United States
Environmental Protection
Agency
xvEPA
Office of Research and
Development
Washington, DC 20460
EPA/540/R-94/515
August 1994
Symposium on
Intrinsic
Bioremediation of
Ground Water
Hyatt Regency Denver
Denver, CO
August 30 - September 1,1994
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EPA/540/R-94/515
August 1 994
Symposium on Intrinsic Bioremediation of Ground Water
Hyatt Regency Denver
DenveV, CO
August 30 to, September 1, 1 994
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Printed on Recycled Paper
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Disclaimer
The projects described in this document have been reviewed in accordance with the peer and
administrative review policies of the U.S. Environmental Protection Agency and the U.S.
Geological Survey, and have been approved for presentation and publication. Mention of trade
names or commercial products does not constitute endorsement or recommendation for use.
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Acknow edgments
The papers abstracted in this book were presen ed atthe Symposium on Intrinsic Bioremediation
of Ground Water, held August 30 to Septsmber 1, 1994, in Denver, Colorado. The
Symposium was a joint effort of the U.S. Environmental Protection Agency's (EPA's) Biosystems
Technology Development Program and the LJ.S. Geological Survey (USGS), with additional
sponsorship from the U.S. Air Force. Fran V. ^remer (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 Lt. Col. Ross N. Miller, Headquarters, U.S. Air Force
Center for Environmental Excellence, Brooks Air Force Base, Texas.
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Contents
Site Characterization: What Should We Measure, Where (When?), and How?
Michael J. Barcelona
Processes Controlling the Distribution of Oil, Air, and Water
John L Wilson
New Tools To Locate and Characterize Oil Sp
Us in Aquifers
Bruce J. Nielsen
Microbiological and Geochemical Degradation Processes
£ Michael Godsy
Field and Laboratory Results: Getting the Whole Picture
Mary Jo BaedecJcer
In Situ Bioremediation at the Seventh Avenue pite in Denver: Remediation of Soils
and Ground Waters
Christopher Nelson
The Role of Intrinsic Bioremediation in Closure of Sites After Cleanup Through
In Situ Bioremediation: The Regulator's Perspective
Mar/c E. Walker and Lisa C. Weers ....
The Importance of Knowledge About Intrinsic Bioremediation for Cost-Effective Site
Closure: The Client's Perspective
Harry E. Moseley
The Role of Intrinsic Bioremediation in Closurs of Sites After Cleanup Through
In Situ Bioremediation: The Role of Mathematical Models
T/ssa H. Illangasekare, David C. Szlag, anc
Intrinsic Bioremediation of JP-4 Jet Fuel
John T. Wilson, Frederick M. Pfeffer, James
Todd H. Wiedemeier, Jerry E. Hansen, and
10
25
35
41
47
49
51
John T. Wilson 53
W. Weaver, Don H. Kampbell,
Ross N. Miller 60
A Natural Gradient Tracer Experiment in a Hoterogeneous Aquifer With Measured
In Situ Biodegradation Rates: A Case for Natiiral Attenuation
Thomas B. Stauffer, Christopher P. Anfworf/j, J. Mark Boggs,
and William G. Maclntyre ,
Traverse City: Distribution of the Avgas Spill
David W. Ostendorf
73
85
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Traverse City: Geochemistry and Intrinsic Bio-remediation of BTX Compounds
Barbara H. Wilson, John T. Wilson, Don H. Kampbell, Bert E. Bledsoe, and
John M. Armstrong 94
Mathematical Modeling of Intrinsic Bioremediation at Field Sites
Hanadi S. Rifai 103
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 Isabelle M. Cozzarelli Ill
Simulation of Flow and Transport Processes at the Bemidji, Minnesota, Crude-Oil
Spill Site
Hedeff I. Essaid 121
An Overview of Anaerobic Transformation of Chlorinated Solvents
Perry L McCarty 135
Contamination of Ground Water With Trichloroethylene at the Building 24 Site at
Picatinny Arsenal, New Jersey
Mary Martin and Thomas E. Imbrigiotta 143
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 154
Poster Session
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 1 63
Wisconsin's Guidance on Naturally Occurring Biodegradation as a Remedial
Action Option
Michael J. Barden I/O
Assessing the Efficiency of Intrinsic Bioremediation
Francis H. Chapelle 171
A Practical Approach to Evaluating Natural Attenuation of Contaminants in
Ground Water
Paul M. McAllister and Chen Y. Chiang 1 72
The Use of Low Level Activities To Assist Intrinsic Bioremediation
Robert D. Norris, Jeffrey C. Dey, and Daniel P. Shine . . 1 73
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Natural Attenuation of Jet Fuel in Ground Wafer
Greg Doyle, Dwayne Graves, and Kandi Brown 1 75
Evaluation of Intrinsic Bioremediation at an Urderground Storage Tank Site in
Northern Utah
R. Ryan Dupont, Darwin L Sorensen, and /v <.
Case Studies of Field Sites To Demonstrate Nctural Attenuation of BTEX Compounds
in Ground Water
Chen Y. Chiang and Paul M. McAllister
arion Kemblowski 1 76
178
Demonstrating Intrinsic Bioremediation of BTEX at a Natural Gas Plant
Keith Piontek, Tom Sale, Steve de Albuquerc
Demonstrating the Feasibility of Intrinsic Bioremediation at a Former Manufactured
Gas Plant
/an D. MacFarlane, Edward J. Bouwer, and Patricia J.S. Colberg 181
Natural and Enhanced Bioremediation of Aronatic Hydrocarbons at Seal Beach,
California: Laboratory and Field Investigations
ue, and John Cruze 1 79
Harold A. Ball, Gary D. Hopkins, Eva Orwir
The Complete Dechlorination of Trichloroetheie to Ethene Under Natural
Conditions in a Shallow Bedrock Aquifer in New York State
David Major, Evan Cox, Elizabeth Edwards,
, and Martin Reinhard 1 83
and Paul W. Hare 187
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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, Ml
Abstract
Site characterization represents the initial phase of the active monitoring process that occurs as
part of intrinsic organic contaminant bioremediation efforts. Initial characterization work sets the
stage for evaluating the progress of the nature transformation of contaminants. The following
have frequently been observed: parent compound disappearance, active microbial populations
with biotransformation capabilities, and the appearance or disappearance of organic and
inorganic constituents that provide evidenc
Quantitative evidence is lacking, however, for
mixtures solely by biological processes. This is
3 of bioremediation at contaminated sites.
net removal of toxic compounds from complex
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 over time.
