopy to analyze the soil
for contamination. The Army has patented a "Device for
Measuring Reflectance and Fluorescence of In SituSo\\"
and is now licensing it.
The typical cone penetrometer is mounted on a 20-ton
truck and driven to the site requiring characterization,
where a conical rod is hydraulically pushed into the
ground to be characterized. The rod is equipped with a
34
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variety of sensors or soil and ground-water sampling
tools. The cone penetrometer can characterize several
aspects of the subsurface, depending on the types of
sensors integrated into the penetrometer. Strain gauges
measure the forces against the tip and sleeve of the
cone tool, allowing determination of soil type (e.g., sand,
silt, clay) and stratification. Other sensors provide elec-
trical resistivity, pore pressure, spectral characteristics,
and other properties of the soil and contamination. The
sensors provide information on hydrogeology and con-
tamination; the samplers verify it. The real-time ability to
receive and assess monitoring data on site without labo-
ratory analysis is critical, facilitating decision-making
during site investigation projects while ensuring accu-
rate and efficient completion of site investigations and
optimization of remedial activities.
Laser Spectrometer System
Each of the services has significant programs for devel-
oping and demonstrating laser spectroscopy and other
sensor systems. One of the key components of the cone
penetrometer is a neodymium:yttrium aluminum garnet
(Nd:YAG) laser, which pumps a dye laser system to
induce fluorescence of fuel products as the cone pene-
trometer probe is advanced into soils. LIF has been
shown to be useful in identifying petroleum, oil, and
lubricant (POL) contamination, such as gasoline and
JP-4 jet fuel. The Armstrong Laboratory's Environics
Directorate, working with North Dakota State University
(NDSU), has developed a tunable laser/fiberoptic spec-
trometer system, which uses laser-generated ultraviolet
light, optical fibers, and spectroscopy for hazardous
waste site monitoring. The basic detection approach
takes advantage of the fact that certain substances
fluoresce when a particular wavelength of light shines
on them. The spectral emission, including fluorescent
lifetime, is somewhat like a fingerprint and therefore is
useful in identifying the contaminant. The fluorescent
intensity indicates concentration of the contaminant.
The transportable laser system is unique because its
output may be tuned to the optimum frequency for de-
tecting the pollutants of interest.
Optical fibers are used to transmit ultraviolet light to
subsurface monitoring points and return resulting light
for the spectroscopic analysis. The system can identify
aromatic hydrocarbons such as benzene, toluene, and
xylene (BTX) and naphthalene by their fluorescent spec-
tra. Jet fuel, which contains naphthalene and BTX, is the
most common contaminant at Air Force sites.
This technology can provide semiqualitative and
semiquantitative information, on site, in minutes. The
LIF response can be correlated to the total petroleum
hydrocarbon (TPH) concentration within the soil. The
system has been tested in the field with detection limits
as low as parts-per-million levels on soil when used with
a cone penetrometer and in the laboratory at parts-per-
billion levels for BTX in water using fiberoptic probes.
Laser spectroscopy technology could also be used to
monitor the progress of site remediation and to provide
baseline data for intrinsic bioremediation modeling studies.
Technology Transition
Armstrong Laboratory and Unisys Corp. signed a Coop-
erative Research and Development Agreement (CRADA)
to commercialize the Air Force-developed laser spec-
trometer system. The laser spectrometer was initially
developed for making ground-water measurements in
monitoring wells or in implanted monitoring points. A
consortia consisting of the CRADA partners, Dakota
Technologies Inc., and NDSU submitted a proposal to
the Advanced Research Projects Agency (ARPA), Tech-
nology Reinvestment Project (TRP). ARPA selected the
proposal to receive a 2-year, $1.6 million grant with two
1-year options for follow-on technology development.
The industry partners provide in-kind contributions and
matching funds.
The Rapid Optical Screening Tool (ROST) is the pro-
posed product from the commercialization of the laser
spectrometer developed for the Air Force by NDSU.
ROST will build upon the previous Environics Director-
ate research by automating the collection and mapping
of data, making equipment components smaller and
more rugged, and developing a more user-friendly inter-
face to allow use by environmental technicians involved
in site characterization and cleanup. ROST also has
potential for process monitoring and for medical diag-
nostics. Initial commercial use will be with cone pene-
trometers for soil characterization.
Use of the proposed ROST technology should result in
substantial savings in costs associated with charac-
terization, monitoring, and remediation of hazardous
waste sites. The participants are committed to commer-
cializing the resulting instrumentation for worldwide
sales by U.S. firms or companies. In short, DOD will
benefit from the technology and knowledge gained; the
private sector will receive a highly transferable and prof-
itable technology; the U.S. economy will be helped; and
all will benefit from a cleaner environment.
Combined Technologies
The combined cone penetrometer and transportable la-
ser spectrometer was demonstrated at Air Force instal-
lations having fuel-contaminated areas. To date, the
laser spectroscopy system's primary function with the
cone penetrometer has been to define the oily-phase
35
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plume. At Tinker AFB, the tunable laser system was
configured to optimize the system for jet fuel and heating
oil, the known petroleum contaminants. Laboratory
fluorescence spectra from these fuels suggest that
naphthalene produces the maximum fluorescence; con-
sequently, a laser excitation wavelength appropriate for
the known fuels was utilized during the field program.
The system is designed to collect data in two different
modes: "push" or "static." In the push mode, laser exci-
tation frequency is fixed, and the LIF signal is monitored
as the cone penetrometer probe is advanced. Operation
in the static mode, or with the probe stopped, allows
collection of LIF multidimensional data sets, typically
fluorescence emission wavelength, intensity, and time of
decay matrices (WTM). WTMs have proven to be very
useful in identifying various fuels.
Results
Studies to demonstrate site amenability towards intrinsic
demonstrations were conducted at Plattsburgh, Patrick,
and Dover AFBs. The sites selected for the demonstra-
tions included closed fire-fighter training areas, gasoline
service stations, and aircraft hydrant fueling systems.
The old fire-fighter training areas were composed of
unlined pits used to ignite fuels such as JP-4, waste oils,
and other flammable substances. The service stations
or hydrant fueling systems had leaking underground
storage tanks or piping. Substances from these opera-
tions have percolated through the vadose zone into the
unconfined aquifer. The penetrometer determined the
areal extent and volume of oily-phase contamination,
and obtained ground-water and soil core samples. To
determine site amenability, acquired data were then fed
into BIOPLUME® II, a computer model for in situ con-
taminant biodegradation. The cone penetrometer sys-
tem rapidly located and defined the leading edge of the
oily-phase petroleum plume. The technology proved
that it can be used to provide timely and accurate data
for intrinsic bioremediation modeling.
The Tri-Services conducted a series of laboratory tests,
and some of the preliminary results are calibration
curves with different fuels on various soil matrices. The
calibration curve obtained in the laboratory for diesel fuel
marine (DFM) on a sand matrix indicates a detection
limit that is lower than 30 mg/kg (ppm) (Figure 1). The
collection of LIF multidimensional data sets (fluores-
cence emission wavelength, intensity, and lifetime) or
WTMs for diesel #2, JP-4, unleaded gasoline, and diesel
fuel marine show how each one has a characteristic
pattern. These patterns make possible reliable fuel-type
identification without the need for bringing samples to
the surface (Figure 2). In the field, the LIF count meas-
urements can be correlated with collected samples and
100,000
10,000
-as- 1,000
Q.
I 10°
10
1
10,000
1 10 100 1,000
Concentration DFM (ppm)
Figure 1. Calibration curve for diesel fuel marine on fisher
sand.
analytical results. To assist in the correlation, several
WTMs were conducted at various depths. Color-coded
WTMs from the North Tank Area (NTA) and Fuel Purge
Area (FPA) at Tinker AFB indicate different fuel types.
The shapes of these spectra identify the contaminants
as fuel oil at the NTA and JP-4 at the FPA (Figure 3).
The fluorescence versus depth profile from push loca-
tion 84L at Plattsburgh AFB indicates narrow bands of
contamination are in the "oily phase," which rests just
above the water table. Note that discrete sampling at 5-ft
intervals (25, 30, 35 ft, etc.) could easily skip over the
contamination (Figure 4). A series of fluorescence ver-
sus depth profiles taken across a north-south transect
at the Plattsburgh AFB fire-fighter training area show the
extent of contamination (Figures 5 and 6). Location 84A
is upgradient, 84D is in the center of the burn pit, and
the remainder are downgradient. The contamination
traveled directly down from the burn pit and then along
the water table.
Conclusions
Currently, these technologies are being further devel-
oped and demonstrated within numerous DOD, Depart-
ment of Energy, and EPA programs. Ongoing research
will develop techniques to monitor contaminants such as
chlorinated solvents, metals, and explosives that do not
naturally fluoresce. Refining this technology is at the
heart of site remediation because it provides cost-effec-
tive characterization before, during, and after remedia-
tion. We can use it to determine if remediation is
needed, what remediation technology we should apply,
whether the remediation is working, and whether the
cleanup effort has been successful—all with a minimum
of risk, time, labor, and cost.
36
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Diesel #2
Diesel Fuel Marine
V
E
300 350 400 450
Wavelength (nm)
JP-4
500
300 350 400 450
Wavelength (nm)
Figure 2. Wavelength time matrices of various fuels.
500
300 350 400 450
Wavelength (nm)
Unleaded Gasoline
500
300 350 400 450
Wavelength (nm)
500
References
1. Bratton, W.L., J.D. Shinn, S.M. Timian, G. Gillispie,
and R. St. Germain. 1993. The Air Force Site Char-
acterization and Analysis Penetrometer System
(AFSCAPS), Vols. I-V. AL/EQ-1993-0009.
2. Gildea, M.L., W.L. Bratton, J.D. Shinn, G. Gillispie,
and R. St. Germain. 1994. Demonstration of the Air
Force Site Characterization and Analysis Pene-
trometer System (AFSCAPS) in support of the intrin-
sic bioremediation (natural attenuation) option
(interim report).
3. Schroeder, J.D., and S.R. Booth. 1991. Cost effec-
tiveness analysis of the Site Characterization and
Analysis Penetrometer System (SCAPS). Los
Alamos National Laboratory, Los Alamos, NM.
37
-------
325 350
375 400 425
Wavelength (nm)
450 475 500
325 350
450 475 500
375 400 425
Wavelength (nm)
Figure 3. Wavelength time matrices at two different sites:
North Tank Area versus Fuel Purge Area.
&
35
40
5 10 15 20 25
Equivalent (% of JP-4 standard)
30
Figure 4. Fluorescent versus depth profile from fire training
facility.
38
-------
CRT
Monitoring Well
Survey Point
Figure 5. Base map of the fire training area showing location of CPT soundings and wells.
260
240
c
o
ro 220
0)
200
180
Fuel Pit
84D-LIF
Section A - A'
84L-LIF
Leading Edge of Oily Phase
84F-LIF 84K-LIF
(J3 GrSand
CJ Sand
E Sand Mix
9 Sample: Water Level (PE)
0 100 200
Vertical Exaggeration: 6
Figure 6. Cross section showing contaminated zone and water table along section A-A' in Figure 5.
39
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Microbiological and Geochemical Degradation Processes
E. Michael Godsy
U.S. Geological Survey, Menlo Park, CA
Introduction
Ground-water contamination is perhaps the most dan-
gerous and intractable environmental impact of toxic
chemical spills and leaks. The prevention of ground-
water contamination is of foremost importance. In terms
of treatment processes, the present technology of exca-
vation and relocation of contaminated soils to "secure
landfills" (which are seldom secure) and "pump and
treatment" of contaminated ground water has proven to
be totally inadequate; these processes just transfer con-
taminants from one environmental phase to another.
Bioremediation, however, achieves contaminant de-
composition or immobilization by exploiting the existing
metabolic potential of microorganisms with novel cata-
bolic functions derived through selection. In response to
the introduction of a toxic contaminant, an indigenous
bacterial population arises that is unique from the stand-
point of physiological capabilities and species diversity.
Conditions that restrict life or inactivate microbial en-
zymes are incompatible with intrinsic bioremediation ef-
forts. Although the physical and chemical characteristics
of the contaminants and the metabolic potential of mi-
croorganisms determine the feasibility of biotransforma-
tion reactions, actually achieving biotransformation also
depends on the prevailing geochemical conditions. The
ecological constraints to bioremediation can be classi-
fied as microbial, chemical, or environmental, and rec-
ognition of these constraints, as required, assesses the
feasibility of intrinsic bioremediation. This article focuses
on the microbiological and geochemical constraints in-
fluencing intrinsic bioremediation processes.
Bacterial Metabolic Diversity
Bacteria comprise a large and diverse group of micro-
organisms that obtain their energy from a variety of
sources, including 1) light (photosynthetic bacteria), 2)
the oxidation or decomposition of reduced organic mat-
ter (heterotrophic bacteria), or 3) the oxidation of re-
duced inorganic compounds (auto-trophic bacteria).
Some bacteria derive energy from more than one
source, such as combinations of light and reduced inor-
ganic or organic compounds; however, heterotrophic
bacteria are the major group responsible for the biode-
gradation of organic compounds.
All living organisms must generate reducing power for
the purpose of replenishing enzymatic systems and
maintaining the oxidation-reduction power cycle. This
involves the reduction of oxidized compounds by the
addition of electrons released from compounds oxidized
during energy production. The electron acceptor can be
either an organic or an inorganic compound. For many
bacteria, most fungi, and all higher organisms, the final
electron acceptor is O2 in the process termed aerobic
respiration. The final reduced substrate in aerobic res-
piration is H2O, and the final oxidized compound re-
spired from energy production is CO2.
In the absence of O2, certain bacterial populations re-
spire other less oxidized inorganic compounds (inor-
ganic respiration) or use only organic compounds
(fermentation). Denitrification is a process promoted by
bacteria that can thrive under either aerobic or anaero-
bic conditions (facultative anaerobes) by using O2 as the
terminal electron acceptor, when available, or, in the
absence of O2, by using NO3- as the terminal electron
acceptor (inorganic respiration). When oxidation-reduc-
tion potentials within soils are even lower, other inor-
ganic compounds are used by specific groups of
bacteria as terminal electron acceptors. Because O2 is
toxic to these bacteria, they are called obligate anaer-
obes. Several common alternative electron acceptors
and associated bacterial groups include ferric iron
(iron reducers), SO42" (sulfate reducers), and CO2
(methanogens).
Incorporating Contaminants Into Bacterial
Ecosystems
The ecosystem as a whole can be thought of as a series
of integrated oxidation-reduction (redox) reactions
driven ultimately by the radiant energy of the sun. Micro-
organisms catalyze many of these reactions and play an
40
-------
essential role in maintaining the electron balance of
complex ecosystems. For a contaminant to be incorpo-
rated into these redox reactions, it must be able to serve
as either an electron donor or an electron acceptor.
Moreover, its tendency to either donate or accept elec-
trons—and thus be oxidized or reduced—depends on
the chemistry of the compound. For example, many
halogenated organic compounds are highly oxidized
relative to their nonhalogenated counterparts, and thus
tend to accept electrons and to be reductively dehalo-
genated. The halogenated compounds must compete
with other physiological electron acceptors in orderto be
incorporated into microbial energy cycles. Thus, the
effectiveness of reductive dehalogenations is often influ-
enced by the presence of other electron acceptors, such
as NO3- or SO42-. This effect would explain why reduc-
tive dehalogenations are more frequently observed un-
der methanogenic conditions, where generally a paucity
of these anions exist.
Bacteria-Contaminant Interactions in
Aquifer Material
A minimum of three conditions must be met before a
contaminant can be degraded or transformed by bacte-
ria: 1) the bacteria must be in the immediate vicinity of
the contaminant; 2) the contaminant must be available
to the bacteria; and 3) the bacteria must have the ca-
pacity to participate in some part of the degradation or
transformation process. Specific bacterial populations
prefer or require particular environmental conditions. If
these conditions do not exist, these populations tend to
become quiescent until more ideal conditions return or
develop, or in some cases they may even die off. The
nature of limiting environmental factors often dictates
which bacterial populations exist. While the subsurface
environment can be modified, modification is often
accomplished with great difficulty and always at great
expense.
Bacterial Nutrition
Bacteria are composed of combinations of elements that
are the components of their genetic material, structural
molecules, enzymes, and intracellular plasma. Because
of the great diversity among bacteria, the proportion of
nutrient elements required for growth varies widely;
however, the major required elements that make up
bacteria are carbon, hydrogen, sulfur, nitrogen, and
phosphorus. In soils and aquifer materials contaminated
with most organic compounds, carbon and hydrogen are
not typically limiting because they are the major compo-
nents of organic compounds, and sulfur is generally
available in sufficient quantities for growth. Thus, the
major limiting elements for growth are nitrogen and
phosphorus. The carbon/nitrogen/phosphorus ratio
(C/N/P) usually considered ideal is 300 to 100:10:1 to
0.05 (1); however, this ratio can vary depending on the
nature of the contaminant(s).
pH and Redox Potential
Near-neutral aquifer pH values are usually optimum for
the biodegradation of contaminant organic material. The
hydrogen ion concentration of the ground water is gov-
erned by the types of compounds produced by bacterial
activity, and is controlled especially by CO32-:HCO3-
:CO2 equilibrium rates (2). Because hydrogen ion trans-
fer is commonly involved in electron transport, pH and
redox potential are interdependent (3).
The redox potential, termed Eh, is extremely important
in the biotransformation schemes of both organic and
inorganic contaminants. Usually, a heavily contaminated
site is anoxic because ongoing bacterial respiration has
depleted all available O2. The resulting anoxic condi-
tions tend to favor different electron acceptors, with the
most oxidized compounds (higher Eh) being used first.
The resultant scheme is NO3- (denitrification) utilization
after O2 depletion, Fe3+ (iron reduction) utilization after
NO3- depletion, SO42- (sulfate reduction) utilization after
Fe3+ depletion, and finally CO2 reduction to CH4 after
depletion of the available SO42-. As a result, bacterial
populations having different degradative potentials can
be operative at different times at the same contamina-
tion site as the redox potential varies.
Although Eh measurements can provide valuable clues
about the functioning of geochemical systems in aquifer
material, they alone will not give definite information
about the chemical species present. These measure-
ments indicate that a potential exists for certain redox
reactions, such as those involving bacterial respiration.
Various organic and inorganic redox reactions cannot be
predicted because of their complexity and different in-
terdependent reaction rates (2).
Temperature
Ground-water temperature is often one of the most im-
portant factors controlling microbial activity and rates of
organic matter decomposition. Generally, rates of enzy-
matic degradation and bacterial metabolism double for
every 10°C increase in temperature until close to inhibi-
tory temperatures, which are usually around 40°C to
50°C for most bacteria. Except forsubfreezing tempera-
tures, however, bacteria are generally capable of degra-
dation at most ambient temperatures (4). Perhaps more
importantly, temperature can influence biodegradation
of a compound or contaminant mixture by changing its
physical properties, bioavailability, ortoxicity to bacteria.
For example, an increase in temperature usually in-
creases the equilibrium vapor concentration, resulting in
an increased volatilization rate, but it can at times also
increase sorption to aquifer particles (5).
41
-------
Physical Deterrents to Biodegradation
Physical or physicochemical factors can also affect the
biodegradation of contaminants. Some molecules are
recalcitrant to degradation because they are too large to
enter bacterial cells, which is usually required for com-
plete degradation by membrane-bound enzymes. Some
substances are difficult to biodegrade because the num-
ber, length, or location of functional groups impede en-
zyme attack. Strong sorption on aquifer material can
greatly hinder the ability of bacteria to attach to, absorb,
or enzymatically attack the molecule (6).
Sorption and solubility of organic contaminants are
complex interdependent phenomena that vary with the
composition of the aquifer material and complex con-
taminant mixtures. For example, aromatic hydrocarbon
concentrations in water extracts of 31 gasoline samples
varied over an order of magnitude (7). Although gasoline
variability could account for this, the solubility of each
component of mixtures can vary from ideal conditions,
with each compound acting as a cosolvent to increase
hydrophobic hydrocarbon solubilities (8). Organic sol-
vents can also affect the sorption of organics on soils in
general (9).
Contaminant Metabolism: Aerobic Versus
Anaerobic
A common misconception is that all organic contami-
nants are biodegraded most rapidly and thoroughly un-
der aerobic conditions. Although this is commonly the
case, anaerobic conditions promote some very impor-
tant degradative processes. For example, compounds
that are highly oxidized, such as polychlorinated
biphenyls or chlorinated solvents, are more susceptible
to reductive processes than to oxidative processes (10)
during the initial stages of mineralization.
Most organic compounds found in crude oil, refined oils,
and fuels are known to degrade under aerobic condi-
tions (4); however, current research efforts have shown
that the biodegradation of many monoaromatic com-
pounds common to most fuels also occurred in the
laboratory under anaerobic conditions. This biodegrada-
tion readily occurs not only with NO3- serving as the
terminal electron acceptor (11) but also under SO42-
reducing conditions (12), Fe3+ reducing conditions (13),
and methanogenic conditions (14,15).
Benzene has been especially recalcitrant to anaerobic
biodegradation in laboratory studies under denitrifying
and sulfate-reducing conditions (10,12). Yet some labo-
ratory (16) and field studies (17) have shown the deple-
tion of all common monoaromatic hydrocarbons found
in gasoline under denitrifying conditions. In 1986, Vogel
and GrbiCEGaliCE(18) demonstrated that benzene and
toluene were degradable to CH4 and CO2 in laboratory
microcosms; however, the confirmation of the degrada-
tion of benzene in field situations under methanogenic
conditions has been rather elusive. One of the few
well-documented instances of the methanogenic degra-
dation of benzene in the field is a crude oil spill in
Bemidji, Minnesota (19), where many of the water-
soluble monoaromatic hydrocarbons present in crude
are undergoing intrinsic bioremediation.
Conclusion
Although bioremediation in general has gained consid-
erable attention, obstacles remain before bacteria can
be used effectively for detoxifying wastes affecting
ground water. A lack of knowledge or misunderstanding
concerning what can and cannot be done with bioreme-
diation has resulted in unrealistic expectations, leading
in turn to disappointments and ultimate failures. Contin-
ued research into and application of sound bioremedia-
tion schemes will undoubtedly prove the viability of
intrinsic bioremediation in the overall remediation efforts
of contaminated ground water.
References
1. Torpy, M.F., H.F. Stroo, and G. Bruebaker. 1989.
Biological treatment of hazardous wastes. Poll.
Eng. 21:80-86.
2. Stumm, W, and J.J. Morgan. 1981. Aquatic chem-
istry. New York, NY: John Wiley & Sons.
3. Grundi, T. 1994. A review of the current under-
standing of redox capacity in natural, disequilibrium
systems. Chemosphere 28:613-626.
4. Leahy, J.G., and R.R. Colwell. 1990. Microbial deg-
radation of hydrocarbons in the environment. Micro-
biol. Rev. 54:305-315.
5. Lyman, W.J., W.F Reehl, and D.H. Rosenblatt.
1982. Handbook of chemical property estimation
methods: Environmental behavior of organic chemi-
cals. New York, NY: McGraw-Hill.
6. Cheng, H.H., K. Haider, and S.S. Harper. 1983.
Catechol and chlorocatechols in soil: Degradation
and extractability. Soil Biol. Biochem. 15:311-317.
7. Cline, P.V., J.F. Delfino, and P.S.C. Rao. 1991. Par-
titioning of aromatic constituents into water from
gasoline and other complex solvent mixtures. Envi-
ron. Sci. Technol. 25:914-920.
8. Groves, F.R. 1988. Effect of cosolvents on the solu-
bility of hydrocarbons in water. Environ. Sci. Tech-
nol. 22:282-286.
9. Fu, J.-K., and R.G. Luthy. 1986. Effect of organic
solvent on sorption of aromatic solutes onto soils.
J. Environ. Eng. 112:346-366.
42
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10. Kuhn, E.R, and J.M. Sulflita. 1989. Dehalogenation
of pesticides by anaerobic microorganisms in soils
and ground water: A review. In: Sawhney, B.L., and
K. Brown, eds. Reactions and movement of organic
chemicals in soils. Special Publication 22. Madison,
Wl: Soil Science Society of America, pp. 111-180.
11. Evans, P.J., D.T. Mang, and L.Y. Young. 1991. Deg-
radation of toluene and m-xylene and transforma-
tion of o-xylene by denitrifying enrichment cultures.
Appl. Environ. Microb. 57:450-454.
12. Edwards, E.A., I.E. Wills, D. GrbiCEGaliCEand M.
Reinhard. 1991. Anaerobic degradation of toluene
and xylene: Evidence for sulfate as the terminal
electron acceptor. In: Hinchee, R.E., and R.F. Olfen-
buttel, ede. In situ bioreclamation: Applications and
investigations for hydrocarbon and contaminated
site remediation. Boston, MA: Butterworth-Heine-
mann. pp. 463-471.
13. Lovely, D.R., and D.J. Lonergan. 1990. Anaerobic
oxidation of toluene, phenol, and p-cresol by the
dissimilatory iron-reducing organism, GS-15. Appl.
Environ. Microb. 56:1,858-1,864.
14. GrbiCEGaliCED., and T.M. Vogel. 1957. Transforma-
tion of toluene and benzene by mixed methano-
genic cultures. Appl. Environ. Microb. 53:254-260.
15. Wilson, B.H., G.B. Smith, and J.F. Rees. 1986.
Biotransformations of selected alkylbenzenes and
halogenated aliphatic hydrocarbons in methano-
genic aquifer material: A microcosms study. Envi-
ron. Sci. Technol. 20:997-1,002.
16. Major, D.W, C.I. Mayfield, and J.F. Barker. 1988.
Biotransformation of benzene by denitrification in
aquifer sand. Ground Water 26:8-14.
17. Berry-Spark, K.L., J.F. Barker, D. Major, and C.I.
Mayfield. 1986. Remediation of gasoline-contami-
nated ground-waters: A controlled experiment. In:
Proceedings of petroleum hydrocarbons and or-
ganic chemicals in ground water: Prevention, de-
tection, and restoration, NWWA/API. Dublin, OH:
Water Well Journal Publishing.
18. Vogel, T.M., and D. GrbiCEGaliCEl 986. Incorporation
of oxygen from water into toluene and benzene
during anaerobic fermentative transformation. Appl.
Environ. Microb. 52:200-202.
19. Cozzarelli, I.M., R.P Eaganhouse, and M.J.
Baedecker. 1990. Transformation of monoaromatic
hydrocarbons to organic acids in anoxic ground-
water environment. Environ. Geol. Water Sci.
16:135-141.
43
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Field and Laboratory Results: Getting the Whole Picture
Mary Jo Baedecker
U.S. Geological Survey, Reston, VA
Abstract
Concern over contamination of ground water in the last
decade has led to an increased awareness of the need
to understand the transport and fate of organic contami-
nants and the geochemical processes that result from
their presence in the subsurface. A large number of
contaminated sites contain petroleum-derived hydrocar-
bons. Many of the hydrocarbons are biodegraded in
ground-water environments, and the extent of their re-
moval by natural processes has been evaluated in field
and laboratory investigations. Both types of investiga-
tions demonstrate that natural biodegradation can be an
important component in remediation strategies for some
contaminated sites.
Results and Discussion
The processes that control the attenuation of organic
compounds in the subsurface are complex, and many
investigations have been undertaken in the field and in
laboratories to understand better the factors that control
degradation reactions. Parts of contaminant plumes
often become anoxic, and the fate of contaminants at
anoxic field sites has been reported in several studies
(1-5). One of the most widespread types of contami-
nants is petroleum-derived hydrocarbons from pipeline
breaks, leaking storage tanks, spills, and disposal of
wastes. Many of these sites, such as those contami-
nated by leaking small underground storage tanks, are
easier to remediate than sites with contaminants such
as chlorinated compounds. The number of such sites is
large, however, and much effort has been spent trying
to understand processes and to develop effective reme-
diation strategies to deal with petroleum-derived hydro-
carbons. Several research efforts have been undertaken
in the field and laboratory to determine processes that
affect the fate and transport of individual hydrocarbons.
In field investigations, degradation of soluble aromatic
hydrocarbons has been shown to occur downgradient
from source areas (6-10). Hydrocarbon-degrading bac-
teria were found and quantified in soil and ground water
at a fuel-oil contaminated site (11). Degradation of pe-
troleum-derived hydrocarbons is generally considered
to occur more rapidly in aerobic or suboxic environ-
ments, where oxygen or nitrate is available as an elec-
tron acceptor (7,12-13), but anaerobic biodegradation
also may remove significant amounts of hydrocarbons
from ground water (14-17).
An investigation of the effects of a crude-oil spill on an
aquifer was undertaken near Bemidji, Minnesota, as
part of the Toxic Substances Hydrology Program of the
U.S. Geological Survey. An underground pipeline carry-
ing crude oil ruptured and sprayed oil over land surface.
Part of the oil was removed during remediation, but part
of it infiltrated through the unsaturated zone and accu-
mulated as an oil body floating on the ground water. A
more detailed description of the processes that occurred
at the site is available in Baedecker et al. (5), Bennett et
al. (18), Eganhouse et al. (9), Baedecker and Cozzarelli
et al. (19), Cozzarelli et al. (20), and Eganhouse et al. (21).
The concentrations of benzene, toluene, ethylbenzene
and o-, m-, and p-xylene (BTEX) in the upper 1.5-m
thickness of the aquifer downgradient from the oil body
are shown in Figure 1. The concentrations of BTEX
decreased with increasing distance from the oil body
and were attenuated under anoxic and oxic conditions.
Where oxygen was encountered, concentrations de-
creased by several orders of magnitude (56 m to 137 m).
The mass of BTEX lost near the oil body (26 mto 56 m),
however, was as high as that lost farther downgradient.
The decrease in concentrations of BTEX near the oil
body (26 m to 56 m) was in the anoxic part of the plume,
where no oxygen was detected in the ground water over
an 8-yr period. Another indicator that this part of the
aquifer was anoxic is that ferrous iron concentrations
were 1.8 mg/L at the water table and 71 mg/L at only 0.3
m deeper in the aquifer (22). Ferrous iron precipitates
where it encounters trace quantities of oxygen. Thus,
the high concentrations of ferrous iron near the oil
body indicate that dissolved oxygen was not trans-
ported to the water table. Degradation of hydrocarbon
44
-------
CD
Q.
o
"5
100
10
1.0
0.1
0.01
CD
£ 0.001
o
O
0.0001
BTEX
50 100 150
Meters Downgradient from Oil Body
200
Figure 1. Concentrations of benzene, toluene, ethylbenzene,
and o-, m-, and p-xylene (BTEX) in the upper 1.5-m
thickness of an aquifer contaminated with crude oil
(19).
vapors in the unsaturated zone most likely consumed
the oxygen by aerobic respiration (23). The downgradi-
ent movement of the BTEX plume averaged about
8 m/yr from 1987 to 1992, but movement has not been
at a steady rate. Near the oil body, microbial degradation
is the primary process of hydrocarbon attenuation. As
the hydrocarbons are transported farther from the oil
body, additional processes such as dispersion, mixing
(24), and sorption become more important.
Laboratory experiments demonstrated that these hydro-
carbons can degrade under controlled laboratory condi-
tions. To verify that the hydrocarbons were degrading
under anoxic conditions, microcosm experiments were
undertaken with sediment and water from the anoxic
part of the plume (5,20). In two separate experiments,
benzene and a mixture of toluene and naphthalene were
added to microcosms under anaerobic conditions. These
compounds were also added to microcosms that were
poisoned and sterilized for controls. In the microbially
active microcosms, benzene decreased in concentra-
tion by 98 percent in 125 d, and toluene decreased by
99 percent in 45 d (5). No loss of naphthalene was
observed during the same period. For the controls, no
loss of benzene, toluene, or naphthalene was observed.
These laboratory experiments and field results provide
strong evidence that hydrocarbons are degrading in an
anoxic environment. By comparing the results from the
microbially active microcosms with results from controls,
sorption and abiological chemical oxidation were elimi-
nated as possible explanations for the loss of benzene
and toluene.
Field and laboratory studies of anaerobic biotransforma-
tion or aromatic hydrocarbons were reviewed by Barker
and Wilson (25) who found evidence at five methano-
genic sites for biodegradation of benzene, toluene,
ethylbenzene, and the xylenes. The estimated half-lives
were 0.5 yr and 3.8 yr. The longest half-lives were for
benzene. Laboratory experiments also have indicated
biodegradation by several pathways (26); however, for
benzene the results have been contradictory. Large con-
centrations of benzene are biodegraded in the subsur-
face (27), yet in some ground-water environments
benzene is the most persistent hydrocarbon among the
monoaromatics.
Natural processes can remove significant concentra-
tions of hydrocarbons and prevent the spreading of a
plume. At sites where the rates of solubilization, volatili-
zation, and biodegradation of hydrocarbons are such
that the plume is either contained or spreading at a slow
rate, these natural processes can be considered in the
design of a site remediation program. Even at sites
where plumes are spreading at a fast rate, knowledge
of the natural processes that are attenuating contami-
nants provides information that can be used to acceler-
ate biodegradation processes.
Note
The data and interpretation in this report were previously
published in Baedeckerand Cozzarelli (19) and Baedecker
et al. (5).
References
1. Baedecker, M.J., and W. Back. 1979. Hydrogeologi-
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Ground Water 17:429-437.
2. Nicholson, R.V., J.A. Cherry, and E.J. Reardon.
1983. Migration of contaminants in ground water at
a landfill: A case study 6. Hydrogeochemistry. J.
Hydrol 63:131-176.
3. Lesage, S., R.E. Jackson, M.W Priddle, and RG.
Riemann. 1990. Occurrence and fate of organic
solvent residues in anoxic ground water at the
Gloucester landfill, Canada. Environ. Sci. Technol.
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4. Godsy, E.M., D.F. Goerlitz, and D. GrbiCEGaliCE
1 992. Methanogenic biodegradaflon of creosote
contaminants in natural and simulated ground-
water ecosystems. Ground Water 30.232-242.
5. Baedecker, M.J., I.M. Cozzarelli, R.R Eganhouse,
D.I. Siegel, and P.C. Bennett. 1993. Crude oil in a
shallow sand and gravel aquifer, III. Biogeochemi-
cal reactions and mass balance modeling in anoxic
ground water. App. Geochem. 8:569-586.
6. Barker, J.F., J.S. Tessman, RE. Plotz, and M. Re-
inhard. 1986. The organic geochemistry of a sani-
tary landfill leachate plume. Contam. Hydrol.
1:171-189.
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7. Major, D.W., C.I. Mayfield, and J.F. Barker. 1988.
Biotransformation of benzene by denitrification in
aquifer sand. Ground Water 26:8-14.
8. Cozzarelli, I.M., R.P. Eganhouse, and M.J.
Baedecker. 1990. Transformation of monoaromatic
hydrocarbons to organic acids in anoxic ground-
water environment. Environ. Geol. Water Sci.
16:135-141.
9. Eganhouse, R.P., M.J. Baedecker, I.M. Cozzarelli,
G.R. Aiken, K.A. Thorn, and T.F. Dorsey. 1993.
Crude oil in a shallow sand and gravel aquifer, II.
Organic geochemistry. Appl. Geochem. 8:551-567.
10. Davis, J.W, N.J. Klier, and C.L. Carpenter. 1994.
Natural biological attenuation of benzene in ground
water beneath a manufacturing facility. Ground
Water 32(2):215-226.
11. Kampfer, P., M. Steiof, and W Dott. 1991. Microbio-
logical characterization of a fuel oil-contaminated site
including numerical identification of heterotrophic
water and soil bacteria. Microb. Ecol. 21:227-251.
12. Kuhn, E.P., J. Zeyer, P. Eicher, and R.P. Schwar-
zenback. 1988. Anaerobic degradation of alkylated
benzene in denitrifying laboratory aquifer columns.
Appl. Environ. Microb. 54:490-496.
13. Hutchins, S.R., G.W Sewell, D.A. Kovacs, and G.A.
Smith. 1991. Biodegradation of aromatic hydrocar-
bons by aquifer microorganisms under denitrifying
conditions. Environ. Sci. Technol. 25:68-76.
14. Wilson, B.H., G.B. Smith, and J.F. Rees. 1986.
Biotransformations of selected alkylbenzenes and
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15. GrbiCEGaliCED., and T.M. Vogel. 1957. Transforma-
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17. Haag, F.M., M. Reinhard, and PL. McCarty. 1991.
Degradation of toluene and p-xylene in anaerobic
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18. Bennett, PC., D.I. Siegel, M.J. Baedecker, and M.F.
Hult. 1993. Crude oil in a shallow sand and gravel
aquifer, I. Hydrology and inorganic geochemistry.
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chemical processes and migration of aqueous con-
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21. Eganhouse, R.P, M.J. Baedecker, and I.M. Cozzarelli.
1994. Biogeochemical processes in an aquifer con-
taminated by crude oil: An overview of studies at
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22. Cozzarelli, I.M., M.J. Baedecker, G. Aiken, and C.
Phinney. 1994. Small-scale chemical heterogeneities
in a crude-oil-contaminated aquifer, Bemidji, Minne-
sota. In: Morganwalp, D.W, and D.A. Aronson, eds.
Proceedings of the Technical Meeting of the U.S.
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gram, Colorado Springs, CO (September 20-24,
1993). Water Res. Invest. Rep. 94-4014. In press.
23. Hult, M.F, and R.R. Grabbe. 1988. Distribution of
gases and hydrocarbon vapors in the unsaturated
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1985). Open-File Report 86-481. pp. C21-C26.
24. Essaid, H.I., M.J. Baedecker, and I.M. Cozzarelli.
1994. Use of simulation to study field-scale solute
transport and biodegradation at the Bemidji, Minne-
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D.A. Aronson, eds. U.S. Geological Survey toxic sub-
stances hydrology program. Proceedings of the Tech-
nical Meeting of the U.S. Geological Survey Water
Resources, Colorado Springs, CO (September20-24,
1993). Water Res. Invest. Rep. 94-4014. In press.
25. Barker, J.F, J.T Wilson. 1992. Natural biological
attenuation of aromatic hydrocarbons under an-
aerobic conditions. Proceedings of the Subsurface
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26. GrbiCEGaliCED. 1990. Anaerobic microbial transfor-
mation of nonoxygenated aromatic and alicyclic
compounds in soil, subsurface, and freshwater
sediments. In: Bollag, J.M., and G. Stotzky, eds. Soil
Biochemistry, Vol. 6. New York, NY: Marcel Dekker,
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27. Hadley, PW, and R. Armstrong. 1991. Where's the
benzene? Examining California ground-water qual-
ity surveys: Ground Water 29(1):35-40.
46
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Case Studies: Public Service's Site
47
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In Situ Bioremediation at the Seventh Avenue Site in Denver:
Remediation of Soils and Ground Water
Christopher Nelson
Groundwater Technology, Inc., Englewood, CO
In situ bioremediation is a cost-effective method for the
remediation of soils and ground waters contaminated
with petroleum hydrocarbons. This paper presents a
case study of the successful application of in situ biore-
mediation at a public utility site in Denver, Colorado.
The site was used as a truck maintenance facility for
almost 30 years. During that time, nonvolatile petroleum
hydrocarbons from used motor oil, diesel, gasoline, and
other automotive fluids were released to a used oil sump
at the site. The local geology of the site includes
claystone and sandstone bedrock, covered with recent
deposits of interbedded alluvial sands and gravels,
as well as silty sands and clays. The principal aquifer at
the site lies within the interbedded alluvial sands and
gravels.
Soil and water sampling confirmed the presence of pe-
troleum hydrocarbons in both the vadose and saturated
zones. Laboratory studies showed that the chemical,
microbiological, and hydrogeological characteristics of
the site were conducive to bioremediation.
Site Assessment/Treatability Study
An extensive site assessment and treatability study was
conducted at the site to determine physical, chemical,
microbiological, and hydrogeological characteristics
controlling the biodegradation of contaminants and the
mass transport of nutrients and oxygen. Nine monitoring
wells were installed at depths of about 7.6 m, using a
3-m screen interval. The site assessment indicated that
the primary contaminant at the site was waste oil located
in the saturated and unsaturated sediments beneath the
former used oil sump. Samples showed high levels of
benzene, toluene, ethylbenzene, and the xylenes
(BTEX); total petroleum hydrocarbons (TPHs); and total
organic carbon (TOC) in ground water, with localized but
detectable levels of chlorinated organics. A relatively
large population of bacteria was found within the
contamination zone; however, its growth appeared to be
restricted by nutrient and oxygen conditions.
Feasibility studies were performed on soil and sediment
samples to determine the biodegradability of contami-
nants under various nutrient loads and aerobic condi-
tions. Aerobic testing was performed to simulate optimal
conditions for the bioremediation of hydrocarbons. Col-
umn studies also were conducted to determine how
nutrient and hydrogen peroxide loading would affect the
hydraulic conductivity of sediments in the subsurface
above the ground-water table.