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 dis
correspondence between contaminant distribu
microbial conditions in the subsurface overtim
tribution of contaminant mass. Then the
ion and favorable physical, geochemical, and
3 provides a basis for net contaminant-removal
estimates. Mere adaptations of detective ground-water monitoring networks are insufficient for
quantitative evaluation of intrinsic bioremediafon technologies.
Introduction
The practice of site characterization for
evolved slowly in the past decade. Early
contamination detection monitoring (i.e., moni
been applied to many sites of potential concern
phases.
This minimal approach has been applied
remediation of subsurface organic contaminants has
guidelines (1 -3) for minimal ground-water
oring wells upgradient and downgradient) have
rom detection through remedial action selection
widely, regardless of the physicochemical
characteristics of contaminant mixtures orthe complexity of hydrogeologic settings. With solubl
inorganic constituents, this approach may be adequate for detection purposes, but assessment
efforts require substantially more comprehensive approaches. For organic contaminant
assessment efforts (i.e., determinations of the rature and extent of contamination), wells alone
have been found to be inadequate monitoring 1ools. Recognition of the value of subsurface soil
vapor surveys for volatile organic components cf fuel and solvent mixtures has generated a flurry
of modified site characterization approaches bused on monitoring wells (4). These approaches
to site characterization and monitoring networljc design suffer also from a failure to identify the
total mass of contaminant in the subsurface.
Symposium on Intrinsic Bioremediation of Ground Water
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This failure occurs for three main reasons. First, although 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 distribution do not
identify the total mass distribution of organic contaminant (6). Second, efforts to correlate
observed soil vapor or ground-water VOC concentrations with those in subsurface solid cores
have often been unsuccessful, because current bulk jar collection/refrigeration at4°C guidelines
for solid core samples for VOC analyses lead to gross negative errors (7). Third, "snapshots"
(i.e., one-time surveys) of background and disturbed ground-water chemistry conditions 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 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, excellent
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
contaminant distributions overtime. Therefore, a dynamic, ongoing site-characterization effort
includes the following objectives:
• Identify the spatial distribution of contaminants, particularly 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, aliphatic, and aromatic hydrocarbon compounds),
because overall microbial activity and disruptions in subsurface geochemical
conditions (and bioremediation 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 downgradient
zones during the first year of characterization and monitoring.
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Derive initial estimates of net
organic matter overtime that i
network design.
rr icrobial transformations of contaminant-related
tmy be built into an efficient long-term monitoring
The first three objectives establish the environment of major contamination and the conditions
under which bioremediation may occur. The later two objectives are vitally important, because
evaluating the progress of intrinsic bioremediation processes depends on distinguishing
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 technical U.S. Air Force (USAF) protocol by Wiedemeier et al. (10).
The latter reference focuses directly on the implementation of intrinsic remediation for dissolved
fuel contamination 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 analytica
• Laboratory microcosm studies
These data, if collected in three dimensions
implement intrinsic remediation successfully
characterization effort (Figure 1) support the d
data in mass balance calculations
using aquifer samples collected from the site
:or an extended period, should be sufficient to
(11). The data collected in the initial site
evelopment of a site-specific conceptual model.
This model is a three-dimensional representatijon of the ground-water flow and transport fields
based on geologic, 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 protocol is quite
comprehensive in identifying important parameters, 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 protocol! for these critical parameters; thus, recent drafts
of the protocol should be even more useful to practitioners.
Typical detective monitoring data sets availab
likely to contain contaminant-related informcjiti
property, hydrogeologic, or geochemical da
recognition of the variability inherent in tl
characterization efforts.
3 prior to in-depth site characterization are more
ion rather than the three-dimensional aquifer
ra needed to formulate a conceptual model. A
lese parameter distributions is critical to site
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Review AvalaWe
SHe Data
4f
Develop Preliminary
Conceptual Modal
Jr
Make PreEminary
Assessment of Potential
For Intrinsic Remediaflon
Based on Existing Sits
Characterization Data
• Ccrtmww* Typ»
KidDcMajten
- Hydmosofagy
• LooCon oJ Rece$(oa
V f
Perform Sfta Characterization 1
In Support of Intrinsic Remediation f
4
Refine Conceptual Model and
Complete Pre-Mode5ng
Cateuiatfons
A
Document Occurrence of
. Intrinsic Remediation and
J Model Intrinsic Remediation
/ Using Nunerieal Models
4
Use Results of Modeling and
Sfta-Spetific Information in
an Exnosura Assessment
.-^\^*^-. ^s
v Evaluate Use of
> Conjunction With
Intrinsic Remediation
^l^\\
^acov8*^ / 111 XL^^ISI^J
rrcfei \
Tmt 1 1 1 Sparging
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iErtuncad
BJorefnediioi
^^Unacceptable Risk ra"*^.
^^s^Potential Recsptorsjx"^
^s^^X^ NO
wv
jv Examine Temporal/
•r Spatial Distributions
Sits Point-O(-Complianca
v Monitorina Weds and
' MonHorinqPlan
d> r__™^™T
Prepare Refined
Monitoring Plan
MBsera Midngs
andLong-Tenn
\ Monitoring Plan To
7 RegUaioiy Agencies
and Reach Agreement
en Monitoring Strategy
I
Assays Potential For
.,.._„,..._.>> IrtJinsir SnrrKvflaffon
^ W* Remedaflon
System Instaled
V
Refine Conceptual Model and
Complete Pre-Modefng
Calculations
V
Model Intrinsic Rsmedatfon
Combined with Remaotal
Option Selected Above
Using Numerical Models
4
Usa Results of Modeling and
Site-Specific Information in
an Exposure Assessment
YES ^X^s^^
^'iJnaccaptabla Risk To^*^.