The results showed that the loading of nutrients and
hydrogen peroxide would be critical to the success of in
situ bioremediation at this site and that the loading
should be minimized in the silty sand zone because of
high reactivity. A nutrient adsorption test indicated that
ammonia and phosphate loading was feasible in the soil
at the site.
System Design and Installation
After reviewing several remedial operations, in situ
bioremediation was selected. The conceptual design
included stimulating indigenous bacterial populations
through the introduction of oxygen and inorganic nutri-
ents. The primary functions of the remediation system
included ground-water recovery, treatment, and reinjec-
tion; vapor extraction and discharge; stimulation of
in situ bioremediation by subsurface inorganic nutrient
and oxygen additions; and phase-separated hydrocarbon
recovery.
Laboratory tests showed that hydrogen peroxide and
nutrient loading worked best in sediments from the
coarse sand interval. Nutrient injections and the addition
of atmospheric oxygen from the vapor extraction system
stimulated bioremediation in the unsaturated zone and
enhanced the desorption of adsorbed hydrocarbons for
recovery in the monitoring wells.
48
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Operation, Monitoring, and Results
Once the bioremediation system was installed, it was
inspected weekly to adjust and maintain the water-table
depression pump, the hydrogen peroxide and nutrient
injection equipment, and the soil vapor extraction sys-
tem. Crews conducted field analyte tests, sampled
ground water, and gauged monitoring wells. The biore-
mediation system began operating in July 1989 and
continued through March 1992.
Approximately 36,000 Ib of hydrocarbons have been
removed. Nearly 94 percent of the contaminant mass
was degraded biologically, as evidenced by the low
concentrations of dissolved oxygen and the relatively
high concentrations of background heterotrophs and
hydrocarbon-utilizing bacteria. Approximately 9 million
gal of ground water were recovered, amended with nu-
trients, and reinfiltrated. The site is currently undergoing
closure with the Colorado Department of Health.
Reference
For more information about the Denver site and support-
ing data, refer to:
Nelson, C.H., R.J. Hicks, and S.D. Andrews. 1994. An
integrated system approach for in situ bioremediation of
petroleum hydrocarbon contaminated soil and ground
water. In: Flathman, P.E., D.E. Jerger, and J.H. Exner,
eds. Bioremediation: Field experience.
49
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The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through
In Situ Bioremediation: The Role of Mathematical Models
Tissa H. Illangasekare and David C. Szlag
Department of Civil, Environmental, and Architectural Engineering, University of Colorado, Boulder, CO
John T. Wilson
Robert S. Kerr Environmental Research Laboratory, U.S. Environmental Protection Agency, Ada, OK
Abstract
This paper discusses the important processes in devel-
oping mathematical models to be used as tools for the
design and evaluation of ground-water remediation
schemes in aquifers contaminated with nonaqueous
phase organic chemicals. The paper also presents the
application of a mathematical model that considers
these processes in developing a methodology for retro-
spective evaluation of bioremediated sites.
Introduction
Mathematical models of water flow and chemical trans-
port have been extensively used in ground-water quan-
tity and quality management. Some of the applications
of these models involve bioremediation. The basic proc-
esses of flow of water and transport of dissolved sub-
stances are fairly well understood, and numerical
models for the solution of the governing equations have
been developed. Field applications of models as predic-
tion and design tools, however, have not been very
successful for many reasons, including the complexities
associated with natural heterogeneities and the inade-
quacy of available field techniques for physical and
chemical characterization; this is especially true in
ground-water contamination situations involving chemi-
cals in the form of separate phase organics. The models
that have been developed to simulate the transport and
entrapment behavior of nonaqueous phase chemicals
and waste products have not been adequately validated
due to the scarcity of laboratory and field data. These
models sometimes fail to simulate flow and entrapment
behavior under heterogeneous soil conditions that are
commonly encountered in the field. Accurate calibration
and prediction become difficult due to the limitations of
field and laboratory techniques that are used to obtain
model parameters. Some of the scaling issues related
to multiphase flow model parameters are not very well
understood. The assumptions that are made in modeling
mass transfer from entrapped chemicals to the aqueous
phase become questionable under some conditions of
ganglia formation and macroscale entrapment.
The movement of nonaqueous phase liquids (NAPLs) in
the subsurface is a complicated process leading to large
amounts of NAPLs becoming trapped in the soil. Most
of the common organic wastes in the form of NAPLs are
only sparingly soluble in water and thus act as long-term
sources of ground-water contamination. In a case study
presented in this paper, we will show that after bioreme-
diation, pockets of NAPLs remained in the soil. These
entrapped fluids have the potential to contribute to
ground-water pollution after active remediation has
been discontinued.
In our research, we have identified two modes of entrap-
ment: microscale and macroscale. Microscale entrap-
ment, which occurs at residual levels, is primarily
governed by fluid properties and pore characteristics
(1). Macroscale entrapment is defined as entrapment at
saturations higher than residual (or irreducible) due to
heterogeneities in the soil (2).
Many models employed in remediation design make use
of the local equilibrium assumption (LEA); in other
words, any water exiting a zone of entrapped NAPL will
be completely saturated with the contaminant, regard-
less of the system parameters. The LEA is conservative
in predicting the maximum concentration observed in
ground water but may lead to gross underprediction of
the contaminant source lifetime. It is our hypothesis that
entrapment itself will change the system parameters
controlling mass transfer into the flowing water, and the
LEA assumption is not a physically realistic way to quan-
tify mass transfer.
50
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Laboratory experiments were conducted to obtain a fun-
damental understanding of the processes that govern
the transport and distribution of organic chemicals in
soils and to generate data for validation of models that
will be used as tools for the design of remediation
schemes and monitoring systems. A detailed investiga-
tion of these processes under controlled conditions was
done in small soil cells, columns, and large flumes.
In our ongoing research, new models and modeling
approaches have been developed. To improve these
models, we have focused on issues related to entrap-
ment, mobilization, and mass transfer associated with
organic waste chemicals in heterogeneous aquifers.
The effectiveness of models as tools to design and
evaluate treatment and remediation technologies de-
pends on their ability to accurately represent the above
processes. A case study conducted in Colorado identi-
fies some of these basic processes that are of impor-
tance in remediating and monitoring sites contaminated
with organic fluids, and demonstrates the use of a
mathematical model.
Study Objectives
Conventional methods for determining the extent of
cleanup at a bioremediation site can often be mislead-
ing. Monitoring wells may show very low or zero levels
of contaminants after active bioremediation, but levels
may increase over time. In most cases, regulatory
authorities require a direct measure of the residual
NAPLs after bioremediation in addition to monitoring
well data. Often the relative composition of the oily
phase is assumed to remain constant during bioreme-
diation. This is a conservative assumption and generally
leads to target levels of total petroleum hydrocarbon
(TPH) concentrations on the order of 10 to 100 mg/kg
aquifer material. Many bioremediation schemes, how-
ever, may preferentially degrade the compounds of
regulatory concern, leaving relatively high TPH levels in
the soil that pose a minimal risk. This modeling study
focuses on developing a methodology to evaluate the
possible risk, if any, associated with benzene, toluene,
ethylbenzene, and xylene (BTEX) sources left in soils
after the implementation of a bioremediation scheme.
This developed methodology will assist in providing an-
swers to the following questions: 1) Will BTEX reappear
in ground water? 2) How long will it take the plume to
reappear? 3) What concentration level may be ex-
pected? The results of the case study will also assist in
providing a technical basis for implementing monitoring
schedules, locating compliance wells, and constructing
rational criteria for site closure.
Problem Description
Atemporary holding tank under a garage in an industrial
area in Denver, Colorado, leaked used crank case oil,
diesel fuel, gasoline, and other material into a shallow
water table aquifer.
Remediation involved removal of separate oily phases,
in situ bioremediation with hydrogen peroxide and min-
eral nutrients, and bioventing. An estimated 2,147 Ib of
hydrocarbons have been removed by pumping hydro-
carbon emulsion from monitoring wells or by volatiliza-
tion through soil aeration. In the treatment system, the
ground water pumped from a downgradient recovery
well was treated and amended with organic nutrients
and hydrogen peroxide. The solution was then injected
into the subsurface system upgradient of the contami-
nated site. The system operated from October 1989 to
March 1992. Table 1 compares the reduction in concen-
tration of benzene and total BTEX compounds in ground
water achieved by this remediation scheme (3,4).
Water from the monitoring wells and the recirculation
well contained low concentrations of contaminants by
March 1992. Active remediation was terminated, and the
site entered a period of postremediation monitoring. In
June 1992, core samples were taken from the aquifer to
determine the extent of hydrocarbon remaining and to
determine whether a plume of contamination could re-
turn once active remediation ceased. The site was cored
along a transect downgradient of the release. The cores
were extracted and analyzed for total petroleum hydro-
carbons and for the concentration of individual BTEX
compounds. Data from one of the boreholes with the
highest concentrations of hydrocarbons are given in
Table 2. At the time of sampling, the elevation of the
Table 1. Reduction in Concentration (ug/L) of Hydrocarbon Contaminants in Ground Water Achieved by In Situ Bioremediation
Benzene Total BTEX
Well
Before
During
After
Before
During
After
MW-1
MW-8
MW-2A
MW-3
RW-1
220
180
—
11
<1
<1
130
11
5
2
<1
16
0.8
2
<1
2,030
1,800
—
1,200
<1
164
331
1,200
820
2
<6
34
13
46
<1
51
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Table 2. Vertical Extent of Total BTEX and TPH (mg/kg) at a Borehole
Elevation (Feet AMSL) TPH BTEX
Benzene
Color and Texture
281.14-5,280.31
5,280.31-5,279.97
5,279.97-5,279.56
5,279.56-5,279.14
5,279.14-5,278.97
5,278.97-5,278.64
5,278.64-5,278.22
5,278.22-5,277.14
<44
227
860
1,176
294
273
<34
<24
<1
5.1
101
206
27
7.4
<1
<1
<0.02
<0.2
<0.2
4.3
0.68
0.26
<0.2
<0.2
Brown sand
Brown sand
Brown sand
Brown sand
Brown sand
Brown sand
Brown sand
Brown/yellow sand
water table was 5,280.5 ft above mean sea level
(AMSL).
To hydraulically characterize the aquifer, selected soil
cores that had previously been used to measure hydro-
carbon saturations were reconstructed in a load cell, and
hydraulic conductivity was determined with a constant
flux apparatus. The hydraulic conductivity varied two
orders of magnitude across the site, with some highly
permeable channels evident.
The results of this field investigation suggest that hydro-
carbon in the form of nonaqueous fluids moved into
preferential flow channels created by the local hetero-
geneities. After direct recovery, the fluids remained en-
trapped at saturations that may be higher than residual.
During remediation, the treating agents did not reach
some of the locations where the chemicals were
entrapped.
A modeling study was conducted at the site to make a
retrospective evaluation of the effectiveness of the re-
mediation scheme. The results of this study are reported
in Szlag et al. (3).
Model Selection
The following observations at the site and in the labora-
tory indicated the need forthree-dimensional simulation:
1) visual inspection of aquifer material indicated the
presence of coarse gravel lenses, clayey sands, and
sands of varying gradation; 2) light nonaqueous phase
liquid (LNAPL) plumes are inherently three-dimensional,
forming thin, pancake-like plumes in the capillary fringe
and just beneath the water table; 3) LNAPL can become
entrapped in coarse lenses that act as preferential flow
channels well beneath the water table; 4) solute plumes
are not vertically homogeneous, and biological activity
will not be uniformly distributed vertically. Bypassing due
to the lowered hydraulic conductivity of the central part
of the LNAPL plume by nutrients and electron acceptors
resulted in high TPH and BTEX levels in some cores.
Coupled with the clear need for a three-dimensional
model are other criteria such as availability, ease of use,
reliability, and cost. We have selected MODFLOW, a
three-dimensional ground-water flow model developed
by the U.S. Geological Survey (USGS), to simulate
ground-water flow. Solute transport is simulated with a
three-dimensional random walk called RAND3D.
Modeling Approach
The problem domain was modeled as a rectangular area
300 ft long and 200 ft wide. Two wells outside of the
modeled domain were used as reference head locations
for general head boundaries. The NAPL contaminant
zone covers approximately 1,600 ft2 of area and a soil
depth of approximately 1.7 ft. To accurately assess the
mass flux from the LNAPL contaminant source, three
layers were chosen in the model. The upper two layers
are 1 ft thick, and the bottom layer is 18 ft thick. The
LNAPL organic contaminant is confined to the uppertwo
layers.
Data for the period June 8, 1989, to April 1, 1990, were
used for calibration. The goodness of fit between the
model and measured well data was characterized by a
mean residual and standard deviation of the mean re-
sidual. The best fit model was obtained by assigning the
pump test average hydraulic conductivity to the bottom
layer, which carried the majority of water.
The main focus of this modeling effort is to determine
how much contaminant mass will be transported from
the remaining residual and whether it will generate a
plume of regulatory concern. A monitoring well screened
in only the upper two layers would see the highest
contaminant concentrations. A pumping well screened
over the entire aquifer thickness is also being consid-
ered; in this case, however, dilution will play a major role
in reducing the maximum concentrations. Two signifi-
cant assumptions are used in the solute transport
modeling: 1) the concentration of BTEX in the source
zone remains constant, and 2) water flowing from the
contaminated cells is in equilibrium with the residual
NAPL. Using the heads generated by MODFLOW,
52
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ground-water velocity through each source cell was cal-
culated. The known NAPL BTEX concentration was then
used to calculate the equilibrium concentration and,
consequently, the mass flux. Estimated benzene mass
fluxes were converted into particle inputs for each layer.
The particle tracking model was used to simulate solute
transport using the velocity field generated from
MODFLOW.
Results and Discussion
Any simulation of solute transport requires specification
of the contaminant source. Forthe case of a NAPL spill,
the source function will consist of a continuous mass flux
of solute from the residual NAPL phase to the aqueous
phase. Many researchers have shown that equilibrium
is quickly reached in spill scenarios if ground-water
velocity is low and the "residence time" of the water in
contact with NAPL is "long enough." From a regulatory
standpoint, the assumption of equilibrium is conserva-
tive, as greater mass fluxes cannot be achieved. The
water flux can vary significantly in the source zone,
which often gives misleading indications that the con-
taminant transport is rate-limited. This primary problem
has been the focus of our work. Preferential flow paths
often develop within the source zone in areas with low
BTEX and TPH concentrations, allowing water flow to
bypass the more highly contaminated areas. Laboratory
determination of the hydraulic conductivity in the
samples containing high amounts of TPH confirms this
observation.
A key result of the modeling study is that the solute
plume emanating from a NAPL source is not homoge-
neous. In general, the solute plume will consist of sub-
plumes at different depth intervals and widely different
concentrations, and moving at different velocities. A
regulatory question posed earlier in this paper is, "How
long should the compliance wells be monitored?" The
answer is when all the subplumes have reached steady-
state. The plumes in the middle and bottom layer have
reached or are close to equilibrium by 330 days. The
plume in the top layer, which has the highest benzene
concentration, has not reached equilibrium at 420 days.
The design of the compliance wells will have tremen-
dous impact on the actual sampled concentration. If the
wells are bailed or pumped so that the well volume is
completely mixed, significant dilution will occur. The ex-
isting monitoring wells at the site are screened over the
top 5 ft of the aquifer. The maximum concentration
achieved in the well screened over a 5-ft interval
reaches a steady-state concentration of 26 ppb. If that
well is screened over the entire saturated thickness, a
concentration of 15 ppb is achieved. Even greater dilu-
tion will occur if the well is pumped.
Several operational considerations for risk assessment
and compliance well monitoring can be made from the
modeling study: 1) a benzene plume will reestablish
itself at the site, but it will be three orders of magnitude
lower than the federal maximum contaminant level—
new standards may be set, however, and the risk from
this plume may be deemed significant; 2) local hydraulic
conductivity plays a significant role in determining the
contaminant mass flux and in creating subplumes of
different concentration and velocity; 3) compliance well
monitoring will have to be continued past August 1993
so that solute plumes in all levels will reach steady-state;
4) retardation coefficients and effective porosity data
would significantly improve the time of arrival estimate
of the solute plume; 5) compliance well design should
be carefully considered when sampling a three-dimen-
sional plume because well design can lead to significant
contaminant dilution.
Conclusions
A modeling methodology for the retrospective evaluation
of bioremediated aquifers contaminated with organic
chemicals was developed. The primary hypothesis on
which the methodology was based is that during a spill,
NAPL contaminant becomes entrapped preferentially in
coarse formations in the saturated zone and fine forma-
tions in the unsaturated zone. This hypothesis is sup-
ported by laboratory experimental (2) and field data.
Flow channels created by naturally occurring aquifer soil
heterogeneities as well as macroscale entrapment of
the NAPL will also produce preferential paths for the
treating agents. The proposed methodology requires
that these local heterogeneities in the contaminant zone
of the spill be captured. Standard pump tests, which
provide regional values for transmissivity, will not have
adequate resolution to capture these spill-site-scale het-
erogeneities. Even though hydraulic conductivity values
determined in the laboratory on disturbed soil samples
were used in this study, a more appropriate charac-
terization method would be well-designed bail tests (or
slug tests) that capture the local layered heterogeneities
more accurately. These local hydraulic conductivity values
allow us to obtain the velocity field in the contaminant zone
and to subsequently determine the contaminant mass flux.
Solute breakthrough curves determined by this method
can then be used to conduct risk analysis and to provide
a rational basis for postremediation well monitoring.
Acknowledgments
The support of the U.S. Environmental Protection
Agency through the Hazardous Substance Research
Center at Kansas State University (agreement R-
815709) is gratefully acknowledged. We would also like
to thank Ms. Lisa Weers of the Colorado Department of
Health for her assistance.
53
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References
1. Szlag, D., and T.H. Illangasekare. 1994. Quantifica-
tion of residual entrapment of nonaqueous phase
organic fluids in soils. Ground Water. In review.
2. Illangasekare, T.H., D. Szlag, J. Campbell, J. Ram-
sey, M. Al-Sherida, and D.D. Reible. 1991. Effect of
heterogeneities and preferential flow on distribution
and recovery of oily wastes in aquifers. Proceedings
of the Conference on Hazardous Waste Research,
Manhattan, KS. Manhattan, KS: Kansas State Uni-
versity.
3. Szlag, D.C., T.H. Illangasekare, and J.T. Wilson.
1993. Use of a three-dimensional ground-water
model for retrospective evaluation of a bioremedi-
ated aquifer contaminated with organic chemicals.
Proceedings of the Ground-Water Modeling Confer-
ence, Golden, CO (June 10).
4. Wilson, J. 1993. Retrospective performance evalu-
ation on /ns/frybioremediation: Site characterization.
In: U.S. EPA. Symposium on bioremediation of haz-
ardous wastes: Research, development, and field
evaluations (abstracts). EPA/600/R-93/054. Wash-
ington, DC (May).
54
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The Importance of Knowledge About Intrinsic Bioremediation for Cost-Effective
Site Closure: The Client's Perspective
Harry E. Moseley
Public Service Company of Colorado, Denver, CO
Public Service Company of Colorado (PSCo) is Colo-
rado's largest electric and gas utility. During an under-
ground storage tank removal and replacement project in
1987 at PSCo's Seventh Avenue Service Center, a fa-
cility used for automotive maintenance, an oil sump was
discovered. Because the sump did not have a concrete
bottom or other type of platform, the soil beneath the
sump was saturated with used oil.
After an extensive study to determine the nature and
extent of the contamination beneath the facility, PSCo
officials decided that in situ bioremediation of the site
would be more cost effective than removing the building
that housed the center. Construction of a new facility
would cost at least $1 million, excluding demolition of
the existing facility and removal or remediation of the
contaminated soil and ground water. After discussions
with the Colorado Department of Health and 18 months
of ground-water monitoring, the site received approval
for final closure in March 1994, 7 years after discovery
of the contamination.
PSCo, a business without extensive expertise in envi-
ronmental restoration, undertook the project with a fairly
direct goal: to remediate the site as efficiently and cost-
effectively as possible. Intrinsic bioremediation was
found to be a key element of the project's success. In
1988, little was known about intrinsic bioremediation,
and some opportunities may have been lost; indeed,
additional knowledge of these fields might have reduced
costs.
For example, in investigating whether bioremediation
was feasible, emphasis was placed on enhancing the
activities of the bacteria that would perform the reme-
diation, rather than on determining how long the en-
hancement could sustain bacterial growth, which
might have revealed a more effective strategy for
delivering nutrients. Once the nutrients penetrated
across the site, batch feeding of the nutrients may have
been more effective for the sustained growth of the
bacteria than the continuous recirculation of the en-
riched ground water. If this fact had been considered
before the start of the project, operation and mainte-
nance costs incurred during the project's life cycle could
have been reduced.
The risk assessment and ground-water data suggested
that the chance of any contaminants appearing in any
ground water being consumed or used by humans was
minimal. The natural ground-water flow across the site
was very slow, and the site might have remediated itself
with the help of a small amount of oxygen and other
nutrients.
To evaluate the progress of the project and to adhere to
Underground Injection Control regulations, many tests
were required that monitored the concentration of the
injected chemicals, or "nutrients"; these tests necessi-
tated costly site visits and report submissions. The
chemicals were not injected at harmful levels, however,
nor was the risk of human exposure significant due to
the location of the injections and the velocity of the
ground water. Three years into the project, the reporting
requirements were reduced to quarterly, but a better
understanding of the bioremediation process would
have reduced these requirements earlier.
The bioremediation process came in on budget and
successfully removed all threat to human health and the
environment (see Table 1).
Information from past projects, sound scientific judg-
ment, and risk analyses will provide an understanding
of intrinsic bioremediation that could increase the num-
ber of cleanups, speed up the closure process, and
reduce testing and reporting requirements, thereby sub-
stantially reducing costs to industry.
55
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Table 1. Resources Used at the Seventh Avenue
Bioremediation Project
1990 Labor hours 2,024 hours total
168 hours per month average
Activities O&M costs
Quarterly and monthly reporting
Testing cost $12,778
Materials cost $16,473
1991 Labor hours
Activities
Testing cost
Materials cost
1,746 hours total
145 hours per month average
Onsite inspections (once or twice a
week)
Risk assessment formulation
O&M costs
Quarterly and monthly reporting
$13,451
$12,000
1992 Labor hours 551 hours total
46 hours per month average
Activities Quarterly reports
Monthly site visits
Closure report submittal
Testing cost $11,539
1993 Labor hours 94.25 hours total
7.85 hours per month average
Activities
Testing
(quarterly)
Closure monitoring
$8,420
56
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A Regulator's Perspective of a Bioremediation Site
Lisa C. Weers
Colorado Department of Health, Denver, CO
The regulator's review and perception of a bioremediation
site differs from those of the person preparing the site
assessment plan/corrective action report and the site
representatives. This paper describes the regulator's role
at the Public Service Facility at 2701 W. 7th Avenue in
Denver, Colorado. At the time of corrective action plan
approval, the Colorado Department of Health (CDH)
reviewed the technology to be used. After 2 years of
postremediation monitoring, Public Service provided infor-
mation indicating that the points of compliance were not
affected, which is a consideration for closure.
The Public Service site is located within 1,000 feet of the
South Platte River, which supplies public water. On the
site is a garage and an office building, while the perime-
ter of the site is secured with a 6-foot fence. The location
of the building and the fence put some constraints on
the placement of the monitoring wells and the bioreme-
diation system.
The information CDH requested for this site is roughly
the same information found in any underground storage
tank site assessment; however, this site used a new
technology, which required that additional information
be provided. The site was first reviewed in 1990, when
there were no other active bioremediation sites in the
Underground Storage Tank Program (Storage Tank Re-
medial Section). Also, no written guidance documents
were available for performing site assessments. Be-
cause the technique at that time was relatively unknown,
finding a coworkerwith prior experience was impossible.
In addition, each staff member was assigned a large
number of sites because only three people were review-
ing sites for the state at the time. As a result, going to
the library during work time was out of the question.
When a new technology is reviewed, the person as-
signed to the site looks for information on the safety of
this remediation in regard to what the effects may be on
human health and the environment, how the technology
affects the sampling and analysis program, and whether
the proposed remediation has special considerations
that need to be incorporated in the approved corrective
action plan. Another question we often ask is, are other
appropriate permits obtained? GTI, the consulting firm
working on the project, was challenged on several oc-
casions by some very tough questions, such as, are the
fatty acids produced during bioremediation interfering
with your laboratory tests? Are any intermediate com-
pounds produced?
For a corrective action plan to be approved, the site
should receive a complete assessment. While reviewing
a site assessment, a technical reviewer examines the
report for completeness. The site assessment should
reveal the following information:
• Is there contamination in the subsurface?
• If so, how much and what type?
• Is the site the source?
• Where is it going?
• Has contamination reached ground water?
• Is ground water used?
• Is the contamination migrating off site?
• What are the potential receptors?
• What is the qualitative risk to potential receptors?
• How will subsurface conditions affect migration?
• Is remediation necessary?
• What are remediation options to consider?
Some of the main points we look for in approving a correc-
tive action plan are determination of the full extent of soil
and ground-water contamination; definition of the hydrol-
ogy of the site, the location of utility corridors, and the
geology of the site; a nearby surface and ground-water
assessment; a risk assessment; and a remediation plan.
After all the information is reviewed, a decision is made to
either approve or deny the corrective action plan.
When the Public Service site corrective action plan was
approved, EPA also evaluated the site for its Superfund
Innovative Technology Evaluation (SITE) program. We
were very excited that the site would be the subject of
research. The additional information that was provided
was useful in helping to determine a monitoring schedule.
57
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The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through
In Situ Bioremediation: The Regulator's Perspective
Mark E. Walker and Lisa C. Weers
Colorado Department of Health, Denver, CO
Because staff are assigned such a large number of
sites, reviewing all of the hard data for each and every
site is not conceptually possible. Therefore, we do not
want to receive all of the technical information and cor-
respondence generated at every site; instead, we prefer
the technical interpretation. Our approach to site evalu-
ation is to review this interpretation, reserving the right
to request any and all hard data generated during the
course of the investigation.
Although the technical staff at the Colorado Department
of Health (CDH) review an abridged version of the total
information generated at a particular site, our version of
the file on the Public Service Company of Colorado
(PSCo) occupied 6 feet of shelf space. Even a review of
the abridged version of this site presented a daunting
task.
I needed to get up to speed quickly on the file to consider
this site for closure. My first task was to verify if an
approved corrective action plan was in place and to
determine what conditions had been placed on this
approval, if any. In any case where ground-water con-
tamination has been documented, we check to see if the
point of compliance (POC) wells have been affected in
excess of state maximum contaminant levels (MCL).
The second task was to conduct an evaluation of the
existing monitoring program to determine if the program
was adequate for detecting contamination emanating
from the release. Based on the calculated ground-water
velocity, was the monitoring period long enough to de-
tect contamination from the source? With respect to
monitoring well construction and location, were these
wells placed downgradient? (Has direction changed with
the seasons?)
Afterthe monitoring plan was completed and the closure
borings had been drilled and analyzed, the next topic for
examination prior to closure was the extent of the re-
maining contamination and its potential for adversely
affecting the environment. This information is consid-
ered with the identified receptors in the immediate vicin-
ity and the potential for these receptors to be affected
by the remaining contamination.
After closure is complete, one might ask, "Well, where
does this site stand now, from a regulatory point of
view?" The following items should be considered: 1)
after closure is complete, CDH does not release the
owner/operator from any liability regarding the contami-
nation remaining in the subsurface; 2) the closure deci-
sion is based solely on the information submitted; and
3) similar to the owner's liability issue, CDH wants to be
informed of any developments that could increase the
potential for this site to affect human health and the
environment adversely, at which point we assess the
need for additional work/remediation.
In conclusion, CDH is willing to consider innovative
remedial technologies. Those who are proposing such
technologies should be prepared to demonstrate that
these technologies can be implemented in a manner
that does not adversely affect human health and the
environment. Proposal of an innovative technology
should be approached from the viewpoint of a regulator:
consider receptors and aspects of the technology that
might be detrimental to the environment, and address
these concerns in a responsible, forthright manner.
58
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Case Studies:
Sites Where Natural Attenuation Has Been Documented
59
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Intrinsic Bioremediation ofJP-4 Jet Fuel
John T. Wilson, Fredrick M. Pfeffer, James W. Weaver, and Don H. Kampbell
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK
Todd H. Wiedemeier
Engineering Science, Denver, CO
Jerry E. Hansen and Ross N. Miller
Air Force Center for Environmental Excellence, Brooks AFB, TX
Introduction
Intrinsic bioremediation is a risk management option
that relies on natural biological processes to contain the
spread of contamination from spills. The option is most
appropriate when the concentration of contaminants is
reduced to regulatory limits before ground water dis-
charges to surface water or is collected by a pumped well.
In the past, remedial action plans have proposed the
intrinsic remediation option based solely on the appar-
ent attenuation of contamination in water from monitor-
ing wells that are distant from the spill. These plans were
often criticized because it was impossible to distinguish
attenuation due to contaminant destruction from attenu-
ation due to simple dilution in the aquifer or in the
monitoring well. Convincing regulators that the wells
with low concentrations of contaminants actually sample
the plume of contaminated ground water has been diffi-
cult. This lack of credibility has led to the "one-more-well"
syndrome, with excessive investment in a monitoring
approach that focuses on the compounds of regulatory
concern but fails to earn the confidence of the regulatory
community.
During characterization of JP-4 jet fuel spills at Eglin Air
Force Base (AFB), Florida, and Hill AFB, Utah, three
approaches were used to distinguish contaminant at-
tenuation due to destruction from attenuation due to
dilution or sorption.
To distinguish attenuation due to biological destruction
of the contaminants from attenuation due to dilution, the
attenuation of the compounds of regulatory concern—
benzene, toluene, ethylbenzene, and the xylenes
(BTEX)—was compared with the attenuation of other
components of the fuel that were relatively recalcitrant.
Tracers have been used successfully to correct for dilu-
tion and sorption of hydrocarbons in ground water. Coz-
zarelli et al. (1) used 1,2,3,4-tetramethylbenzene to
normalize the concentrations of other alkylbenzenes
and their anaerobic degradation products in ground
water that had been contaminated by a spill of light
crude oil. Wilson et al. (2) used 2,3-dimethylpentane to
normalize the concentrations of BTEX compounds in
ground water contaminated with gasoline from an un-
derground storage tank. In both cases, the tracer was a
component of the spilled fuel.
To distinguish attenuation due to biological destruction
from attenuation due to sorption, core samples were
analyzed for the total quantity of individual BTEX com-
pounds and for total petroleum hydrocarbons. Partition-
ing theory was used to predict the concentration of
individual hydrocarbons in ground water in contact with
the core material. The predicted concentrations were
compared with concentrations in water from monitoring
wells to determine if the ground water in the plume was
in sorptive equilibrium with the spilled fuel.
To prove that the attenuation was due to biological ac-
tivity, the geochemistry of the ground water was exam-
ined. Microbial metabolism of petroleum hydrocarbons
has predictable geochemical consequences. The hydro-
carbons can be respired, resulting in the consumption
of oxygen, nitrate, sulfate, or iron II minerals in the
aquifer matrix and in the production of water, dinitrogen,
sulfide, or iron II. Microbiologists often refer to the sub-
strates for microbial respiration as electron acceptors.
Alkylbenzenes can be fermented, resulting in the produc-
tion of methane. Simple stoichiometry can be used to
predict the quantity of electron acceptors consumed orthe
quantity of methane produced during biotransformation
60
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of a given concentration of petroleum-derived hydro-
carbons.
Intrinsic Bioremediation of Ground Water
at Eglin AFB, Florida
Leaking distribution pipes from an underground storage
tank released JP-4 jet fuel to the water table aquifer
under the petroleum, oil, and lubricants storage depot
(POL) at Eglin AFB. The POL is located over sands and
silty peats characteristic of a barrier island complex. At
the time of the study, a plume of contaminated ground
water moved away from the residual JP-4 and dis-
charged to a small creek approximately 300 ft down-
gradient (Figure 1). The elevation of the water table is
approximately 8.4 ft above sea level in the area with
residual JP-4. The watertable in the creek is 1.4 ft above
mean sea level. Hydraulic conductivity determined by
pumping tests in monitoring wells varied from 48 ft to
102 ft per day. Based on these data, and assuming an
effective porosity of 30 percent, the residence time along
the flow path from the spill to the creek is on the order
of 10 wk. Water samples were acquired with a geoprobe,
using an 18-in. screen. Seasonally, the temperature of
ground water at the site varies from 19°C to 28°C, and
the pH varies from 5.6 to 6.7. Samples producing the
data in Table 1 were taken in August 1993, when ground-
water temperatures varied from 24°C to 28°C.
a Geoprobe Location
A Soil Boring
* 2" Monitoring Well
•f^ Line of Equal
Water Table Elevation
(Feet above Mean
Sea Level)
Scale
0' 50'
A — A Line of Hydrogeologic Section
Figure 1. Flow path of ground water from a JP-4 spill at Eglin
AFB to the point of discharge to surface water: plan
view.
Correcting for Dilution at Eglin AFB
Table 1 presents the changes in concentrations of BTEX
compounds and the three trimethylbenzenes (TMB)
along the flow path from the spill to the creek (Figures
1 and 2). Sample 80H-3 is from a location just outside
Table 1. Bioattenuation in Methanogenic Ground Water: Changes in Ground-Water Chemistry Along a Flow Path From a JP-4
Spill to the Point of Discharge to Surface Water
Location Along Flow Path to Surface Water
Compound
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
1 ,3,5-TMB
1,2,4-TMB
1 ,2,3-TMB
BTEX and TMB
Oxygen
Nitrate + Nitrite - N
Sulfate
Methane
Iron II
80 H -3
100
5,150
1,700
3,120
6,750
5,480
327
1,090
406
24.1
0.4
0.14
1.6
3.7
2.3
83Z-2
153
18.3
227
594
1,270
<1
114
420
182
2.97
0.2
0.12
0.62
16.8
7.8
83 U -3
(i-ig/L)
198
1.1
1.4
13.8
13.5
<1
70.9
172
115
(mg/L)
0.59
0.3
O.05
5.6
12.5
3.3
83U-2
6.9
<1
1.4
25.1
39.8
<1
119
299
187
0.68
0.6
<0.05
1.76
14.2
2.8
83U-1
<1
<1
<1
<1
<1
<1
2.3
<1
<1
0.002
3.8
O.05
<0.5
0.7
<0.5
61
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5-20
Figure 2.
Legend
x Sample Location
-*- Water Table
Flow path of ground water from a JP-4 spill at Eglin
AFB to the point of discharge to surface water: cross
section.
the JP-4 spill and appears to be in chemical equilibrium
with the weathered residual fuel. Samples 83H-1 and
83Z-2 are from locations approximately 150 ft and 300 ft
downgradient from the spill. Samples 83U-2 and 83U-3
were taken 0.5 ft and 4.1 ft below the sediments of the
creek receiving discharge from the plume. Sample 83U-1
is water from the creek at the sediment boundary, taken
when the tide was going out and the plume was actively
discharging to the creek.
When trimethylbenzene concentrations in samples 80H-3
and 83U-3 were compared, the reduction in the concen-
tration of these compounds was found to be remarkably
uniform. Concentrations under the creek were 36, 27,
and 46 percent of the concentrations nearthe spill, while
the concentrations of toluene, ethylbenzene, p-xylene,
m-xylene, and o-xylene were 0.02, 0.08, 0.44, 0.20,
and less than 0.02 percent of the initial concentrations,
respectively.
Apparently, concentrations were reduced from one-half
to one-fourth of the initial concentration due to dilution,
with further reductions due to biological activity. Ben-
zene was not degraded in the anaerobic portion of the
flow path.
As the plume moved up into the sediments of the creek,
the concentration of benzene was reduced more than
20-fold (compare 83U-2 with 83U-3 in Table 1). If we
assume that the trimethylbenzenes are recalcitrant, the
dilution of 1,2,4-trimethylbenzene can be used to correct
for the dilution of benzene and determine the true re-
moval due to biodegradation. The corrected concentra-
tion of a biologically transformed compound in a
downgradient well would be the measured concentra-
tion in the downgradient well, multiplied by the meas-
ured concentration of 1,2,4-trimethylbenzene in the
upgradient well, and divided by the measured concen-
tration of 1,2,4-trimethylbenzene in the downgradient
well.
The concentration of benzene measured in sample 83U-2
was 6.9 |j,g/L. Compared with sample 83Z-2, the con-
centration corrected for dilution would be 9.7 |j,g/L. Com-
pared with 153 |j,g/L in 83Z-2, there was a 15-fold
attenuation in benzene concentration between 83Z-2
and 83U-2. The benzene attenuation at 83U-2 com-
pared with 83U-3 would be 50-fold.
The concentration of oxygen in the sample 83U-2, taken
0.5 ft below the sediment surface, was higher than the
concentration in sample 83U-3, taken 4.5 ft below the
sediment surface. Tidal action may reverse the hydraulic
gradient in the area proximate to the creek. This would
produce a reciprocating flow of oxygenated creek water
into the sediments and mix oxygen into the contami-
nated ground water. Benzene may have been degraded
aerobically. In any case, benzene and the other BTEX
compounds did not discharge to the stream at detect-
able concentrations (see sample 83U-1).
Kinetics of Bioattenuation in Ground
Water at Eglin AFB
First-order rate constants were calculated by correcting
the downgradient concentration for dilution. The rate
constants were calculated as:
Rate =
ln(corrected cone, downgradient/cone, upgradient)
residence time
Based on this relationship, rates were calculated for flow
path segments from samples 80H-3 to 83Z-2 and 83Z-2
to 83U-3. The residence time in each segment was
assumed to be 5 wk.
The rates of anaerobic bioattenuation (Table 2) were
some of the fastest that have ever been encountered by
staff of the Robert S. Kerr Environmental Research
Laboratory, probably due to the high water tempera-
tures. There seemed to be preferential removal of tolu-
ene and o-xylene in the segment close to the spill.
Rates of removal of ethylbenzene and m+p-xylene in-
creased after toluene and o-xylene were depleted.
Table 2. First-Order Rate Constants for Bioattenuation of
BTEX Compounds in a Plume of Ground Water
Contaminated by JP-4
Compound
80H-3 to 83Z-2
83Z-2 to 83U-3
(per week)
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
None
-0.94
-0.21
-0.14
-0.14
-1.5
None
-0.38
-0.38
-0.57
-0.73
Cannot calculate
62
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Stoichiometry of Bioattenuation at
EglinAFB
There was very little oxygen, nitrate, orsulfate available
for respiration of the BTEX compounds in the plume of
contaminated ground water (see Table 1). There was an
increase in the concentration of methane and iron and
a corresponding decrease in aromatic fuel hydrocar-
bons along the flow path, however. Assuming the follow-
ing Stoichiometry for methanogenesis from BTEX
compounds,
2 CH + 3/2 H20 -» 5/4 CH4 + 3/4 CO2,
approximately 1.0 mg of methane is produced for each
1.3 mg of BTEX destroyed. If 80H-3 and 83Z-2 are
compared, the concentration of methane at 83Z-2, cor-
rected for dilution, would be 43.5 mg/L. The increase in
methane from 80H-3 to 83Z-2 would be 39.8 mg/L.
Corrected for dilution, the concentration of BTEX plus
TMB in 83Z-2 is 7.69 mg/L, a decrease of 16.4 mg/L
compared with 80H-3. This decrease in aromatic petro-
leum-derived hydrocarbons would be expected to pro-
duce 12.6 mg/L of methane. Some of the methane
sampled at 83Z-2 may have come from the natural
degradation of the peat. In any case, the accumulation
of methane is sufficient to rationalize the destruction of
the aromatic fuel hydrocarbons along the flow path.
Intrinsic Bioremediation of Ground Water
at Hill AFB, Utah
Hill AFB is situated on a bird's foot delta formed by the
Weber River in Pleistocene Lake Bonneville. Leaking
distribution pipes from an underground storage tank
released JP-4 jet fuel to the water table aquifer under
the POL. An area with oily phase JP-4 extends approxi-
mately 1,000 ft downgradient of the spill (Figure 3). The
oily phase hydrocarbons and the plume of contaminated
ground water are confined to the channel sands of the
delta deposits (Figure 4). The elevation of the water
table drops 56 ft across the length of the spill. Slug
testing of monitoring wells indicates a hydraulic conduc-
tivity near 8.5 ft per day, corresponding to an interstitial
seepage velocity of 1.6 ft per day.
Continuous cores were taken near the source of the spill
(82! in Table 3), near the midpoint (82D), at the lower
edge of the spill (82C), and just beyond the edge of the
spill (82B). The continuous cores started in clean mate-
rial above the spill and extended through the spill into
clean material underneath the interval containing hydro-
carbon. The continuous cores were subdivided into core
samples representing 0.3 vertical ft of the subsurface.
Near the spill, JP-4 appeared in the capillary fringe
(Table 3). The concentration maximum of 14,800 mg/kg
was located 0.5 ft above the water table. A second
interval that contained JP-4 extended from 3.8 to 5.2 ft
o
IW-10
Extent of JP-4 Spill
Scale
0 300 600
Feet
Storm
- Drain
MW-10 Monitoring Well Location
B-B' Line of Hydrogeologic Section
Figure 3. Flow path of ground water through a JP-4 spill at Hill
AFB to the point of discharge in a storm drain: plan
view.
SWB
LU
Legend
I Poorly Sorted Sands
| Silty or Clayey Sands
. Ground Water Level
I Silt or Silty Clay
Storm Sewer
0 150 300
Feet
Figure 4. Flow path of ground water from a JP-4 spill at Eglin
AFB to the point of discharge in a storm drain: cross
section.
below the water table, with TPH concentrations that
ranged from 1,290 to 3,830 mg/L. At the midpoint of the
spill (82D), only one core sample contained significant
concentrations of hydrocarbons. That core sample
came from 1.0 ft below the water table. At the lower
edge, two core samples representing 0.6 vertical ft had
significant concentrations of hydrocarbons. Again these
core samples were at or just below the water table. At
the midpoint and lower edge of the spill, hydrocarbons were
not detected in core material collected above or below the
samples reported in Table 3 (detection limit 10 mg/kg).
Concentrations of hydrocarbons in the midpoint and lower
edge of the spill were less than 700 mg/kg.
63
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Table 3. Vertical Distribution of Oily Phase Hydrocarbons at Hill AFB
Core
821-5
821-4
821-3
821-2
821-1
82l-27to82l-19
821-17
821-37
821-32
82D-24
82D-23
Elevation
(ft)
4,665.7
4,665.4
4,665.0
4,664.6
4,664.3
4,663.9 to
4,660.7
4,659.3
4,659.0
4,631 .0
4,630.7
TPH
821, near the spill
4,330
3,770
1 4,800
5,870
398
4,661 .0 <300
1,290
1,370
3,830
82D, downgradient
77.1
572.0
Benzene
area, water table
0.0326
0.517
4.55
0.401
<0.01
—
0.653
0.712
0.0136
Toluene
(mg/kg)
elevation 4,664.53 ft
0.0266
0.235
2.73
12.6
0.142
—
0.591
0.182
0.032
Ethyl-
benzene
14.5
4.8
47.7
17.5
0.556
—
3.39
2.72
1.24
1,2,4-
TMB
49.9
42.7
167
69.8
4.43
—
2.34
5.38
8.37
of spill, water table elevation 4,631.7 ft
—
0.271
82C, lower end of spill, water table
82C-20
82C-19
82B-20
82B-19
4,603.5
4,603.2
4,608.0
4,607.7
593.0
638.0
82B, below the
0.7
1.0
O.01
0.0062
<0.01
elevation 4,603.4 ft
0.0176
<0.01
1.48
0.00618
0.0180
3.11
1.03
1.16
spill, water table elevation 4,608.3 ft
O.01
O.01
O.01
O.01
O.01
O.01
O.01
O.01
Monitoring wells were installed in the boreholes used to
acquire the core samples. Additional monitoring wells
were installed downgradient of the lower edge of the
spill.
Correcting for Dilution at Hill AFB
The TMBs were remarkably persistent in ground water
from the area containing oily phase hydrocarbons (Ta-
ble 4). Concentrations of the TMBs in water from the
lower edge of the spill varied from 50 percent to 147
percent of the concentration at the source, while the
concentrations of benzene, toluene, ethylbenzene, p-
xylene, m-xylene, and o-xylene were reduced to 0.02,
0.9, 5.6, 5.5, 3.4, and 0.23 percent of the initial con-
centration, respectively. The saturated thickness of
the channel sands was at most a few feet, and, there-
fore, there was little opportunity for dispersive dilution
of the contaminated ground water into clean ground
water underneath the plume. Once the ground water
moved past the spill, it was remediated rapidly. Concen-
trations of all the aromatic hydrocarbons are low in
well 82B just past the lower edge of the spill.
Stoichiometry of Bioattenuation at
Hill AFB
Similar to the plume at Eglin AFB, neither oxygen nor
nitrate is available for respiration of the BTEX com-
pounds within the spill area (Table 4). Oxygen concen-
trations as high as 8.0 mg/L occur outside the spill.
Unlike the case at Eglin, there was little accumulation of
iron II and practically no methane production in the spill.
Sulfate concentrations were high throughout the spill.
Assuming the following Stoichiometry,
8 CH + 5 SO4_2 -> 5 S'2 + 8 CO2 + 4 H2O,
4.6 mg of sulfate would be required to degrade 1.0 mg
of BTEX and TMB. The 7.7 mg/L of BTEX and TMB
present at the source would have a total theoretical
demand of 35 mg/L of sulfate. Concentrations in excess
64
-------
Table 4. Changes in Ground-Water Chemistry Along a Flow Path Through a JP-4 Spill Undergoing Sulfate Reduction
Location Along Flow Path to Surface Water
Compound
821
MW-11
82D
82C
82B
(Mfl/L)
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
1,3,5-TMB
1,2,4-TMB
1 ,2,3-TMB
BTEX and TMB
Oxygen
Nitrate + Nitrite - N
Sulfate
Methane
Iron II
2,740
327
486
784
1,370
1,140
162
495
240
7.7
—
—
—
0.68
—
336
90
139
230
635
204
71
165
69
2.1
0.1
0.4
98
0.022
0.05
96
10
147
149
383
103
129
183
89
(mg/L)
1.3
1.3
0.5
193
<0.001
1.7
4.9
3.1
27
43
47
2.6
238
324
120
2.1
0.5
0.1
50
0.002
0.84
<1
4.3
<1
<1
<1
<1
1.1
1.4
<1
0.001
1.2
0.4
64
<0.001
0.11
of the theoretical demand remain in the water after it has
moved away from the spill. There is an adequate supply
of sulfate to remediate the plume through sulfate reduc-
tion alone.
Concentrations of sulfate were higher in ground water near
the source. It is possible that the lower concentrations of
sulfate nearthe lower edge of the spill represent the sulfate
demand exerted. It is just as possible that they represent
natural variations in sulfate concentrations.
Hydrogen sulfide did not accumulate in ground water
undergoing intrinsic remediation at Hill AFB. Unim-
pacted sediments at the site are tan or brown colored,
while contaminated samples are black. Iron minerals in
the aquifer matrix must act to precipitate sulfide. Sulfide
very likely precipitated as iron II sulfide.
The concentrations of electron acceptors were com-
pared in ground water at various distances from the
lower edge of the spill. Well 82B is just outside the lower
edge; wells 82F, 82H, and 82E are 300, 400, and 500 ft
downgradient of the spill. These wells are screened in
the water-table aquifer (Figure 4); well 82A is adjacent
to 82F but is screened in the first confined aquifer un-
derneath the water table aquifer.
None of the wells outside the spill had significant concen-
trations of aromatic hydrocarbons (Table 5). Water adja-
cent to the spill (82B) was depleted in oxygen and nitrate.
Intrinsic Bioremediation of Oily Phase
JP-4 Jet Fuel at Hill AFB
Two samples of floating oil from monitoring well MW-10,
near the midpoint of the spill, were analyzed by gas
chromatography/mass spectrometry (GC/MS) to deter-
mine the number-average mean molecular weight of the
weathered fuel. Values of 156 and 160 daltons were
determined on duplicate samples.
The concentration of individual hydrocarbons in ground
water in contact with oily phase JP-4 was estimated
using Raoult's Law. With data from Table 3, the concen-
tration of an individual petroleum hydrocarbon (mg/kg)
was divided by its molecular weight to express its con-
centration in moles/kg core material. The concentration
of total petroleum hydrocarbons (TPHs) was divided by
160 to express TPH in moles/kg core material. The mole
fraction of the individual hydrocarbon was calculated by
dividing moles of individual hydrocarbon per kilogram by
moles of TPH per kilogram. Then the mole fraction was
multiplied by the water solubility of the individual petro-
leum hydrocarbons to estimate the equilibrium concen-
tration in ground water.
Concentrations of individual petroleum hydrocarbons in
water from monitoring well 82I, near the source of the
spill, were in reasonable agreement with the concentra-
tions predicted from the cores that had the highest TPH
65
-------
Table 5. Changes in Ground-Water Chemistry Along a Flow Path Downgradient of a JP-4 Spill
Location Along Flow Path to Surface Water
Compound
82B
82F
82E
82H
82A
BTEX
TMB
(Mfl/L)
(mg/L)
0.9
2.0
Oxygen
Nitrate + Nitrite - N
Sulfate
Methane
Iron II
1.2
0.4
72.2
O.001
0.11
1.1
5.1
55.5
0.006
0.04
2.7
5.6
65.8
O.001
0.02
5.4
2.0
55.7
O.001
0.19
0.3
0.1
60.5
0.001
0.1
values (see cores 82I-3, 821-17, and 821-37 in Table 6
and Table 3). The JP-4 in the capillary fringe appeared
to be aerobically weathered. There was extensive re-
moval of benzene and toluene in the more shallow cores
82I-5 and 82I-4, while ethylbenzene and 1,2,4-TMB
were not removed.
Concentrations in water from the well at the midpoint of
the spill (82D) were considerably lower than concentra-
tions predicted from the core material. Oily-phase material
at this location could not remove individual hydrocar-
bons from ground water through partitioning processes.
Core material at the lower edge of the oil spill (82C) was
very highly weathered with respect to benzene, toluene,
and ethylbenzene, while weathering of 1,2,4-TMB was
less extensive. When core 821-37 near the source and
82C-19 at the lower edge were compared, the mole
fraction of benzene was reduced by a factor of 53, while
the mole fraction of 1,2,4-TMB was reduced by a factor
of 2.2. When concentrations in monitoring well 82! near
the source and 82C near the lower edge were com-
pared, the concentration of benzene was reduced by a
factor of 560 while the concentration of 1,2,4-TMB was
reduced by a factor of 1.5.
Within an order of magnitude, the concentrations in
ground water were in equilibrium with the mole fraction
in the oily phase. Concentrations predicted for ground
water were higher than measured concentrations, indi-
cating that sorption or partitioning could not be respon-
sible for removal of the individual hydrocarbons from
ground water.
Kinetics of Bioattenuation at Hill AFB
There is no straightforward approach to calculate the
kinetics of bioattenuation in ground water in contact with
oily-phase material. Forthe ground-water plume at Eglin
AFB, the distance between wells along a flow path and
the estimated interstitial seepage velocity of ground
water was used to estimate residence time of the con-
taminants. At Hill AFB, the contaminated ground water
is in contact with oily-phase material. The major fraction
of individual contaminant hydrocarbons in the aquifer is
partitioned to the oily phase, which is moving slowly if it
is moving at all. Travel time of ground water is not related
to residence time of contaminants, making it impossible
to determine the kinetics of attenuation from data at
different locations collected at the same time. Kinetics
must be inferred from long-term monitoring data, which
are not available at the present time.
If we assume that the spill started 30 years ago, that the
JP-4 at the lower edge was the first oil spilled, and that
the mole fraction of benzene in JP-4 used at Hill AFB
has not changed appreciably over time, a comparison
of the mole fraction of benzene in 83I-3 near the source
and 83C-19 near the lower edge of the spill indicates a
reduction in the mole fraction of benzene to 3 percent of
the original concentration. If kinetics are first order on
time, the rate would be -0.11/year.
References
1. Cozzarelli, I.M., R.P. Eganhouse, and M.J.
Baedecker. 1990. Transformation of monoaromatic
hydrocarbons to organic acids in anoxic ground-
water environment. Environ. Geol. Water Sci.
16(2):135-141.
2. Wilson, J.T., D.H. Kampbell, and J. Armstrong. 1994.
Natural bioreclamation of alkylbenzenes (BTEX)
from a gasoline spill in methanogenic ground water.
In: R.E. Hinchee, B.C. Alleman, R.E. Hoeppel, and
R.N. Miller, eds. Hydrocarbon bioremediation. Ann
Arbor, Ml: Lewis Publishers.
66
-------
Table 6. Comparison of the Concentration of Benzene, Toluene, and Ethylbenzene Measured in Ground Water to the
Concentration Predicted Using Partitioning Theory From the Mole Fraction of Those Compounds in Extracts
of Core Material
Core
11/93
82I-5
82I-4
82I-3
82I-2
821-1
821-27 to 821-19
821-17
821-37
821-32
Elevation
(ft)
—
4,665.7
4,665.4
4,665.0
4,664.6
4,664.3
4,663.9 to 4,661
4,660.7
4,659.3
4,659.0
Benzene
821, near the spill area, water
2,740*
22.8
513
1,150
254
103
.0 <300
1,870
1,940
13.2
Toluene
Predicted
table elevation
372*
5.5
55.6
165
1,923
320
—
410
119
7.5
Ethylbenzene
Concentration (ug/L)
4,664.53 ft
486*
769
289
740
684
321
—
603
456
74.3
1,2,4-TMB
495*
900
884
881
929
869
—
135
307
171
82D, downgradient of spill, water table elevation 4,631.7 ft
7/93
11/93
82D-23
7/93
11/93
82C-20
82C-19
4,631.7
4,631.7
4,630.7
4,603.4
4,603.0
4,603.5
4,603.2
96*
174*
1,769
82C, lower end of spill, water
4.9*
-------
A Natural Gradient Tracer Experiment in a Heterogeneous Aquifer With Measured
In Situ Biodegradation Rates: A Case for Natural Attenuation
Thomas B. Stauffer and Christopher P. Antworth
Armstrong Laboratory, Tyndall AFB, FL
J. Mark Boggs
Engineering Laboratory, Tennessee Valley Laboratory, Morris, TN
William G. Maclntyre
School of Marine Science, College of William and Mary, Gloucester Point, VA
Biodegradation rates of organic compounds have been
measured in a heterogeneous unconfined aquifer at
Columbus Air Force Base (AFB), Columbus, Missis-
sippi, during a pulse release experiment. Reaction rate
calculations were based on a kinetic model that includes
the hydrologic characteristics of the aquifer. Degrada-
tion kinetics were approximately first order with the fol-
lowing rate constants: benzene, 0.0066 d~1; p-xylene,
0.0141 d'1; naphthalene, 0.0063d'1; o-dichlorobenzene,
0.0059 d'1.
Introduction
Biodegradation rates of organic contaminants in aqui-
fers are needed for use in fate and transport models.
Field experiments for determination of in situ biodegra-
dation rates are desirable because laboratory measure-
ments may not relate to conditions in an aquifer. Madsen
(1) notes in a recent review "that determination of micro-
bial activity in disturbed, displaced environmental
samples incubated in the laboratory is likely to be quan-
titatively, even qualitatively, different from the same de-
termination in situ." Reliable estimation of an in situ
biodegradation rate requires introduction of a known
mass of contaminant at a defined time zero, and obser-
vation of contaminant concentration variation in space
and time. Confirmation of biodegradation requires main-
tenance of mass balances and determination of organic
compound degradation products. To our knowledge, no
experiments to accurately measure biodegradation
rates and confirm biodegradation in an aquifer are found
in the literature.
The objective of this research was to measure in situ
biodegradation rates of organic compounds. Accordingly,
an experimental pulse injection of tritiated water and
organic solutes has been conducted at Columbus AFB
at the macrodispersion experiment (MADE) site. The
injection was into the saturated zone of the heterogene-
ous unconfined aquifer formed by fluvial sedimentation.
Kinetics of in situ biodegradation of benzene, p-xylene,
naphthalene, and o-dichlorobenzene in the Columbus
aquifer are reported here and related to the structural
and hydrologic properties of the aquifer. The Columbus
aquifer material has a wide range of particle sizes and
large spatial variation in horizontal hydraulic conductivity
(Kh), from <10"4 to 1 cm/s. The heterogeneity is in con-
trast to the generally lower (10"4 to 10"3 cm/s) values and
less variable Kh field at the Borden Canadian Forces
Base, Ontario (2).
A summary of hydrogeologic properties of the Columbus
aquifer and a description of the MADE site are given by
Boggs et al. (3). Rehfeldt et al. (4) measured the spatial
distribution of hydraulic conductivity at the MADE site
using borehole flow meters and othertechniques. Figure
1 shows the Kh distribution over a vertical section di-
rected along the test plume axis, and indicates large-
scale heterogeneity and structures that Rehfeldt et al.
(4) refer to as channels. Close to the injection wells K^
is relatively low, with values near 10"3 cm/s from the
phreatic surface to the lower confining layer and extend-
ing about 40 m downgradient from the injection location.
Immediately beyond this region, Kh increases to values
near 10"1 cm/s in the uppers m of the aquifer, which are
maintained out to 200 m, while Kh in the lower portion
of the aquifer remains low. Thus, a near surface channel
crosses the intermediate and far field at the site. These
observations suggest that solute transport and distribution
68
-------
65
60 -
55
(0
I 50
45
-10° 10~1 10~2 10"3 10"4 (cm/s)
Flow-
0 25 50 75 100 125 150 175 200 225 250 275
Distance (m)
Figure 1. Distribution of hydraulic conductivity over a vertical
section containing the center line of the plume.
might be analogous to the behavior of a hypothetical
leaking reactor located at the injection wells, with the
near field region of the aquifer representing the reactor
vessel and the upper portion of the far field serving as
the leak. A leaky reactor kinetic model including radio-
active decay and biodegradation was developed for use
with field data to obtain in situ biodegradation rates for
organic compounds in the Columbus aquifer. Radiocar-
bon measurements using 14C-labeled p-xylene were
done to confirm biodegradation of p-xylene to its degra-
dation products.
Experimental Methods
The experiment (MADE2) is generally similar in design
to the Borden site test described by Mackay et al. (5). A
2-day pulse injection of water containing 3H2O and or-
ganic solutes was begun on June 26, 1990. Injection
wells were closely spaced on a line normal to the flow
direction at locations given in Boggs et al. (3). The
intersection line formed by the well plane and a vertical
plane along the flow path is shown in Figure 1. Locations
of injection and sampling wells are shown in plan view
in Figure 2.
Experimental parameters for the MADE2 test were in-
jection volume, 9,600 L; injection time, 47.5 hr; injection
wells screened over a 0.6-m interval at 4 m below the
phreatic surface in the saturated zone of the aquifer; and
five injection wells in a line normal to the hydraulic
gradient and spaced at 1 -m intervals with equal flow to each
well. Injection concentrations were: tritium, 55.6 nCi/mL;
benzene, 68.1 mg/L; naphthalene, 7.23 mg/L; p-xylene,
51.5 mg/L containing ring-labeled 14C p-xylene,
2.77 nCi/mL; o-dichlorobenzene, 32.8 mg/L. Tritium
and 14C were analyzed by liquid scintillation counting of
water samples. Concentrations of the organic com-
pounds in water were determined by solvent extraction
into pentane containing toluene as an internal standard,
and analysis of the extract was made by gas chroma-
tography (GC) with flame ionization detection. Most
sampling wells contained multilevel samplers, but BAR-
CAD water samplers were used in a few wells. Water
was collected from the sampling wells and analyzed to
provide three-dimensional snapshots of solute concen-
300
250
200
150
100
50
Barcad a
Multilevel Sampler •
Injection Well a
Inset
Injection
Point ""
20
15
10
5
0
-15-10 -50 5 10 15
-50
-100
I
-50
50
x(m)
-100
150
200
Figure 2. Plan view of the injection and sampling wells at the
MADE2 site.
trations at 27, 132, 224, 328, and 440 d after injection.
Statistical moments for each snapshot were calculated
by establishing a triangular grid followed by vertical
integration of concentrations at each well, and spatial
integration was calculated over the grid by the method
of Garabedian (6), which was previously applied at the
MADE site by Adams and Gelhar (7). The agreement
between the calculated zeroth moments and the mass
of 3H2O injected confirmed mass balance for3H2O. Se-
lected wells from high and low solute concentration
regions of the plume were analyzed for p-xylene degra-
dation by 14C counting methods 421 d after injection.
Each water sample was subsampled in the field. A2-mL
subsample was mixed with cocktail and counted to ob-
tain total 14C in the water as either p-xylene or its
degradation products. A 20-mL subsample was ex-
tracted with 2 mL of unlabeled p-xylene, and 1 mL of the
xylene layer was transferred to scintillation cocktail and
counted for 14C p-xylene, which remained in the organic
layer. Finally, a 10-mL subsample was made basic by
adding 1 mL 3N NaOH and 1 mL 3N Ca(NO3)2, and
centrifuged to isolate a precipitate containing Ca14CO3.
The washed precipitate and 2 mL of supernatant water
were placed in separate scintillation vials and counted.
These analyses determined the amount of 14C carbon-
ate produced by complete degradation of 14C p-xylene
and the amount of 14C in water-soluble organic interme-
diate products, respectively.
Leaky Reactor Model
Organic solutes remaining in the low Kh region near the
injection wells may be regarded as semiconfined in a
69
-------
INJECTION
WELL
GROUND
WATER
FLOW
REACTOR ZONE
KB[B]
KX[X]
KC[C]
DEGRADATION
• RATES
CHANNEL
K (LEAK RATE CONSTANT)
Figure 3. Diagram of the leaky reactor model.
reactor zone that is stirred by advection and dispersion.
This reactor is considered to leak by advection through
a near surface channel with relatively high Kh, with
replacement water supplied by ground-water flow
across the upstream boundary of the reactor zone. The
conceptual model of this situation is shown in Figure 3.
This model compares well with the Y^ distribution shown
in Figure 1. Reactions occurring in the reactor zone are
tritiated water (Tr) decay, with a rate constant (kd) of
1.548 x 10"3 d"1, and biodegradation of organic com-
pounds, assumed to be first order with rate constants kt,,
kx, kn, and ko for benzene (B), xylene (X), naphthalene
(N), and o-dichlorobenzene (C), respectively. The leak
rate for each solute from a well-mixed reactor should be
first order, with the same constant, kh for all solutes.
Sorption-desorption processes at aquifer material sur-
faces are assumed to be very fast relative to the other
processes considered here, and to be rate-limited by
physical transport of organic solute molecules to and
from the solid-solution interface. Biodegradation and
leakage occur simultaneously in the reactor zone by the
scheme given in Figure 3. The overall kinetics can be
expressed by the following set of linear differential equa-
tions with initial conditions: at t = IQ = 0, [Tr]0 = 0.539 Ci,
[B]0 = 660 g, [X]0 = 402 g, [N]0 = 70 g, [C]0 = 318 g, where
square brackets represent solute mass and t is time
following injection.
Analytic solutions of equations 1 to 5 with these condi-
tions are the integrated rate laws (equations 6 to 10)
describing solute concentrations as functions of time.
The leak rate constant k, is calculated by solving equa-
tion 6, using the known kd of tritium. Constants kt,, kx, kn,
and kc are then determined by substitution of k, and
solving equations 7 through 10.
d[Ti\
dt
•*£-*«•
k,[E\
(1)
(2)
dlX\
dt
dM.
dt
d[C]
dt
= kn[N\ + k,[N\
= kc[C] + k,[C]
(3)
(4)
(5)
In
k, = -
[7>]
-k,.
»$) "ffl
V - + ^—L+kd
t
t
mfWo) mfm
Mo m\[X\
kx =
t
t
+ kri
In
In
t
t
(6)
(7)
(8)
(9)
(10)
Organic solutes were retarded slightly relative to 3H2O
by weak sorption on the aquifer material (8), but this
effect was negligible when compared with concentration
changes due to biodegradation. Laboratory batch sorp-
tion coefficients on a composite sample of Columbus aqui-
fer material for naphthalene, o-dichlorobenzene, p-xylene,
and benzene were 0.085, 0.065, 0.048, and 0.059 L/kg,
respectively. Accordingly, this model does not incorporate
terms for sorption of solutes in the reactor zone.
Results
The distribution of 3H2O in sampling wells on the plume
axis 224 d after injection is presented in Figure 4a. It is
apparent that a large fraction of the 3H2O mass re-
mained near the injection wells after 224 d. Biodegrada-
tion of organic solutes occurred primarily in a reactor
zone approximately delineated by the p-xylene distribu-
tion at 224 d after injection. Figure 4b shows that
p-xylene was approximately confined within 20 m of the
injection wells. The same distribution pattern also held
for the other organic solutes. Confinement of organic
solutes to the reactor zone occurred because their
biodegradation was rapid relative to the leak rate from
the reactor zone.
70
-------
65
60
55
50
45
Tritium
C/Co
(a)
25
50
75
100
125
150
175
200
225
250
275
65
60
55
50
45
p-Xylene
10"
10
C/Co
ID'2
10"
10
(b)
25
50
75
100
125
150
175
200
225
250
275
Figure 4. Distribution of (a) normalized tritium and (b) normalized p-xylene concentration zone over a vertical section containing
the center line of plume motion.
The spatial distributions of Kh in Figure 1 and of 3H2O in
Figure 4a are similar, which is expected because hy-
draulic conductivity is the controlling parameter for
transport of an unretained solute. High 3H2O concentra-
tions remained in the neighborhood of the injection wells
224 d after injection, with lower concentrations in the
upper portion of the aquifer from about 30 m to 250 m
downgradient. There was little 3H2O downgradient in the
lower portion of the aquifer between 30 m and 175 m
depth. This distribution indicates that solutes were trans-
ported downgradient from the low Kh near the injection
wells through an upper channel whose size, location, and
Kh govern the solute loss rate from the source region. Very
little transport occurred in the lower portion of the aquifer.
The observed transport of 3H2O during this experiment
was consistent with the leaky reactor model.
Leakage of solutes was determined by the loss of tritium
from the reactor zone, defined as the portion of the
plume volume within 10 m downstream from the injec-
tion wells. The amount of tritium in the reactor zone at
a given time was calculated by spatial integration. Total
tritium in this zone decreased exponentially with time,
as shown in Figure 5. The tritium values have been
corrected for radioactive decay. This plot indicates that
0.
-2-
0
100
200
300
400
500
Days After Injection
Figure 5. Tritiated water content in the reactor organic as a
function of time.
solute leakage from the reactor is first order, with leak-
age rate constant, k, = 5.45 x 10'3 d'1. A mass balance
for tritium calculated using spatial integration over the
entire plume showed that the injected tritium mass was
accounted for at the times given in Figure 5. This implies
that the tritium decrease in the reactor zone was indeed
due to a process analogous to leakage.
71
-------
Biodegradation rates within the reactor zone are ap-
proximately first order, as shown in Figure 6. Departure
of the curves from linearity is attributed to microbial
processes. There is an apparent lag period for microbial
activity soon after the injection, during which degrada-
tion rates are low. This lag period is followed by degra-
dation at a maximum rate, which is characteristic of the
microbial metabolism. Finally, the rate decreases late in
the reaction, as the solutes (substrates) are depleted.
Biodegradation rate constants for the organic solutes in
the reactor zone (kmax(reactor)), determined from the
maximum slopes in Figure 6, are presented in Table 1.
The maximum rate constants (kmax(corr)(reactor)) in the
second column have been corrected for organic solute
leakage from the reactor zone. Maximum biodegrada-
tion rate constants for each organic solute were also
calculated by spatial integration overthe entire well field,
and are included in Table 1 as kmax(whole field). These
values are independent of the leaky reactor model.
Degradation rate calculations for p-xylene are based on
analyses of p-xylene by GC. Ring-labeled 14C p-xylene
was included in the injection solution to demonstrate
that reductions in p-xylene concentration were a result
of biodegradation. Microorganisms mineralize 14C p-
xylene to water-soluble labeled intermediates and14CO2
(predominantly as H14CO3- at ground-water pH). 14C
counting of whole water samples does not distinguish
between degraded and intact p-xylene, and thus does
not measure degradation of 14C-labeled organic com-
pounds. Detection of these products provides a strong
indication that p-xylene has biodegraded in the aquifer.
Results of measurement of 14C p-xylene degradation
are given in Table 2. Total 14C in the water sample and
the amount of 14C in the water after extraction were used
-1-
-4-
-5'
fj Benzene
A Naphthalene
O p-Xylene
A o-Dichlorobenzene
100 200 300 400
Days After Injection
500
Figure 6. Degradation curves for the compounds in the reactor
zone.
Table 1. Maximum Biodegradation Rate Constants From the
MADE2 Site
max(corr)
Benzene
p-Xylene
Naphthalene
o-Dichloro-
benzene
kmax(reactor)
0.0120
0.0196
0.0118
0.0114
(reactor)
(d1)
0.0066
0.0141
0.0063
0.0059
kmax(whole
field)(d1)
0.0104
0.0187
0.0104
0.0100
Table 2. Degradation of p-Xylene in Water Samples Taken
After 421 Days, Expressed as Weight Percent
Converted
% 14C
p-Xylene
Converted
to All
Products
p-Xylene
Converted
to CO2
Mass-
Balance-
Based %
p-Xylene
Converted
[xylene]>1 ppm 85.1 (±6.3) 73.3 (±11.1) 98
(n=8)
[xylene]>1 ppm 82.6 (±6.1) 74.2 (±6.2) 98
(n=10)
to calculate the fraction of p-xylene converted to all
products. Total 14C in the water sample and the 14C in
the carbonate precipitate were used to calculate the
fraction of p-xylene converted to CO2.
These p-xylene conversion figures compare well with
mass-balance-based conversions calculated from the
total p-xylene remaining in the plume at day 421, which
were interpolated from snapshot data. Mass-balance-
based conversions are based on GC analysis and on
the known mass of p-xylene injected. They are included
in the last column of Table 2.
The values in the second column are on the high, but
normal, side for 14CO2 release. This implies that most of
the p-xylene went to energy production and was not
converted to biomass. The difference between degrada-
tion in the high and low concentration regions of the
aquifer is not significant. The difference of the means for
p-xylene converted to all products and p-xylene con-
verted to CO2 may imply that some intermediate prod-
ucts are present, but this difference may also be
attributable to analytical anomalies (i.e., loss of Ca14CO3
during precipitate transfer prior to counting). Agreement
between the first and third columns of Table 2 indicates
the consistency of p-xylene degradation measurements
by GC and 14C counting methods. The small difference
between these numbers is due to incorporation of 14C
into biomass and insoluble carbonates. Most p-xylene
degradation products apparently remained in the local
ground water.
72
-------
Discussion
Biodegradation rates given in Table 1 are based on field
observations of solute behavior and are an essential
input in modeling organic contaminant fate and transport
in this aquifer material. The relative rates are as ex-
pected, with p-xylene most rapidly biodegraded. Figure
6 indicates that biodegradation in the Columbus aquifer
is a first-order process. Simkins and Alexander (9) have
indicated that biodegradation can be expected to follow
Monod kinetics for a microbial population not limited by
nutrients and with a sufficient substrate concentration,
but first-order kinetics are observed at low substrate
concentrations. Organic solute concentrations in the
MADE2 test were quite low, so this conclusion is con-
sistent with experimental results. Larson (10) notes that
the rate of decrease in concentration as a function of
substrate concentration can often be expressed as a
first-order equation, and that first-order kinetics are gen-
erally expected for biodegradation at low organic sub-
strate concentrations. Applicability of the leaking reactor
model is site specific, and its success with the MADE2
test data is a fortuitous circumstance dependent on the
arbitrarily selected positions of the injection wells in the
Columbus aquifer. It is encouraging that the two meth-
ods of calculating biodegradation rates gave similar re-
sults, thus providing confidence that these rates are
useful for predictive purposes.
Organic solutes in MADE2 biodegraded under aerobic
conditions in the aquifer. The organic solute concentra-
tions injected were chosen to be too low to significantly
deplete dissolved oxygen in the reactor zone of the
aquifer. Maintenance of oxic conditions was confirmed
by monitoring dissolved oxygen and redox potential of
water samples during the experiment. Thus the biodeg-
radation rates reported here were not affected by oxy-
gen limitation. For very large releases of similar organic
compounds, the conditions near the source might rap-
idly become reducing due to the biological oxygen de-
mand. Kinetic models to determine biodegradation rates
in this situation would be complex due to the need to
include oxygen transport terms and the potential for
degradation by anaerobic bacteria. It is therefore wise
to use small injection amounts in field experiments to
determine biodegradation rates.
In MADE2, the organic solutes degraded quickly. These
results suggest that, for similar solutes, aquifer remedia-
tion activities should be restricted to the source region,
which might include pumping to remove nonaqueous-
phase liquids or excavation of contaminated aquifer ma-
terial. The source would be reduced but not eliminated
by this treatment. In an aquifer with approximately
steady flow, the plume of organic solutes from the re-
duced source would reach a steady state, with the
boundary determined by the hydrology of the site,
sorption, in situ biodegradation, and oxygen and nutrient
supply.
Conclusions
Controlled-release experiments similar to the MADE2
test are needed to determine accurate biodegradation
rates for use in ground-water contaminant fate and
transport models of aquifer situations. The MADE2
study has demonstrated the practicality of these experi-
ments and obtained in situ degradation rates for four
organic contaminants in the Columbus aquifer. These
rates will be used in the design and modeling stages of
a new field test at the MADE site, which is now in
preparation.
References
1. Madsen, E.L. 1991. Determining in situ biodegra-
dation: Facts and challenges. Environ. Sci. Technol.
25(10):1,663-1,673.
2. Robin, M.J.L., E.A. Sudicky, R.W. Gillham, and R.G.
Kachanoski. 1991. Spatial variability of strontium
distribution coefficients and their correlation with
hydraulic conductivity in the Canadian Forces Base
Borden aquifer. Water Resour. Res. 27(10):2,619-
2,632.
3. Boggs, J.M., S.C. Young, and L.M. Beard. 1992.
Field study of dispersion in a heterogeneous aqui-
fer, 1. Overview and site description. Water Resour.
Res. In press.
4. Rehfeldt, K.R., J.M. Boggs, and L.W Gelhar. 1992.
Field study of dispersion in a heterogeneous aqui-
fer, 3. Geostatistical analysis of hydraulic conduc-
tivity. Water Resour. Res. In press.
5. Mackay, D.M., D.L. Freyberg, and P.V. Roberts.
1986. A natural gradient experiment on solute
transport in a sand aquifer, 1. Approach and over-
view of plume movement. Water Resour. Res.
22(13):2,017-2,029.
6. Garabedian, S.P., D. LeBlanc, L.W. Gelhar, and
M.A. Celia. 1991. Large-scale natural gradient
tracer test in sand and gravel, Cape Cod, Massa-
chusetts, 2. Analysis of spatial moments for a non-
reactive tracer. Water Resour. Res. 27(5):911-924.
7. Adams, E.E., and L.W. Gelhar. 1992. Field study of
dispersion in a heterogeneous aquifer, 2. Spatial
moments analysis. Water Resour. Res. In press.
8. Maclntyre, W.G., T.B. Stauffer, and C.R Antworth.
1991. A comparison of sorption coefficients deter-
mined by batch, column, and box methods on a low
carbon aquifer material. Ground Water 29(6):908-
913.
73
-------
9. Simkins, S., and M. Alexander. 1984. Models for 10. Larson, R.J. 1979. Role of biodegradation kinetics
mineralization kinetics with the variables of sub- in predicting environmental fate. In: Maki, A.W., et
strate concentration and population density. Appl. al., eds. Biotransformation and fate of chemicals in
Environ. Microbiol. 47(6):1,299-1,306. the aquatic environment. Washington DC: Ameri-
can Society for Microbiology.
74
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Traverse City: Distribution oftheAvgas Spill
David W. Ostendorf
Civil and Environmental Engineering Department, University of Massachusetts, Amherst, MA
Abstract
The capillary tension/liquid saturation equations of
Parker and Lenhard (1) and Lenhard and Parker (2)
provide a reasonably accurate description of the vertical
distribution of the separate phase aviation gasoline
(avgas) in solid core samples taken from the U.S. Coast
Guard Air Station in Traverse City, Michigan.
Introduction
The depth of a continuous liquid below the ground sur-
face is directly related to its capillary tension in the
absence of vertical velocity, since the sum of the poten-
tial and pressure heads forms a vertically uniform hy-
draulic head under static conditions. Thus, the theory (1)
describing capillary tension/saturation relations for im-
miscible, continuous fluids also finds expression as ver-
tical profiles of water and avgas in the absence of
infiltration. Lenhard and Parker (2) recognize this
equivalence and propose profiles for total liquid and
avgas saturation that are reasonably borne out by Os-
tendorf et al. (3) in their analysis of solid core data from
five stations in the avgas plume at Traverse City (Figure
1). We summarize the theory here and discuss some
implications of the distribution.
Vertical Distribution of Free Avgas
The vertical distribution of water and free avgas in the
contaminated soil is idealized in Figure 2. The water is
more strongly attracted to the solid grains than the
avgas and fills the smaller pores in the soil, forming an
interface with free avgas. The free avgas in turn faces
the soil gas in the large pores, which control the overall
saturation of the soil as a consequence. Residual avgas
may be trapped as discontinuous bubbles within the
water phase due to hysteresis and a fluctuating water
table.
The volumetric water 0W, residual avgas 0LR, and irre-
ducible water OWR contents combine to define an effec-
tive apparent water saturation Sw and comprise the total
X
Interdiction
Well Line
/
/Flow
Direction
/
Building
'/\ *5
Smith XxvXsOCM
HallV^/y
/
50BT
50BS
Plume
50 m , /
/ SUBi / •-•••-
/5OCL* 50CE Boundary
Tanks/p
Hangar
Administration
Building
Spill
Origin
Figure 1. Site plan, U.S. Coast Guard Air Station.
saturation S when they are added to the free avgas
content 0LF (1).
S =
9|/i/+ QLF+
with porosity n. The total saturation is controlled by the
tension across the avgas/air interface and varies with
depth b below the ground surface in the absence of
dynamic effects associated with infiltration
S=S
W
S =
/aH
S=1
(bM>b)
(bL >b >bM)
(b>bL)
(2a)
(2b)
(2c)
75
-------
Ground
Surface
Air
.... Water
volumetric Liquid Content
Monitoring
Well
Figure 2. Vertical distribution of liquids in soil and a monitoring
well.
Lenhard and Parker (2) derive equation 2 as an appli-
cation of van Genuchten's (4) pore size distribution to
separate phase contamination in soil. Figure 2 displays
the minimum depth bM of free avgas occurrence, along
with the avgas table bL and water table bw existing in a
hypothetical monitoring well. Note that the avgas does
not occupy all the pores at a given elevation and extends
above the avgas table in the soil due to capillary tension.
Thus, the monitoring well levels, though necessary to
determine liquid tensions, do not explicitly determine the
vertical distribution of avgas in the soil. The pore size
distribution of the soil is also needed to specify the
profile.
These latter data are characterized by the pore size
uniformity exponent a and the scaling factor (3L appear-
ing in equation 2b. The latter parameter may be ex-
pressed in terms of the mean pore radius r by noting
that the avgas/air surface tension OLA relates fluid ten-
sion to interfacial radius
PL
(3)
with gravitational acceleration g, avgas density PL, and
an assumed zero contact angle. Table 1 lists parameter
values calibrating the data from five stations at Traverse
City (3). Observations (circles) and theory (curve) for a
typical total saturation profile are sketched in Figure 3.
A mean pore radius of 5.4 x10"5m is implied by equation
3. A simple extension of a classical grain size based
model of r checks this estimate. We note that r/2 is the
Table 1. Total and Avgas Saturation Profile Parameters at
Traverse City
Symbol
Parameter
Value
n
QWR
PL
a
PL
PW
A
Y
Porosity
Irreducible moisture content
Avgas density
Pore size uniformity
Avgas scaling factor
Water scaling factor
Water table amplitude
Trapping factor
0.367
0.059
707 kg/m3
3.00
8.20 m'1
1 .53 m'1
0.35 m
40
2 -
Figure 3. Total saturation, Core 50BT.
hydraulic radius of the mean pore size flowing full, and
equate this to a modified Fair and Hatch (5) estimate of
the quantity
_r VVOIDS
2~ A
'SOLIDS
=
3(1 -
(4a)
(4b)
^ ;
The observed mean grain size of 3.8 x 10"4 m at Trav-
erse City leads to a mean pore size value of 5.7 x 1CT5 m,
in excellent agreement with the equation 3 value.