^•s^Potential ReceptofsTX'^
NO ^^^^>^
Figure 1. Intrinsic remediation flow chart.
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Table 1. Site-Specific Parameters To Be Determined During Site Characterization (10)
Fractionation and
Spatial Extent of
Contamination
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
Hydrogeologic
and Geochemical
Framework
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-s
Aquifer
ze distribution (sand versus silt versus clay)
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
Sampling in Space
The initial site characterization phase should I
of critical data over volumes corresponding to
flow paths. If the flow path intersects a
should be scaled accordingly. For example, if
volume would be 1 yr of travel time. The '
hydrogeologic and geochemical parameters
derived from data sets that are large enough
mean, median, correlation distance, and
variance
ie designed to provide spatially dense coverage
1 0-yrto 1 00-yrtravel times along ground-water
discharge zone in less than 1 00 yr, then the volume
the flow path discharges after 10 yr, the critical
/olume-averaged" values of the contaminants,
within zones along the flow path(s), should be
to permit estimation of statistical properties (e.g.,
. In general, this means that the data sets for
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derived mass loadings of contaminants, aquifer properties, 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.
Table 2. Target Constituents for Site Characterization in Support of Intrinsic Bioremediation
Contamination
Area
Source
Downgradient
Upgradient/
Far-field
downgradient
Apparent/
Geochemical
Redox Zone
Reducing
Transitional/
Suboxic
Oxic
Contaminant
Mixture
Fuels
Chlorinated
solvents
Fuels
Chlorinated
solvents
Fuels
Chlorinated
solvents
Inorganic
Constituents
O2, CO2, H2S;
pH, Fe2+,
HS-/S-, NCy,
NH3, alkalinity
02, C02, H2S;
pH, Fe2+,
alkalinity, NO2",
NCy, NH3,
HS-/S-
02, C02, H2S;
alkalinity,
Fe2+, NO3-,
NCy, NH3
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
Organic carbon,
CH4, organic acids,
phenols
As above and:
chlorinated
metabolites,
ethylene, ethane
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 estimates 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.
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The second question pertains to the depth intsrval over which "planar" data points might be
averaged. With fuel-related aromatic contaminants, the depth interval above and below the
capillary fringe/water table interface typically exhibits order-of-magnitude differences in solid-
associated concentrations. In this situation, averaging 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 solids 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 sampl
characterization. Push technologies rely on hycraulic 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
ess cost than drilling techniques.
The approach to site characterization forchloririati
Very few models of site characterization for
in specific media. Many of the previously
detection, assessment, and quantitation,
sampling than intuition based on experience.
"these
'referenced
ng devices present very useful tools for site
ed hydrocarbons is significantly more difficult.
contaminants have estimated mass loadings
methods may work satisfactorily. Free-phase
however, may be more a matter of luck and exhaustive
Sampling Over Time
VOC compounds (e.g., aromatic hydrocarbons and chlorinated solvents) are among the target
contaminants that are considered constituents of concern in remedial investigations. Their
aqueous solubility and demonstrated associatiDn with aquifer solids require sampling of these
media during the site characterization phase. Tr
of complex organic mixtures (e.g., ethylene,
is suggestion also applies to organic metabolites
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 characterizati*
physical, geochemical, and microbial determi
ffort. This should also be the case for th
lotions.
Initially, conventional nested monitoring wells with screened lengths of 1 m or more will be
useful for estimating the spatial extent of fie dissolved plume, for delineating apparent
geochemical zones, and for providing data on
water level and aquifer property (e.g., slug- and
particularly multilevels appropriately designed
course of the long-term monitoring program. S
pump-test derived hydraulic conductivity estimates). Semiannual or annual sampling of wells
and completed, should be quite useful over the
impling 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. Unless
biotransformation can be shown to be a major
over extended periods, it will remain an area
Because very few contamination situations have
several years, it is difficult to define specific sanr
oss mechanism for contaminants mainly in solids
of research rather than practice.
been monitored intensively for periods exceeding
pling frequencies forthe range of hydrogeologic
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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 contamination situations.
Acknowledgements
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. Sea If, M.R., J.F. McNabb, W.J. Dunlop, R.L Cosby, and J.S. Fryberger. 1 981. Manual
of ground-water sampling procedures. National Water Well Association.
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 Contract Report 374 U.S.
Environmental Protection Agency, Ada, OK.
3. U.S. EPA. 1 986. RCRA technical enforcement guidance document, OSWER-9950.1.
Washington, DC.
4. Eklund, B. 1 985. Detection of hydrocarbons in ground water by analysis of shallow soil
gas/vapor. API Publication No. 4394. Washington, DC.
5. Plumb, R.H. 1 987. 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 detection, mapping, and modeling. Ground Water 2:1 55-1 62.
7. U.S. EPA. 1992. RCRA ground-water monitoring: Draft technical guidance document.
EPA/530/R-93/001. Washington, DC.
8. U.S. EPA. 1 994. 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: National Academy Press.
pp. 160-184.
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 remediation
(natural attenuation) with long-term monitoring option for dissolved-phase fuel
contamination in ground water. Air Force Center for Environmental Excellence, Brooks
AFB, San Antonio, TX (March).
-------
1 1. National Research Council. 1 993.
Washington, DC: National Academy
n situ bioremediation—when does it work?
ress.
12. Journel, A.G. 1 986. Geostatistics: Models and tools forthe earth sciences. Math. Geol.
18:119-140.
1 3. Hoeksema, R.J., and P.K. Kitanidis. 1 985. Analysis of the spatial structure of properties
of selected aquifers. Water Resour. Res. 21:563-572.
14. Gilbert, R.O., and J.C. Simpson. 19B5. Kriging for estimating spatial patterns of
contaminants: Potential and problems.! Environ. Monitor. Assess. 5:113-135.
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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,
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.
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, intrinsic 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 distribution of the NAPL—the plume's
source—or determining whether that source is still moving.