The water/avgas interface controls the water saturation
in the presence of the free avgas
(bw>b>bM) (5a)
(b>bw) (5b)
(5c)
/= PL-
GWL
76
-------
with water density p and scaling factor Pw predicted on
the water/avgas surface tension OWL. The upper extent
bM of free avgas may be estimated by equating the total
saturation (equation 2b) and the water saturation (equa-
tion 5a) at this elevation. A water/air interface with a
surface tension OWA governs the water saturation above
this depth
(6a)
O|/|/— I
,-bw+-^(bw-b]}ar/a)"1 (bM>b)
PL
(6b)
The free avgas saturation SLF, in view of equation 1, is
simply given by
— S — Si
i/i/
(7)
Figure 4 shows a typical free avgas profile at the site,
based upon the scaling factors of Table 1.
Vertical Distribution of Residual Avgas
Residual avgas may be attributed to hysteretical trap-
ping of product as it rises and falls through the water wet
soil over a fluctuating water table. Free liquids at a given
depth b experience progressively stronger capillary ten-
sions as the water table falls due to their higher position
above this reference level. After the table attains a maxi-
mum depth and begins to rise, the water/avgas interface
becomes steadily larger due to a decreasing capillary
tension. The encroaching water occludes some of the
avgas, giving rise to a discontinuous residual fraction
within the water phase.
Historical maximum and minimum effective water satu-
rations SWMAX, SWMIN are established by corresponding
minimum (bvwiiN) and maximum (bwwiAx) water table
depths, respectively
WM,N =
WMAX
= (1 +
JWMIN - UWMAX
-A
(8a)
(8b)
(8c)
Ostendorf et al. (3) infer the historical water table excur-
sion amplitude A cited in Table 1 from water level vari-
ations in nearby Lake Michigan. Parker and Lenhard (1)
suggest that the extreme saturations induced by ex-
treme water positions trap residual avgas saturation SLR
that is given by
_ 1 -SWMIN
1
> b >bwMiN)
with empirical trapping factor Y. Ostendorf et al. (3)
calibrate residual saturation profiles from their five sta-
tions with the trapping factor cited in Table 1, with the
typical results sketched in Figure 5.
Discussion
The vertical distribution of free and residual avgas has
important implications for the estimation of separate
phase contamination from monitoring well observations.
The depth integrated mass MLF of free avgas may be
estimated formally by integrating equations 2b, 5a, 6a,
5.5
E
-Q
0.0
Figure 4. Free avgas, Core 50CE.
0.00 0.01 0.02 0.03
S
LR
Figure 5. Residual avgas, Core 50CL.
77
-------
and 7 from the water table to the upper limit of free avgas
occurrence, with the result
(10a)
^
]}
MLF=PLJ SLFdb
t>M
'=•5- ®w(bw-bL)-(1-(-jr)i[
Pl/l/ °L
Pw
PL
(10b)
(10c)
The integral function I(y) varies with the uniformity ex-
ponent of the pore sizes, while its argument is deter-
mined by the scaling factors and the product thickness
in the monitoring well. Figure 6 displays the equilibrium
variation of the depth integrated mass predicted at Trav-
erse City as a function of the monitoring well avgas
thickness. We note the exclusion of the residual mass
from this estimate, since the trapped avgas is not con-
tinuously connected to product in the monitoring well.
For a point of reference, Figure 5 implies a depth inte-
grated residual avgas mass of about 5.3 kg/m2; if this
mass were free avgas, we would observe a 0.3-m prod-
uct thickness in a monitoring well.
The methylene chloride extract chromatograms are
composed primarily of known compounds (6), so that
the composition of the avgas can be examined at a
given depth. We note distillation of avgas as a function
of depth in the solid core samples, as suggested by a
profile from station SOBS (Figure 7). The open circles
correspond to relatively volatile avgas compounds, with
pure phase vapor densities of 0.27 to 1.00 kg/m3 (at
12°C). This fraction becomes more important in the
0.2
b -b
W L
0.4
m
_c
-ir->
Q.
Q
O High v-
• Medium v • O
V Low
i i i
Figure 6. Depth integrated free avgas mass.
10 20 30 40
Avgas Fraction, %
Figure 7. Avgas composition, SOBS.
lower elevations due presumedly to decreased volatili-
zation losses in the wetter region of the soil. The mid-
range volatiles (closed circles), with vapor densities of
0.14 to 0.27 kg/m3, are quite uniformly distributed, while
the heavy compounds (triangles), with vapor densities
less than 0.14 kg/m3 in magnitude, tend to dominate the
blend of hydrocarbons in higher, drier soil due to strip-
ping away of the lighter fraction.
The partitioning of avgas into free and residual fractions
has important implications on its downgradient fate and
transport as well. Figure 2 suggests that the free avgas
occupies a relatively high-permeability, saturated soil
region below the avgas table and a lower-permeability,
unsaturated zone above the table. Since both these
regions share a common horizontal gradient due to the
slope of the avgas table, we anticipate slower and faster
zones of horizontal separate phase transport. This ver-
tical profile of horizontal specific avgas discharge is very
nonuniform for sands with high values like that at the
site. The phenomenon can be approximated as a
pseudosorptive process, with a linear balance between
a mobile fraction and a reversible, immobile fraction
"sorbed" by capillary tension (7). Pursuing this analogy
further, the residual avgas can be thought of as an
irreversibly sorbed partition, lost from the mobile fraction
by hysteretical trapping. The net effect of these mecha-
nisms is that the avgas travels with the underlying
ground water, but at a retarded velocity.
The free avgas is relatively easy to strip out of the soil
due to its direct interface with air. The residual avgas is
much more difficult to remediate, since it is surrounded
by essentially immiscible water. We accordingly expect
an initial period of relatively rapid remediation in re-
sponse to soil venting or air sparging, followed by an
78
-------
asymptotically lower removal rate exacerbated by avgas
distillation.
References
1. Parker, J.C., and R.J. Lenhard. 1987. A model for
hysteretic constitutive relations governing multi-
phase flow, 1. Saturation pressure relations. Water
Resour. Res. 23:2,187-2,196.
2. Lenhard, R.J., and J.C. Parker. 1990. Estimation of
free hydrocarbon volume from fluid levels in monitor-
ing wells. Ground Water 28:57-67.
3. Ostendorf, D.W, R.J. Richards, and P.P. Beck. 1993.
LNAPL retention in sandy soil. Ground Water 31:285-
292.
4. van Genuchten, M.T. 1980. A closed form equation
for predicting the hydraulic conductivity of unsatu-
rated soils. Soil Sci. Soc. Am. J. 44:892-898.
5. Fair, G.M., and L.P. Hatch. 1933. Fundamental fac-
tors governing the streamline flow of water through
sand. J. Am. Waterworks Assoc. 25:1,551-1,565.
6. Ostendorf, D.W, I.E. Leach, E.S. Hinlein, and Y.F.
Xie. 1991. Field sampling of residual aviation gaso-
line in sandy soil. Ground Water Monitor. Rev.
11:107-120.
7. Ostendorf, D.W. 1990. Long-term fate and transport
of immiscible aviation gasoline in the subsurface
environment. Water Sci. Tech. 22:37-44.
79
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Traverse City: Geochemistry and Intrinsic Bioremediation ofBTX Compounds
Barbara H. Wilson, John T. Wilson, Don H. Kampbell, and Bert E. Bledsoe
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK
John M. Armstrong
The Traverse Group, Inc., Ann Arbor, Ml
Introduction
Loss of petroleum products from underground storage
tanks, pipelines, and accidental spills is a major source
of contamination of unsaturated soils, aquifer solids, and
ground water. Volatile aromatics such as benzene, tolu-
ene, ethylbenzene, and the xylenes (BTEX) are more
soluble in water than the aliphatic and higher molecular-
weight aromatic constituents of petroleum products (1).
Once released to the subsurface, petroleum compounds
are subject to aerobic microbial processes. The low-
molecular-weight alkanes and aromatics are readily
biodegraded in oxygenated ground water, depleting the
ground water of available oxygen (2,3). Reoxygenation
of the ground water may occur through reaeration from
soil gases, ground-water recharge, and usually ineffi-
cient mixing with surrounding oxygenated ground wa-
ters (4,5). Although ground waters near the perimeter of
the contaminant plume may be reoxygenated, the inte-
rior of the plume will remain anoxic for a distance down-
gradient. Anaerobic biological processes can account
for most of the removal of BTEX from the plume (6,7).
The biogeochemical mechanisms that contribute to an-
aerobic processes in the subsurface, however, are not
well understood.
The impact of anaerobic microbial processes on the fate
of monoaromatics, substituted aromatics, and chlorin-
ated hydrocarbons in anoxic subsurface environments
has been studied in laboratory and field situations (8-
12). Field evidence of biotransformation of o- m-, and
p-xylene was observed in methanogenic landfill
leachate with their preferential removals compared with
other alkylbenzenes present (13). Methanogenesis has
been observed at two sites with ground water contami-
nated with creosote (14,15). Denitrifying, iron-reducing,
sulfate-reducing, and methanogenic activities found in
ground water at the first site were highly correlated with
the biodegradation of creosote; methane was only
detected in the ground water that had been contami-
nated with creosote. Intermediate products of methane
fermentation, formate and acetate, were found in the
ground water at the second site.
In 1969 the unsaturated soil and ground water underlying
the U.S. Coast Guard Air Station at Traverse City, Michi-
gan, were contaminated with an estimated 25,000 gal of
aviation gasoline when a flange in an underground stor-
age tank failed. Dissolution of the aromatics in the
ground water resulted in concentrations of 36 mg to
40 mg of total alkylbenzenes per liter near the center of
the plume. The subsurface contamination existed as
residual-phase hydrocarbon, dissolved-phase aromat-
ics, and gaseous hydrocarbons resulting from volatiliza-
tion (16-18). The plume of dissolved-phase aromatics
extended from the air station into the East Arm of Grand
Traverse Bay and affected numerous drinking water
wells. The area near the leaking underground storage
tank was used to store degreasing solvents and to
conduct degreasing operations during aircraft mainte-
nance. A small plume of chlorinated solvents lies adja-
cent to the gasoline plume.
Geochemical Characterization
Geochemical analyses of the water samples collected
at Traverse City revealed waters of four distinct geo-
chemistries: 1) the heart of the plume, 2) an anaerobic
zone of treatment, 3) an aerobic zone of treatment, and
4) a pristine or renovated zone. The water from the heart
of the plume contained high concentrations of methane
and BTEX, with no detectable oxygen. These waters
were surrounded by an anaerobic zone of treatment with
greatly reduced concentrations of dissolved aromatics,
no oxygen, and substantial concentrations of methane.
Surrounding the anaerobic zone of treatment was an
aerobic zone of treatment with measurable oxygen, small
quantities of methane, and very low concentrations of
80
-------
the alkylbenzenes. The perimeter of the plume was
surrounded by a renovated or pristine zone with high
concentrations of oxygen, no detectable alkylbenzenes,
and no methane.
Gas chromatography/mass spectrometry (GC/MS) analy-
ses of the waters confirm the presence of BTEX at Trav-
erse City. Also found were the chlorinated compounds
1,2-dichloroethane (1,2-DCA), tetrachloroethylene (PCE),
and 1,1,1-trichloroethane (TCA). The GC/MS analyses of
waters from wells R, S, and Q (Figure 1a) identified phe-
nols and aromatic acids indicative of anaerobic microbial
action on the soluble aromatic constituents of petroleum
products (7,19,20). These compounds are found in por-
tions of the plume with substantial concentrations of meth-
ane and no detectable oxygen, and are probably
precursors of the methane. Complete information on the
geochemical characterization of the ground waters at Trav-
erse City, collection of aquifer material, analytical methods
used, and microcosm construction may be found in Wilson
etal. (21).
Laboratory Studies
To confirm field evidence of intrinsic bioremediation,
laboratory microcosm studies were conducted on aqui-
fer material from the U.S. Coast Guard Air Station.
Material for aerobic and anaerobic fate studies was
collected from three locations in the plume (Figure 1a).
Aquifer material from site A (11.7 m to 12.3 m below land
surface), the zone of anaerobic treatment, was used to
construct the microcosms for the anaerobic fate study.
Aquifer material from site B (9.6 m to 10.2 m below land
surface), the aerobic zone of active biological treatment,
and from site C (6.6 m to 7.2 m below land surface), the
pristine or renovated zone, was used to prepare the
aerobic fate studies. Autoclaved controls were prepared
from the site C material. The compounds added to the
microcosms were benzene, toluene, p-xylene, o-xylene,
TCA, TCE, and chlorobenzene.
The initial compound concentrations and results of the
fate studies at various incubation times are shown in
Table 1. Benzene, toluene, p-xylene, and o-xylene were
biodegraded in both aerobic and anaerobic aquifer ma-
terial. The removals were quite rapid for all compounds
in each of the three geochemical zones studied, whether
in the anaerobic zone of treatment, the aerobic zone of
treatment, or the renovated (pristine) material. By the
end ofSwk of incubation of the compounds in the anaero-
bic aquifer material, the concentrations of benzene, tolu-
ene, p-xylene, and o-xylene had been reduced one order
of magnitude. The biotransformation of the four com-
pounds in the aerobic zone of biological treatment oc-
curred even more rapidly. By the end of2wk of incubation,
the concentrations for all four compounds were decreased
by two orders of magnitude. Similar losses were seen in
material from the renovated/pristine zone.
U.S. Coast Guard
Air Station Boundary
Purge Well Field
D & Q
150 Meters
Figure 1a. Locations of wells in an aviation gasoline plume at
the U.S. Coast Guard Air Station at Traverse City.
Well A is located in the plume below the purge field;
well B is located in the aerobic zone of treatment;
well C is located in the pristine region surrounding
the plume; and wells D, P, Q, R, and S are located
in the plume above the purge field.
Toluene
Benzene
Xylenes
ug/L
Figure 1b. Concentration of BTX in monitoring wells along a
flow path down the central axis of the plume in the
first quarter of 1986.
At the end of 4 wk of incubation, the respective concen-
trations for benzene, toluene, p-xylene, and o-xylene in
the autoclaved samples were 57 percent, 61 percent,
46 percent, and 49 percent of the original concentra-
tions. Due to the removals of BTX in the controls, a
second set of autoclaved samples was prepared. The
concentrations after 16 wk of incubation in the dupli-
cated controls were benzene, 83 percent; toluene, 66
percent; m- + p-xylene, 51 percent; o-xylene, 50 per-
cent; TCA, 86 percent; TCE, 73 percent; and chloroben-
zene, 58 percent. The cause for the removal of organics
in the controls has not been determined; however, sorp-
tion to aquifer solids probably occurred. Chlorobenzene
was also biodegraded in each of the three geochemical
zones studied. Decreases of one order of magnitude
were observed for chlorobenzene in both the aerobic
and anaerobic zones of biological treatment after 4 wk
81
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Table 1. Behavior of Benzene, Alkylbenzenes, TCAa, TCEb, and Chlorobenzene in Aquifer Material From an Aviation Gasoline
Plume
Compound (ug/L Pore Water)
Subsurface
Material
A, anaerobic
treatment, 11.7 m
to 12.3 m below
land surface
B, active aerobic
treatment, 9.6 m
to 10.2 m below
land surface
C, aerobic
renovated, 6.6 m
to 7.2 m below
land surface
C, autoclaved,
6.6 m to 7.2 m
below land
surface
Week
0
4
8
95C
0
2
4
14
0
2
4
14
0
2
4
Benzene
450
12
6
ndd
450
2
5
2
420
4
1
—
420
380
240
Toluene
420
56
40
nd
420
2
5
2
380
3
3
—
380
290
230
m- + p-
Xylene
440
78
17
nd
390
1
2
1
370
3
1
—
370
200
170
o-Xylene
410
41
6
nd
390
2
1
nd
370
3
1
—
370
190
180
TCA
570
420
580
73
600
440
540
430
600
580
620
650
600
590
560
TCE
540
260
340
54
650
440
480
260
540
540
600
500
650
540
430
Chlorobenzene
500
66
34
3
500
53
50
30
500
100
106
67
500
300
210
a1,1,1-Trichloroethane
bTrichloroethylene
Concentrations at this time interval were determined by GC/MS; concentrations at other time intervals were determined by GC. All values are
means of triplicate analyses.
dNot detected, detection limit of 0.1 |ig/L
of incubation. Similar removal was seen in the reno-
vated/pristine material after 14 wk of incubation. In the
autoclaved samples, 44 percent of the Chlorobenzene
remained after 4 wk of incubation.
No significant biotransformation of TCA or TCE, com-
pared with the controls, was observed at the end of
8 wk of incubation. Evidence of reductive dechlorination
of both compounds, however, was indicated by GC/MS
analyses of anaerobic microcosms at 95 wk of incuba-
tion by the identification of 1,1-dichloroethylene (1,1-
DCE) and 1,1-dichloroethane (1,1-DCA) (22,23).
Headspace concentrations of methane were measured
immediately before sampling to determine the mainte-
nance of methanogenic conditions in the microcosms.
Methane was found in all the anaerobic samples, with
concentrations ranging from 50 ppm to 100 ppm; no
methane was found in the headspace of the aerobic or
autoclaved samples.
Correspondence Between Laboratory and
Field Data
As part of a settlement with the State of Michigan, the
U.S. Coast Guard monitors alkylbenzene concentra-
tions in selected monitoring wells quarterly. Three of the
monitoring wells (M30 near site S, M31 near site Q, and
M2 near site A in Figure 1) lie along a flow path. The
time required for water to move from one well to the next
can be estimated by dividing the distance between the
wells by the flow velocity (approximately 1.5 m per day).
This value was determined directly from tracer tests, and
is confirmed by calculations based on the hydraulic
conductivity of the aquifer and its hydraulic gradient (24).
Water takes 10 wk to flow from S to Q, and 24 wk to flow
from S to A. The first-order rate of biodegradation along
a segment of aquifer between the monitoring wells can
be estimated by dividing the concentration in the well
distal to the spill by the concentration in the proximate
well, taking the natural logarithm, then dividing by the
time required for water to flow between the wells.
Table 2 portrays the depletion of total BTX between S
and Q and between S and A for the years 1984 through
1987. The rate constants are surprisingly consistent. A
purge field was installed and put on line in mid-1985 to
prevent further migration of the plume from Coast Guard
property. As soon as the purge field was put into opera-
tion, the water behind the field near site A (Figure 1a)
became stagnant, and concentrations of BTX began to
drop. Solution concentrations of BTX dropped to low
values by late 1985 (data not shown). Because the
water was not moving, the decline in concentration over
time could be used to estimate the first-order rate con-
stant for anaerobic BTX biotransformation in the part of
the aquifer near site A (Table 3).
Sediment samples for the microcosm study were ac-
quired from site A in late 1985, and this study was
conducted in the first quarter of 1986. Figure 1 b depicts
the concentration of BTX in monitoring wells along a flow
path down the central axis of the plume at that time.
Table 3 compares the depletion of BTX along the aquifer
82
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Table 2. First-Order Rates of Anaerobic Biotransformation (per Week) of Total BTX Along Segments in the Aquifer
S to Q, Quarter of Year S to A, Quarter of Year
Year
First
Third
First
Third
1984
1985
1986
1987
0.11
0.02
0.43
0.34
0.17
0.27
0.14
0.10
0.10
0.07
0.20
0.33
0.09
cnca
cnc
cnc
aCannot calculate: the plume was intercepted by a purge well field and did not reach site A.
Table 3. Comparison of the First-Order Rates (per Week) of
Anaerobic Biotransformation of BTX
Compound
Benzene
Toluene
m- +
p-Xylene
o-Xylene
All xylenes
Microcosms3
0.5
0.3
0.4
0.5
—
Aquifer
Segment13
S to Q
0.05
1.3
—
0.03
Change Over
Time at Ac,
6/85 to 10/85
0.17
0.47
—
0.10
laboratory microcosm studies
bAlong flow path segments in the aquifer
cAt a site with stagnant ground water behind a purge well field
segment S to Q during the first quarter of 1986 and the
depletion in the static water at site A in the last half of
1985 with the rates of anaerobic BTX biotransformation
in the microcosm study.
The rate of disappearance of BTX compounds as meas-
ured in aerobic microcosms does not compare well with
actual rates of aerobic BTX degradation in the field. The
rate of degradation in the field is controlled by mass-
transport limitations for oxygen (4,5), while laboratory
studies are limited by reaction rates. The anaerobic fate
study compared quite favorably with those rates meas-
ured by field data. Methanogenesis and other anaerobic
processes are not limited by the availability of oxygen in
either microcosms or subsurface materials. Mass-trans-
port considerations are therefore not as critical to the
comparison of microcosm and field data. Anaerobic mi-
crocosms might prove to be a valuable tool to evaluate
intrinsic biorestoration of aquifers contaminated with pe-
troleum products.
Conclusions
The results of the laboratory study confirm field evidence
of both aerobic and anaerobic transformation of alkyl-
benzenes and suggest that intrinsic aerobic and anaero-
bic in situ biorestoration of ground water contaminated
with petroleum products can occur. The anaerobic trans-
formations seen at this site and confirmed by the labo-
ratory study provide an attractive alternative to aerobic
restoration. The removals of the alkylbenzenes in the
anaerobic material were quite rapid and compared fa-
vorably with removals seen in the aerobic zone of treat-
ment. Comparison of first-order rates of disappearance
in anaerobic microcosms with those calculated from
field data show acceptable agreement. Anaerobic proc-
esses in the subsurface are probably limited by in situ
reaction rates rather than by mass-transport limitations
for nutrients. Potentially, anaerobic microcosm studies
could be useful in the evaluation of intrinsic bioremedia-
tion of petroleum-contaminated subsurface materials.
The aerobic degradation of alkylbenzenes in subsurface
environments has been well documented (25,26) and is
currently the state-of-the-art for restoration of petroleum
contamination. Anaerobic biotransformation, however,
can enhance in situ biorestoration in oxygen-depleted
regions of a plume where heavily contaminated ground
water has excessive oxygen demand. Naturally occur-
ring anaerobic biological processes can potentially re-
mediate ground water contaminated with petroleum
products and significantly increase the reliability of ex-
isting remediation technologies.
References
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Ringhand, and F.C. Kopfler. 1984. The identification
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2. Atlas, R.M. 1981. Microbial degradation of petro-
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7. Cozzarelli, I.M., R.P. Eganhouse, and M.J.
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9. Lovley, D.R., M.J. Baedecker, D.J. Lonergan, I.M.
Cozzarelli, E.J.P Phillips, and D.I. Siegel. 1989.
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oxidation of toluene, phenol, and p-cresol by the
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11. Suflita, J.M., S.A. Gibson, and R.E. Beeman. 1988.
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13. Reinhard, M., N.L. Goodman, and J.F. Barker.
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14. Ehrlich, G.G., D.F. Goerlitz, E.M. Godsy, and M.F.
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in ground water by anaerobic bacteria: St. Louis
Park, Minnesota. Ground Water 20(6):703-710.
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18. Rifai, H.S., P.B. Bedient, J.T. Wilson, K.M. Miller,
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21. Wilson, B.H., J.T. Wilson, D.H. Kampbell, B.E. Bled-
soe, and J.M. Armstrong. 1990. Biotransformation
of monoaromatic and chlorinated hydrocarbons at
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23. Klecka, G.M., S.J. Gonsior, and D.A. Markham.
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25. Lee, M.D., J.M. Thomas, R.C. Borden, P.B. Bedient,
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84
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Mathematical Modeling of Intrinsic Bioremediation at Field Sites
Hanadi S. Rifai
Energy and Environmental Systems Institute, Rice University, Houston, TX
Introduction
Intrinsic bioremediation is an important attenuation mecha-
nism at contaminated field sites because it limits pollutant
migration and reduces contaminant mass in the subsur-
face. Quantifying the impact of intrinsic bioremediation at
a field site involves conducting a comprehensive field
sampling program in conjunction with extensive modeling
of contaminant transport and fate. This paper reviews the
basic data requirements for modeling intrinsic bioremedia-
tion at field sites and discusses some of the available
models that can be used for that purpose. The limitations
of existing models for simulating biodegradation are pre-
sented, as well as the difficulties in validating and verifying
these models. Finally, a few case studies of modeling
intrinsic bioremediation are reviewed.
Intrinsic Bioremediation Processes
Intrinsic bioremediation refers to the reduction of contami-
nant mass at a field site due to biodegradation. This reduc-
tion can occur under aerobic or anaerobic conditions.
When oxygen is utilized as the electron acceptor, the
process is referred to as aerobic respiration. When oxygen
is not present (anoxicconditions), microorganisms can use
organic chemicals or inorganic anions as alternate
electron acceptors. Anaerobic biodegradation refers to
fermentative, denitrifying, iron-reducing, sulfate-reducing,
or methanogenic processes. To quantify the impact of
intrinsic bioremediation on contaminant concentrations at
a field site, one needs to develop an accurate picture of
the distribution of electron acceptors both in pristine and
contaminated areas. The concentrations of the electron
acceptor in pristine areas provide an indication of the
biodegradation potential at the site. The disappearance or
decline of electron acceptor concentrations in contami-
nated areas provides an indication that biodegradation
may be occurring.
Existing Biodegradation Models
The biodegradation of contaminants in ground water is
mainly controlled by the rate of the reaction and the
availability of the electron acceptor. A mathematical ex-
pression that represents the chemical reaction can be
written to account for the effect that the rate of the
reaction has on biodegradation. This mathematical ex-
pression can then be combined with the transport equa-
tion to account for the electron acceptor limitation effect
on the biodegradation process in the subsurface.
Many biodegradation models have been developed in
recent years, most of which utilize some kinetic expres-
sion for biodegradation (see Table 1). The models listed
in Table 1 simulate a number of aerobic and anaerobic
biodegradation processes subject to specified condi-
tions and assumptions. The difficulties involved in apply-
ing these models include 1) the data required as input
to the model is lacking, such as the kinetic rate parame-
ters or estimates of the hydraulic conductivity; 2) the
majority of these models are proprietary and very few
are public domain models; and 3) modeling in general
is complicated and time-consuming, and requires a cer-
tain level of expertise.
Modeling of Intrinsic Bioremediation at
Field Sites
Many case studies of simulating biodegradation at field
sites exist in the general literature. In this paper, four
case studies are reviewed that combine field and labo-
ratory investigation programs with modeling studies.
Conroe Superfund Site, Texas
Borden et al. (2) applied the first version of the
BIOPLUME model to simulate biodegradation at the
Conroe Superfund site in Texas. The United Creosoting
Company (UCC) site was operated as a wood preserv-
ing facility from 1946 to 1972. Wastes disposed in two
unlined ponds were composed of predominantly polycy-
clic aromatic hydrocarbons and pentachlorophenol
(PCP). Monitoring of the site has shown elevated levels
of organic contaminants in the soil and ground water, as
well as elevated levels of chloride in the ground water.
85
-------
Table 1. Biodegradation Models
Name Description
Author(s)
1-D, aerobic, microcolony, Monod
BIOPLUME 1-D, Monod
- 1-D, analytical first-order
BIO1D 1-D, aerobic, anaerobic, Monod
1-D, co-metabolic, Monod
- 1-D, aerobic, anaerobic, nutrient limitations, microcolony,
Monod
- 1-D, aerobic, co-metabolic, multiple substrates,
fermentative, Monod
BIOPLUME II 2-D, aerobic, instantaneous
2-D, Monod
BIOPLUS 2-D, aerobic, Monod
ULTRA 2-D, first-order
- 2-D, denitrification
- 2-D, Monod, biofilm
Molz et al. (1)
Borden et al. (2)
Domenico (3)
Srinivasan and Mercer (4)
Semprini and McCarty (5)
Widdowson et al. (6)
Celia et al. (7)
Rifai et al. (8)
MacQuarrie et al. (9)
Wheeler et al. (10)
Tucker et al. (11)
Kinzelbach et al. (12)
Odencrantz et al. (13)
The ground-water velocity at the UCC site is approxi-
mately 5 m/yr.
Oxygen exchange with the unsaturated zone was simu-
lated by Borden et al. (2) as a first-order decay in
hydrocarbon concentration. The loss of hydrocarbon
due to horizontal mixing with oxygenated ground water
and resulting biodegradation was simulated by generat-
ing oxygen and hydrocarbon distributions independently
and then combining them by superposition. Simulated
oxygen and hydrocarbon concentrations closely
matched the observed values. The Conroe Superfund
site was one of the first sites to be modeled using a
biodegradation expression in a transport model.
Traverse City Site, Michigan
The Traverse City field site is a U.S. Coast Guard Air
Station located in Grand Traverse County in the north-
western portion of the lower peninsula of Michigan. The
ground water at the site is contaminated with organic
chemicals from a leaking underground storage tank. The
main contaminants at the site are benzene, toluene, and
xylenes (BTX). The contaminant plume ranges from
150 ft to 400 ft wide and is about 4,000 ft long. A
pumping wellfield system was installed at the down-
gradient end of the dissolved plume to control offsite
migration.
A modeling effort of natural attenuation at the site was
completed by Rifai et al. (8) with the BIOPLUME II
model. Modeling was performed for the period before
the pumping wells were installed and also for the period
after the wells were turned on. The data in Figure 1 show
the results of the model simulation along the center line
of the plume for the period before the wellfield was
80 _
0 „ Observed Data
*-__„ BIOPLUME II Model
M28
M26 TP4 TP3 M30
M31 IN2 M2
IN4M4
Figure 1. BIOPLUME II model predictions for the Traverse City
field site (8).
turned on. The model predictions by Rifai et al. (8)
matched the observed concentrations at the monitoring
wells reasonably well except in the vicinity of well M31.
Rifai et al. (8) indicated that this was because the simu-
lation did not account for anaerobic biodegradation,
which was occurring in the interior of the plume.
Gas Plant Facility in Michigan
Soluble hydrocarbon and dissolved oxygen (DO) were
characterized in a shallow aquifer beneath a gas plant
facility in Michigan by Chiang et al. (14). The distributions
86
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of BTX in the aquifer had been monitored in 42 wells for
a period of 3 years. The site geology is characterized as
a medium to coarse sand with interbeds of small gravel
and cobbles. The general direction of ground-water flow
is northwesterly. The depth to water table ranges from
10 ft to 25 ft below land surface, and the slope of the
water table was estimated as 0.006. Based on ground-
water and soil sampling data, Chiang et al. (14) con-
cluded that the flare pit was the major source of the
hydrocarbons found in the aquifer, while the slope oil
tank was a secondary source.
Chiang et al. (14) evaluated a first-order decay biode-
gradation approach and the BIOPLUME II model for
simulating biodegradation at the gas plant facility. Using
the model and assuming first-order decay, several simu-
lations were made to match the observed benzene con-
centration distribution of 1/22/85 by setting the observed
concentration distribution of 11/1/84 as the initial condi-
tion. The variables involved included the distribution of
the leakage/spill rates between the flare pit and the
slope oil tanks and macrodispersivities of the aquifer.
The BIOPLUME II model was used to simulate the July
1987 data by setting the observed concentration distri-
bution of February 1987 as an initial condition. The data
in Figure 2 show the comparison between the measured
and the simulated soluble BTX concentrations of July
1987. As can be seen from Figure 2, the model predic-
tions for BTX were reasonable. The model predictions
for oxygen concentrations, however, were not as similar.
The authors attributed the differences to the fact that the
BIOPLUME II model assumes a requirement of 3 ppm
of oxygen for 1 ppm of benzene, whereas the actual
requirement is in the range of 1 ppm to 3 ppm.
Cliffs-Dow Superfund Site
Ground water at the Cliffs-Dow site is contaminated with
low levels of phenolic and polycyclic compounds. The
aquifer sediments at the site consist of mostly coarse
sands and gravels. The hydraulic conductivity ranges
between 3.5 x 10'3 to 4.6 x 10'2 cm/s. The principal
contaminants found at the site near the source area
include phenol, several methyl-substituted phenols, and
naphthalene at concentrations ranging from 220 u,g/Lto
860 u,g/L. Based on the analysis of samples obtained
from monitoring wells, Klecka et al. (15) found that the
levels of organic contaminants are reduced to near or
below the detection limit within a distance of 100 m
downgradient from the source. Further analyses of the
ground-water chemistry were used to verify that biode-
gradation was occurring at the site and causing the
disappearance of the contaminants.
The migration of organic constituents in the aquifer was
simulated using the BIO1D model and assuming a first-
order decay expression. Half-lives for the contaminants
at the site were estimated from the results of soil
Retardation factor = 5
Ground-water velocity = 0.2 m/d
t1/2 = 2 days
*1/2 = 10 days
t1/2 = 20 days
50
100 150
Distance (m)
(a)
200
250
Retardation factor = 5
Ground-water velocity = 0.46 m/d
t1/2 = 2 days
*1/2 = 10 days
t1/2 = 20 days
50
100 150
Distance (m)
(b)
200
250
Figure 2. BIO1D model predictions for the Cliffs-Dow Super-
fund site (15).
microcosm experiments based on the time required for
50-percent disappearances of the parent compound.
The velocity was varied over a range from 0.2 to 0.46 m/d,
which is representative of the range of ground-water
flow rates at the site.
Figure 3 illustrates the impact of biodegradation on con-
taminant concentrations at the site. Model simulations
performed using a half-life of 2 d indicated that levels of
the phenolic components were reduced by greater than
99 percent within a distance of 30 m downgradient of
the source. When the half-life was increased by a factor
of 10, the concentrations were reduced to a similar
extent within 75 m. Because of the dominance of biode-
gradation, increases in ground-water velocity from 0.2
to 0.46 m/d had minor effects on the level of attenuation
predicted with the model.
Conclusions
A number of biodegradation models have been devel-
oped over the last 10 years. These models are generally
87
-------
,5317
0||1S03|3421
10690] *
10544] 0 g 0
4 1387 5876
|o||157315024
11301
0 0 1398
Figure 3. BIOPLUME II model predictions for the gas plant fa-
cility in Michigan (14)—top number: observed data;
bottom number: simulated data.
similar in that they simulate the transport and biodegra-
dation of a number of components in the ground water.
The models differ in the mathematical biodegradation
expressions that they use and in the numerical proce-
dures used to solve the complicated system of equa-
tions. Application of these models to field sites has
proven to be complicated due to the lack of biodegrada-
tion parameters that can be measured in the field for
model input. As a result, most modeling applications at
the field scale have resorted to first-order decay or
instantaneous representation of the biodegradation
process.
References
1. Molz, F.J., M.A. Widdowson, and L.D. Benefield.
1986. Simulation of microbial growth dynamics cou-
pled to nutrient and oxygen transport in porous
media. Water Resour. Res. 22(8):1,207-1,216.
2. Borden, R.C., P.B. Bedient, M.D. Lee, C.H. Ward,
and J.T. Wilson. 1986. Transport of dissolved hy-
drocarbons influenced by oxygen-limited biodegra-
dation, 2. Field application. Water Resour. Res.
13:1,983-1,990.
3. Domenico, P.A. 1987. An analytical model for mul-
tidimensional transport of a decaying contaminant
species. J. Hydrol. 91:49-58.
4. Srinivasan, P., and J.W Mercer. 1988. Simulation
of biodegradation and sorption processes in ground
water. Ground Water 26(4):475-487.
5. Semprini, L, and PL. McCarty. 1991. Comparison
between model simulations and field results tor in situ
biorestoration of chlorinated aliphatics, 1. Biostimula-
tion of methanotrophic bacteria. Ground Water
29(3):365-374.
6. Widdowson, M.A., F.J. Molz, and L.D. Benefield.
1988. A numerical transport model for oxygen- and
nitrate-based respiration linked to substrate and nu-
trient availability in porous media. Water Resour.
Res. 24(9): 1,553-1,565.
7. Celia, M.A., J.S. Kindred, and I. Herrera. 1989.
Contaminant transport and biodegradation, 1. A nu-
merical model for reactive transport in porous me-
dia. Water Resour. Res. 25(6):1,141-1,148.
8. Rifai, H.S., P.B. Bedient, J.T. Wilson, K.M. Miller,
and J.M. Armstrong. 1988. Biodegradation model-
ing at aviation fuel spill site. J. Environ. Eng.
114(5):1,007-1,029.
9. MacQuarrie, K.T.B., E.A. Sudicky, and E.O. Frind.
1990. Simulation of biodegradable organic contami-
nants in ground water, 1. Numerical formulation in
principal directions. Water Resour. Res. 26(2):207-
222.
10. Wheeler, M.F., C.N. Dawson, P.B. Bedient, C.Y.
Chiang, R.C. Borden, and H.S. Rifai. 1987. Numeri-
cal simulation of microbial biodegradation of hydro-
carbons in ground water. Proceedings of the
Solving Ground Water Problems With Models Con-
ference, Denver, CO (February 10-12). Dublin, OH:
National Water Well Association (NWWA).
11. Tucker, W.A., C.T Huang, J.M. Bral, and R.E. Dick-
inson. 1986. Development and validation of the un-
derground leak transport assessment model
(ULTRA). Proceedings of Petroleum Hydrocarbons
and Organic Chemicals in Ground Water: Preven-
tion, Detection, and Restoration, Houston, TX (Oc-
tober-November). Dublin, OH: National Water Well
Association (NWWA). pp. 53-75.
12. Kinzelbach, W, W Schafer, and J. Herzer. 1991.
Numerical modeling of natural and enhanced deni-
trification processes in aquifers. Water Resour. Res.
27(6):1,123-1,135.
13. Odencrantz, J.E., A.J. Valocchi, and B.E. Rittman.
1990. Modeling two-dimensional solute transport
with different biodegradation kinetics. Proceedings
of Petroleum Hydrocarbons and Organic Chemicals
in Ground Water: Prevention, Detection, and Res-
toration, Houston, TX (October-November). Dublin,
OH: National Water Well Association (NWWA).
14. Chiang, C.Y, J.P Salanitro, E.Y. Chai, J.D. Colthart,
and C.L. Klein. 1989. Aerobic biodegradation of
benzene, toluene, and xylene in a sandy aquifer:
Data analysis and computer modeling. Ground
Water 6:823-834.
15. Klecka, G.M., J.W. Davis, D.R. Gray, and S.S. Mad-
sen. 1990. Natural bioremediation of organic con-
taminants in ground water: Cliffs-Dow Superfund
site. Ground Water 4:534-543.
88
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Biogeochemical Processes in an Aquifer Contaminated by Crude Oil:
An Overview of Studies at the Bemidji, Minnesota, Research Site
Robert P. Eganhouse, Mary Jo Baedecker, and Isabella M. Cozzarelli
U.S. Geological Survey, Reston, VA
Abstract
Crude oil inadvertently released from a pipeline in a
remote area of north-central Minnesota has altered the
geochemistry of a shallow aquifer. Part of the oil was
sprayed over a large area to the west of the pipeline,
and another portion accumulated in an oil body that now
floats on the water table to the east of the point of
discharge. Dissolution of oil components into the ground
water and microbial degradation of the oil have resulted
in the formation of distinct geochemical zones in which
a variety of natural biogeochemical reactions can be
observed. Upgradient from the oil body in the "spray
zone," concentrations of total dissolved organic carbon
(TDOC), Ca, Mg, and HCO3- are greater and pH is lower
than measurements observed in native ground water.
These differences reflect the transport of oil constituents
to the water table by recharge, oxidation under aerobic
conditions, and dissolution of carbonates. Beneath and
just downgradient from the oil body, oxygen is depleted,
and anaerobic degradation reactions (Fe and Mn reduc-
tion, methanogenesis) dominate. This is evidenced by
increased concentrations of Fe2+, Mn2+, CH4, and TDOC
in ground water. Volatile hydrocarbons, mainly benzene
and its alkylated derivatives, represent 26 percent of the
TDOC in this zone. Microbially mediated removal of
these compounds within the anoxic zone is indicated by
the presence of structurally related oxygenated interme-
diates (for example, alkylated benzenecarboxylic acids)
and differences in the removal rates of isomeric alkyl-
benzenes. Downgradient from the anoxic zone, mixing
of oxygenated water with plume constituents leads to
removal of iron (via precipitation) and virtually all of the
oil-derived organic constituents. Within a distance of
200 m downgradient from the oil body, the geochemistry
of the ground water is virtually indistinguishable from
that of native ground water. Data collected over an 8-yr
period demonstrate that while the contaminant plume has
become increasingly reducing in character, its size has not
changed significantly. This attests to the efficiency of
natural processes in removing/attenuating the oil-derived
contaminants of primary concern at this site.
Introduction
Contamination of ground water by intentional or inadver-
tent releases of crude oil or refined petroleum products
is a widespread problem, and a great deal of effort (and
money) is presently being devoted to remediation ef-
forts. The efficacy of current engineering approaches is
subject to considerable debate. One thing is clear, how-
ever. For future remediation efforts to be effective and
useful, we must improve our understanding of how na-
ture responds to such impacts. In principle, this should
permit the manipulation of natural processes for pur-
poses of contaminant removal. Investigations that have
been carried out at the Bemidji, Minnesota, crude oil spill
site by the U.S. Geological Survey since 1984 are aimed
at developing such an understanding. The present paper
discusses some of the major findings of those studies.
Methods
Sampling and Analysis
Wells were installed at the study site using a hollow stem
augerwithout lubricants orgrease on the equipment (1).
Augers for the drill rig and stainless steel well screens
were steam cleaned prior to use. The polyvinyl chloride
casings were 5 cm in diameter. Two types of wells were
installed. Watertable wells were emplaced so that stain-
less steel screens (1.5 m long) intersected the water
table at the approximate midpoint of the screen; loca-
tions are shown in Figure 1. Deeper wells were installed
below the water table with screen lengths of 0.15 m or
0.61 m (Figure 2).
Samples of water, sediment, and oil were collected and
subjected to a variety of chemical analyses. Water sam-
ples to be used for determination of inorganic constitu-
ents and dissolved organic carbon were collected with
89
-------
Sampling Sites
Ground Water
Sediment
Oil
Oil and Sediment
Oil and Ground Water
Figure 1. Map of the field site near Bemidji showing location
of the ruptured pipeline, approximate location of the
oil body, area over which oil was sprayed, and sam-
pling locations (2).
435
430
425
420
^ Sn«"d
N^^^ce^ ,<^r^^
^ ^ «?*, ^
• i Screened / "
I interval of 7 , v. Zone V
_
-
Water Table '
'}'.'--
Direction of
. sampling sites ^u"u.' (' Ground-Water Flow" -
-200 -150 -100 -50 0 50 100 150
Distance From Center of Oil Body (Meters)
200
Figure 2. Cross section of the aquifer near Bemidji along main
sampling transect. Locations of water table and deep
wells indicated as filled bars (2). For a detailed de-
scription of zonation, see Baedecker et al. (8).
submersible pumps, whereas a Teflon bailer was used
to collect water for determination of organic constitu-
ents. The analyses included 1) water—dissolved oxy-
gen (DO), volatile dissolved organic carbon (VDOC),
nonvolatile dissolved organic carbon (NVDOC), meth-
ane, pH, Eh, major cations (Ca, Mg, Na, K), alkalinity,
NH4+, NO3-, Cr, SO42-, sulfide, iron (total 8 Fe2+), Mn2+,
Si, Ba, Al, Sr, S13Cmethane, §13CTic 0"IC=total dissolved
inorganic carbon), volatile hydrocarbons (VHCs), ex-
tractable hydrocarbons, XAD resin isolates, and low-
molecular-weight organic acids; 2) oil—813C, elemental
analysis (C,H,N,S), hydrocarbons, 13C-NMR, and heavy
metals; and 3) sediment—813C, hydrocarbons, and ele-
mental analysis (C,H,N,S). Details of the methods of
sample collection and analysis are given by Eganhouse
et al. (2,3) for hydrocarbons; by Leenheer and Huffman
(4), Huffman and Stuber (5), and Thorn and Aiken (6) for
XAD resin isolates; by Baedecker and Cozzarelli (7),
Baedecker et al. (8), and Bennett et al. (9) for inorganic
constituents, methane, and elemental analyses; and by
Cozzarelli etal. (10,11) for low-molecular-weight organic
acids.
Site Description
The study site is in north central Minnesota near the
town of Bemidji (Figure 1). The aquifer is a pitted and
dissected glacial outwash underlain by a poorly perme-
able till at about 24 m below land surface. The outwash
sediments are heterogeneous and composed of moder-
ately calcareous (6 percent carbonates), moderately to
poorly sorted sands consisting primarily of quartz and
feldspar of fine-to-medium grain size (1,12). Coring
studies have revealed that the sands are variably inter-
bedded with gravel deposits and clay lenses. The water
table is 6 m to 10 m below land surface, and ground-
waterflow is to the east-northeast (Figure 1), discharging
into an unnamed lake approximately 300 m downgradient
of the spill site. Estimates of the flow velocities near the
water table range from 0.05 meters/day (m/d) to 0.5 m/d
(9) for fine-grained and coarse-grained sediment, re-
spectively.
A pipeline rupture occurred in August 1979, spilling
1,670 m3 of crude oil. A portion of the 410 m3 of crude
oil unaccounted for after the cleanup effort is present in
a body of oil floating on the water table. The oil body
described here is irregularly distributed over a 7-m to
8-m vertical interval of unconsolidated sediment above
the water table; by 1990, it had spread to a length of 70 m
to 80 m in the direction of ground-water flow (13). An
area upgradient of this oil body received oil spray during
the pipeline rupture, and crude oil coated only the sur-
face sediment. This area, extending 140 m to 180 m to
the west-southwest of the pipeline and encompassing
approximately 6,500 m2, is hereafter referred to as the
"spray area" (see Figure 1).
90
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Results and Discussion
Effects of the Crude Oil on Ground-Water
Geochemistry
Formation of Geochemical Zones. The inadvertent
introduction of crude oil to this aquifer has resulted in
marked alteration of geochemical conditions in the
ground water. The driving force for these changes has
been the microbial degradation of metabolizable organic
matter, in this case represented by soluble constituents
of the oil. After several years of research during the mid
to late 1980s, it became evident that the saturated zone
of this aquifer could be characterized by geochemically
distinct zones. These zones, depicted in a cross section
centered on the main contaminant plume in the direction
ofground-waterflow(Figure2), correspond to I—uncon-
taminated native ground water, II—ground water up-
gradient of the oil body but within the "spray area,"
III—anoxic ground water immediately beneath and
downgradient of the oil body, IV—suboxic transitional
zone where anoxic ground water from zone III mixes
with more oxygenated ground water further downgradi-
ent, and V—oxygenated ground water that increasingly
approaches conditions of the native ground water. The
geochemistry of the ground water in each zone reflects
that of the water upgradient from it and processes (e.g.,
sorption, dilution, degradation, dissolution, gas exchange)
occurring within that zone.
The native ground water (zone I) is a dilute Ca-Mg-
HCO3" water with total dissolved solutes <400 mg/L,
a Ca:Mg ratio of 2.2:1, a median DO concentration of
7.68 mg/L, and a pH of 7.6 to 7.8. TDOC concentrations
are approximately 2 mg/L to 3 mg/L. The chemical com-
position is controlled by carbonate equilibria and, to a
lesser extent, by dissolution of quartz, feldspar, and clay
minerals and by the degradation of naturally occurring
organic material (8).
Ground water within zone II is influenced by the aerobic
degradation of soluble oil constituents originally depos-
ited at land surface by the spraying of oil from the
ruptured high pressure pipeline. In this area, oil is known
to have penetrated the upper 4 cm to 6 cm of the soil.
The quantities were insufficient, however, for formation
of a discrete oil phase. Ground water in zone II exhibits
elevated concentrations of TDOC, Ca, Mg and HCO3-;
lower pH (about 0.5 to 1 pH unit); and a slightly reduced
DO concentration. Carbon dioxide partial pressures
(PCo) calculated by Bennett et al. (9) are as high as 10"15
atm (as compared with ambient atmospheric values of
10"35 atm). The TDOC concentrations are nearly an
order of magnitude greater than those found in the
native ground water. While the additional TDOC in zone
II is undoubtedly derived from the oil, virtually all of it
represents partially oxidized transformation products,
not hydrocarbons (2). No VHCs were detected. The few
petroleum hydrocarbons that are present have
compositions that indicate extensive biodegradation.
Presumably, some of the degradation/oxidation oc-
curred in the unsaturated zone followed by transport of
soluble transformation products to the water table during
recharge. The slight depression of oxygen concentration
in ground water in this zone indicates that some degra-
dation is likely to be occurring within the ground water
as well. If removal of the NVDOC is occurring via aerobic
respiration, however, the rate of supply of NVDOC must
exceed its rate of removal because NVDOC concentra-
tions tend to increase with approach to zone III. Bennett
and others (9) have suggested that the large increases
in Ca, Mg, and HCO3- with increasing PCo2, as well as
the similarity of the [Ca+Mg]/[HCO3-] mole ratio of
ground water in this zone to that of the native ground
water, indicate that the principal reaction responsible for
generating the plumes of alkali-earth solutes is dissolu-
tion of carbonates by reaction with carbonic acid. Be-
cause these processes are occurring under oxic (as
opposed to anoxic) conditions, the primary chemical
signatures of the contaminant plumes reflect the prod-
ucts of organic remineralization (HCCy) and transfor-
mation (TDOC) and those resulting from acidification
(pH, Ca, Mg), not redox reactions.
Zone III encompasses ground water near, downgradient
from (about 75 m), and in immediate contact with the oil
to a depth of about 3 m below the water table. Dramatic
changes in ground-water geochemistry result from the
dissolution of soluble oil constituents and their metabo-
lism, leading to complete consumption of oxygen and
the dominance of anoxic degradation reactions, includ-
ing iron and manganese reduction and methanogene-
sis. Sulfate reduction and denitrification are not
important in this system because of the very low con-
centrations of SO42- and NO3- found in the native
ground water. In this zone, TDOC concentrations rise to
a maximum of 48 mg/L, and a significant fraction of this
is VDOC (42 percent). The volatile compounds are
dominated (approximately 63 percent) by a mixture of
saturated, aromatic, and alicyclic hydrocarbons derived
from the oil, the most important constituents of which are
benzene and a complex assemblage of alkylbenzenes.
Methane is found in the ground water, and field meas-
urements of Eh indicate a strongly reducing environ-
ment. In 1987 the average stable carbon isotope ratios
for TDOC (813CT|C) and methane (S13Cmethane) were -8.23
and -55.45 per mil, respectively. The average 813C value
for TIC in native ground water is -12.55 per mil. The
heavier ratio for TIC in zone III, thus, reflects fractiona-
tion resulting from methanogenesis. The most dramatic
changes in inorganic chemistry are seen as large in-
creases in the concentrations of dissolved iron and man-
ganese that result from mobilization of these redox
species in response to the microbially mediated
oxidation of hydrocarbons. Silica concentrations also
91
-------
increase due to enhanced dissolution of silicate miner-
als. Other compounds present in the ground water of
zone III, but not found in the oil or (except in trace
amounts) ground water in zones I and II, are a complex
mixture of oxygenated products of hydrocarbon degra-
dation, including low-molecular-weight organic acids
(10,11). The acids are structurally related to coexisting
monoaromatic hydrocarbons from the oil. Those organic
acids that have been identified correspond to aromatic
hydrocarbons whose concentrations decrease most
rapidly within the anoxic zone, whereas no potential acid
intermediates were found for aromatics which appear to
be more stable within zone III. These results signal the
partial oxidation of hydrocarbons to more soluble meta-
bolites.
Zone IV is a transition zone characterized by small but
detectable quantities of oxygen. Dissolved iron and sil-
ica decrease to below detection limits at the boundary
of this zone due to precipitation reactions. Low-molecu-
lar-weight organic acids are at or below detection limits,
and the concentrations of all oil-derived hydrocarbons
are much lower than in the ground water of zone III. Ca,
Mg, and Sr also decrease but rather gradually. Bennett
and others (9) have hypothesized that this is the result
of dispersive mixing and that transport of these constitu-
ents is conservative.
Evolution of the Contaminant Plume. Figure 3 depicts
the concentrations of dissolved Fe2+, Mn2+, methane,
and 813CT|C from ground water taken from zone III near
thedowngradientedgeofthe oil body for the years 1984
to 1992 (8,14). At this site, concentrations of methane
and Fe2+ increased by factors of 100 and 25, respec-
tively, during the first 5 years. Thereafter, the concentra-
tions of methane and Fe2+ have virtually leveled off.
Over the same period, manganese concentrations first
increased and then declined, whereas 813CT|C has in-
creased continuously. These variations in the contaminant
10
0.1
0.01
0.001
-5
OI^
-10 O"
H
O
^
-15 9
|
-20
-25
1984 1986
1988 1990 1992
Figure 3. Concentrations of dissolved ferrous iron, manga-
nese, methane (in millimoles), and 513C in per mil of
TDOC for the years 1984 to 1992 (14).
plume chemistry reflect evolutionary changes in the bi-
ology and geochemistry of this perturbed system with
continued supply of hydrocarbons and depletion of ter-
minal electron acceptors.
In particular, it would appear that manganese reduction
was an important degradative process at the earliest
stages of plume evolution, but as time has progressed
iron reduction and methanogenesis have become domi-
nant, presumably due to depletion of the supply of re-
ducible manganese. As of 1992, the dominant process
at this location appears to be methanogenesis. This is
supported by the steady increase in 813CT|C. During this
same period, the distribution and size of the contaminant
plume, as measured by the concentrations of monoaro-
matic hydrocarbons (2), have not changed appreciably.
While our understanding of the factors controlling plume
evolution remains limited and the size and shape of the
contaminant plume itself has not changed, it is clear that
the biogeochemistry within the plume is dynamic.
Transport and Fate of Oil-Derived
Contaminants
The following discussion considers the transport and
fate of the monoaromatic hydrocarbons within zones III
and IV. Benzene is the dominant individual VHC in
anoxic ground water near the leading edge of the oil
body (70 percent), reaching concentrations in excess of
9,300 u,g/L. The C-\.A alkylated benzenes are second in
abundance, representing approximately 14 percent of
the total VHCs. Alkanes, consisting of a complex mixture
of normal and branched hydrocarbons with fourto seven
carbon atoms, account for 12 percent of the VHCs.
Cyclic hydrocarbons (cyclopentanes and cyclohexanes),
cycloaromatics (e.g., alkylated indans), and heteroa-
tomic species (e.g., tetrahydrothiophenes) are relatively
minor constituents (total less than 4 percent). Thirty-one
meters farther downgradient from this site (but still within
zone III), the VHC composition is markedly different. The
relative abundance of alkanes (and toluene; see Figure
4) is reduced, and benzene plus monoaromatic hydro-
carbons with two to four alkyl carbon atom substituents
dominate (96 percent). Benzene represents 90 percent
of the total VHCs. These compositional differences are
also found among individual C^ (alkylated) benzenes.
There are four isomeric C2-benzenes, eight isomeric
C3-benzenes, and 22 isomeric C4-benzenes. All of these
compounds, with the exception of t-butylbenzene, which
was not detected in the crude oil, are present in the
contaminated ground water within zone III. The compo-
sition of the monoaromatic hydrocarbons changes sys-
tematically with distance downgradient from the oil body
and with increasing depth in the saturated zone (15,16).
Most importantly, isomeric alkylbenzenes show dramati-
cally different apparent removal rates downgradient
from the oil body. Because these isomers have similar
physical properties, the attenuation of VHCs (and
92
-------
0)
c
o
10-
0)
o
c
o
O
10'
10"
10"
Toluene
20 40 60 80 100 120 140 160
Distance From Center of Oil Body (meters)
Figure 4. Concentrations of benzene, toluene, and ethylben-
zene (ug/L) along the flowpath of the contaminant
plume. Also shown is the concentration of a conser-
vative solute with an assumed starting concentration
of 10,000 ug/L. Model assumptions discussed in
Baedecker et al. (8); figure from Mallard and
Baedecker (17).
therefore VDOC) is attributable to biological, rather than
physical, processes.
This hypothesis is reinforced by data illustrated in Figure
4. Here concentrations of benzene, toluene, ethylben-
zene, and a conservative tracer (see Baedecker et al.
[8] for discussion) are shown as a function of distance
downgradient from the center of the oil body. Two fea-
tures of these data are readily apparent. First, the con-
centration changes observed for these aromatic
hydrocarbons greatly exceed what would be predicted
on the basis of conservative transport. Second, because
benzene, toluene, and ethylbenzene form an homolo-
gous series with increasing KQWS (octanol-water parti-
tion coefficients), one would expect that a systematic
pattern of removal rates would be found if sorption was
the dominant process. This obviously is not the case, as
toluene is very rapidly removed by comparison with
either benzene or ethylbenzene. Clearly, biodegradation
is the most important process limiting the transport of
these hydrocarbons.
Conclusions
At the Bemidji field site, the introduction of crude oil to
the subsurface has resulted in dramatic changes in the
geochemistry of the ground water. Even so, natural
processes, with biodegradation being most important,
have effectively limited the transport of oil-derived con-
taminants to a distance of 200 m from the source.
Continued monitoring of the contaminant plume has
revealed the dynamic nature of the coupled biological
and geochemical processes operative at this site. An
understanding of these processes and the factors that
affect plume evolution will be essential if we are to
develop environmentally responsible remediation meth-
ods in the future.
References
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of the U.S. Geological Survey Toxic Substances
Hydrology Program, Monterey, CA (March 11-15).
pp. 614-620.
14. Baedecker, M.J., and I.M. Cozzarelli. 1994. Biogeo-
chemical processes and migration of aqueous con-
stituents in ground water contaminated with crude
oil. In: Dutton, A.R., ed. Toxic substances and the
hydrologic sciences. Minneapolis, MN: American
Institute of Hydrology, pp. 69-70.
15. Cozzarelli, I.M., R.P. Eganhouse, and M.J.
Baedecker. 1989. The fate and effects of crude oil
in a shallow aquifer, II. Evidence of anaerobic deg-
radation of monoaromatic hydrocarbons. In: Mal-
lard, G.E., and S.E. Ragone, eds. In: Proceedings
of the technical meeting of the U.S. Geological Sur-
vey Toxic Substances Hydrology Program, Phoe-
nix, AZ (September 26-30, 1988). Water Res.
Invest. Rep. 88-4220. pp. 21-34.
16. Eganhouse, R.P, T.F. Dorsey, C.S. Phinney, M.J.
Baedecker, and I.M. Cozzarelli. 1987. Fate of
monoaromatic hydrocarbons in an oil-contaminated
aquifer: Evidence for the importance of microbial
activity. In: Proceeding of the Geological Society of
America 1987 annual meeting, Phoenix, AZ.
19:652.
17. Mallard, G.E., and M.J. Baedecker. 1993. Hydro-
carbon transport and degradation in ground water:
U.S. Geological Survey investigations. In: Pare,
K.M., ed. Proceedings of the Air Combat Command
environmental quality 1993 symposium, Langley
AFB, VA (March 1-5) pp. 102-108.
94
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Simulation of Flow and Transport Processes at the Bemidji, Minnesota,
Crude-Oil Spill Site
Hedeff I. Essaid
U.S. Geological Survey, Menlo Park, CA
Abstract
The Bemidji, Minnesota, field site has provided an op-
portunity to study in detail the natural processes that
occur following a crude-oil spill. Detailed field studies
have characterized the subsurface oil distribution and
the characteristics of the contaminated ground-water
plume. Numerical models that simulate flow and trans-
port processes are useful tools for integrating informa-
tion collected in the field, for testing hypotheses, and for
studying the relative importance of simultaneously oc-
curring processes in complex, real-world systems. Nu-
merical modeling of multiphase flow at the Bemidji site
has illustrated the importance of spatial variability on the
movement and distribution of oil in the subsurface. Sol-
ute-transport modeling that includes aerobic and an-
aerobic degradation processes is being used as a tool
to study the field-scale, naturally occurring solute-trans-
port and degradation processes occurring at the site.
Introduction
On August 20, 1979, a buried oil pipeline near Bemidji
broke, spilling about 1.7 x 106 L (11,000 barrels) of crude
oil (Figure 1). The site is located in a pitted and dissected
glacial outwash plain. Depth to the water table ranges
fromO mto8 m below land surface, and the flow through
the aquifer is generally horizontal and northeastward
towards an unnamed lake 300 m downgradient from the
point of pipeline rupture. An estimated 1.2 x 106 L (7,800
barrels) of the spilled oil was removed by pumping from
surface pools, trenching, burning, and excavating soil
(1). The petroleum in the pipeline was under pressure,
causing oil to be sprayed over approximately 6,500 m2
when the pipeline broke. The oil collected in topographic
depressions and trenched areas where large volumes
of oil infiltrated into the subsurface, forming two main
bodies of oil floating on the water table. The subsurface
oil bodies provide a long-term, continuous source of
hydrocarbon components that dissolve in and are trans-
ported with the flowing ground water.
Numerical models that simulate flow and transport proc-
esses are useful tools for integrating information col-
lected in the field, for testing hypotheses, and for studying
the relative importance of simultaneously occurring proc-
esses in complex, real-world systems. Many researchers
working at the Bemidji site have focused considerable
effort on characterizing the subsurface distribution of the
oil and the nature of the resulting contaminated ground-
water plume. This paper reviews this work, then briefly
summarizes the results of numerical simulations of mul-
tiphase flow and of ground-water transport and biode-
gradation at the site.
The Observed Subsurface Oil Distribution
The Bemidji field site has provided an opportunity to
study in detail the subsurface oil distribution following a
95 5'4"
47 34'14"
Figure 1.
The north oil pool at the Bemidji crude-oil spill site.
Plus symbols are locations of boreholes where
samples were collected for saturation analyses, and
circles are wells sampled for ground-water concen-
trations. The line A-A' is the trace of the multiphase
flow simulation section, and B-B' is the trace of the
transport simulation section.
95
-------
spill. Field cores were collected for the purpose of de-
termining the oil-saturation distribution (the fraction of
the pore space that is occupied by oil) in the subsurface.
Determination of the subsurface fluid-saturation distribu-
tions required the implementation of a sampling tech-
nique that could recover relatively undisturbed core
samples from the unsaturated and saturated zones
while maintaining the in situ pore-fluid distribution (2). To
improve field sample collection, a freezing-tip core bar-
rel was developed and used for sample collection (3).
To allow visual inspection of the cores in the field, clear
polycarbonate liners, 47 mm in diameter and 1.5 m long,
were used within the core barrel.
Following retrieval, the cores were frozen and cut into
78-mm long subsamples using a circular saw fitted with
a masonry blade. The oil saturation of each core was
determined in the laboratory using a porous polyethyl-
ene (PPE) technique (2,4). In this process, strips of
hydrophobic PPE are placed into a slurry created by
adding water to a core sample. The PPE absorbs the oil
from the sample but does not take up water. The amount
of oil present in the core is calculated from the change
in weight of the oily PPE strips. The sample is then dried
in an oven, and water saturation is determined gravimet-
rically. Air saturation can then be calculated by subtract-
ing the sum of oil and water saturation from unity.
Following the saturation analysis, each sample was
sieved to obtain the particle-size distribution.
To date, samples have been collected and analyzed
from 10 boreholes (550 subsamples) at the site of the
north oil pool (Figure 1). The three-dimensional distribu-
tion of oil saturation at the north pool obtained by kriging
the observed data is shown in Figure 2 (5,6). The oil
distribution at this site is complex. A considerable
amount of oil remains in the unsaturated zone at loca-
tions where oil infiltrated from a trench excavated follow-
ing the spill. The maximum oil saturation measured at
this site (0.74) was located downgradient from the zone
of oil infiltration. The body of oil floating on the water
table is not lens-shaped. The oil-saturation contours
follow a sinuous path that is roughly parallel to the
direction of ground-water flow. Thin silt lenses at the
north pool site appear to have considerable influence on
the observed oil-saturation distribution. These layers
result in high residual oil saturations in the unsaturated
zone and cause the shape of the oil body floating on the
water table to be complex and irregular rather than
lens-shaped.
Geostatistical and multiphase flow simulation have been
used to assess the effect of spatial variability of hydrau-
lic properties on the oil-saturation distribution at the site.
Because of the difficulty of measuring sediment hydrau-
lic properties for oil-contaminated samples, the particle-
size distribution data were used to estimate the
permeabilities (k) of the field samples and the retention
curves (7). The mean and variance ofthe log(k) distribution
(k in m2) were -11.2 and 0.25, respectively, at the north
pool. Figure 3 shows a cumulative probability plot of
log(k) at the site. A linear relation indicates a lognormal
distribution of k, and the slope of the line is related to
the variance of k. The north pool plot of log k appears
to consist of two linear segments, a distribution that
suggests that the permeability distribution at this site
consists of two lognormal populations: a coarse fraction
(log(k) >-11.64) and a fine fraction (log(k) <-11.64).
To obtain a regular grid of k values needed for the
multiphase flow simulations, geostatistical simulation
techniques (8) were used to generate k distributions for
the north pool site that were conditioned on the values
estimated from the core samples. A permeability reali-
zation was obtained that reproduced the geometry ofthe
fine and coarse fractions and also reproduced the vari-
ability structure within these fractions. The details of this
process are explained by Dillard (9) and Dillard et al. (10).
Kriged North Pool Oil Saturations
430
Elevation
(m)
422
20
-/W
Distance (m)
0 Oil Saturation .75
t-"*?•»%.
-3
90
Figure 2. Three-dimensional oil saturation distribution at the
north oil pool site. The water table is at an elevation
of about 423.5 m above sea level.
99.99
Cumulative Probability
o coco
b PP -^Njcj^aicn^ioococococo t°!°
i -^ -MVJ -nvJoiooooaoooooiooco ooco
•
^
,/
••
i
/
/
f
/
•
••
/
r
/
r
5. -14. -13. -12. -11. -10. -9
Log (k)
Figure 3. Cumulative probability plot of log (permeability) at the
north pool site.
96
-------
Multiphase Flow Modeling
A two-dimensional numerical model was developed and
used to simulate multiphase flow along a longitudinal
vertical transect parallel to the direction of flow at the
north oil pool. The model solves a mass balance equa-
tion for the oil and water phases, assuming that the air
phase is maintained at atmospheric pressure. An impor-
tant feature of the model is that it incorporates hysteretic
relations between capillary pressures and fluid satura-
tions. The details of the model and the approaches used
are given by Essaid et al. (7).
The model was used to simulate subsurface flow from
the time of the spill in August 1979 until the samples
were collected in June 1990. The simulated section was
120 m long and 10 m deep. The initial condition for the
simulation was a hydrostatic water pressure distribution
corresponding to the measured water-table elevation at
430
Elevation
(m)
420
15
Distance (m)
100
430
the time of sampling. The lateral and bottom boundaries
were assumed to be hydrostatic water-pressure bounda-
ries. Oil was assumed to infiltrate through five constant
oil pressure nodes at the top boundary representing the
trench that was excavated following the spill. Oil infiltra-
tion was stopped when the cumulative oil infiltration was
equal to the estimated oil mass in the observed transect.
More details of the north pool simulation are given by
Dillard (9) and Dillard et al. (10).
Data from the first sample transect (Figure 4a) were com-
pared with the two-dimensional simulated oil-saturation
distribution. In the first simulation, a uniform mean k of
5 x 10-12 m2 and uniform mean capillary pressure/satu-
ration function parameters were used. A uniform k re-
sulted in a symmetric, lens-shaped oil body (Figure 4b)
that did not reflect the features of the observed oil
distribution. In the second simulation, spatial variability
of hydraulic properties was introduced, resulting in vari-
ability in oil saturations within the lens (Figure 4c). Very
little oil was trapped in the unsaturated zone. The shape
of the oil body floating on the water table became com-
plex and irregular, with zones of low oil saturation cor-
responding to the low permeability silt layers. In the third
simulation, hysteresis was introduced, causing more oil
to become entrapped in the unsaturated zone (Figure
4d). The modeling results suggest that the silt layers and
spatial variability exert a strong control on the oil distri-
bution in the subsurface.
Elevation
(m)
420
430
Elevation
(m)
420
d 430
Elevation
(m)
420
Distance (m)
120
Distance (m)
120
Distance (m)
, 'I , *-"! 4 "' .
0 Oil Saturation
120
1.0
Figure 4. Oil saturation distributions along the sample tran-
sect: a) observed oil saturations, b) simulated oil
saturations with uniform mean properties, c) simu-
lated oil saturation with spatially variable properties,
d) simulated oil saturation with spatially variable
properties and hysteresis.
The Observed Ground-Water Plume
Numerous researchers working at the Bemidji crude-oil
spill site have documented evidence for microbial activ-
ity and degradation of petroleum hydrocarbons in the
field (11-18). As a result of these studies, five geo-
chemical zones in ground water (Figure 5) have been
identified at the site (11,12,19). Zone 1 consists of
oxygenated, uncontaminated native ground water. Zone
2, which is below the area where the land surface was
sprayed by oil following the pipeline rupture, is charac-
terized by reduced oxygen concentrations and the pres-
ence of refractory high molecular-weight hydrocarbons.
Zone 3, beneath and immediately downgradient from
the separate-phase oil body, consists of an anoxic
plume of ground water that contains high concentrations
of hydrocarbons and methane. Zone 4 is the transition
zone from anoxic conditions to fully oxygenated condi-
tions, and concentrations of hydrocarbons decrease
rapidly as a result of aerobic degradation processes.
Zone 5 consists of oxygenated water downgradient of
the contamination plume that contains slightly elevated
concentrations of dissolved constituents. Long-term
monitoring of the plume since 1984 has shown that,
near the water table, the concentration of total dis-
solved organic carbon (TDOC) and dissolved oxygen
(DO) downgradient from the oil body has remained
relatively stable with time. In the anoxic zone (Zone 3),
97
-------
Recharge Zones
424
,- 420
416
Zone 1
Explanation
X Midpoint of Well
Screen
Datum Is Sea Level
Oil Body
100
200
Distance, in Meters
300
400
Figure 5. The simulated cross section showing the five geochemical zones in ground water.
concentrations of reduced manganese (Mn) and iron
(Fe) and of methane have increased with time, indicat-
ing a sequence of Mn reduction followed by Fe reduction
and methanogenesis.
Ground-Water Transport and
Biodegradation Modeling
A two-dimensional, multispecies solute-transport model
that incorporates biodegradation is being developed
and applied to the ground-water system at the Bemidji
site. The model is being used to quantify the field-scale
degradation processes and to identify the important fac-
tors affecting the distribution of solute species in the
field. The model simulates the aerobic and anaerobic
degradation processes that have been observed in the
contaminated ground-water plume at the spill site. The
U.S. Geological Survey's Method of Characteristics
transport model (20,21) was expanded to handle multi-
ple solutes and to include biodegradation terms. The
approach of Kindred and Celia (22) was used to repre-
sent the biodegradation terms in the transport equation.
Details of the model are given by Essaid et al. (23).
A vertical cross section of unit width that is approxi-
mately parallel to the direction of ground-water flow was
simulated using the transport model for the period from
the time of the spill in 1979 until September 1990 (Figure
5). Ground-water samples from numerous wells along
this section have been analyzed overtime (11,12,15).
Steady-state flow, no sorption, and isothermal condi-
tions were assumed. Four solutes and two microbial
populations were modeled. TDOC was split into two
fractions: degradable dissolved organic carbon (DDOC)
and refractory dissolved organic carbon (RDOC). The
remaining two solutes modeled were DO and methane.
To represent the aerobic and anaerobic degradation
processes, aerobic and methanogenic populations of
bacteria were included in the simulations. Competitive
inhibition was used to represent the suppression of
methanogenesis by oxygen. In this manner, as oxygen
in the ground water is consumed and an anoxic zone
develops, the methanogens begin to flourish, resulting
in increased methane production. Iron and manganese
reductions were not included because of the complexity
of incorporating the rock-water interactions of dissolu-
tion and precipitation into the transport model.
For simulation purposes, the system was represented
by an initially clean aquifer with background dissolved
organic carbon concentrations and fully oxygenated
water. Following the oil spill, it was assumed that DDOC
and RDOC dissolved and entered the aquifer with re-
charge water. The estimated values of initial concentra-
tions and recharge water concentrations for each solute
are given in Table 1. The oil present in the pore space
within the oil body reduces water flow through this zone.
The magnitude of reduction of water flow is a complex
function of the oil distribution. As a first approximation of
this effect, the hydraulic conductivity and recharge rate
in the zone of the oil body were reduced to 25 percent
of the aquifer values. No measurements or estimates
are available for many of the transport and biodegrada-
tion parameters under natural field conditions. There-
fore, reasonable estimates of these values were used in
the simulations. The details of the simulation parameters
and boundary conditions are given by Essaid et al. (23).
Transport Simulation Results
Observed and simulated profiles of TDOC, DO, and
methane near the water table are plotted in Figure 6.
The observed and simulated concentration profile of
TDOC is shown in Figure 6a. The observed points show
a concentration distribution at the water table that is
relatively stable with time. The simulation has captured
this feature, as can be seen by the similarity between
the simulated 1986 and 1990 concentration profiles.
98
-------
Table 1. Initial and Recharge Water Concentration (mg/L)*
Recharge Zone
Solute
DDOC
RDOC
TDOC
DO
Methane
Initial
Concentration
0.0
2.0
2.0
9.0
0.0
A
0.0
2.0
2.0
9.0
0.0
B
10.0
20.0
30.0
3.0
0.0
c
100.0
30.0
130.0
0.0
0.0
D
0.0
2.0
2.0
0.0
0.0
E
0.0
2.0
2.0
9.0
0.0
'Recharge zones A through E are shown in Figure 5.
TDOC
Methane
80
g. 60
1
ro
m
= 40
c
o
I
o
0
20
100 200 300
Distance, in Meters
400
30
0
CL
| 20
ro
D)
10
I
0
o
c
o
O
100 200 300
Distance, in Meters
400
DO
10
0 8
Q_
(/)
- 4
c
o
•
0
o
c
o
O
100 200 300
Distance, in Meters
400
Explanation
— 1986 Simulated
— 1990 Simulated
A 1986 Observed
1987 Observed
1988 Observed
1990 Observed
Figure 6. Graphs of simulated and observed concentrations at the water table: a) TDOC, b) DO, c) methane.
99
-------
There is an increase in TDOC concentration in the up-
gradient spray zone, followed by a rapid increase in
TDOC concentration in the zone of the oil body. Down-
gradient from the oil body, the TDOC concentration
decreases gradually to the background concentration.
This decrease is a result of microbial and physical proc-
esses. There is anaerobic degradation of DDOC within
the anoxiczone near the oil body and aerobic degrada-
tion of DDOC at the margins of the plume, where oxy-
genated recharge and ground water are encountered.
Also, there is dilution of TDOC downgradient from the
oil body as a result of the physical processes of dis-
placement and mixing of flowing ground water with re-
charge water.
The simulated DO concentration (Figure 6b) decreases
in the spray zone because of the assumed decrease in
DO concentration in recharge water, caused by degra-
dation of hydrocarbons in the unsaturated zone and by
the consumption of oxygen by degradation in the ground
water. Near the oil body and immediately downgradient
from it, the DO of the recharge water is assumed to have
been completely consumed in the unsaturated zone. An
anoxic zone develops in this area. Farther downgradi-
ent, DO begins to increase as oxygenated recharge
water enters the system.
Methane is produced in the anoxic zone that develops
in the immediate vicinity and downgradient of the oil
body. The methane peak is displaced downgradient
from the center of the oil body because of the input of
methane-free ground water from the upgradient area.
The predicted decline in methane concentration at a
distance of 230 m (Figure 6c) is a result of the upwelling
of oxygenated water caused by the upward bending of
flow lines around the oil body. The simulated profiles
show a marked increase in methane production from
1986 to 1990 as the population of methanogens in-
creases. This increase in methanogenesis results in a
slight decrease in TDOC concentrations from 1986 to
1990 (Figure 6a). The rate of increase in methane pro-
duction was quite sensitive to the biodegradation pa-
rameters used in the simulation.
To examine the effect of degradation on DDOC in the
aquifer, two-dimensional distributions of DDOC forthree
different simulations are shown in Figure 7. In the first
simulation, there is no degradation (Figure 7a); in the
second simulation, degradation occurs (Figure 7b); and
in the third simulation, degradation occurs and the hy-
draulic conductivity distribution is heterogeneous (Fig-
ure 7c). The distribution of DDOC for the case with no
degradation (Figure 7a) reflects the physical processes
of dispersion, diversion of flow around the oil body, and
the depression of the plume beneath the water table
because of the deflection of flow lines by incoming
recharge water.
In the second simulation, the anaerobic and aerobic
degradation processes result in a contaminant plume
that is narrower than the plume in the first simulation and
whose concentration gradients are comparatively sharp
at the edges (Figure 7b). In this simulation, 46 percent
of the total DDOC mass entering the aquifer is de-
graded: 14 percent by anaerobic degradation and 32
percent by aerobic degradation.
Previous work has shown that the hydraulic properties
of the aquifer are spatially variable (7,9). To make the
simulation more realistic, a heterogeneous hydraulic
conductivity distribution was created using the methods
of Dillard et al. (10) and was used in the transport model.
Because of the complex flow field, an irregularly shaped
plume develops (Figure 7c). The variability in flow paths
and flow velocities results in increased mixing and dis-
persion of ground water. This, in turn, results in in-
creased biodegradation. In this simulation, of the total
DDOC mass entering the aquifer, 60 percent is de-
graded: 21 percent by anaerobic degradation and 39
percent by aerobic degradation.
The simulations represent a highly simplified repre-
sentation of the true field conditions and neglect Fe and
Mn reduction. Also, the parameters used in the simula-
tions are highly uncertain. Nevertheless, the results do
reproduce the general features of the observed contami-
nated ground-water plume. In addition to the kinetics of
the biodegradation processes, important factors that af-
fect the magnitude of degradation and the distribution of
the solutes in the field are the recharge influx and the
degree of dispersion and mixing in the ground-water
system caused by heterogeneity of the hydraulic
conductivity.
Summary
The Bemidji crude-oil spill site has provided an opportu-
nity to study in detail the processes that occur following
a spill of an organic immiscible fluid that is slightly
soluble in water. Detailed field studies have charac-
terized the subsurface oil distribution and the charac-
teristics of the contaminated ground-water plume.
Numerical modeling of multiphase flow has illustrated
the importance of spatial variability on the movement
and distribution of oil in the subsurface. Solute-transport
modeling that includes aerobic and anaerobic degrada-
tion processes is being used as a tool to study the
field-scale solute-transport and degradation processes.
In addition to the kinetics of the biodegradation proc-
esses, important factors that affect the distribution of the
solutes in the field are the recharge influx and the degree
of dispersion and mixing in the ground-water system.
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crude oil at the Bemidji, Minnesota, research site:
100
-------
Recharge Zones
424
C —. 420
0
T3
.i
416
Explanation
K Midpoint of Well Screen
Zone Boundary
- 1 Isoconcentration Line
Datum Is Sea Level ,
Direction of
Ground-Water Flow
100
200
Distance, in Meters
300
400
Figure 7. Simulated two-dimensional distributions of DDOC: a) with no degradation, b) with degradation, c) with degradation and
spatial variability.
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1991. Field validation of conceptual models of mo-
bilization and transport of volatile petroleum deriva-
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sota. In: Mallard, G.E., and D.A. Aronson, eds.
Proceedings of the Technical Meeting of the U.S.
Geological Survey Toxic Substances Hydrology
Program, Monterey, CA (March 11-15,1991). Water
Res. Invest. Rep. 91-4034. pp. 621-626.
18. Lovley, D.R., M.J. Baedecker, D.J. Lonergan, I.M.
Cozzarelli, E.J.P Phillips, and D.I. Siegel. 1989.
Oxidation of aromatic contaminants coupled to mi-
crobial iron reduction. Nature 339:297-299.
19. Baedecker, M.J., D.I. Siegel, P. Bennett, and I.M.
Cozzarelli. 1989. The fate and effects of crude oil
in a shallow aquifer, I. The distribution of chemical
species and geochemical fades. In: Mallard, G.E.,
and S.E. Ragone, eds. Proceedings of the Techni-
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Substances Hydrology Program, Phoenix, AZ (Sep-
tember 26-30, 1988). Water Res. Invest. Rep. 88-
4220. pp. 13-20.
20. Goode, D.J., and L.F. Konikow. 1989. Modification
of a method-of-characteristics solute-transport
model to incorporate decay and equilibrium-control-
led sorption or ion exchange. Water Res. Invest.
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21. Konikow, L.F, and J.D. Bredehoeft. 1978. Com-
puter model of two-dimensional solute transport
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Survey Techniques of Water Resources Investiga-
tions. Book 7.
22. Kindred, J.S., and M.A. Celia. 1989. Contaminant
transport and biodegradation, 2. Conceptual model
and test simulations. Water Resour. Res. 26(6):1,149-
1,160.
23. Essaid, H.I., M.J. Baedecker, and I.M. Cozzarelli.
1994. Use of simulation to study field-scale solute
transport and biodegradation at the Bemidji, Minne-
sota, crude-oil spill site. In: Morganwalp, D.W, and
D.A. Aronson, eds. Proceedings of the Technical
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102
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Natural Attenuation of Trichloroethylene and
Similar Compounds:
Case Studies
103
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An Overview of Anaerobic Transformation of Chlorinated Solvents
Perry L. McCarty
Department of Civil Engineering, Stanford University, Stanford, CA
Abstract
Intrinsic cometabolic transformation of chlorinated sol-
vents commonly occurs at sites where co-contaminants
are present as primary substrates to support the energy
needs and metabolic activities of transforming bacteria.