Introduction
Oil or another NAPL, such as gasoline or trichloroethylene (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 ds 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 discussion to the contamination issue, and aquifer heterogeneity, 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 ;n 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 la). Above the water table, in the vadose
10
Symposium on Intrinsic Bioremediation of Ground Water
-------
zone (see Figure 2), air is also present within thi
control both the movement and distribution of
forces, and gravity or buoyancy forces (1).
between the solid phase and each of the fluids.
inversely proportional to the pore size. In the va
; pore space (see Figure 1 c). Three major forces
h of the fluid phases: capillary forces, viscous
eac
Capillarity is the result of the cohesive forces within each fluid phase and the adhesive forces
\capillaryforce is proportional to the interfacial
tension at the fluid-fluid interface and the strength of fluid wetting to the solid surface, and
dose 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 la). The
exception occurs beneath a fluctuating water table, where gas bubbles may become entrapped
by capillary forces (see Figure 1 b and Figu'e 3; Figure 3 is a photomicrograph from a
visualization experiment (see also the appendix to this paper). For most aquifer materials, water
is the wetting fluid; that is, the solid aquifer ma
and the air or gas is almost always the nonwetting phase (see Figures 1 b and 1 c). Water occurs
as an interconnected film or layer of wetting
vadose zone. The nonwetting gas phase occupies the larger pores.
Viscous or flowing forces within a fluid phase
erial has a greater affinity for water than for air,
quid covering and connecting the solids in the
require an expenditure of potential energy. For
in
example, in the saturated zone the ground-voter flow rate is proportional to the hydraulic
gradient. The flow rate is also a function of tre aquifer material and the structure of its pore
space, as represented by the effective permeability. If more than one fluid phase is present, as
Figures Ib and 1 c, the interconnected paths in each phase are more tortuous. The
relative permeability closer to zero than to one,
occupies the pore space.
ho the density difference between two fluids. The
permeability of each phase is reduced, with a
the value that applies when only a single fluid
Buoyancy is a gravitational force proportional
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 watertable, 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, j/ith 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 concentration varies enough td 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 tanb. NAPLs often share the pore space with both gas and water, in the vadose zone,
or water alone, in the saturated zone. The NAP^_ 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 heavierthan air, so it easily moves downward through the vadose zone. Once it reaches
the vicinity of the watertable, 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.
11
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water air solid
•js^sN
S!s&
(a) Water Saturated (b) Residual Gas (c) Continuous Gas
Figure 1. Diagram of fluid saturation in a porous media (4).
hazardous waste site
ground surface
vapor -^ *jw-A
phase \ ''*?:** <;'
organic
VADOSE
ZONE
floating NAPL
phase liquid4$
saturation
SATURATED
ZONE
capillary fringe
•M
water table
residual
non-aqueous*'
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).
12
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Most NAPLs appear to be of intermediate wetta
jility in typical aquifer materials; that is, they are
nonwetting relative to water but wetting relative to gas. Thus, depending 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 hternal cohesion and will spread at a gas-water
interface, presumably forming a film between the water phase, which because of capillarity
preferentially occupies the smallest pores, and t le gas phase, which preferentially fills the largest
pores. In the vadose zone, this film interconne :ts 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, h this case, the sum of the interfacial tensions
between the NAPL and the water and air, <7OW + aao, exceed the surface tension between the gas
and water, aaw (see Figure 5, 0>0). These NAPLs will not spread as films. On a flat water
surface (the gas-water interface), nonspreadind 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 between the NAPL and the solid surfaces. After the NAPL moves on, it leaves
behind a residual saturation trapped by capilla -y forces, which occupies the larger pores. NAPL
blobs occupying one ortwo pore bodies are shown in Figure 7. A similar process occurs in the
vadose zone, when a rising water table entraDS air bubbles, as shown in Figures 2b and 3.
Because of lower interfacial tension, NAPL blob
forming interconnected groups of larger pore
> are often much more complex than air bubbles,
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 altered wetting and
significant wetting hysteresis. The residual NAPi. is not trapped in the pore bodies, the
characteristic location fora nonwetting fluid. Some of it is in the pore throats, some in the pore
bodies, some has an end in both. Because of
wet conditions prevailed during the invasion
he large wetting hysteresis of this surface, water
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 locations, clearly water wet, we can infer that
the interface was not disturbed as the water 'centered. In other locations, 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 facies. Many formations, even those
containing clay, are fractured and sometimes faulted. Heterogeneities provide preferential paths
for fluid and chemical migration in an interp ay of the geology 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 ;uch 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.
13
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Figure 3. Photograph of a micromodel with entrapped air bubbles. The bubbles are trapped
in the pore bodies, which are roughly 200 jum in diameter, while the pore throats are
50/im in diameter (4).
Figure 4. A micromodel after a spreading NAPL (Soltrol 130) is drained with air. Air fills the
pore bodies and is interconnected 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).
14
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Figure 5. Cross-sectional diagram of spread! ig potential for a drop of NAPL floating on the
air (gas)-water interface, after Adamson (1 0) and others. The water is wetting, the gas
is nonwetting, and the NAPL is intermediate wetting.
Water
N&BLpoefc
Figure 6. A micromodel after a nonspreading DNAPL (mineral oil) is drained with air. The
DNAPL is not connected. As it leave's the photo at the bottom, the air is surrounded
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 trunc ates somewhere in between. In this area, the air-
water interface is dimpled, with sir all lenses of mineral oil. These lenses are not
interconnected. The various pockets! of oil, such as on the right side of the figure, are
also generally discontinuous and not connected (1 1).
15
-------
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
fim across, and the model was 1 0 cm x 1 5 cm in area (8).
Figure 8. A micromodel with a large, branched, nonaqueous, nonwetting phase blob (dark
gray) in an irregular pore network (3).
16
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Figure 1 0 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 material (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
bypassed. Capillary forces are too strong, and the viscous forces in the moving watertoo 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 highe
was occurring is shown in Figure 1 1. A signifi
- water flow rate. A photograph taken while this
:ant 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 because, 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 h
dissolved plume, we would like to be able to prc
of the geology and the fluid flow history. We use mathematical models to assist with this kind
of prediction. There is a problem, however. Th
3 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
of residual.