The extent of transformation that occurs depends upon
the relative concentration of primary substrates and the
microorganisms and environmental conditions present.
Reduction of tetrachloroethene (PCE) and trichlo-
roethene (TCE) to ethene has occurred at many sites,
although transformations are often not complete.
Evidence for intrinsic biotransformation of chlorinated
aliphatic hydrocarbons (CAHs) is provided by the pres-
ence of CAM transformation products and indicators of
anaerobic biogical activity, such as the disappearance
of dissolved oxygen, nitrates, and sulfates, and the pro-
duction of methane and soluble iron (II).
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 CAHs, have been used widely for
degreasing of aircraft engines, automobile parts, elec-
tronic components, and clothing. Only during the past
15 years has it become recognized that CAHs can be
transformed biologically (1). Such transformations some-
times occur under the environmental conditions present in
an aquifer in the absence of planned human intervention,
a process called intrinsic biotransformation (2). Condi-
tions under which this is likely to occur with CAHs and
the end products that can be expected are discussed in
this paper.
The major chlorinated solvents are carbon tetrachloride
(CT), PCE, TCE, and 1,1,1-trichloroethane (TCA).
These compounds can be transformed by chemical and
biological processes in soils to form a variety of other
CAHs, including chloroform (CF), methylene chloride
(MC), cis- and trans-1,2-dichloroethene (c-DCE, t-DCE),
1,1-dichloroethene (1,1-DCE), vinyl chloride (VC),
1,1-dichloroethane (DCA), and chloroethane (CA). In
CAH transformation, the microorganisms responsible
cannot obtain energy for growth from the transforma-
tions. The transformations are brought about through
co-metabolism or through interactions of the CAHs with
enzymes or cofactors produced by the microorganisms
for other purposes. In co-metabolism, other organic
chemicals must be present to serve as primary sub-
strates to satisfy the energy needs of the microorgan-
isms. Chemical transformations of some CAHs can also
occur within the timeframe of interest in ground water.
Transformations that are likely, and the environmental
conditions required, are discussed below.
Chemical Transformation
TCA is the only major chlorinated solvent that can be
transformed chemically in ground water under all condi-
tions likely to be found and within the one- to two-decade
time span of general interest, although chemical trans-
formation of CT through reductive processes is a possi-
bility. TCA chemical transformation occurs by two
different pathways, leading to the formation of 1,1-DCE
and acetic acid (HAc):
(elimination)
1,1-DCE
(1)
CH3CCI3 TCA
+H2O
CH3COO +3/-/++3C/~ (hydrolysis)
HAc
(2)
The rate of each chemical transformation is given by the
first-order reaction:
C = C0e
,-kt
(3)
104
-------
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:
k = Ae
.-E/0.008314K
(4)
where A and E are constants and K is the temperature
in degrees Kelvin. Table 1 provides a listing of values
reported for A and E for TCA abiotic transformation by
various investigators, and 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 ti/2 = 0.69/k. The temperature
effect on TCA half-life is quite significant.
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, however, is of great sig-
nificance because 1,1-DCE is considered more toxic
than TCA, with a maximum contaminant level of 7 |j,g/L
compared with 200 |j,g/L for TCA. 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.
Chloroethane, 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 nonharmful products (6).
Biological Transformations
CAHs can be oxidized or reduced, generally through
co-metabolism, as noted in Table 2. In ground waters,
intrinsic reductive transformations are most often noted,
perhaps because the presence of intermediate products
that are formed provide strong evidence that reductive
transformations are taking place. Intrinsic aerobic trans-
formation of TCE is also possible, although if it did occur,
the intermediate products are unstable and more diffi-
cult, analytically, to measure. Thus, convincing evidence
for the latter is difficult to obtain. Also, aerobic co-meta-
bolism of TCE would only occur if sufficient dissolved
oxygen and a suitable electron donor, such as methane,
ammonia, or phenol, were present. Since circumstances
under which the proper environmental conditions for
significant aerobic co-metabolism are unlikely to occur
often, intrinsic aerobic co-metabolism of TCE is prob-
ably of little significance. Evidence is ample, however,
that anaerobic reductive transformations of CAHs occur
frequently, and this process is important to the transfor-
mation of all chlorinated solvents and their transforma-
tion products. The major environmental requirement is
the presence of sufficient concentrations of other or-
ganics that can serve as electron donors for energy
metabolism, which often is the case in aquifers. Indeed,
the extent to which reductive dehalogenation occurs
may be limited by the amount of such co-contaminants
present. Theoretically, only a 0.4-g chemical oxygen
demand (COD) equivalent of primary substrate would
be required to convert 1 g of PCE to ethene (7), but
much more is actually required because of the co-
metabolic nature of the transformation.
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 completed reduction to ethane. Table
3 indicates that while some transformations, such as CT
to chloroform and carbon dioxide, may take place under
mild reducing conditions such as those associated with
denitrification, complete reductive transformation to in-
organic end products and of PCE and TCE to ethene
generally requires conditions suitable for methane fer-
mentation. Extensive reduction, although perhaps not
complete, can also occur under sulfate-reducing condi-
tions. For methane fermentation to occur in an aquifer,
the presence of sufficient organic co-contaminant is
required to reduce the oxygen, nitrate, nitrite, and
sulfate present. Some organics will be required to re-
duce the CAHs, and perhaps Fe(ll) as well, if present
Table 1. Reported First-Order TCA Abiotic Transformation Rates (kTCA)
krcA (yr'1)
A
yr1
3.47 (10)20
6.31 (10)20
1.56 (10)20
Average half-life (yr)
E
kJ
118.0
119.3
116.1
10
°c
0.058
0.060
0.058
12
15
°C
0.137
0.145
0.137
4.9
20
°C
0.32
0.34
0.31
0.95
References
Haag and Mill (3)
Cline and Delfino (4)
Jeffers et al. (5)
105
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Table 2. Biodegradability of Chlorinated Solvents Under Aerobic or Anaerobic Conditions and Through Use as a Primary
Substrate for Energy and Growth or Through Co-metabolism
1,1,1-
Carbon Tetrachloroethylene Trichloroethylene Trichloroethane
Tetrachloride (CT) (PCE) (TCE) (TCA)
\erobic Biotransformation
3rimary substrate No No No No
}o-metabolism No No Yes Perhaps
\naerobic Biotransformation
3rimary substrate No Perhaps Perhaps No
^o-metabolism Yes Yes Yes Yes
Hazardous intermediates Yes Yes Yes Yes
Chemical Transformation Perhaps No No Yes
CCI2-CCI2| PCE
CHCI=CCI2| TCE g CH3CCI3
± ^^ ±
CHCI = CHCI CH2=CCI2 g/CH3fCHCI2
CH =CHCI / CH CH Cl /
\ \ ' J 7
CH2=CH2\ CHgCOOH3 * s
CH =CH, X T •»
CT 4 dehaloaenation had occurred. Other dichloroethylenes
/ i (1,1-DCE and trans-DCE) were not significant in con-
/ CHCI3 centration, indicating that cis-DCE was the major trans-
/ ? formation intermediate. Microcosm studies also supported
/ CH2CI2 that biotransformation was occurring at the site, with com-
f plete disappearance of PCE, TCE, and cis-DCE and
CH3ci production of both VC and ethene. The conversions
' were accompanied by significant methane production,
indicating that suitable redox conditions were present for
the transformation.
co2 + H2o + cr
Figure 1. Anaerobic chemical and biological transformation
pathways for chlorinated solvents.
insignificant amounts. If the potential for intrinsic trans-
formation of CAHs is to be evaluated, then the concen-
trations of nitrate, nitrite, sulfate, Fe(ll), and methane,
and of organics (as indicated by COD or total organic
carbon [TOC]) should be determined. Unfortunately,
such analyses are not considered essential in remedial
investigations, but it is evident that they should be.
Case Studies
Major et al. (10) reported field evidence for intrinsic
bioremediation of PCE to ethene and ethane at a chemi-
cal transfer facility in North Toronto. PCE was stored at
the site 10 years prior to the study and contaminated the
ground water below with both free and dissolved PCE.
In addition to high concentrations of PCE (4.4 mg/L),
high concentrations of methanol (810 mg/L) and acetate
(430 mg/L) were found in the contaminated ground
water; methanol and acetate are co-contaminants that
served as the primary substrates for the transforming
organisms. Where high PCE was 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) was found. At one
downgradient well, however, no PCE or TCE was found,
but cis-DCE (76 mg/L), VC (9.7 mg/L), and ethene
Fiorenza et al. (11) reported on PCE, TCE, TCA, and
dichloromethane (DCM) contamination of ground water
at two separate locations at a carpet-backing manufac-
turing plant in Hawkesbury, Ontario. The waste lagoon
was the major contaminated area, with ground water
containing 492 mg/L of volatile fatty acids and 4.2 mg/L
of methanol, organics that appeared to provide the co-
contaminants that served as primary sources of energy
for the dehalogenation reactions. Here, the sulfate con-
centration 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 well above
the upgradient concentration of 2.1 mg/L. Methane was
present, although quite low in concentration (0.06 mg/L).
These parameters are all supportive of conditions suit-
able for intrinsic biodegradation of the chlorinated sol-
vents. While some chemical transformation of TCA was
indicated (0.4 mg/L), biotransformation was quite exten-
sive, as indicated by a 1,1-DCA concentration of 7.2 mg/L
compared with the TCA concentration of 5.5 mg/L.
Some CAwas also present (0.19 mg/L). Transformation
was also indicated for PCE and TCE, which remained
at concentrations of only 0.016 mg/L and 1.5 mg/L,
respectively, while the cis-DCE, VC, and ethene
concentrations were 56, 4.2, and 0.076 mg/L, respec-
tively. Only traces of ethane were found. Trans-DCE
concentration was only 0.57 mg/L, again providing evi-
dence that cis-DCE is the most common transformation
intermediate from TCE and PCE. Downgradient from the
106
-------
Table 3. Environmental Conditions Generally Associated With Reductive Transformations of Chlorinated Solvents
Redox Environment
Chlorinated Solvent
All
Denitrification
Sulfate Reduction
Methanogenesis
Carbon tetrachloride ~*
,,,-,-.., ,. TCA-M.1-DCE + CHgCOOH —
1 ,1,1-Trichloroethane J
Tetrachloroethylene
Trichloroethylene
CT-
TCA
PCE
TCE
-» 1,1-DCA
-M.2-DCE
-M.2-DCE
TCA -^ CO2 + Cl-
PCE ^> ethene
TCE ^ethene
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 intrinsic biotransformation is provided for
this site, the ethene and ethane concentrations appear
very low compared with VC concentration, suggesting
that biotransformation was not eliminating the chlorin-
ated solvent hazard at the site, although it was produc-
ing 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 (12) of data from Charnley et al. (13) indicated
average gaseous concentrations in parts per million by
volume from eight refuse landfills to be 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 presence of high
VC 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 (7,14-17). Ground-
water concentrations of TCE as high as 100 mg/L were
found present, 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 percent to 25 percent of the TCE had
been converted to ethene, and that up to 15 percent of
the reduction in COD in this zone was associated with
reductive dehalogenation (15). Through more extensive
analysis of ground water farther downgradient from the
contaminating source, Wilson et al. (17) found a 24-fold
reduction in CAHs across the site. 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.
References
1. McCarty, P.L., and L. Semprini. 1994. Ground-water
treatment for chlorinated solvents. In: Morris, R.E.,
ed. Handbook of bioremediation. Boca Raton, FL:
Lewis Publishers, Inc. pp. 87-116.
2. Council, N.R. 1993. In situ bioremediation: When
does it work? Washington, DC: National Academy
Press.
3. Haag, W.R., and T Mill. 1988. Effect of subsurface
sediment on hydrolysis of haloalkanes and epox-
ides. Environ. Sci. Technol. 22:658-663.
4. Cline, P.V, and J.J. Delfino. 1989. Transformation
kinetics of 1,1,1-trichloroethane to the stable prod-
uct 1,1-dichloroethene. In: R.A. Larson, ed. Biohaz-
ards of drinking water treatment. Chelsea, Ml:
Lewis Publishers, Inc. pp. 47-56.
5. Jeffers, P., L. Ward, L. Woytowitch, and L. Wolfe.
1989. Homogeneous hydrolysis rate constants for
selected chlorinated methanes, ethanes, ethenes,
and propanes. Environ. Sci. Technol. 23(8): 965-
969.
6. Vogel, T.M., C.S. Griddle, and PL. McCarty. 1987.
Transformations of halogenated aliphatic com-
pounds. Environ. Sci. Technol. 21:722-736.
7. McCarty, PL., and J.T Wilson. 1992. Natural an-
aerobic treatment of a TCE plume, St. Joseph,
Michigan, NPLsite. In: U.S. EPA. Bioremediation of
hazardous wastes. EPA/600/R-92/126. Cincinnati,
OH. pp. 47-50.
8. Freedman, D.L., and J.M. Gossett. 1989. Biological
reductive dechlorination of tetrachloroethylene and
trichloroethylene to ethylene under methanogenic
conditions. Appl. Environ. Microbiol. 55(9):2,144-
2,151.
9. de Bruin, WP, et al. 1992. Complete biological
reductive transformation of tetrachloroethene to
107
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ethane. Appl. Environ. Microbiol. 58(6):1,996-
2,000.
10. Major, D.W., W.W. Hodgins, and B.J. Butler. 1991.
Field and laboratory evidence of in situ biotransfor-
mation of tetrachloroethene to ethene and ethane
at a chemical transfer facility in North Toronto. In:
Hinchee, R.E., and R.F. Olfenbuttel, eds. Onsite
bioreclamation. Boston, MA: Butterworth-Heine-
mann. pp. 147-171.
11. Fiorenza, S., et al. 1994. Natural anaerobic degra-
dation of chlorinated solvents at a Canadian manu-
facturing plant. In: Hinchee, R.E., A. Leeson, L.
Semprini, and S.K. Ong, eds. Bioremediation of
chlorinated and polycyclic aromatic hydrocarbon
compounds. Boca Raton, FL: Lewis Publishers, Inc.
pp. 277-286.
12. McCarty, PL., and M. Reinhard. 1993. Biological
and chemical transformations of halogenated ali-
phatic compounds in aquatic and terrestrial environ-
ments. In: Oremland, R.S., ed. The biogeochemistry
of global change: Radiative trace gases. New York,
NY: Chapman & Hall, Inc.
13. Charnley, G., E.A.C. Crouch, L.C. Green, and T.L.
Lash. 1988. Municipal solid waste landfilling: A
review of environmental effects. No. Meta Systems,
Inc.
14. Hasten, Z.C., PK. Sharma, J.N. Black, and PL.
McCarty. 1994. Enhanced reductive dechlorination
of chlorinated ethenes. In: U.S. EPA. Bioremedia-
tion of hazardous wastes. San Francisco, CA.
15. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T
Wilson. 1993. Natural anaerobic bioremediation of
TCE at the St. Joseph, Michigan, Superfund site.
In: U.S. EPA. Symposium on bioremediation of haz-
ardous wastes: Research, development, and field
evaluations (abstracts). EPA/600/R-93/054. Wash-
ington, DC (May). Cincinnati, OH. pp. 47-50.
16. McCarty, PL., et al. 1991. In situ methanotrophic
bioremediation for contaminated ground water at
St. Joseph, Michigan. In: Hinchee, R.E., and R.G.
Olfenbuttel, eds. Onsite bioreclamation processes
for xenobiotic and hydrocarbon treatment. Boston,
MA: Butterworth-Heinemann. pp. 16-40.
17. 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. Pre-
sented at the U.S. EPA Symposium on Intrinsic
Bioremediation of Ground Water, Denver, CO (Au-
gust 30 to September 1).
108
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Contamination of Ground Water With Trichloroethylene at the Building 24 Site at
Picatinny Arsenal, New Jersey
Mary Martin and Thomas E. Imbrigiotta
U.S. Geological Survey, West Trenton, NJ
Abstract
Ground water at the Building 24 site at Picatinny Arsenal
in Morris County, New Jersey, is contaminated with
trichloroethylene (TCE). Estimated average linear
ground-water flow velocities are 0.3 to 1.0 m/d, and
estimated travel time from the source area to Green
Pond Brook is 2 y to 5 y. The total mass of dissolved
TCE within the 130,000-m2 plume area is estimated to
be 970 kg. About 65 percent of the mass is in areas
where TCE concentrations exceed 10,000 |j,g/L,
whereas about 30 percent is in areas where TCE con-
centrations are 1,000 to 10,000 |j,g/L.
The average flux of TCE discharged to Green Pond
Brook from the plume area, estimated from measured
TCE concentrations in samples of water from the uncon-
fined aquifer and measured base-flow discharge in
Green Pond Brook, is 1 to 2 mg/s. Biotransformation is
the most important mechanism by which dissolved TCE
leaves the ground-water system.
Introduction
Picatinny Arsenal, a U.S. Army armament research and
development center, is located in a glaciated valley in
north-central New Jersey (Figure 1). In 1960, Building
24 was remodeled, and a new metal-plating facility and
industrial wastewater treatment plant were installed.
From 1960 to 1981, the wastewater treatment system
discharged tens of thousands of liters of wastewater
daily into two 2.5-m deep, sand-bottomed settling la-
goons behind the building (1). The metal-plating waste-
water contained trace metals, such as cadmium,
chromium, copper, lead, nickel, tin, vanadium, and zinc,
and other ions used in plating solutions, such as so-
dium, potassium, sulfate, chloride, and cyanide (2).
From 1973 to 1985, an improperly installed relief system
of the degreasing unit allowed pure chlorinated solvents
to condense in an overflow pipe and discharge to a 1-m
deep dry well in front of Building 24. The condensate
from the overflow system contained TCE and, after
1983, 1,1,1-trichloroethane (2).
The infiltration of wastewater from the lagoons and of
chlorinated solvents from the dry well has created a
plume of contaminated ground water downgradient from
Building 24. Most of the contamination is limited to
unconfined sediments, where estimated ground-water
flow velocities (based on estimated horizontal hydraulic
conductivities, measured head gradients, estimated po-
rosities, and results of calibrated solute-transport model
simulations) generally range from 0.3 to 1.0 m/d in the
plume area. On the basis of these estimated velocities,
the residence time within the unconfined aquifer of a
conservative solute entering the aquifer near Building 24
and discharging at Green Pond Brook is 2 y to 5 y.
Ground-water contamination in the unconfined aquifer
has been discussed previously by Sargent et al. (3),
Fusillo et al. (2), and Imbrigiotta et al. (4,5). Results of
water-quality sampling during 1986 to 1991 confirm that
TCE remains the dominant contaminant in the uncon-
fined aquifer and that the extent of the plume of TCE-
contaminated ground water has changed little since
September 1986. These water-quality measurements
also show that 1) Building 24 is the source of the TCE
plume; 2) the highest TCE concentrations near Building
24 are found at depths of less than 6 m and within 10 m
downgradient from the dry well; 3) the plume extends
500 m from Building 24 to Green Pond Brook and fol-
lows the ground-water head gradients as it flows down-
ward through the unconfined aquifer, then upward
toward Green Pond Brook; 4) the plume disperses as it
moves downgradient, is about 350 m wide where it
enters Green Pond Brook, and has an areal extent of
about 130,000 m2; 5) the highest concentrations of TCE
are found at the base of the unconfined aquifer midway
between Building 24 and Green Pond Brook; and 6) the
highest TCE concentrations at Green Pond Brook are
found at depths of less than 6 m (Figure 1).
109
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A. Areal extent of trichloroethylene
contaminant plume
EXPLANATION
I :-:::::::| Area in which trichloroethylene concentration
li:::::S:'vil exceeds 10 micrograms per liler
10 LINE OF EQUAL TRICHLOROETHYLENE
CONCENTRATION-Shows trichloroethylene
concentration, in micrograms per liter.
Dashed where approximate
A —- A' Line of section
•41-9 Sampling site location and local identifier
Bear Swamp Brook
620
600 400 200 0 200 400 600 800 1,0001.2001.4001.6001.8002,000
DISTANCE FROM BUILDING 24 (B-24), IN FEET
B. Vertical distribution of trichloroethylene
concentrations
EXPLANATION
—100 LINE OF EQUAL TRICHLOROETHYLENE
CONCENTRATION-Shows trichloroethylene
concentration, in micrograms per liter.
Dashed where approximate
iWcll screen and trichloroethylene concentration,
100 in micrograms per liter
NS Not sampled
< Less than
CAF-7 Location of well and local identifier
Figure 1. (A) Location of Building 24 study area at Picatinny Aresenal, and areal extent of TCE plume and altitude of water table,
January 1993, and (B) vertical distribution of TCE concentrations, October to November 1991. (Location of section A-A'
is shown in Figure 1A.)
110
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The U.S. Geological Survey is conducting an interdisci-
plinary research study of ground-water contamination by
the chlorinated solvents and other contaminants at the
Building 24 site at Picatinny Arsenal. The objectives of
the study are to 1) describe the chemical, physical, and
biological processes that affect the movement and fate
of these contaminants, particularly TCE, in the subsur-
face; 2) determine the relative importance of these proc-
esses; and 3) develop predictive models of contaminant
transport.
This paper describes a conceptual model of the proc-
esses that affect the fate and transport of chlorinated
solvents at Picatinny Arsenal research site. A prelimi-
nary solute mass balance of TCE in the unconfined
aquifer at the site is presented and compared with re-
sults of numerical solute-transport simulations.
Conceptual Model
A conceptual model of the physical, chemical, and bio-
logical processes that affect the transport and mass
balance of TCE within the plume was developed by
Imbrigiotta and Martin (6) (Figure 2). TCE is present at
the site in several forms: solute phase (dissolved in
water), vapor phase (soil gas), and matrix phase (sorbed
onto solid surfaces or associated with biota). TCE also
may be present as a dense nonaqueous-phase liquid
(DNAPL). Nonaqueous-phase flow of TCE from Building
24 to the base of the unconfined aquifer is hypothesized
to be the cause of high concentrations of TCE at the
base of the unconfined aquifer (5).
Results of chemical analyses of samples of ground
water, soil gas, and aquifer sediments from the plume
area indicate that the fate and transport of TCE are
affected by the physical, chemical, and biological proc-
esses at the site. Advection and dispersion in the satu-
rated zone affect the movement of dissolved TCE and
cause it to be removed from the system in the discharge
to Green Pond Brook. TCE also is removed from the
system by anaerobic biotransformation (reductive deha-
logenation), volatilization at the water table, and sorption
to saturated-zone sediments. Volatilization of TCE to the
unsaturated-zone soil gas was measured by Smith and
others (7). Transport in the unsaturated zone was
determined to be driven principally by molecular diffu-
sion. The detection of the biotransformation products
Volatilization
(0.1 mg/s)
Advective
Transport to
Green Pond Brook Lgnd
(1-2-mg/s) Surfgce
Trichloroethylene Plume
Snort-Term •',•
Adsorption
(Not Estimated),-'.;
Long-Term
Desorption
(15-85 mg/s)
Short-Term
Desorption
(Not Estimated)
Anaerobic
Biotransformation
(1 -30 mg/s)
Dissolution
(Not Estimated)
/
Estimated Top of Confining Unit
Not to Scale
Gains
Trichloroethylene Mass Balance Components
(mg/s, milligrams per second; <, less than)
Losses
Long-Term Desorption
Infiltration
Dissolution
Short-Term Desorption
15-85 mg/s
< 0.1 mg/s
Not Estimated
Not Estimated
Anaerobic Biotransformation
Advective Transport to
Green Pond Brook
Volatilization
Short-Term Absorption
1 -30 mg/s
1 -2 mg/s
0.1 mg/s
Not Estimated
Figure 2. Conceptual model and preliminary mass-balance estimates of TCE fluxes resulting from processes that affect the fate
and transport of TCE in the ground-water system at Picatinny Arsenal.
111
-------
cis-1-2-dichloroethylene (cis-DCE) and vinyl chloride
(VC) in water from 75 percent of the wells screened
within the plume is indicative of biologically mediated,
reductive dehalogenation of TCE. The occurrence of
methanogenesis is indicated by the detection of dis-
solved methane in water from 85 percent of the wells in
which cis-DCE was detected and in water from 94 per-
cent of the wells in which VC was detected.
Desorption from contaminated sediments is a source of
TCE in the ground-water system. If TCE is present as a
DNAPL, dissolution will result in a source of dissolved
TCE in the ground water. Relatively constant TCE con-
centrations measured near Building 24 and at the base
of the unconfined aquifer from 1986 to 1991 indicate that
slow desorption of TCE from contaminated aquifer sedi-
ments, dissolution of DNAPL TCE in the aquifer, or both
are a continuing source of dissolved TCE in the ground-
water system.
Estimated Mass Distribution of TCE
For a given sample set, the total mass of dissolved TCE
below the water table is estimated to be 970 kg or about
660 L of liquid TCE. This estimate was calculated by
using results of six sets of water-quality analyses made
from 1986 to 1991 (Figure 3). Each measured TCE con-
centration is assumed to represent the TCE concentration
of a volume of ground water extending half the horizon-
tal and vertical distance to the adjacent sampling points.
These volumes are limited vertically by the extent of the
unconfined aquifer and horizontally by the 10-|j,g/L maxi-
mum TCE concentration line. Aquifer porosity is
assumed to be 30 percent. The total volume of dissolved
TCE within the ground water outside the 10-|j,g/L
maximum TCE concentration line is estimated to be
about 1 L.
The estimated total volume of TCE appears to be unre-
lated to the total number of samples collected. Foursets
of data plotted in Figure 3 (A, C, E, and F) show that
about 65 percent of the total volume of TCE is found in
ground water that contains TCE concentrations greater
than 10,000 |j,g/L, and about 30 percent is found in
areas where TCE concentrations range from 1,000 to
10,000 |j,g/L. The estimate of the volume of TCE in the
water appears to depend mostly on the number of sam-
ples that contain relatively high concentrations of TCE
and the volume of water each sample is assumed to
represent.
The relative amounts of TCE in the dissolved, vapor, and
matrix phases in a block of aquifer and unsaturated zone
immediately downgradient from Building 24 were esti-
mated on the basis of measured TCE concentrations in
samples of all three phases (8) (Figure 4). Most of the
o
Sample Set A
Sample Set
Number of Samples
1,319
Ooo
Explanation
[|ig/L, micrograms per liter; <, less than or equal to; >, greater than]
Volume of trichloroethylene
calculated using samples
with trichloroethylene
concentrations of:
• 1 to<100|ig/L
M >100to<1,000|ig/L
•i >1,000 to <10,000 ng/L
CU >10,000 ng/L
Sample
Set
CO
E
CD
^3
2
o
^
.c
o
Date of Collection
A April and August-September 1986
B August 1987
C June 1989
D November-December 1989
E April-March 1990
F October-November 1991
Figure 3. Estimated volume of dissolved TCE in ground water at the Building 24 site at Picatinny Arsenal, 1986-91.
112
-------
Building 24
Mass distribution of trichloroethylene near Building 24
[kg, kilogram; <, less than]
Unsaturated
zone
Saturated
zone
Phase
Vapor
Matrix
Solute
Matrix
Mass (kg)
0.001
3
1
4
Percent
of Total
<0.1
99.9
20
80
Not to Scale
Figure 4. Mass distribution of TCE in the saturated and unsaturated zones immediately downgradient from Building 24 at Picatinny
Arsenal.
mass of TCE in the system near Building 24 is in the
matrix phase and is associated with the sediments in
both the unsaturated and saturated zones. The mass
distribution shown in Figure 4 is based on a repre-
sentative sample of each TCE phase in the area of
Building 24. The ratio of the mass of TCE sorbed to the
soil to the mass of dissolved TCE in six sets of samples
collected near the water table throughout the site ranged
from 2:1 to 5:1. One sample, collected downgradient
from the wastewater lagoons, had an unusually high
ratio of about 20:1. The amount of DNAPL TCE at the
site has not been estimated.
Preliminary Solute Mass Balance
Desorption of TCE from soils that have undergone long-
term adsorption (years) and dissolution of DNAPL TCE
probably are the processes by which most TCE enters
the ground-water system, because the direct release of
TCE from Building 24 stopped in 1985. Because the
amount of DNAPL TCE has not been estimated, the rate
of TCE dissolution cannot be estimated. Three first-
order rate constants of TCE desorption from shallow
aquifer sediments at the arsenal calculated by Koller et
al. (8) ranged from 0.003 to 0.015 per week. The rate
constants were measured in laboratory flowthrough col-
umns using uncontaminated water as the influent fluid.
Desorption rates in the field, where ground water con-
taining TCE is flowing past the desorbing sediments,
probably would be lower. By using an estimated mass
of TCE sorbed to the aquifer sediments of three to four
times the mass of TCE in the dissolved state, the esti-
mated flux of TCE into the ground-water system through
desorption is 15 to 85 mg/s. This flux estimate is made
by assuming that, over long periods (years), the short-
term desorption rate (weeks and months) is equal to the
short-term adsorption rate, and that soils previously
have undergone long-term adsorption, which is no
longer occurring.
Preliminary estimates of the flux of TCE into and out of
the ground-water system at the Building 24 site for each
of the mass-balance components are shown in Figure 2.
The estimated flux of TCE discharged to Green Pond
Brook from the plume area, calculated on the basis of
measured TCE concentrations in ground water and
measured base-flow discharge in the brook, is 1 to 2 mg/s.
The flux of TCE volatilized from the water table is esti-
mated to be about 0.1 mg/s on the basis of measured
soil-gas TCE concentration gradients and estimates of
the physical characteristics of the unsaturated zone.
Biotransformation probably is the mechanism by which
most of the dissolved TCE leaves the ground-water
system. First-order rate constants for TCE transforma-
tion ranging from less than 0.001 to 0.02 per week were
estimated by Wilson et al. (9) on the basis of results of
laboratory microcosm studies of soil from five sites
within the plume area. By using these rate constants and
the estimated mass of dissolved TCE, the rate of TCE
loss from the plume through biotransformation is calcu-
lated to be about 1 to 30 mg/s. Analogous first-order rate
constants for TCE biotransformation calculated by Ehlke
et al. (10) from field-measured TCE concentrations and
time-of-travel data generally were higher than those
measured in the laboratory experiments. Thus, the ac-
tual flux of TCE lost through biotransformation may be
greater than that shown in Figure 2.
A reactive multispecies transport model of a two-dimen-
sional vertical section along the central axis of the plume
is being used to analyze the laboratory and field esti-
mates of the physical and chemical transport charac-
teristics and the estimated TCE mass balance. The
model design, calibration, and sensitivity analysis have
been described by Martin (11). Transport, desorption,
volatilization, and microbial degradation of TCE are
simulated. The formation and transport of the degrada-
tion products cis-DCE and VC also are simulated.
Specified-flux boundary conditions are used to repre-
sent ground-water recharge and flows across the hori-
zontal and vertical boundaries of the cross-sectional
area. Constant-concentration nodes and desorption
within the plume area are solute sources to the simulated
system. The constant-concentration solute sources are
assumed to represent high rates of desorption or
113
-------
(A) Trichloroethylene
200 100 0 100 200 300 400 500 600
Distance From Building 24 (B-24), in Meters
(B) cis-1-2-Dichloroethylene
Bfar Swalnp Sroo*
200 100 0 100 200 300 400 500 600
Distance From Building 24 (B-24), in Meters
Explanation
100 Line of Equal Solute Concentration
Shows solute concentration, in micrograms per liter.
0 Location of Well Screen
CAF-7 Location of Well and Local Identifier
Figure 5. Simulated concentrations from calibrated model with
simulated desorption, volatilization, and microbial
degradation: (A) TCE and (B) cis-DCE. (Location of
section A-A' is shown in Figure 1A.)
dissolution of TCE near the settling lagoons, at the
overflow dry well, and near the base of the unconfined
aquifer 230 m downgradient from Building 24.
The simulations were designed to represent average
steady-state flow conditions and virtually steady-state
transport conditions after 1985. Simulated concentra-
tions of TCE and cis-DCE from the calibrated model are
shown in Figure 5. Most concentrations were simulated
to within an order of magnitude of average concentra-
tions measured in micrograms per liter in water samples
from each well. Because the degradation of VC was not
simulated, concentrations are higher than measured
concentrations and are not shown.
The calibrated model does not provide a unique esti-
mate of the magnitudes of the various mass-balance
components of the plume of TCE-contaminated ground
water at the Building 24 site; however, sensitivity simu-
lations were used to test hypotheses concerning the fate
of TCE in ground water. Results of a series of sensitivity
simulations discussed by Martin (11) support the general
conceptual model as defined by the estimated solute
mass balance presented above.
Results of sensitivity simulations made with various
desorption and degradation rates typically showed that
use of the laboratory estimates resulted in reasonable
simulated concentrations. Although volatilization is not a
major mass-balance component, this process was
shown to be an important mechanism for removing sol-
utes and thereby affecting solute concentrations near
the water table. The overall flux of TCE into and out of
the system was not simulated well because the total
simulated mass of TCE in the system was too low.
Increasing the simulated solute-source area near the
base of the unconfined aquifer might result in a reason-
able simulated TCE mass balance.
Summary
Infiltration of wastewater from the lagoons and chlorin-
ated solvents from the dry well at Building 24 at Picat-
inny Arsenal, New Jersey, has created a plume of
contaminated ground water downgradient from the
building. TCE is the predominant contaminant in the
130,000-m2 plume, which extends 500 m to Green Pond
Brook. Ground-water velocities typically range from 0.3
to 1.0 m/d.
Results of water-quality sampling conducted from 1986
through 1991 show that the TCE contaminant plume has
changed little since September 1986, and that the high-
est TCE concentrations are found near the water table
near Building 24 and near the base of the unconfined
aquifer about midway between Building 24 and Green
Pond Brook. TCE is present at the site in several
phases: dissolved in water, as a vapor in the soil gas,
and sorbed onto solid surfaces or associated with biota.
TCE also may be present as a DNAPL. The total volume
of dissolved TCE below the water table is estimated to
be about 660 L, but most of the mass of TCE in the
system is associated with the sediments in the saturated
and unsaturated zones.
A conceptual model of the physical, chemical, and bio-
logical processes that affect the transport and mass
balance of TCE within the plume includes 1) transport
of TCE from the Building 24 source area to Green Pond
Brook by advection and dispersion; 2) loss of dissolved
TCE from the ground-water system by discharge to
Green Pond Brook, biotransformation, and volatilization;
and 3) gain of dissolved TCE by slow desorption from
contaminated aquifer sediments and possibly from
114
-------
dissolution of DNAPL TCE in the aquifer. Preliminary
estimates of the flux of dissolved TCE discharged to
Green Pond Brook from the plume is 1 to 2 mg/s. Most
dissolved TCE leaves the ground-water system by
means of biotransformation. The estimated flux of TCE
out of the system by this process may be about an order
of magnitude greater than the flux of TCE discharged to
Green Pond Brook. Although the estimated flux of TCE
out of the ground-water system by volatilization is esti-
mated to be about an order of magnitude less than the
flux of TCE discharge to Green Pond Brook, volatiliza-
tion is an important mechanism for removing solutes and
thereby affecting solute concentrations near the water
table.
References
1. 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. Descriptions of and
sampling plans for remedial investigation sites. Ar-
gonne, IL: Argonne National Laboratory, Environ-
mental Assessment and Information Sciences
Division, pp. 22-1 to 22-24.
2. Fusillo, TV., T.A. Ehlke, M. Martin, and B.P Sar-
gent. 1987. Movement and fate of chlorinated sol-
vents in ground water: Preliminary results and
future research plans. In: Franks, B.J., ed. Proceed-
ings of the U.S. Geological Survey Program on
Toxic Waste—Ground-Water Contamination, Pen-
sacola, FL (March 23-27). U.S. Geological Survey
Open File Rep. 87-109. pp. D5-D12.
3. Sargent, B.P, J.W Green, P.T. Harte, and E.F. Vow-
inkel. 1986. Ground-water-quality data for Picatinny
Arsenal, New Jersey, 1958-85. U.S. Geological
Survey Open File Rep. 86-58. 66 pp.
4. Imbrigiotta, T.E., M. Martin, B.P. Sargent, and L.M.
Voronin. 1989. Preliminary results of a study of the
chemistry of ground water at the Building 24 re-
search site, Picatinny Arsenal, New Jersey. In: Mal-
lard, G.E., and S.E. Ragone, eds. Proceedings of
the U.S. Geological Survey Toxic Substances Hydrol-
ogy Program, Phoenix, AZ (September 26-30,1988).
Water Res. Invest. Rep. 88-4220. pp. 351-359.
5. Imbrigiotta, T.E., T.A. Ehlke, and M. Martin. 1991.
Chemical evidence of processes affecting the fate
and transport of chlorinated solvents in ground
water at Picatinny Arsenal, New Jersey. In: Mallard,
G.E., and D.A. Aronson, eds. Proceedings of the
U.S. Geological Survey Toxic Substances Hydrol-
ogy Program, Monterey, CA (March 11-15). Water
Res. Invest. Rep. 91-4034. pp. 681-688.
6. Imbrigiotta, T.E., and M. Martin. 1991. Overview of
research activities on the movement and fate of
chlorinated solvents in ground water at Picatinny
Arsenal, New Jersey. In: Mallard, G.E., and D.A.
Aronson, eds. Proceedings of the U.S. Geological
Survey Toxic Substances Hydrology Program, Mon-
terey, CA (March 11-15). Water Res. Invest. Rep.
91-4034. pp. 673-680.
7. Smith, J.A., C.T Chiou, J.A. Kammer, and D.E. Kile.
1990. Effect of soil moisture on sorption of trichlo-
roethene vapor to vadose-zone soil at Picatinny
Arsenal, New Jersey. Environ. Sci. Technol.
24(5):676-683.
8. Koller, D., T.E. Imbrigiotta, A.L. Baer, and J.A.
Smith. 1994. Desorption of trichloroethylene from
aquifer sediments at Picatinny Arsenal, New Jersey.
In: Morganwalp, D.W, and D.A. Aronson, eds. Pro-
ceedings of the U.S. Geological Survey Toxic Sub-
stances Hydrology Program, Colorado Springs, CO
(September 20-24, 1993). Water Res. Invest. Rep.
94-4014. In press.
9. Wilson, B.H., T.A. Ehlke, T.E. Imbrigiotta, and J.T
Wilson. 1991. Reductive dechlorination of trichlo-
roethylene in anoxic aquifer material from Picatinny
Arsenal, New Jersey. In: Mallard, G.E., and D.A.
Aronson, eds. Proceedings of the U.S. Geological
Survey Toxic Substances Hydrology Program, Mon-
terey, CA (March 11-15). Water Res. Invest. Rep.
91-4034. pp. 704-707.
10. Ehlke, T.A., B.H. Wilson, J.T. Wilson, and T.E. Im-
brigiotta. 1994. In situ biotransformation of trichlo-
roethylene and cis-1,2-dichloroethylene at Picatinny
Arsenal, New Jersey. In: Morganwalp, D.W, and D.A.
Aronson, eds. Proceedings of the U.S. Geological
Survey Toxic Substances Hydrology Program, Colo-
rado Springs, CO (September 20-24, 1993). Water
Res. Invest. Rep. 94-4014. In press.
11. Martin, M. 1994. Simulation of transport, desorption,
volatilization, and microbial degradation of trichlo-
roethylene in ground water at Picatinny Arsenal,
New Jersey. In: Morganwalp, D.W, and D.A. Aron-
son, eds. Proceedings of the U.S. Geological
Survey Toxic Substances Hydrology Program,
Colorado Springs, CO (September 20-24, 1993).
Water Res. Invest. Rep. 94-4014. In press.
115
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Intrinsic Bioremediation of TCE in Ground Water at an NPL Site in
St. Joseph, Michigan
John T. Wilson, James W. Weaver, and Don H. Kampbell
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK
Introduction
The ground water at the St. Joseph, Michigan, National
Priority List (NPL) site is contaminated with chlorinated
aliphatic compounds (CACs) at concentrations in the
range of 10 mg/L to 100 mg/L. The chemicals are
thought to have entered the shallow sandy aquifer either
through waste lagoons that were used from 1968 to
1976 or through disposal of trichloroethene (TCE) into
dry wells at the site (1). The contamination was deter-
mined to be divided into eastern and western plumes,
as the suspected sources were situated over a ground-
water divide. Both plumes were found to contain TCE,
cis- and trans-1,2-dichoroethene (c-DCE and t-DCE),
1,1-dichloroethene (1,1-DCE), and vinyl chloride (VC).
Previous investigation of the site indicated that natural
anaerobic degradation of the TCE was occurring, be-
cause transformation products and significant levels of
ethene and methane (2,3) were present. The purpose of
this presentation is to present the results of later sam-
pling of the western plume near Lake Michigan, to esti-
mate the contaminant mass flux, and to estimate
apparent degradation constants. The estimates are
based on visualization of the data that represent each
measured concentration by a zone of influence that is
based on the sample spacing. The presentation of the data
is free from artifacts of interpolation, and extrapolation of
the data beyond the measurement locations is controlled.
Data Summary
In 1991, three transects (1, 2, and 3 on Figure 1) were
completed near the source of the western plume (2).
The three transects consisted of 17 borings with a slot-
ted auger. In 1992, two additional transects (4 and 5 on
Figure 1) were completed, consisting of 9 additional
slotted auger borings. In each boring, water samples
were taken at roughly 1.5 m (5 ft) depth intervals. Onsite
gas chromatography was performed to determine the
width of the plume and find the point of highest concen-
tration. Three of the transects (2, 4, and 5) are roughly
perpendicular to the contaminant plume. Of the remain-
ing transects, transect 1 crosses the plume at an angle
and transect 3 lies along the length of the plume. The
perpendicular transects form logical units for study of
TCE biotransformation.
The site data from the transects are visualized as sets
of blocks that are centered around the measurement
point. The blocks are defined so that the influence of a
particular measured concentration extends halfway to
the next measurement location both horizontally and
vertically. Thus, presentation of the data is simple and
direct. The visualization of the data is performed on a
Silicon Graphics Indigo workstation using a two-dimen-
sional version of the fully three-dimensional field-data
analysis program called SITE-3D that is under develop-
ment at the Robert S. Kerr Environmental Research
Laboratory.
The mass of each chemical per unit thickness and the
advective mass flux of each chemical are calculated by
summing over the blocks. By following this procedure,
the measured chemical concentrations are not extrapo-
lated into the clay layer under the site, nor are they
extrapolated beyond a short distance from the measure-
ment locations (5 ft vertically and 50 ft to 100 ft horizon-
tally). Other interpolation schemes such as inverse
distance weighting or kriging could also be used to
estimate the concentration field and perform the mass
estimates. Figures 2 and 3 show the distributions of the
VC and TCE at transect 5 using a logarithmic, black-
and-white "color" scale. Notably, the maximum VC con-
centration at transect 4 was 1,660 |j/L and at transect 5
was 205 |j,g/L. The maximum TCE concentration at tran-
sect 4 was 8,720 |j,g/L and at transect 5 was 163 |j,g/L.
As noted previously for other portions of the site (2,4),
the contamination is found near the bottom of the aqui-
fer. The highest concentrations of VC and TCE do not
appear to be co-located. In Table 1, mass estimates are
presented for the perpendicular transects ordered from
116
-------
Tram
Legend
Slotted Auger Boring
Ground-water Elevation
8/23/87
Figure 1. Site plan, St. Joseph, Michigan, NPL site.
St. Joseph, Michigan
Vinyl Chloride
Transect: 5
Mass: 0.4811E-01 kg/m
Concentration
151 152
154 153 155
Ground Surface
10 Feet
> Approx. N
Figure 2. VC distribution at transect 5.
100 Feet
250,000. 1
25,000.1
2,500.1
250.0 1
25.00 1
2.500
0.2500
0.0250 1
St. Joseph, Michigan
Trichloroethene
Transect: 5
Mass: 0.2821 E-01 kg/m
151 152 154 153 155
Ground Surface
10 Feet J
100 Feet
fr Approx. N | |
Figure 3. TCE distribution at transect 5.
Concentration
ng/L
250,000.
25,000.
2,500.
250.0
25.00
2.500
0.2500
0.0250 I
farthest upgradient (transect 2) to farthest downgradient
(transect 5). The data in Table 1 represent the mass in
a volume of aquifer that has an area equal to the cross-
sectional area of the transect and is 1.0 m thick in the
direction of ground-water flow.
Advective Mass Flux Estimates
Results from the calibrated MODFLOW model of Tiede-
man and Gorelick (4) were used to estimate the ground-
water flow velocity at each transect. The estimate is an
upper bound because the modeled vertical component
of flow was neglected in the present analysis. The head
drop from one location to the next was assumed to
generate horizontal flow only. Tiedeman and Gorelick (4)
also represented the aquifer by single values of hydraulic
conductivity and porosity. They gave, however, 95 percent
confidence limits for the hydraulic conductivity. Well
yields estimated for each sample location indicate declin-
ing hydraulic conductivity toward the west (i.e., towards
Lake Michigan and transects 4 and 5). Thus, using the
single parameter values from the MODFLOW simulations
may overestimate the flux of water into the lake.
As would be expected, the advective mass fluxes de-
cline toward the downgradient edge of the plume (Table
2). There the concentrations are lower, due to either
transient flow or degradation of the TCE. Notably the
mass fluxes using the average hydraulic conductivity
result in a total flux of 13 kg/y of TCE, c-DCE, t-DCE,
1,1-DCE, and VC at transect 5. This value contrasts with
the total flux of these CACs of 310 kg/y at transect 2
117
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Table 1. Mass per Unit Thickness (kg/m) at St. Joseph, Michigan
Chemical
Vinyl chloride
1,1-DCE
t-DCE
c-DCE
TCE
Methane
Ethene
Ethane
TOC
Chloride
Sulfate
NO3-Nitrogen
NH4-Nitrogen
TKN-Nitrogen
Table 2. Mass Flux
Chemical
Vinyl chloride
1,1-DCE
t-DCE
c-DCE
TCE
Methane
Ethene
Ethane
TOC
Chloride
Sulfate
NO3-Nitrogen
NH4-Nitrogen
TKN-Nitrogen
2
0.457
0.0713
0.170
3.70
3.20
1.76
0.205
No data
No data
39.0
11.1
0.871
0.551
0.896
(kg/y) at St Joseph, Michigan
2
18.81
2.934
6.995
152.1
131.7
72.29
8.453
No data
No data
1,604
457.4
35.85
22.66
36.88
1
0.569
0.0245
0.152
1.53
1.67
1.64
0.268
No data
No data
44.6
10.3
0.741
0.768
1.15
1
36.03
1.551
9.609
97.11
106.0
104.1
16.95
No data
No data
2,826
652.9
46.93
48.64
72.85
Transect
4
0.146
0.00435
0.0109
0.567
0.419
1.39
0.0524
0.0626
3.79
63.9
28.7
1.33
0.137
0.191
Transect
4
10.69
0.3185
0.7963
41.48
30.67
101.4
3.836
4.577
277.2
4,678
2,102
97.05
10.01
13.95
5
0.0144
0.00031
0.00211
0.0850
0.00846
0.412
0.00147
0.00051
2.49
46.9
19.9
2.47
0.0677
0.109
5
1.676
0.03648
0.2453
9.868
0.9829
47.86
0.1708
0.05885
289.7
5,444
2,306
287.4
7.861
12.70
near the source of contamination. Thus, there is a 24.4-
fold decrease in mass flux of CACs across the site.
Using the 95 percent confidence limits on the hydraulic
conductivity determined by Tiedeman and Gorelick (4),
the range total of mass flux of these five chemicals
ranges from 205 kg/y to 420 kg/y at transect 2 and from
8.4 kg/y to 17 kg/y at transect 5. The range effluxes at
transect 5 is an upper bound on, and best estimate of,
the flux into Lake Michigan.
Apparent Degradation Constants
The mass per unit thickness of TCE at transects 2, 4,
and 5 was used to estimate apparent first-order degra-
dation constants. The constants are estimated by apply-
ing the first-order rate equation
\=kAt
(1)
118
-------
Table 3. Chemical and Hydraulic Values Used in Estimating Apparant Degradation Rates
Transect
2
Area With
Nonzero TCE
Concentration
(m2)
1,592
Mass Per
Unit
Thickness
From
SITE-3D
(kg/m)
10.67
Average TCE
Concentration
in the
Transect
(kg/m3)
Cj and Cj+i in
Equation 1
6.70e-3
Distance
Between
Transects
(m)
Gradient
Estimated
From
Tiedeman
and
Gorelick
(1993)
Retarded
Seepage
Velocity
for TCEa
(m/d)
Estimated
Travel
Time
Between
Transects
(weeks)
At in
Equation 1
2,774
1,943
1.397
0.0282
5.04e-4
1.44e-5
260
160
7.3e-3
0.11
0.156
340
145
Constants used in seepage velocity calculation:
Hydraulic conductivity: 7.5 m/d
Retardation factor for TCE: 1.78 = 1 + KoCfoc Pb/6
Porosity, 6: 0.30
Bulk density pb: 1.86 g/cm3
KOC: 126 mL/g, foc: 0.001
to the site data, where q is the average concentration in
the transect j, cj+1 is the average concentration in the
downgradient transect j+1, At is the advective travel time
for TCE to move between the transects, and A, is the
apparent degradation constant. The mass per unit thick-
ness data for TCE and the cross-sectional area were
used to determine the average concentrations q and cj+1
in the up- and downgradient transects. The porosity,
bulk density, fraction organic carbon, organic carbon
partition coefficient (5), ground-water gradient, and dis-
tance between the transects were used to determine the
advective travel times. The values used in equation 1
are given in Table 3. From these quantities, the apparent
degradation constant for TCE was determined to be
-0.0076/wk from transect 2 to 4 and -0.024/wk from
transect 4 to 5.
References
1. Engineering Science, Inc. 1990. Remedial investiga-
tion and feasibility study, St. Joseph, Ml, phase I
technical memorandum. Liverpool, NY.
2. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T
Wilson. 1993. Natural anaerobic bioremediation of
TCE at the St. Joseph, Michigan, Superfund site. In:
U.S. EPA. Symposium on bioremediation of hazard-
ous wastes: Research, development, and field evalu-
ations. EPA/600/R-93/054. Washington, DC (May).
pp. 57-60.
3. McCarty, PL., and J.T. Wilson. 1992. Natural anaero-
bic treatment of a TCE plume, St. Joseph, Michigan,
NPLsite. 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 ground-water contaminant
capture design. Water Resour. Res. 29(7):2,139-
2,153.
5. U.S. EPA. 1990. Subsurface remediation guidance
tables. EPA/540/2-90/011b.
119
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Posters
121
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Technical Protocol for Implementing the Intrinsic Remediation With
Long-Term Monitoring Option for Natural Attenuation of
Fuel-Hydrocarbon Contamination in Ground Water
Todd H. Wiedemeier
Engineering-Science, Inc., Denver, CO
John T. Wilson and Donald H. Kampbell
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory, Ada, OK
Ross N. Miller and Jerry E. Hansen
U.S. Air Force Center for Environmental Excellence, Technology Transfer Division, Brooks AFB, TX
This paper presents a brief overview of the technical
protocol, currently under development by the U.S. Air
Force Center for Environmental Excellence (AFCEE),
Technology Transfer Division, for data collection,
ground-water modeling, and exposure assessment in
support of intrinsic remediation (natural attenuation) with
long-term monitoring for restoration of fuel-hydrocarbon
contaminated ground water (1). The material presented
herein was prepared through the joint effort of AFCEE,
the Bioremediation Research Team at the U.S. Environ-
mental Protection Agency's Robert S. Kerr Environ-
mental Research Laboratory in Ada, Oklahoma, and
Engineering-Science, Inc., to facilitate implementation
of intrinsic remediation at fuel-hydrocarbon contami-
nated sites. Specifically, this protocol is designed to
evaluate the fate of dissolved-phase fuel hydrocarbons
having regulatory maximum contaminant levels (MCLs).
The intended audience for this document is U.S. Air
Force (USAF) personnel, scientists, consultants, regu-
latory personnel, and others charged with remediating
fuel-hydrocarbon contaminated ground water at USAF
facilities.
Intrinsic remediation is achieved when naturally occur-
ring attenuation mechanisms, such as biodegradation
(aerobic and anaerobic), bring about a reduction in the
total mass of a contaminant dissolved in ground water.
During intrinsic remediation, contaminants are ultimately
transformed to innocuous byproducts (e.g., carbon diox-
ide and water), not just transferred to another phase or
location within the environment. Intrinsic remediation
results from the integration of several subsurface at-
tenuation mechanisms that are classified as either de-
structive or nondestructive. Destructive processes in-
clude biodegradation, abiotic oxidation, and hydrolysis.
Nondestructive attenuation mechanisms include sorp-
tion, dilution (caused by dispersion and infiltration), and
volatilization.
In some cases, intrinsic remediation reduces dissolved-
phase contaminant concentrations to below MCLs be-
fore the contaminant plume reaches potential receptors,
even if little or no source removal or reduction takes
place. In situations where intrinsic remediation will not
reduce contaminant concentrations to below regulatory
MCLs, in an acceptable time frame, less stringent
cleanup goals may be implemented. This is especially
likely if it can be demonstrated that intrinsic remediation
will result in a continual reduction in contaminant con-
centrations over time such that calculated risk values
are reduced.
Intrinsic remediation is gaining regulatory acceptance
and has been implemented at several sites over the past
few years (2-5). In addition to bringing about complete
mineralization of contaminants, intrinsic remediation is
nonintrusive and allows continuing use of infrastructure
during remediation. The main limitation is that intrinsic
remediation is subject to natural and institutionally in-
duced changes in local hydrogeologic conditions. In
addition, aquifer heterogeneity may complicate site
characterization as it will with any remedial technology.
Evaluating the effectiveness of intrinsic remediation re-
quires the quantification of ground-water flow and solute
transport and transformation processes, including rates
of natural attenuation. Quantification of contaminant
122
-------
migration and attenuation rates, and successful imple-
mentation of the intrinsic remediation option, requires
completion of the following steps, each of which is dis-
cussed in the following sections and outlined in Figure 1:
1. Review existing site data.
2. Develop a preliminary conceptual model for the site,
and assess the potential significance of intrinsic re-
mediation.
3. Perform site characterization in support of intrinsic
remediation.
4. Refine the conceptual model based on site charac-
terization data, complete premodeling calculations,
and document indicators of intrinsic remediation.
5. Model intrinsic remediation using numerical fate and
transport models that allow incorporation of a biode-
gradation term (e.g., BIOPLUME II or BIOPLUME III).
6. Conduct an exposure assessment.
7. Prepare a long-term monitoring plan and site long-term
monitoring wells and point-of-compliance monitoring
wells.
8. Present findings to regulatory agencies, and obtain
approval for the intrinsic remediation with long-term
monitoring option.
Collection of an adequate database during the iterative
site characterization process is an important step in the
documentation of intrinsic remediation. At a minimum,
the site characterization phase should provide data on
the location and extent of contaminant sources; data on
the location, extent, and concentration of dissolved-
phase contamination; ground-water geochemical data;
geologic data on the type and distribution of subsurface
materials; and hydrogeologic parameters such as hy-
draulic conductivity, hydraulic gradients, and potential
contaminant migration pathways to human or ecological
receptors. Contaminant sources include nonaqueous-
phase liquid (NAPL) hydrocarbons present as mobile
NAPL (NAPL occurring at sufficiently high saturations to
drain, underthe influence of gravity, to a well) or residual
NAPL (NAPL occurring at immobile residual saturations
that are unable to drain to a well by gravity).
The following analytical protocol should be used for
analysis of soil and ground-water samples. This analyti-
cal protocol includes all of the parameters necessary to
document intrinsic remediation of fuel hydrocarbons,
including the effects of sorption and biodegradation
(aerobic and anaerobic) of fuel hydrocarbons. Soil sam-
ples should be analyzed for total volatile and extractable
hydrocarbons, aromatic hydrocarbons, and total organic
carbon. Ground-water samples should be analyzed for
dissolved oxygen, oxidation-reduction potential, pH,
temperature, conductivity, alkalinity, nitrate, sulfate, sul-
fide, ferrous iron, carbon dioxide, methane, chloride,
total petroleum hydrocarbons, and aromatic hydrocar-
bons. The extent and distribution (vertical and horizon-
tal) of contamination and electron acceptor and
metabolic byproduct concentrations and distributions
are of paramount importance in documenting the occur-
rence of biodegradation of fuel hydrocarbons and in
numerical model implementation. Dissolved oxygen
concentrations below background in an area with fuel-
hydrocarbon contamination are indicative of aerobic hy-
drocarbon biodegradation. Similarly, nitrate and sulfate
concentrations below background in an area with fuel-
hydrocarbon contamination are indicative of anaerobic
biodegradation through denitrification and sulfanogene-
sis. Elevated concentrations of metabolic byproducts
such as ferrous iron and methane in areas with fuel
hydrocarbon contamination are indicative of hydrocar-
bon biodegradation by the processes of ferric hydroxide
reduction and methanogenesis. Contour maps can be
used to provide visible evidence of these relationships.
To support implementation of intrinsic remediation, the
property owner must scientifically demonstrate that deg-
radation of site contaminants is occurring at rates suffi-
cient to be protective of human health and the environ-
ment. Three lines of evidence can be used to support
intrinsic remediation: 1) documented loss of contami-
nants at the field scale, 2) the use of chemical analytical
data in mass balance calculations of microbial metabo-
lism, and 3) laboratory microcosm studies using aquifer
samples collected from the site.
The first line of evidence involves using measured dis-
solved-phase concentrations of biologically recalcitrant
tracers found in fuels in conjunction with aquifer hydro-
geologic parameters such as seepage velocity and dilu-
tion to show that a reduction in the total mass of
contaminants is occurring at the site. The second line of
evidence involves the use of chemical analytical data in
mass balance calculations to show that a decrease in
contaminant and electron acceptor concentrations can
be directly correlated to increases in metabolic bypro-
duct concentrations. This evidence can be used to show
that electron acceptor concentrations are sufficient to
degrade dissolved-phase contaminants. Numerical
models can be used to aid mass-balance calculations
and to collate information on degradation. The third line
of evidence, the microcosm study, involves studying site
aquifer materials under controlled conditions in the labo-
ratory to show that indigenous biota are capable of
degrading site contaminants and to confirm rates of
contaminant degradation measured at the field scale.
The primary objective of the intrinsic remediation inves-
tigation is to determine if natural processes of degrada-
tion will reduce contaminant concentrations in ground
water to below regulatory standards before potential
exposure pathways are completed. This requires that a
projection of the potential extent and concentration of
123
-------
Review Available
Site Data
Develop Preliminary
Conceptual Model
Make Preliminary
Assessment of Potential
for Intrinsic Remediation
Based on Existing Site
Characterization Data
- Contaminant Type
and Distribution
- Hydrogeology
- Location of Receptors
Perform Site Characterization
in Support of Intrinsic Remediation
Refine Conceptual Model and
Complete Premodeling
Calculations
Document Occurrence of
Intrinsic Remediation and
Model Intrinsic Remediation
Using Numerical Models
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
Evaluate Use of
Other Remedial
Options in
Conjunction With
Intrinsic Remediation
Assess Potential for
Intrinsic Remediation
With Remediation
System Installed
Refine Conceptual Model and
Complete Premodeling
Calculations
Model Intrinsic Remediation
Combined With Remedial
Option Selected Above
Using Numerical Models
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
Yes
Is There
< Unacceptable
Risk to Potential
,> Receptors
No ?.
Site Point-of-Compliance
Monitoring Wells and
Prepare Long-Term
Monitoring Plan
Present Findings
and Long-Term
Monitoring Plan to
Regulatory Agencies
and Reach Agreement
on Monitoring Strategy
Figure 1. Intrinsic remediation flow chart.
124
-------
the contaminant plume in time and space is made based
on governing physical, chemical, and biological proc-
esses. This projection should be based on historic vari-
ations in—and the current extent and concentration
of—the contaminant plume, as well as on the measured
rates of contaminant attenuation.
The data collected during site characterization can be
used to model the fate and transport of contaminants in
the subsurface. Such modeling allows an estimate of the
future extent and concentration of the dissolved-phase
plume to be made. Several models, including
BIOPLUME II (6), have been used successfully to model
dissolved-phase contaminant transport and attenuation.
Additionally, a new version of the BIOPLUME model,
BIOPLUME III, is under development by AFCEE. The
intrinsic remediation modeling effort has three primary
objectives: 1) to estimate the future extent and concen-
tration of a dissolved-phase contaminant plume by mod-
eling the combined effects of advection, dispersion,
sorption, and biodegradation; 2) to assess the possible
risk to potential downgradient receptors; and 3) to pro-
vide technical support for the natural attenuation reme-
dial option at postmodeling regulatory negotiations.
Microorganisms generally utilize dissolved oxygen and
nitrate in areas with dissolved-phase fuel-hydrocarbon
contamination at rates that are instantaneous relative to
the average advective transport velocity of ground
water. This results in the consumption of these com-
pounds at a rate approximately equal to the rate at which
they are replenished by advective flow processes. For
this reason, the use of these compounds as electron
acceptors in the biodegradation of dissolved-phase fuel
hydrocarbons is a mass-transport-limited process (7,8).
The use of dissolved oxygen and nitrate in the biodegra-
dation of dissolved-phase fuel hydrocarbons can be
modeled using BIOPLUME II. Microorganisms generally
utilize sulfate, iron III, and carbon dioxide (used during
methanogenesis) in areas with dissolved-phase fuel-hy-
drocarbon contamination at rates that are slow relative
to the advective transport velocity of ground water. This
results in the consumption of these compounds at a rate
slower than the rate at which they are replenished by
advective flow processes. Therefore, the use of these
compounds as electron acceptors in the biodegradation
of dissolved-phase fuel hydrocarbons is a reaction-lim-
ited process that can be approximated by first-order
kinetics. The BIOPLUME II model utilizes a first-order
rate constant to model such biodegradation. First-order
decay constants can be determined by simple calcula-
tions based on ground-water chemistry or through the
use of laboratory microcosm studies. In addition, the use
of radiolabeled materials in a microcosm study can be
used to provide evidence of the ultimate fate of the
contaminants.
The results of the modeling effort are not in themselves
sufficient proof that intrinsic remediation is occurring at
a given site. The results of the modeling effort are only
as good as the original data input into the model and the
model itself. Because of the inherent uncertainty asso-
ciated with such predictions, it is the responsibility of the
proponent to provide sufficient evidence to demonstrate
that the mechanisms of intrinsic remediation will reduce
contaminant concentrations to acceptable levels before
potential receptors are reached. This requires the use of
conservative input parameters and numerous sensitivity
analyses so that consideration is given to all plausible
contaminant migration scenarios. When possible, both
historical data and modeling should be used to provide
information that collectively and consistently supports
the natural reduction and removal of the dissolved-
phase contaminant plume. In some cases, simple calcu-
lations of contaminant attenuation rates are all that are
required to successfully support intrinsic remediation.
Upon completion of the fate and transport modeling
effort, model predictions can be used in an exposure
assessment. If intrinsic remediation is sufficiently active
to mitigate risks to potential receptors, the proponent of
intrinsic remediation has a reasonable basis for negoti-
ating this option with regulators. The exposure assess-
ment allows the proponent to show that potential
exposure pathways will not be completed.
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 moni-
tor plume migration over time and to verify that intrinsic
remediation is occurring at rates sufficient to protect
potential downgradient receptors. The long-term moni-
toring plan should be developed based on the results of
a numerical model such as BIOPLUME II.
Point-of-compliance (POC) monitoring wells are wells
that are installed at locations downgradient of the con-
taminant plume and upgradient of potential receptors.
POC monitoring wells are generally installed along a
property boundary or at a location approximately 5 yr
downgradient of the current plume at the seepage ve-
locity of the ground water of 1 yr to 2 yr upgradient of
the nearest downgradient receptor, whichever is more
protective. The final number and location of POC moni-
toring wells depends on regulatory considerations.
Long-term monitoring wells are wells that are placed
upgradient of, within, and immediately downgradient of
the contaminant plume. These wells are used to monitor
the effectiveness of intrinsic remediation in reducing the
total mass of contaminant within the plume. The final
number and location of long-term monitoring wells de-
pends on regulatory considerations. Figure 2 shows a
hypothetical long-term monitoring scenario. The results
of a numerical model such as BIOPLUME II can be used
125
-------
Direction of
Plume Migration
Anaerobic
Treatment
Zone Extent of Dissolved-
Phase BTEX Plume *
Aerobic
Treatment
Zone
Legend
K Point-of-Compliance Monitoring Well
o Long-Term Monitoring Well
Not to Scale
Figure 2. Hypothetical long-term monitoring strategy.
to help site both the long-term and POC monitoring
wells.
References
1. Wiedemeier, T.H., D.C. Downey, J.T. Wilson, D.H.
Kampbell, R.N. Miller, and J.E. Hansen. 1994. Draft
technical protocol for implementing the intrinsic re-
mediation with long-term monitoring option for natu-
ral attenuation of dissolved-phase fuel contamination
in ground water. U.S. Air Force Center for Environ-
mental Excellence, Technology Transfer Division.
2. Klecka, G.M., J.W. Davis, D.R. Gray, and S.S. Mad-
sen. 1990. Natural bioremediation of organic con-
taminants in ground water—Cliffs-Dow Superfund
site. Ground Water 28(4):534-543.
3. Downey, D.C., and M.J. Gier. 1991. Supporting the
no-action alternative at a hydrocarbon spill site. In:
Proceedings of the USAF Environmental Restoration
Technology Symposium, San Antonio, TX (May 7-8).
Section U. pp. 1-11.
4. Wiedemeier, T.H., PR. Guest, R.L. Henry, and C.B.
Keith. 1993. The use of Bioplume to support regula-
tory negotiations at a fuel spill site near Denver,
Colorado. In: Proceedings of the Petroleum Hydro-
carbons and Organic Chemicals in Ground Water:
Prevention, Detection, and Restoration Conference.
NWWA/API. pp. 445-459.
5. Wiedemeier, T.H., B. Blicker, and PR. Guest. 1994.
Risk-based approach to bioremediation of fuel hy-
drocarbons at a major airport. In: Proceedings of the
1994 Federal Environmental Restoration III and
Waste Minimization Conference and Exhibition, New
Orleans, LA. Hazardous Materials Control Re-
sources Institute, pp. 51-60.
6. Rifai, H.S., PB. Bedient, J.T. Wilson, K.M. Miller, and
J.M. Armstrong. 1988. Biodegradation modeling at
aviation fuel spill site. J. Environ. Eng. 114(5):1,007-
1,029.
7. Borden, R.C., and PB. Bedient. 1986. Transport of
dissolved hydrocarbons influenced by oxygen limited
biodegradation—theoretical development. Water
Resour. Res. 22(13):1,973-1,982.
8. Wilson, J.T, J.F. McNabb, J.W. Cochran, T.H. Want,
M.B. Tomson, and PB. Bedient. 1985. Influence of
microbial adaptation on the fate of organic pollutants
in ground water. Environ. Toxicol. Chem. 4:721-726.
126
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Wisconsin's Guidance on Naturally Occurring Biodegradation as a Remedial
Action Option
Michael J. Barden
Wisconsin Department of Natural Resources, Emergency and Remedial Response Section, Madison, Wl
In February 1993, the Wisconsin Department of Natural
Resources issued an interim guidance on naturally oc-
curring biodegradation as a remedial action option for
contaminated sites. The focus of this guidance was
primarily on soil contamination by petroleum hydrocar-
bons and on the requirements for site characterization
and monitoring necessary to use this approach. Sub-
sequent implementation of the interim guidance has
resulted in further refinement to make this an effective
approach for both soil and ground-water contamination.
The application of naturally occurring biodegradation as
a remedial action requires that the site be adequately
characterized and that an adequate monitoring program
be developed and implemented. This is a long-term
remediation option, likely requiring years or decades to
effect adequate cleanup. From a regulatory perspective,
the primary concerns are that the site conditions are
amenable to naturally occurring biodegradation and that
the process is effective in reducing contaminant concen-
trations to acceptable levels within a reasonable period
with respect to potential contaminant migration or im-
pacts to receptors.
Adequate site characterization is required during the site
investigation so that naturally occurring biodegradation
can be evaluated with other possible remedial action
options. This also provides baseline information for the
potential application of enhanced bioremediation as
well, because the basic site characterization require-
ments are essentially the same. Characterization in-
volves identification of 1) the contaminants present and
their concentrations and biodegradability, 2) physical
and chemical parameters affecting availability of oxygen
and alternative electron acceptors, 3) nutrients, and 4)
microbiological parameters indicating the presence and
viability of appropriate microbial populations. A sufficient
number of samples should be used to represent the
extent of contamination and site heterogeneity. The
guidance provides a framework for interpretation and
evaluation of the results.
If site conditions are favorable, a monitoring plan must
be developed and implemented. Monitoring indicates
that contaminant concentrations are decreasing over
time, ensures that no unexpected contaminant migration
is occurring, and provides information regarding the na-
ture and rate of biodegradation at the site. A variety of
monitoring approaches and techniques are available for
soil and ground water. In general, monitoring changes
in contaminant concentrations and/or concentrations of
co-reactants are appropriate.
Experience with implementation indicates that many re-
sponsible parties are unlikely to select this option due to
the long time frame involved. This suggests that natu-
rally occurring biodegradation is more viable as an op-
tion for stable entities where time is not an issue. The
availability of the guidance, however, has encouraged
consideration of bioremediation in general as a viable
remedy due to perceived regulatory acceptance of the
technology and because the required consideration of
biodegradation potential provides baseline site informa-
tion that can be used in the evaluation and design of
enhanced bioremediation systems.
127
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Assessing the Efficiency of Intrinsic Bioremediation
Francis H. Chapelle
U.S. Geological Survey, Columbia, SC
Abstract
The efficiency of intrinsic bioremediation to contain con-
taminant migration in ground-water systems can be
quantitatively assessed by comparing rates of contami-
nant transport with rates of biodegradation. If transport
rates are fast relative to rates of biodegradation, con-
taminants can migrate freely with ground-water flow and
possibly reach a point of contact with human or wildlife
populations. Conversely, if transport rates are slow rela-
tive to biodegradation rates, contaminant migration will
be more confined and less likely to reach a point of
contact. In either case, the efficiency of intrinsic biore-
mediation can be assessed by evaluating the presence
or absence of contaminant transport to predetermined
points of contact. Thus, this assessment includes hydro-
logic (rates of ground-water flow), microbiologic (rates
of biodegradation), and sociopolitical (points of contact)
considerations.
The U.S. Geological Survey, in cooperation with the
Naval Facilities Engineering Command, has developed
a framework for assessing the efficiency of intrinsic
bioremediation that is based on these three considera-
tions. In this framework, hydrologic and microbiologic
information is synthesized using a solute-transport code
(SUTRA) and used to estimate rates of contaminant
transport to predefined points of contact (adjacent water
supply wells or surface water bodies). This framework
is applied to two sites, in Beaufort and Hanahan, South
Carolina, contaminated with aviation fuel. At the
Beaufort site, rates of biodegradation are slow due to
anaerobic conditions (Kbio -0.01 d'1), but because rates
of ground-water flow are low (~0.02 ft/d), soluble con-
taminants are effectively contained and are not trans-
ported to adjacent points of contact. At the Hanahan site,
biodegradation rates are similarly slow under the ambi-
ent anaerobic conditions (Kbio ~0.01 d"1), but because
rates of ground-water flow are relatively high (~1.0 ft/d),
contaminants are transported to multiple points of con-
tact with humans. These examples illustrate the com-
plex interplay that develops between hydrologic,
microbiologic, and sociopolitical considerations, and
show that the efficiency of intrinsic bioremediation can
only be assessed on a site-by-site basis.
128
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A Practical Approach to Evaluating Natural Attenuation of Contaminants in
Ground Water
Paul M. McAllister and Chen Y. Chiang
Shell Development Company, Houston, TX
The extent of natural attenuation is an important consid-
eration in determining the most appropriate corrective
action at sites where ground-water quality has been
affected by releases of petroleum hydrocarbons or other
chemicals. The objective of this presentation is to pro-
vide guidelines for evaluating natural attenuation based
on easily obtainable field and laboratory data.
The primary indicators that can be used to evaluate
natural attenuation include dissolved oxygen (DO) lev-
els in ground water and plume characteristics. Back-
ground DO levels greater than 1 to 2 mg/L and inverse
correlation between DO and soluble hydrocarbon con-
centration have been identified through laboratory and
field studies as key indicators of aerobic biodegradation.
Several unique plume characteristics include 1) plumes
migrate more slowly than expected; 2) plumes reach a
steady state; and 3) plumes decrease in extent and
concentration, which may indicate the effects of natural
attenuation.
When DO is depleted in an aquifer, anaerobic conditions
prevail. For biodegradation to occur, an alternative electron
acceptor such as nitrate, carbonate, or iron III must be
available. Between aerobic and anaerobic conditions (i.e.,
0.1 ppm to 2 ppm), there is a region sometimes labeled the
hypoxic zone. Studies in the hypoxic zone have indicated
that biodegradation of benzene, toluene, ethylbenzene, and
the xylenes (BTEX) may occur at relatively low DO levels
provided a secondary electron acceptor is available.
Other secondary indicators (e.g., geochemical data) and
more intensive methods (e.g., contaminant mass bal-
ances, laboratory microcosm studies, and detailed
ground-water modeling) can be applied to demonstrate
natural attenuation as well. The recommended ap-
proach for evaluating natural attenuation is to design site
assessment activities so that required data such as DO
levels and historical plume flow path concentrations are
obtained. With the necessary data, the primary indica-
tors should be applied to evaluate natural attenuation.
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The Use of Low Level Activities To Assist intrinsic Bioremediation
Robert D. Morris
Eckenfelder Inc., Nashville, TN
Jeffrey C. Dey
Resource Control Corporation, Rancogas, NJ
Daniel P. Shine
Sun Company, Inc., Aston, PA
Intrinsic bioremediation as discussed in the recent re-
port of the National Academy of Sciences Committee on
Bioremediation (1) can reduce the concentrations of
some common contaminants to levels generally consid-
ered protective of human health. Since the observations
of the role of biodegradation in limiting the extent of the
ground-water plume at the Conroe, Texas, wood-pre-
serving site (2), many other sites have been observed
to have undergone natural attenuation at sufficient rates
to limit the size of contaminant plumes; in several in-
stances, contaminant levels decreased to below cleanup
levels.
The use of intrinsic bioremediation, while generally at-
tractive from a cost perspective, may actually be desir-
able environmentally because secondary effects of
active remediation are avoided. Intrinsic bioremediation,
however, has several costs associated with its use.
These include some level of documentation that biologi-
cal degradation of constituents of concern is taking
place and costs associated with monitoring and man-
agement of the site. Managing a small plume, such as
is frequently found at service stations, includes sampling
and reporting to the responsible environmental agency.
At one typical site (3), the annual environmental man-
agement costs have been approximately $11,900 per
year, consisting of site visits and sampling ($1,000),
chemical analysis ($7,200), reporting ($2,400), and con-
sulting ($1,300). At other sites, costs have exceeded
these values by 50 percent or more. Documenting that
biodegradation is occurring adds substantially to these
costs.
Intrinsic bioremediation will be effective where the elec-
tron acceptor requirements are relatively small. While
oxygen may reach the affected zones at rates sufficient
to prevent and shrink contaminant plumes, and thus
eventually achieve remedial goals, the time frame may
be unacceptably long from the site owner's perspective
because of long-term monitoring costs and manage-
ment burdens.
Addition of appropriate electron acceptors would accel-
erate reduction in constituents such as monoaromatic
hydrocarbons. In some cases where intrinsic bioreme-
diation is technically feasible, it may not be the most
cost-effective approach. To evaluate the concept of ap-
plying limited engineering solutions at sites where intrin-
sic bioremediation appears to be slowly reducing the
contaminant mass, we are testing the use of air sparging
wells (3). The air sparging wells are placed immediately
outside the plumes and operated intermittently at low
flow. The cost of installing three shallow air sparging
wells, routine maintenance, limited additional sampling,
and reporting was budgeted at $8,500 per site for each
of the three sites. If the time to reach closure is short-
ened by 1 yr or more, the cost of treatment will have
been less than the cost that would have been incurred
by only monitoring and managing the site.
Another approach to providing electron acceptors is to
add aqueous solutions of hydrogen peroxide or nitrate
to wells located within or immediately upgradient of the
plume; however, nitrate addition has not been shown to
be effective for degradation of benzene, and introduction
of hydrogen peroxide is more labor intensive and thus
more costly than air sparging. Alternatively, a slow-
release oxygen compound such as magnesium perox-
ide can be placed in wells and oxygen allowed to diffuse
into the formation.
Regulatory acceptance for natural attenuation may be
more easily attained if a migration barrier is created
130
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along the downgradient edge of the plume. Migration
barriers can be created through a series of low-flow air
sparging wells. Alternatively, a row of wells containing a
slow-release oxygen compound can be placed perpen-
dicularto the ground-water gradient nearthe downgradi-
ent edge of the plume. As demonstrated by Bob Borden
(4) of the University of North Carolina at a commercial
site in North Carolina and by Doug Mackay (5) of Stan-
ford University in tests conducted at the Borden Landfill
in Canada, this approach can successfully prevent mi-
gration of monoaromatic hydrocarbons.
Intrinsic bioremediation should be viewed as one ap-
proach in a continuum of methods of utilizing biodegra-
dation processes to remediate soil and ground water. It
should be used alone, in combination with other ap-
proaches, or as a polishing step based on evaluation of
the site conditions, regulatory issues, technical feasibil-
ity, implementability, and cost.
This poster session will present cost and analytical data
from three New Jersey service stations during the moni-
toring-only phase, site maps, diagrams of the air
sparging systems that have been installed, costs of
installation and operation, and analytical data available
at the time of the meeting. The cost of adding hydrogen
peroxide at two similar sites will be discussed. Addition-
ally, the costs and relative advantages of the use of
slow-release oxygen compounds and air sparging will
be presented.
References
1. National Research Council. 1993. In situ bioremedia-
tion: When does it work? Washington, DC: National
Academy Press.
2. Wilson, J.T., J.F. McNabb, J. Cochran, T.H. Wang,
M.B. Tomson, and RB. Bedient. 1985. Influence of
microbial adaptation on the fate of organic pollutants
in ground water. Environ. Toxicol. Chem. 4:721-726.
3. Norris, R.D., J.C. Dey, and D.R Shine. 1993. The
advantages of concerted bioremediation of lightly
contaminated sites compared to intrinsic bioremedia-
tion. Presented at the American Chemical Society
I&E Special Symposium, Atlanta, GA.
4. Kao, C.-M., and R.C. Borden. 1994. Enhanced aero-
bic bioremediation of a gasoline contaminated aqui-
fer by oxygen-releasing barriers. In: Hydrocarbon
bioremediation. Boca Raton, FL: Lewis Publishers.
5. Bianchi-Mosquera, G.C., R.M. Allen-King, and D.D.
Mackay. 1994. Enhanced degradation of dissolved
benzene and toluene using a solid oxygen-releasing
compound. Ground Water Monitor. Rev. pp. 120-128.
131
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Natural Attenuation of Jet Fuel in Ground Water
Greg Doyle, Dwayne Graves, and Kandi Brown
International Technologies Corporation, San Bernardino, CA
Natural attenuation is a minimum action remedial strat-
egy that permits the biodegradation of organic contami-
nants under natural, in situ conditions. Mechanisms that
act to affect contaminant biodegradation include aerobic
biodegradation at the plume boundary and various
anaerobic processes within the plume. Nitrate, iron,
manganese, and sulfate reduction and methanogenesis
are thought to support contaminant biodegradation.
Naturally occurring levels of nitrate, sulfate, oxidized
iron and manganese, and carbonates support anaerobic
biodegradation.
Natural attenuation was proposed as a feasible remedial
alternative for the dissolved jet fuel plume at George Air
Force Base in Victorville, California. The plume, cover-
ing approximately 1.1 million square feet, was located in
a poorly yielding perched aquifer, 120 ft below the
ground surface. Ground-water flow rate was approxi-
mately 20 ft/yr, and the edge of the plume was about
2,000 ft from the property line. The ground water in the
affected aquiferwas not being used. Because of subsur-
face conditions, slow migration, poor water yield, and
lack of use, natural attenuation represented the most
efficient and cost-effective approach for remediating dis-
solved jet fuel at this site.
A separate-phase layer of jet fuel floated on the water
table near the center of the plume. An aggressive skim-
ming operation was employed to remove all recoverable
separate-phase jet fuel from the subsurface. Bioventing
will be used to further remediate contaminated vadose
zone soil in areas where jet fuel was detected and
removed.
Natural attenuation for aerobic and aerobic/anaerobic con-
ditions was modeled using BIOPLUME II. The BIOPLUME
II model simulates the transport of dissolved hydrocarbons
under the influence of oxygen-limited biodegradation.
Using BIOPLUME II with site-specific parameters, vari-
ous treatment scenarios were evaluated. Assuming that
60 percent of the separate-phase product was removed,
both anaerobic and aerobic biodegradation occurred,
and residual jet fuel diffused into the water based on
Pick's Law of Diffusion, a remediation time of 42 yr was
determined. This treatment time was adequate to reme-
diate the plume before it migrated off site.