These photographs were taken from experiment
to stratigraphic bedding planes—for instance,
story. Because the residual is the source of the
diet its location and magnitude from knowledge
issue of predicting the amount and distribution
; that simulated the movement of DNAPL parallel
as it migrates along the bottom of an aquifer.
What happens as it moves perpendicular to tie bedding, for example, during its movement
downward from the water table toward the bo torn of the aquifer? Figure 1 2 shows the same
kind of heterogeneous lenses, except for this ch Jnge 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 horizontally because of c apillary forces. As soon as it fills a lens, it begins
to spill downward, but because of the 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 finge
displacement, with lots of water bypassed by the
it is difficult to directly measure DNAPLs in the
remediation measures.
ring. The process leads to a very inefficient
DNAPL, and illustrates some of the reasons why
field, predict their behavior, or design effective
17
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Figure 9. Residual nonaqueous-phase saturation in a micromodel treated with an alkoxysilane
of proprietary composition, GlassClad 1 8 (Huls Americal Petrarch, Bristol, PA). This
treated surface has significant contact angle hysteresis. The water is light gray, and
the NAPL is darker gray in this image (9).
Figure 10. Water has swept through a heterogenous micromodel 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).
18
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Figure 1 1. 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 displacement was still underway. The bypassing is complete on the
right side of the model and has not yet occurred at the left side (3).
Figure 1 2. DNAPL is moving downward through a water-saturated heterogeneous micromodel
(9).
19
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Influence on Intrinsic Bio remediation
The distribution of fluid and solid phases controls the mass transfer and transport of chemicals
and their availability for biotransformation by the microbial population. 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-contaminated 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 interfacial forces that attract abiotic colloids to solid-liquid or fluid-fluid
interfaces. Figure 13 shows fluorescing colloids that are repulsed from the solid surface by
electrostatic 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 important 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 indirect influences on intrinsic bioremediation.
Consider mass transfer of chemical components between a multicomponent residual NAPL and
the water phase in the saturated zone. Many spilled NAPLs—from gasoline and other fossil fuels
to transformer oils containing polychlorinated 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 comparatively quickly, leaving behind less soluble components to leach
out more slowly.
There are also cosolvency effects to consider. Interaction between components can either
enhance the solubility 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 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 1 0. 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 bioremediation
can do little to speed up this mass-transfer process, except to possibly increase concentration
gradients between 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.
Consequently, it is difficult to supply the nutrients necessary for growth (3, 8).
20
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Figure 13.
Negatively charged hydrophilic pclysty
bubbles entrapped in a micromcd
surface at this ionic strength (4).
rene particles attached to the surface of gas
el. There is almost no sorption on the glass
Figure 14. A colony of bacteria (Arfhrobacfersp
after 48 hr of growth. The blob h
gray (5).
. ZALOOl) in the vicinity of an iso-octane blob
trapped in a pore body. The bacteria are light
21
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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 basic concern with the
NAPL is as a source of that contamination.
To understand the plume, we must know something about the spacial distribution of the source
and determine 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 photographs that even a small amount of liquid may be
spread 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 mafor 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 forthe 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 pathways and fluid distributions.
Intrinsic bioremediation within a NAPL-contaminated zone is probably of limited effectiveness.
Mass-transfer limitations constrain the rate at which dissolved components reach the aqueous
phase. There are toxicity concerns with the high concentrations that are found nearer the NAPL
blobs orfilms. 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 determine whether that source is still moving. It is a direct
answerto the "where" question. The "how much" question depends on the effective mass-transfer
rate between 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.
Characterization of this distribution and movement is difficult and may be beyond the current
state of the art.
22
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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 fuLed together (3, 8, 9). The resulting pores have
complex three-dimensional structures, although the network is only two dimensional. The
micromodel in Figure 3 shows pore bodies connected together by narrower pore throats. The
fused glass located 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 together and described in more detail in
Wilson (9). Greater detail is given in the original references.
Acknowledgements
The photographs presented in this paper were taken from work sponsored by the Subsurface
Science Program of the Department of Energy', or from work sponsored by the Robert S. Kerr
Environmental Research Laboratory of the U.S.
New Mexico Institute of Mining and Technology
References
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2.
3.
4.
5.
6.
Bear, J. 1972. Fluid flow in porojs media. New York: Elsevier.
Mercer, J.W., and R.M. Cohen
subsurface: Properties, models,
Hydrol. 6:107-163.
Environmental Protection Agency. M. Flinsch of
assisted with the preparation of this manuscript.
1990. A review of immiscible fluids in the
:haracterization and remediation. J. Contam.
Wilson, J.L., S.H. Conrad, E. Hagan, W.R. Mason, W. Peplinski, and E. Hagan.
1 990. Laboratory investigation of lesidual liquid organics. ReportCR-81 3571. U.S.
Environmental Protection Agency,
Wan, J., and J.L Wilson. 1 994. V
on the fate and transport of collo
23.
Ada, OK.
sualization of the role of the gas-water interface
ds in porous media. Water Resour. Res. 30:1 1 -
Wan, J., J.L. Wilson, and T.L. Kiett. 1 994. The gas-water interface as an influence
on the transport of microorganisms through unsaturated porous media. Appl.
Environ. Microbiol. 60:509-516
Miller, C.T., M.M. Poirier-McNeil
nonaqueous phase liquids: Ma
26:2783-2796.
, and A.S. Mayer. 1 990. Dissolution of trapped
s transfer characteristics. Water Resour. Res.
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7.
8.
9.
10.
11.
Powers, S.E., C.O. Louriero, L.M. Abriola, and WJ. Weber. 1991. Theoretical
study of the significance of nonequilibrium dissolution of NAPLs in subsurface
systems. Water Resour. Res. 27:463-478.
Conrad, S.H., J.L Wilson, W.R. Mason, and W. Peplinsld. 1 992. Visualization of
residual organic liquid trapped in aquifers. Water Resour. Res. 28:467-478.
Wilson, J.L. 1 994. Visualization of flow and transport at the pore level. In: Dracos,
T.H., and F. Stauffer, eds. Transport and reactive processes in aquifers. Rotterdam:
Balkema. pp. 1 9-36.
Adamson, A.W. 1 982. Physical chemistry of surfaces, 4th ed. New York: Wiley and
Sons.
Wilson, J.L. 1 992. 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. 107-1 14.
24
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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 partnership between the Department of Defense (DOD), academia, and private industry
evolves into a combined technology that can :>ave millions of dollars in long-term hazardous
waste site cleanup costs.
man
DOD has about 20,000 contaminated sites,
also may require monitoring for more than 3C
The cost of site characterization and monitoring
remediation costs.
eth
-------
This paper briefly describes the Tri-Services program for developing this cone penetrometer
technology and recent 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 characterizing contaminated DOD
sites. The Army has provided 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 Situ Soil" 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 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 electrical resistivity, pore pressure, spectral characteristics,
and other properties of the soil and contamination. The sensors provide information on
hydrogeology and contamination; the samplers verify it. The real-time ability to receive and
assess monitoring data on site without laboratory analysis is critical, facilitating decision-making
during site investigation projects while ensuring accurate and efficient completion of site
investigations and optimization of remedial activities.
Laser Spectrometer System
Each of the services has significant programs for developing 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 penetrometer 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
spectrometersystem, 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 detecting the pollutants of interest.
Optical fibers are used to transmit ultraviolet light to subsurface monitoring points and return
resulting lightforthe spectroscopic analysis. The system can identify aromatic hydrocarbons such
as benzene, toluene, and xylene (BTX) and naphthalene by their fluorescent spectra. Jet fuel,
which contains naphthalene and BTX, is the most common contaminant at Air Force sites.
26
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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 penetrometerand in the laboratory at
parts-per-billion levels for BTX in water using f beroptic probes. Laser spectroscopy technology
could also be used to monitorthe progress of site remediation and to provide baseline data for
intrinsic bioremediation modeling studies.
Technology Transition
Armstrong Laboratory and Unisys Corp. sigred a Cooperative Research and Development
Agreement (CRADA) to commercialize the Air
:orce-developed laser spectrometer system. The
laserspectrometerwas 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 proDosal to the Advanced Research Projects Agency
(ARPA), Technology Reinvestment Project (TRP)
$ 1,600,000 grant with two 1 -year options for
partners provide in-kind contributions and me
of the laser spectrometer developed for the
ARPA selected the proposal to receive a 2-year,
follow-on technology development. The industry
tching funds.
The Rapid Optical Screening Tool (ROST) is the proposed product from the commercialization
Air Force by NDSU. ROST will build upon the
previous Environics Directorate research by automating the collection and mapping of data,
making equipment components smaller and more rugged, and developing a more user-friendly
interface to allow use by environmental technicians involved in site characterization and cleanup.
ROST also has potential for process monitoring and for medical diagnostics. Initial commercial
use will be with cone penetrometers for soil characterization.
Use of the proposed ROST technology shoulc
result in substantial savings in costs associated
with characterization, monitoring, and remediation of hazardous waste sites. The participants
are committed to commercializing 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 profitable technology; the U.S. economy will
be helped; and all will benefit from a cleaner
Combined Technologies
The combined cone penetrometerand transpo
Force installations having fuel-contaminated
environment.
table laserspectrometerwas demonstrated at Air
areas. To date, the laser spectroscopy system's
primary function with the cone penetrometer h 3s been to define the oily-phase 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
wavelength appropriate for the known fuels vas utilized during the field program.
The system is designed to collect data in two
mode, laser excitation frequency is fixed,
penetrometer probe is advanced. Operation
allows collection of LIF multidimensional data
intensity, and time of decay matrices (WTM).
various fuels.
luorescence; consequently, a laser excitation
different modes: "push" or "static." In the push
and the LIF signal is monitored as the cone
in the static mode, or with the probe stopped,
sets, typically fluorescence emission wavelength,
V/TMs have proven to be very useful in identifying
27
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Results
Studies to demonstrate site amenability towards intrinsic demonstrations were conducted at
Plattsburgh, Patrick, and Dover AFBs. The sites selected for the demonstrations included closed
fire-fightertraining areas, gasoline service stations, and aircraft hydrant fueling systems. The old
fire-fightertraining areas were composed of unlined pits used to ignite fuels such as JP-4, waste
oils, and otherflammable substances. The service stations or hydrant fueling systems had leaking
underground storage tanks or piping. Substances from these operations 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®!!, a computer model
for in situ contaminant biodegradation. The cone penetrometer system 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
(fluorescence 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 measurements can be correlated
with collected samples and 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 location 84-L 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-tt intervals (25, 30, 35 ft, etc.) could easily skip over the contamination (Figure
4). A series of fluorescence versus depth profiles taken across a north-south transect at the
Plattsburgh AFB fire-fightertraining 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 developed and demonstrated within numerous
DOD, Department 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-
effective characterization before, during, and after remediation. We can use it to determine if
remediation is needed, what remediation technology we should apply, whetherthe remediation
is working, and whether the cleanup effort has been successful—all with a minimum of risk,
time, labor, and cost.
28
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100,000
10,000
1,000
c
U)
55
100
10
i i I I I I 1 1 L
10
o-
•o-
-o-
100
1,000 10,000
Concentration DFM (ppm)
Figure 1. Calibration curve for diesel fuel mcirine on fisher sand.
29
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Diesel #2
Diesel Fuel Marine
on
70
^ 60
8.50
o 40
.1 30
H 20
10
0
3(
A
)0 350 400 450 5C
IZU
100
80
60
40
20
0
)0 3(
•*•».,
)0 350 400 450 500
Wavelength (nm) Wavelength (nm)
80
70
^ 60
c 50
o 40
.1 30
H 20
10
0
3C
JP-4
'-
)0 350 400 450 5C
I
80
70
60
50
40
30
20
10
0
10 3(
Unleaded Gasoline
ite*
)0 350 400 450 500
Wavelength (nm)
Wavelength (nm)
Figure 2. Wavelength time matrices of various fuels.