Based on this prediction and the cost savings realized
by applying natural attenuation, a 5-yr performance pe-
riod was established prior to the issuance of a finalized
record of decision. Joint efforts by the Air Force, the U.S.
Environmental Protection Agency's Robert S. Kerr Envi-
ronmental Research Laboratory, and International Tech-
nologies Corporation will provide data verifying the
accuracy of the BIOPLUME II model predictions and
demonstrating the level of natural biological activity oc-
curring in the ground water. These efforts are expected
to lead to regulatory acceptance of natural attenuation
forthe full-scale remediation of the jet fuel plume on site.
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Evaluation of Intrinsic Bioremediation at an Underground Storage Tank Site in
Northern Utah
R. Ryan Dupont, Darwin L. Sorensen, and Marion Kemblowski
Utah Water Research Laboratory, Utah State University, Logan, UT
A 2-yr field study was initiated by the Utah Water Re-
search Laboratory for the U.S. Environmental Protection
Agency's Office of Underground Storage Tanks at two
former underground storage tank (UST) sites in northern
Utah: a U.S. Air Force site at Hill Air Force Base (AFB)
near Ogden, Utah, and a private site in Layton, Utah.
The study sought to evaluate rapid site assessment
techniques and data collection and summary methods
that could be used to provide a comprehensive descrip-
tion of the potential for intrinsic bioremediation at UST
sites. This poster presentation focuses on the Hill AFB
site, where tank and line leaks from an 18,000 gal UST,
removed in 1989, were the probable sources of ob-
served ground-water contamination. Total dissolved
petroleum hydrocarbon contaminant mass at the begin-
ning of the study was estimated to be 950,000 mg, while
approximately 6,000 mg of benzene and 8,000 mg of
toluene were quantified in the plume underlying the site.
The site was covered with a permeable gravel surface
layer during the study, and dissolved oxygen concentra-
tions at a number of locations in uncontaminated areas
surrounding the plume were over 2 mg/L.
Initial site assessment activities utilized cone pene-
trometry for rapid collection of subsurface soil strati-
graphic data and for the placement of more than 60
small-diameter ground-water sampling points at the two
shallow field sites. Ambient temperature headspace
analyses were conducted using a polyethylene bag
method developed by the Utah Water Research Labo-
ratory in addition to a commercial Lag-in-a-Bag appara-
tus to provide rapid field-determined measurements of
ground-water total hydrocarbons. Field total petroleum
hydrocarbon (TPH) measurements were collected on a
near real-time basis to guide the initial placement of
ground-water monitoring points along and perpendicular
to the axis of the contaminant plume at each site. These
field methods proved that the original conceptual model
of the nature and extent of contamination and of the
potential contaminant migration pathways at the Hill
AFB site, based on conventional site assessment tech-
niques (soil gas survey, collection of limited soil core
samples, and placement of more limited numbers of
large-diameter ground-water monitoring wells), was
greatly in error. The model could be improved signifi-
cantly with these cost-effective field techniques, which
are now widely available to the consulting community.
Seven field sampling events, beginning in April 1992,
were conducted over the 2-yr study period as part of a
proposed intrinsic bioremediation strategy developed in
the project. Ground-water quality data were collected from
the small-diameter sampling points and existing ground-
water monitoring wells to assess the distribution and trans-
port of contaminants, along with the predominant microbial
reactions taking place within the contaminant plume.
Ground-water quality data collected included field meas-
urements of pH, dissolved oxygen, temperature, and
ambient headspace total hydrocarbon concentrations;
laboratory determinations included nitrate-N, sulfate,
dissolved iron and manganese, total hydrocarbons
(purge and trap and semivolatile constituents), specific
organic contaminants (C-6 to C-15 alkanes and ben-
zene, toluene, ethylbenzene, p-xylene, naphthalene,
and methylnaphthalene), and contaminant boiling point
split concentrations.
Ground-water data were used to generate total and
specific compound dissolved contaminant and dis-
solved electron acceptor mass data for each sampling
period. The location of the center of the mass of con-
taminants and electron acceptors were evaluated, and
changes in these parameters over time, with respect to
bulk ground-water flow, were used to assess the rate
and magnitude of natural degradation processes taking
place at each site.
Over the 2-yr study period, the mass of TPH showed
exponential decay, while all specific constituents dis-
played zero-order decay rates. Dissolved TPH mass
declined to less than 1,000 mg (99.9 percent removal)
133
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by the end of the 630-d monitoring period, while
dissolved benzene and toluene mass remaining in
the contaminant plume declined to less than 200 mg
(97 percent removal). TPH, benzene, and toluene mass
decay rates were found to be -0.013/d (p = 0.03,
r2 = 0.933), -11.2 mg benzene/d (p = 0.01, r2 = 0.976),
and -14.2 mg toluene/d (p = 0.01, r2 = 0.998), respec-
tively. Mass center data indicate that while ground-water
velocities at the site through the study period averaged
0.45 ft/d, the net movement of contaminants was attenu-
ated significantly, with measured center of mass veloci-
ties of TPH, benzene, and toluene being 0.03 ft/d, 0.05
ft/d, and 0.07 ft/d, respectively. Corresponding utilization
of oxygen and other terminal electron acceptors oc-
curred across the plume.
Intrinsic bioremediation of the dissolved plume at the Hill
AFB site successfully attenuated and removed the re-
sidual hydrocarbon mass existing at the site at the be-
ginning of the study. In January 1994, only 1 of 34
monitoring well/piezometer ground-water samples con-
tained a benzene concentration (20.7u,g/L) above regu-
latory concern, and the site is expected to be eligible
for closure at the next routine, semiannual sampling
event.
This poster session will detail the physical/chemical
characteristics of the field sites and the rapid site as-
sessment techniques and typical results collected, as
well as highlight the data collection/reduction/interpreta-
tion methodology developed in this study. Finally, more
complete results demonstrating natural degradation of
hydrocarbon contaminants at this site will be presented,
along with a summary of a natural attenuation decision
support system to aid investigators in assessing the
viability of intrinsic bioremediation for the selection of a
"no action'Ynatural attenuation monitoring alternative at
their sites.
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Case Studies of Field Sites To Demonstrate Natural Attenuation ofBTEX
Compounds in Ground Water
Chen Y. Chiang and Paul M. McAllister
Shell Development Company, Houston, TX
The most definitive indicators of natural attenuation
such as plume characteristics and dissolved oxygen
(DO) concentrations are based on actual concentrations
obtained during periodic monitoring events. Based on
appropriate data from monitoring wells, the following
parameters can be used to indicate and demonstrate
that natural attenuation is occurring: 1) the mass of
benzene, toluene, ethylbenzene, and the xylenes
(BTEX) present and 2) the extent and rate of migration
and distribution of BTEX concentrations. Data collected
from two field sites will be used to demonstrate natural
attenuation mechanisms.
The first site is characterized by 42 monitoring wells to
show the relationship between soluble BTEX and DO
plumes. Results from 10 sampling periods over 3 years
show a significant reduction in total BTEX mass with
time in ground water. These reduction and leakage rates
from sources are determined from material balance and
nonlinear least-squares analyses. The natural attenu-
ation rate is calculated to be 0.95 percent/d. Spatial
relationships between DO and total BTEX are shown to
be strongly correlated by statistical analyses and solute
transport modeling. In addition, laboratory microcosm
biodegradation experiments are performed to determine
possible threshold limits for aromatic hydrocarbon oxi-
dation under varying levels of DO. The results are re-
markably consistent with field data on the presence of
high or low levels of BTEX and DO in several monitoring
well-water samples.
The second site data will be used to demonstrate natural
attenuation from a cost-effectiveness perspective
through evaluation of plume characteristics over time.
The benzene concentrations along the primary flow path
at this site are observed to decrease from 2,600 ppb at
the source to 2.7 ppb at a distance 1,425 ft downgradi-
ent. The decrease in concentrations with distance from
the source is a direct indication that some degree of
natural attenuation is occurring. If no natural attenuation
was occurring, then concentrations would remain rela-
tively constant out to the leading edge, where a sharp
front would be observed. It is emphasized that natural
attenuation also includes other mechanisms than biode-
gradation: dispersion, sorption, volatilization, and chemical
transformation.
135
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Demonstrating Intrinsic Bioremediation of BTEX at a Natural Gas Plant
Keith Piontek, Tom Sale, and Jake Gallegos
CH2M Hill, St. Louis, MO, and Denver, CO
Steve de Albuquerque and John Cruze
Phillips Petroleum Company, Bellaire, TX, and Bartlesville, OK
Intrinsic bioremediation is being characterized at the site
of a former natural gas plant. Gas plant operations
resulted in the release of nonaqueous-phase liquid
(NAPL) to the eolian deposits beneath the site. Previous
investigation of the site had established that NAPL
(natural gas condensate) had been released to the sub-
surface, and had delineated to the extent of benzene,
toluene, ethylbenzene, and xylene (BTEX) in ground
water beneath the site. The BTEX biodegradation that
occurs under natural conditions is being characterized
to provide for more realistic assessment of potential
risks posed by the site, and to support long-term site
management decision-making.
The technical protocol proposed by Dr. John Wilson of
the U.S. Environmental Protection Agency's Robert S.
Kerr Environmental Research Laboratory was used as
a guide in planning the intrinsic bioremediation demon-
stration. Key parameters assessed included seasonal
variations in ground-water flow direction and velocity,
and changes in the concentrations of hydrocarbons and
geochemical parameters (e.g., electron acceptors)
across the site. The ground-water monitoring revealed
significant changes in ground-water geochemistry at lo-
cations upgradient, within, and downgradient of the
NAPL zone. These data, together with knowledge of the
stoichiometry of hydrocarbon biodegradation under vari-
ous redox conditions, were used to characterize the extent
of naturally occurring hydrocarbon biodegradation.
It was found that while oxygen flux into the plume oc-
curs, the majority of contaminant mass removal occurs
under anoxic conditions. Site data suggest that sulfate
is the most significant electron acceptor in terms of
hydrocarbon mass removal, with over 90 percent of the
hydrocarbon mass removal attributable to sulfate reduc-
tion. Additional hydrocarbon mass removal is attribut-
able to reduction of nitrate, ferris iron, and oxygen; and
methanogenic activity.
The project provided evidence that naturally occurring
contaminant biodegradation plays a significant role in
limiting contaminant migration from the site, as summa-
rized below:
• There is a significant flux of electron acceptors, par-
ticularly sulfate, into the NAPL zone. Depressed con-
centrations of these electron acceptors within the NAPL
zone are indicative of contaminant biodegradation.
• Elevated levels of bicarbonate, a byproduct of con-
taminant biodegradation, are found in monitoring
wells located downgradient of the NAPL zone. This
indicates that ground water from the NAPL zone has
reached the monitoring wells. BTEX is not present at
detectable concentrations in these wells.
• Contaminant transport calculations indicate that, in
the absence of naturally occurring contaminant
biodegradation, benzene would likely have migrated
to the downgradient monitoring wells by this time. As
stated above, benzene has not been detected in
these monitoring wells.
• The presence of oxygen and nitrate in the downgradi-
ent wells indicates that flux of electron acceptors into
ground water downgradient of the NAPL zone pro-
vides additional capacity for contaminant biodegrada-
tion, in the event of further lateral BTEX migration.
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Demonstrating the Feasibility of intrinsic Bioremediation at a Former
Manufactured Gas Plant
Ian D. MacFarlane
EA Engineering, Science, and Technology, Sparks, MD
Edward J. Bouwer
The Johns Hopkins University, Baltimore, MD
Patricia J.S. Colberg
University of Wyoming, Laramie, WY
A former manufactured gas plant (MGP) in Baltimore,
Maryland, is the subject of an investigation to assess
natural, /ns/fry biodegradation for the purposes of basing
remediation decisions on intrinsic bioremediation and
engineered, enhanced bioremediation. Tar, a byproduct
of the former gas manufacturing operation, is found in
the site's subsurface to depths as great as 100 ft. The
tar is a dense nonaqueous-phase liquid (DNAPL) that
contains monocyclic aromatic hydrocarbons (MAHs),
such as benzene and toluene, and polycyclic aromatic
hydrocarbons (PAHs), such as naphthalene and
benzo(a)pyrene. The tar DNAPL acts as a long-term
source of aromatic hydrocarbons to ground water, and
concentrations in ground water over much of the 70-acre
site are close to theoretical effective solubilities.
Laboratory and field data were collected to evaluate
biodegradation. The first phase of the laboratory inves-
tigation, consisting of tests on 16 soil samples from
various depths, was to discern whether or not microbes
existed in the subsurface, if they could use MGP-tar as
a carbon source, and if relationships could be estab-
lished between hydrogeology, contaminant distribution,
and microbiological characteristics. Aerobic and anaero-
bic enumerations were performed, followed by fatty acid
analyses to identify the microbes. Viable aerobic bacte-
ria were detected in all subsurface samples. Bacteria
were grown in the samples plated under anaerobic con-
ditions, but at counts of 10 percent to 50 percent less
than the corresponding aerobic counts. Tar-degrading
bacteria were detected in 7 of the 16 samples.
The second laboratory phase consisted of more detailed
microcosm studies performed by the Johns Hopkins
University (JHU) and the University of Wyoming (UW).
JHU used 49 soil samples from five boreholes with site
ground-water and individual radiolabeled target sub-
strates (benzene, naphthalene, phenanthrene, and ace-
tic acid) in sealed vials to make microcosms that
mimicked in situ redox conditions (i.e., oxygenated or
unoxygenated). Viable bacteria were enumerated, and
total cell counts were performed by the acridine orange
direct-count method. Similar microcosm studies are be-
ing performed by UW under sulfidogenic, iron-reducing,
and methanogenic conditions using phenol, benzene,
toluene, naphthalene, and phenanthrene as the tar-
geted radiolabeled compounds. Benzene, naphthalene,
and phenanthrene were observed to mineralize 6 percent
to 24 percent, 8 percent to 43 percent, and 3 percent to
31 percent, respectively, in JHU aerobic microcosms
over a 4-week incubation period. Anaerobic naphtha-
lene mineralization (7 percent to 13 percent) was ob-
served in two JHU samples in the presence of NO3.
Half-lives calculated from first-order degradation rates
typically ranged from tens to hundreds of days, with a
lower half-life in the initial stages of incubation followed
by a slower rate presumably indicative of oxygen- or
nutrient-limiting conditions. Under sulfate-reducing con-
ditions, phenol mineralized 13 percent to 18 percent in
200 d, but benzene, naphthalene, and phenanthrene
showed less than 1 percent mineralization in the same
period, and toluene showed less than 1 percent miner-
alization in 165 d. The apparent limited transformation
may, in fact, be due to the small inoculum sizes used, a
phenomenon documented by others.
Field investigations for natural in situ biodegradation
included aqueous phase redox conditions, biogenic
product analyses, and apparent attenuation of model
137
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contaminants. Ground-water quality data generally
showed reduced conditions with little or no measurable
oxygen, low redox potential (-70 mV average), high
biochemical oxygen demand in source zones (>200 mg/L),
elevated sulfate (2,200 mg/L average), and elevated
iron (570 mg/L average). Biogenic gases (CO2, H2S,
and CH4) were detected at levels greater than atmos-
pheric in 11 of the 16 wells measured.
Apparent degradation rates were calculated using the
first-order model by regressing the natural log of con-
stituent concentrations (adjusted for dilution) by esti-
mated in situ travel time. Because no nonreactive,
conservative tracers are unique to the tar sources, dilu-
tion was estimated by considering dissolved carbon
(organic plus inorganic) as a tracer. While the use of
carbon as a tracer is prone to error due to its reactive
nature, its use is conservative in that observed carbon
concentrations would tend to be less than actual anthro-
pogenic carbon, resulting in overestimates of dilution
and underestimates of degradation rates. Half-lives for
benzene, toluene, ethylbenzene, xylenes, and naphtha-
lene were calculated to be 729, 660, 877, 855, and
2,166 d, respectively. Half-lives for the three-ring and
greater PAHs were not calculated due to the poor re-
gression correlation coefficients. Degradation estimates
showed that toluene is the most preferred aromatic sub-
strate studied and that naphthalene appears to degrade
the slowest as predictable from the literature. Surpris-
ingly, the benzene rate was only slightly less than the
toluene rate.
Laboratory investigations have shown that 1) microbes
exist in the subsurface, 2) microbes are capable of using
tar as a carbon source, 3) various redox conditions can
be established with site consortia, and 4) site bacteria
can degrade selected aromatic constituents under aero-
bic and anaerobic conditions. Field evidence showing
various redox conditions and biogenic products of or-
ganic degradation gives clues to the possible fate of
aromatic hydrocarbons, but this indirect evidence can
only be used to support more definitive evidence in
demonstrating intrinsic bioremediation. Although esti-
mates of in situ constituent decay are based on numer-
ous assumptions and are fraught with uncertainty, this
evidence is needed to show real attenuation of aromatic
hydrocarbons. In this case, enough geochemical and
hydrogeologic data were available to segregate dilution
(an important "attenuation" process) from degradation
processes. The task of estimating apparent degradation
was facilitated by the relatively simple hydraulics and
the aged system to allow assumption for negligible sorp-
tion. The attenuation assessment demonstrated that
contaminant loss was observed over time along the
aqueous plume travel path (i.e., travel time, rather than
at one point overtime) and degradation, probably biotic,
was measured for target contaminants.
The combined laboratory and field evidence point to
natural /ns/fry biodegradation as an active process in the
site's low oxygen, subsurface environment. Intrinsic
bioremediation, albeit slow due to mass transfer limita-
tions from the tar to the aqueous phase, may be techni-
cally viable for controlling aqueous aromatic hydrocarbon
contamination emanating from MGP-tar sources. Labora-
tory studies are continuing to explore compound-spe-
cific biodegradation under various conditions, and plans
are being formulated now for an in situ biodegradation
pilot study.
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Natural and Enhanced Bioremediation of Aromatic Hydrocarbons at Seal Beach,
California: Laboratory and Field Investigations
Harold A. Ball, Gary D. Hopkins, Eva Orwin, and Martin Reinhard
Western Region Hazardous Substance Research Center, Stanford, CA
Introduction
The objective of this study was to develop our under-
standing of environmental factors that are important for
intrinsic anaerobic biodegradation of aromatic hydrocar-
bons in contaminated ground-water aquifers, and to
determine methods to enhance this process. The focus
of the investigation was a site at the Seal Beach Naval
Weapons Station in southern California, where a signifi-
cant gasoline spill resulted in contamination of the
ground-water aquifer (1). In the field, nitrate was present
at about 5 mg/L in background wells and approached
detection levels (mg/L) in the contaminated wells. There
was a high natural background sulfate concentration of
about 85 mg/L in the ground water, and methane was
detected in the contaminated well headspace. The dis-
tribution of aromatics present in the contaminated ground
water differs from that expected from dissolution of pure
gasoline (2). This suggests that natural biotransforma-
tion of several organic species is occurring at the site.
The project was divided into laboratory and field compo-
nents, which were interrelated. The goals of both the
laboratory and field experiments were to determine the
capability of the native aquifer microbial community to
transform aromatic hydrocarbon compounds under an-
aerobic conditions and to understand the effect of alter-
nate environmental conditions on the transformation
processes. Field experiments were carried out on site at
Seal Beach. At the field site, experimental monitoring of
biotransformation under natural (presumed sulfate-
reducing) and nitrate-enhanced conditions have been
carried out.
Approach and Results
Laboratory Study
In a laboratory microcosm experiment (3), we evaluated
several factors which were hypothesized to influence in
situ biotransformation processes. Individual monoaromatic
compounds (e.g., benzene, toluene, ethylbenzene, and
m-, p-, and o- xylene) were the primary substrates. In
replicate bottles during the first 52 d of the study, toluene
and m+p-xylene (here, m-xylene and p-xylene were
measured as a summed parameter) were biotrans-
formed in the unamended ground-water samples under
presumed sulfate-reducing conditions. Addition of ni-
trate to the ground water increased rates of toluene
biotransformation coupled to nitrate reduction, stimu-
lated biotransformation of ethylbenzene, and inhibited
the complete loss of m+p-xylene that was observed
when nitrate was not added and sulfate-reducing condi-
tions prevailed. Addition of the nutrients ammonia and
phosphate had no effect on either the rate of aromatics
transformation or the distribution of aromatics trans-
formed. When Seal Beach sediment was placed into
nitrate-reducing media, ethylbenzene was transformed
first, followed by toluene. When the sediment was
placed into sulfate-reducing media, lag times were in-
creased, but toluene and m-xylene were ultimately
transformed just as in the microcosms with ground water
alone. Although methane had been detected in the field,
there appeared to be no transformation of aromatic
compounds in the methanogenic microcosms during the
period of the experiment.
Bioreactor Study
A pilot-scale facility consisting of 90-L reactors was con-
structed at the Seal Beach site (4-6). The facility was
designed for the operation of three anaerobic in situ
bioreactors. The reactors consisted of aquifer-sediment-
filled, stainless steel cylindrical vessels with the capabil-
ity to control and monitor both hydrodynamic flow and
supplements to the composition of the native ground-
water influent. Initial operation of the three anoxic/an-
aerobic reactors focused on evaluating anaerobic biore-
mediation strategies for aromatic hydrocarbons under
natural (presumed sulfate-reducing) and enhanced de-
nitrifying conditions.
139
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Bioreactor results were consistent with the laboratory
microcosm experiments. Toluene and m+p-xylene were
degraded in both the unamended and nitrate-amended
bioreactors. Degradation of ethylbenzene was stimu-
lated by nitrate addition. There was no evidence that
benzene or o-xylene was transformed in either reactor.
The final percentage removal efficiency appeared to be
higher in the unamended bioreactor, where flow was
slower.
Field Study
Field experiments have been conducted to assess aro-
matic biotransformation in a test zone within the con-
taminated aquifer at the Seal Beach site. Initial work
focused on evaluation of intrinsic bioremediation as evi-
denced by the distribution of aromatic species in back-
ground wells. Subsequent experiments to determine our
ability to enhance this biotransformation have been con-
ducted using a slug test experimental design in which a
single well was used for the injection of the "slug" or test
pulse and the same well was used to extract the test
pulse. Since the native ground water contained a variety
of electron acceptors and the water used for the injected
pulses was water that had previously been extracted
from the test zone, the ground water was treated to
control the concentration of all electron acceptors and
organics during the injection of the test pulse. Before
injection, the desired salts were added back to the de-
oxygenated injection stream, and the stream was me-
tered into the injection well. Sodium bromide was added
as a conservative tracer. Under this scenario, the differ-
ent electron acceptors investigated (e.g., nitrate and
sulfate) could be added as desired. During initial tracer
studies, the injection water was organics free, and thus
the source of the organics was desorption from the in
situ aquifer solids. In subsequent and ongoing bioreme-
diation studies, benzene, toluene, ethylbenzene, m-
xylene, and o-xylene were added with the injection pulse
at a concentration of approximately 200 u,g/L each.
The initial bromide tracer data showed stable tracer
concentrations and indicated no substantial encroach-
ment of native ground water detected in the first 0.4 pore
volumes. There was a very small hydraulic gradient at
the site, hence recovery of the bromide mass from the
test wells ranged from 93 percent to 99 percent with the
extraction of three pore volumes over a 103-d period.
During the tracer test, the equilibrium desorption con-
centrations for the aromatic hydrocarbons when the
electron acceptors nitrate and sulfate were absent from
the ground water were evaluated. Benzene, ethylben-
zene, and o-xylene concentrations remained relatively
stable and thus appeared to be at an equilibrium. The
toluene and m+p-xylene concentrations had a down-
ward trend relative to benzene once the native ground
water encroached after approximately 0.4 pore vol-
umes, suggesting that the nitrate and sulfate concentra-
tions available in the native ground water supported
some intrinsic biological activity in the latter part of the
experiment for toluene and m+p-xylene removal.
In a nitrate augmentation experiment, nitrate and aro-
matics were added to the injection pulse, resulting in
complete consumption of toluene and m-xylene followed
by ethylbenzene within the first 2 wk. o-Xylene was
slowly degraded, and its concentration approached zero
by day 60. There was no apparent loss of benzene when
compared with the inert tracer. The addition of nitrate to
the test region appeared to enhance the natural anaero-
bic denitrifying population. This would confirm that there
was already an active nitrate-reducing population in the
aquifer whose activity was enhanced by the addition of
nitrate. With the exception of o-xylene transformation,
these results were comparable with those from the ni-
trate-amended microcosm and bioreactor experiments,
wherein toluene, ethylbenzene, and m-xylene were
transformed under denitrifying conditions.
During the tracer study, methane was detected in the
test wells. With the encroachment of the native ground
water and associated increase in nitrate and sulfate
concentrations, the methane concentration decreased
to values close to zero, suggesting that nitrate and
sulfate inhibit methanogenesis at this site.
Acknowledgment
Funding for this study was provided by the U.S. Envi-
ronmental Protection Agency's Office of Research and
Development, under agreement R-815738-01 through
the Western Region Hazardous Substance Research
Center. The content of this study does not necessarily
represent the views of the Agency. Additional funding
was obtained from the Chevron Research and Technol-
ogy Company, Richmond, California.
References
1. Schroeder, R.A. 1991. Delineation of a hydrocarbon
(weathered gasoline) plume in shallow deposits at
the U.S. Naval Weapons Station, Seal Beach, Cali-
fornia. Water Res. Invest. Rep. 89-4203.
2. Cline, P.V., J.J. Delfino, and P.S.C. Rao. 1991. Par-
titioning of aromatic constituents into water from
gasoline and other complex solvent mixtures. Envi-
ron. Sci. Technol. 25(5):914-920.
3. Ball, H.A., and M. Reinhard. 1994. Laboratory study
of monoaromatic hydrocarbon degradation under an-
aerobic conditions at Seal Beach, California. In
preparation.
4. Ball, H.A., and M. Reinhard, M. 1994. Pilot-scale
study of monoaromatic hydrocarbon degradation un-
der anaerobic conditions at Seal Beach, California.
In preparation.
140
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5. Huxley, M.R, C. Lebron, M. Reinhard, H. Ball, H.F. 6. Reinhard, M., I.E. Wills, H.A. Ball, T. Harmon, D.W.
Ridgway, and D. Phipps. 1992. Anaerobic and aero- Phipps, H.F. Ridgway, and M.P. Eisman. 1991. Afield
bic degradation of aromatic hydrocarbons using in experiment for the anaerobic biotransformation of
situ bioreactors at an unleaded gasoline spill site. aromatic hydrocarbon compounds at Seal Beach,
Presented at the 18th Environmental Symposium of California. In: Hinchee, R.E., and R.F. Olfenbuttel,
the American Defense Preparedness Association, eds. In s/'fry bioreclamation: Applications and investi-
Alexandria, VA. gations for hydrocarbon and contaminated site reme-
diation. Boston, MA: Butterworth-Heinemann.
141
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The Complete Dechlorination of Trichloroethene to Ethene Under Natural
Conditions in a Shallow Bedrock Aquifer Located in New York State
David Major, Evan Cox, and Elizabeth Edwards
Beak Consultants Limited, Guelph, Ontario
Paul W. Hare
General Electric Company, Corporate Environmental Programs, Albany, NY
Introduction
In anaerobic environments, chlorinated ethenes can act
as electron acceptors in a process called reductive de-
halogenation (specifically, reductive dechlorination).
The extent of dechlorination, however, has been shown
to vary in anaerobic dechlorination studies depending
upon the flow and availability of electrons within the
anaerobic microbial community. For example, dechlori-
nated intermediates such as dichloroethene (DCE) and
vinyl chloride (VC) were found to accumulate during the
dechlorination of tetrachloroethene (PCE) and trichlo-
roethene (TCE) (1-3). Freedman and Gossett (4), how-
ever, were the first to observe the complete dechlorination
of PCE to ethene in a methanogenic enrichment culture,
and DeBruin et al. (5) showed that PCE could be re-
duced to ethane. The first field observation of the com-
plete dechlorination of PCE to ethene was documented
by Major et al. (6). Their laboratory and field study
showed that anaerobic microorganisms in a low-perme-
ability aquifer were capable of naturally dechlorinating
PCE in the presence of methanol. This paper docu-
ments that microorganisms in a bedrock aquifer unit are
also capable of completely dechlorinating TCE to
ethene.
Study Site Conditions
The study site is located in the Finger Lakes region of
central New York. The property was used from 1951 to
1990 as a manufacturing plant for a variety of electrical
components, including high-voltage semiconductors. In
the early to mid-1960s waste solvents were disposed of
in an unlined evaporation pit. TCE, which was often
mixed with acetone or methanol, was among the sol-
vents disposed of in the unlined evaporation pit. As a
result, these chemicals are now found in the overburden
and bedrock units beneath the study site.
Results
Our study involved collecting representative ground-
watersamples from 21 existing ground-water monitoring
wells, mostly in the shallow bedrock unit, forgeochemi-
cal and microbiological analyses. 1,2-DCE and VC were
detected in ground-water samples, which indicated that
TCE was being biodegraded in the subsurface at the
site. These TCE degradation products were not used or
produced at the site, and thus their presence can only
be attributed to the dechlorination of TCE. In addition,
the detection of ethene provides evidence that VC is
being dechlorinated at the site. Three observations of
the relative distribution of TCE and its dechlorination
products suggested that the migration of the volatile
organic compounds (VOCs) in the shallow bedrock unit
is being controlled by biodegradation. First, the distribu-
tion of TCE is much less extensive than the observed
distributions for 1,2-DCE, VC, and ethene. Second, the
distribution of VC, which should be greater than that of
1,2-DCE as predicted by its mobility in ground water,
was essentially the same as the distribution of 1,2-DCE.
Third, VC and ethene migrate at similar rates relative to
ground-water velocity and should have had similar dis-
tributions, but the distribution of VC was less than
ethene. The distribution of acetone and methanol is
considerably limited in comparison to the distribution of
the chlorinated VOCs and ethene. The mobility of ace-
tone and methanol should be approximately the same
as the average linear ground-water flow; however, their
distribution was less than the distribution of the VOCs.
This suggests that acetone and methanol are also being
biodegraded.
In addition to the distributions of VOCs, the distributions
of inorganic anions, methane and methane isotopes,
and volatile fatty acids (e.g., acetate) were used as
indirect measures of the activity of functional groups of
142
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microorganisms in the bedrock. The depletion of sulfate
and the production of methane and acetate indicated
that sulfate-reducing, methanogenic, and acetogenic
bacteria were active in the bedrock aquifer. Isotopic
analysis of methane indicated that the methane was
produced biotically. Furthermore, the distribution of
methane and methane isotopes clearly showed that the
microorganisms were active in the bedrock.
Microbial biomass, composition, and nutritional status
were assessed by extracting and analyzing phos-
pholipid fatty acids (PLFAs) and respiratory quinones
from microorganisms that were trapped onto 0.22-u.m
membranes. Analysis of respiratory quinones indicated
that the microbial populations at the site are strictly
anaerobic. The microbial biomass in the ground water
ranged from 1.6 x 102 cells/ml to 4.2 x 104 cell/ml. The
total biomass appeared generally to correlate with the
presence of VOCs and other nutrients, and was found
to be higher in the areas containing acetone and metha-
nol. The total biomass was orders of magnitude higher
near VOC source areas, as well as in areas with meas-
urable concentrations of acetone and methanol, than at
downgradient or background (upgradient) locations. The
microbial biomass distribution suggested that a biologi-
cally active zone (BAZ) has developed in response to
the presence of VOCs, acetone, and methanol. The
microbial populations in the samples generally demon-
strated nutritional or environmental stress, as indicated
by the ratio of specific PLFAs. Stress may be due to an
inadequate supply of nitrogen and phosphorous to sup-
port ideal growth. Cluster analysis of the PLFA data
showed that three population groups exist at the site.
The population groups appear to coincide with observed
changes in the concentration and types of VOCs and
other geochemical parameters.
Conclusions
This study provides evidence that microbial populations
can exist and function in bedrock. Furthermore, these
populations possess an intrinsic capability to anaerobi-
cally dechlorinate TCE to ethene when suitable sub-
strates are present to support their growth. At this study
site, an active and diverse anaerobic microbial commu-
nity, consisting of sulfate-reducing, methanogenic, and
acetogenic bacteria, has been established and is being
maintained by acetone and methanol. This anaerobic
microbial community is affecting the distribution and
migration of TCE, TCE biodegradation products, and
other chemicals at the site.
References
1. Bouwer, E.J., and PL. McCarty. 1983. Transforma-
tion of 1- and 2-carbon halogenated aliphatic organic
compounds under methanogenic conditions. Appl.
Environ. Microbiol. 45(4):1,286-1,294.
2. Parsons, F., PR. Wood, and J. DeMarco. 1984.
Transformations of tetrachloroethylene and trichlo-
roethylene in microcosms and ground water. J. Am.
Water. Assoc. 76:56-59.
3. Wilson, B.H., G.B. Smith, and J.J. Rees. 1986.
Biotransformations of selected alkylbenzenes and
halogenated aliphatic hydrocarbons in methano-
genic aquifer material: A microcosm study. Environ.
Sci. Technol. 20(10):997-1,002.
4. Freedman, D.L., and J.M. Gossett. 1989. Biological
reductive dechlorination of tetrachloroethylene and
trichloroethylene to ethylene under methanogenic
conditions. Appl. Environ. Microbiol. 55(9):2,144-
2,151.
5. De Bruin, WP, M.J.J. Kotterman, M.A. Posthumus,
G. Schraa, and A.J.B. Zehnder. 1992. Complete bio-
logical reductive transformation of tetrachloroethene
to ethane. Appl. Environ. Microbiol. 58(6):1,996-
2,000.
6. Major, D.M., E.W Hodgins, and B.J. Butler. 1991.
Field and laboratory evidence of in situ biotransfor-
mation of tetrachloroethene to ethene and ethane at
a chemical transfer facility in North Toronto. In:
Hinchee, R.E., and R.F. Olfenbuttel, eds. On-site
bioremediation, pp. 147-171. Boston, MA: Butter-
worth-Heinemann.
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Appendix A:
Regulatory Meeting Participants
145
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U.S. Environmental Protection Agency
Office of Research and Development
Symposium on Natural Attenuation of Ground Water
Regulatory Meeting on State and Federal Issues
Impacting Natural Attenuation
Hyatt Regency Denver
Denver, CO
August 29, 1994
Final Participant List
Michael Barden
Hydrogeologist
Wisconsin Department of Natural Resources
101 South Webster Street
P.O. Box 7921 (SW/3-ERR)
Madison, Wl 53707
608-264-6007
Fax: 608-267-2768
Dolloff Bishop
Chief, Biosystems Branch
Water & Hazardous Waste Research Division
U.S. Environmental Protection Agency
26 West Martin Luther King Drive
Cincinnati, OH 45268
513-569-7629
Fax:513-569-7105
Francis Chapelle
Hydrologist
U.S. Geological Survey
720 Gracern Road - Suite 128
Columbia, SC 29210
803-750-6116
Fax: 803-750-6181
Matthew Charsky
Environmental Scientist
Hazardous Site Control Division
U.S. Environmental Protection Agency
401 M Street, SW (5203-G)
Washington, DC 20460
703-603-8777
Fax:703-603-9100
Daniel Cozza
Remedial Project Manager
U.S. Environmental Protection Agency
77 West Jackson Boulevard (HSRW-6J)
Chicago, IL 60604
312-886-7252
Fax: 312-353-5541
Marty Faile
Environmental Engineer
Air Force Center for Environmental Excellence
U.S. Air Force
8001 Arnold Drive
Brooks AFB, TX 78235-5357
210-536-4342
Fax:210-536-4330
Felix Flechas
Environmental Engineer
U.S. Environmental Protection Agency
999 18th Street - Suite 500 (HWM-HW)
Denver, CO 80202
303-293-1524
Fax: 303-293-1724
Rutherford Hayes
Remedial Project Manager
South Carolina Section
North Superfund Branch
U.S. Environmental Protection Agency
345 Courtland Street, NE (4WD-NSRB)
Atlanta, GA 30365
404-347-4103
Fax:404-347-1695
146
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Michael Jawson
Chief, Subsurface Processes Branch
Robert S. Kerr Environmental Research Laboratory
U.S. Environmental Protection Agency
P.O. Box 1198
Ada, OK 74820
405-436-8560
Fax: 405-436-8703
Robin Jenkins
Hydrogeologist
Division of Environmental Response & Remediation
Utah Department of Environmental Quality
168 North 1950, W- 1st Floor
Salt Lake City, UT 84116
801-536-4100
Fax:801-359-8853
Fran Kremer
Chairperson, Biosystems Technology
Development Program
U.S. Environmental Protection Agency
26 West Martin Luther King Drive
Cincinnati, OH 45268
513-569-7346
Fax: 513-569-7585
John Kuhns
Superfund Project Manager
U.S. Environmental Protection Agency
77 West Jackson Boulevard (HSRW-6J)
Chicago, IL 60604
312-353-6556
Fax: 312-353-5541
Tim Larson
Engineer
Bureau of Waste Cleanup
Florida Department of Environmental Protection
2600 Blair Stone Road
Tallahassee, FL 32399-2400
904-488-3935
Fax: 904-922-4368
Darryl Luce
Remedial Project Manager
U.S. Environmental Protection Agency
JFK Federal Building (HSN-CAN5)
Boston, MA 02203
617-573-5767
Fax:617-573-9662
Sriram Madabhushi
Hydrologist
Division of Ground Water Protection
South Carolina Department of Health & Environmental
Control
2600 Bull Street
Columbia, SC 29201
803-734-0723
Fax: 803-734-3604
Gail Mallard
Chief, Branch of Toxic Substances Hydrology
U.S. Geological Survey
12201 Sunrise Valley Drive - 412 National Center
Reston, VA 22092
703-648-6872
Fax: 703-648-5295
Robert Menzer
Laboratory Director
Gulf Breeze Environmental Research Laboratory
U.S. Environmental Protection Agency
One Sabine Island Drive
Gulf Breeze, FL 32561-5299
904-934-9208
Fax: 904-934-9201
Amy Mills
Chief, Superfund Technical Liaison Program
Office of Research & Development
U.S. Environmental Protection Agency
401 M Street, SW(8105)
Washington, DC 20460
202-260-0569
Fax: 202-260-0507
Read Miner
Hydrogeologist
South Carolina Department of Health & Environmental
Control
2600 Bull Street
Columbia, SC 29201
803-734-5335
Fax: 803-734-3604
Roland Ramsey
Soil Scientist
Petroleum Storage Tank Division
Texas Natural Resource Conservation Commission
P.O. Box13087
Austin, TX 78711-3087
512-239-6276
Fax: 512-239-2216
147
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John Rogers
Research Microbiologist
Athens Environmental Research Laboratory
U.S. Environmental Protection Agency
860 College Station Road
Athens, GA 30605-2720
706-546-3592
Fax: 706-546-3252
Alison Thomas
Research Project Engineer
Armstrong Laboratory
U.S. Air Force
139 Barnes Drive - Suite 2 (AL/EQW)
Tyndall AFB, FL 32403-5323
904-283-6303
Fax: 904-283-6286
Guy Tomassoni
Hydrogeologist
Office of Solid Waste
U.S. Environmental Protection Agency
401 M Street, SW (5303-W)
Washington, DC 20460
703-308-8622
Fax: 703-308-8617
Katrina Varner
Environmental Scientist
Analytical Science Division
Environmental Monitoring Systems Laboratory
U.S. Environmental Protection Agency
P.O. Box 93478
Las Vegas, NV 89193-3478
702-798-2645
Fax: 702-798-2692
John Wadhams
Environmental Health Scientist
Montana Department of Health
2209 Phoenix Avenue - P.O. Box 200901
Helena, MT 59620-0901
406-444-1420
Fax:406-444-1901
Mark Walker
Industrial Hygienist
Colorado Department of
Public Health & Environment
4300 Cherry Creek Drive, S (HMWMD-SWIM-B2)
Denver, CO 80222-1530
303-692-3449
Fax: 303-759-5355
Lisa Weers
Engineer
Colorado Department of Public Health & Environment
4300 Cherry Creek Drive, S (HMWMD)
Denver, CO 80222-1530
303-692-3451
Fax: 303-759-5355
Candida West
Research Environmental Scientist
Robert S. Kerr Environmental Research Laboratory
U.S. Environmental Protection Agency
P.O. Box1198
Ada, OK 74820
405-436-8551
Fax: 405-436-8703
John Wilson
Senior Research Microbiologist
Robert S. Kerr Environmental Research Laboratory
U.S. Environmental Protection Agency
P.O. Box1198
Ada, OK 74820
405-436-8532
Fax: 405-436-8703
148
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