30
-------
120
100
1 80
E 60
40
20
0
300
300
NTA
350 I 400 450
Wavelength (nm)
350
400
Wavelength (nm)
450
Figure 3. Wavelength time matrices at two di
Area.
500
500
ferent sites: North Tank Area versus Fuel Purge
-------
0
-------
Figure 5. Base map of the fire training area showing location of CPT soundings and wells.
Fuel Pit
84D-LIP
•p •
1
SECTION A-A-
Leading Edge of Oily Phase
J=rLIF BAF.LIP
03 GrSand
S«nd
CD Sand Mix
x Sampl*: Water L»v«l (PE)
O 1OO 2OO
VERTICAL EXAGGERATION: 6
Figure 6. Cross section showing contaminated zone and water table along section A-A' in
Figure 5.
33
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References
1. Bratton, W.L., J.D. Shinn, S.M. Timian, G. Gillispie, and R. St. Germain. 1 993. The
Air Force Site Characterization 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. 1 994.
Demonstration of the Air Force Site Characterization and Analysis Penetrometer
System (AFSCAPS) in support of the intrinsic bioremediation (natural attenuation)
option (interim report).
3. Schroeder, J.D., and S.R. Booth. 1991. Cost effectiveness analysis of the Site
Characterization and Analysis Penetrometer System (SCAPS). Los Alamos National
Laboratory, Los Alamos, NM.
34
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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
impact of toxic chemical spills and leaks. The
foremost importance. In terms of treatment pro
nost dangerous and intractable environmental
prevention of ground-water contamination is of
:esses, the present technology of excavation 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 contaminants from one environmental phase to another. Bioremediation, however,
achieves contaminant decomposition or immobilization by exploiting the existing metabolic
potential of microorganisms with novel catabolic functions derived through selection. In response
to the introduction of a toxic contaminant, an indigenous bacterial population arises that is
unique from the standpoint of physiological capabilities and species diversity.
Conditions that restrict life or inactivate microbial enzymes are incompatible with intrinsic
bioremediation efforts. Although the physical and chemical characteristics of the contaminants
and the metabolic potential of microorganisms determine the feasibility of biotransformation
reactions, actually achieving biotransformation also depends on the prevailing geochemical
conditions. The ecological constraints to bioremediation can be classified as microbial,
of these constraints, as required, assesses the
article focuses on the microbiological and
chemical, or environmental, and recognition
feasibility of intrinsic bioremediation. This
decomposition of reduced organic matter (heh
geochemical constraints influencing intrinsic bioremediation processes.
Bacterial Metabolic Diversity
Bacteria comprise a large and diverse group :>f microorganisms that obtain their energy from
a variety of sources, including 1) light (j/iotosynfhef/c bacteria), 2) the oxidation or
-otrophic bacteria), or 3) the oxidation of reduced
inorganic compounds (autotrophic bacteria). Some bacteria derive energy from more than one
source, such as combinations of light and reduced inorganic or organic compounds; however,
heterotrophic bacteria are the major group
compounds.
responsible for the biodegradation of organic
All living organisms must generate reducing
systems and maintaining the oxidation-
oxidized compounds by the addition of elect;
energy production. The electron acceptor can
For many bacteria, most fungi, and all higher
process termed aerobic respiration. The final
and the final oxidized compound respired
from
power for the purpose of replenishing enzymatic
reducfon power cycle. This involves the reduction of
ons released from compounds oxidized during
either an organic or an inorganic compound.
organisms, the final electron acceptor is O2 in the
duced substrate in aerobic respiration is H2O,
energy production is CO2.
Symposium on Intrinsic Bioremediation of Ground Water
35
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In the absence of O2, certain bacterial populations respire other less oxidized inorganic
compounds (inorganicrespiration) or use only organic compounds (fermenfaf/onj.Denitrification
is a process promoted by bacteria that can thrive under either aerobic or anaerobic 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-reduction potentials within soils are even lower, other inorganic compounds are used
by specific groups of bacteria as terminal electron acceptors. Because O2 is toxic to these
bacteria, they are called obligate anaerobes. 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. Microorganisms catalyze many of
these reactions and play an essential role in maintaining the electron balance of complex
ecosystems. For a contaminant to be incorporated 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 electrons—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
dehalogenated. The halogenated compounds must compete with other physiological electron
acceptors in order to be incorporated into microbial energy cycles. Thgs, the effectiveness :of
reductive dehalogenations is often influenced by the presence of other electron acceptors, such
as NO3" or SO42". This effect would explain why reductive dehalogenations are more frequently
observed under methanogenic conditions, where generally a paucity of these onions exist.
Bacteria-Contaminant Interactions in Aquifer Material
A minimum of three conditions must be met before a contaminant can be degraded or
transformed by bacteria: 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 capacity
to participate in some part of the degradation or transformation process. Specific bacterial
populations preferor require particularenvironmental 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 components of organic
36
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compounds, and sulfur is generally available ir
limiting elements for growth are nitrogen and
ratio (C/N/P) usually considered ideal is 300
vary depending the nature of the contaminant
sufficient quantities for growth. Thus, the major
phosphorus. The carbon/nitrogen/phosphorus
fo 1 00:10:1 to 0.05 (1); however, this ratio can
(s).
pH and Redox Potential
Near-neutral aquifer pH values are usually
organic material. The hydrogen ion concentrat
of compounds produced by bacterial activity, a
equilibrium rates (2). Because hydrogen ion transfer is
pH and redox potential are interdependent (3
O3timum for the biodegradation of contaminant
on of the ground water is governed by the types
id is controlled especially by CO32":HCO3":CO2
commonly involved in electron transport,
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 conditions tend
to favor different electron acceptors, with the most oxidized compounds (higher Eh) being used
first. The resultant scheme is NGy (denitrification) utilization after O2 depletion, Fe3+ (iron
reduction) utilization after NGy 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 degrade ive potentials can be operative at different times
at the same contamination site as the redox potential varies.
Although Eh measurements can provide valuaDle clues about the functioning of geochemical
systems in aquifer material, they alone will not give definite information about the chemical
species present. These measurements indicate
such as those involving bacterial respiration.
hat a potential exists for certain redox reactions,
Various organic and inorganic redox reactions
activity and rates of organic matter decompos
temperatures, however, bacteria are generc
temperatures (4). Perhaps more importantly,
cannot be predicted because of their complexit/ and different interdependent reaction rates (2).
Temperature
Ground-water temperature is often one of the most important factors controlling microbial
tion. Generally, rates of enzymatic degradation
and bacterial metabolism double for every 10°C increase in temperature until close to inhibitory
temperatures, which are usually around 40°C to 50°C for most bacteria. Except for subfreezing
lly capable of degradation at most ambient
•emperature 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 increases the equilibrium vapor
concentration, resulting in an increased vola
sorption to aquifer particles (5).
Physical Deterrents to Biodegradation
ilization rate, but it can at times also increase
Physical or physicochemical factors can also effect the biodegradation of contaminants. Some
molecules are recalcitrant to degradation because they are too large to enter bacterial cells,
which is usually required for complete degradation by membrane-bound enzymes. Some
substances are difficult to biodegrade because the number, length, or location of functional
groups impede enzyme attack. Strong sorption
of bacteria to attach to, absorb, or enzymatically attack the molecule (6).
on aquifer material can greatly hinderthe ability
37
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Sorption and solubility of organic contaminants are complex interdependent phenomena that
vary with the composition of the aquifer material and complex contaminant 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 forthis, the solubility
of each component of mixtures can vary from ideal conditions, with each compound acting as
a cosolventto increase hydrophobic hydrocarbon solubilities (8). Organic solvents can also
affect the sorption of organics on soils in general (9).
Contaminant Metabolism: Aerobic Versus Anaerobic
A common misconception is that all organic contaminants are biodegraded most rapidly and
thoroughly under aerobic conditions. Although this is commonly the case, anaerobic conditions
promote some very important 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 conditions (4); however, current research efforts have shown that the biodegradation of
many monoaromatic compounds common to most fuels also occurred in the laboratory under
anaerobic conditions. This biodegradation readily occurs not only with NO3" serving as the
terminal electron acceptor (1 1) but also under SO42' reducing conditions (12), Fe3+ reducing
conditions (13), and methanogenic conditions (14,15).
Benzene has been especially recalcitrantto anaerobic biodegradation in laboratory studies under
denitrifying and sulfate-reducing conditions (1 0,12). Yet some laboratory (1 6) and field studies
(1 7) have shown the depletion of all common monoaromatic hydrocarbons found in gasoline
under denitrifying conditions. In 1 986, Vogel and Grbic-Galic (1 8) demonstrated that benzene
and toluene were degradable to CH4 and CO2 in laboratory microcosms; however, the
confirmation of the degradation of benzene in field situations under methanogenic conditions
has been rather elusive. One of the few well-documented instances of the methanogenic
degradation of benzene in the field is a crude oil spill in Bemidji, Minnesota (1 9), where many
of the water-soluble monoaromatic hydrocarbons present in crude are undergoing intrinsic
bioremediation.
Conclusion
Although bioremediation in general has gained considerable 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 bioremediation
has resulted in unrealistic expectations, leading in turn to disappointments and ultimate failures.
Continued research into and application of sound bioremediation schemes will undoubtedly
prove the viability of intrinsic bioremediation in the overall remediation efforts of contaminated
ground water.
38
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References
1. Torpy, M.F., H.F. Stroo, and G. Bruefc
2.
3.
4.
5.
6.
7.
8.
9.
11.
12.
13.
wastes. Poll. Eng. 21:80-86.
aker. 1 989. Biological treatment of hazardous
Stumm, W., and J.J. Morgan. 1981. Aquatic chemistry. New York, NY: John Wiley &
Sons.
Grundi, T. 1 994. A review of the currsnt understanding of redox capacity in natural
disequilibrium systems. Chemosphere ~~
Leahy, J.G., and R.R. Colwell. 1990.
environment. Microbiol. Rev. 54:305-
28:613-626.
Microbial degradation of hydrocarbons in the
315.
Lyman, W.J., W.F. Reehl, and D.H. Rosenblatt. 1982. Handbook of chemical property
estimation methods: Environmental behavior of organic chemicals. New York, NY:
McGraw-Hill.
Cheng, H.H., K. Haider, and S.S. Harper. 1 983. Catechol and chlorocatechols in soil:
Degradation and extractability. Soil Bipl. Biochem. 15:31 1-317.
Cline, P.V., J.F. Delfino, and P.S.C. Rod. 1 991. Partitioning of aromatic constituents into
waterfrom gasoline and other complex solvent mixtures. Environ. Sci. Technol. 25:914-
920.
Groves, F.R. 1988. Effect of cosolve its on the solubility of hydrocarbons in water.
Environ. Sci. Technol. 22:282-286.
Fu, J.-K., and R.G. Luthy. 1 986. Effect of organic solvent on sorption of aromatic
solutes onto soils. J. Environ. Eng. 1 12:346-366.
10. Kuhn, E.P., 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. Ill -1 80.
Evans, P.J., D.T. Mang, and L.Y. Yourjg
and transformation of o-xylene by
Microb. 57:450-454.
^. 1991. Degradation of toluene and m-xylene
cenitrifying enrichment cultures. Appl. Environ.
Edwards, E.A., LE. Wills, D. Grbic-Galic, and M. Reinhard. 1991. Anaerobic
degradation of toluene and xylene: Evidence for sulfate as the terminal electron
acceptor. In: Hinchee, R.E., and R.F. Olfenbuttel, eds. In situ bioreclamotion:
Applications and investigations for hydrocarbon and contaminated site remediation.
Boston, MA: Butterworth-Heinemann.
pp. 463-471.
Lovely, D.R., and DJ. Lonergan. 1 990. Anaerobic oxidation of toluene, phenol, and p-
cresol by the dissimilatory iron-reducing organism, GS-15. Appl. Environ. Microb.
56:1,858-1,864.
39
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14. Grbic-Galic, D., and T.M. Vogel. 1987. Transformation of toluene and benzene by
mixed methanogenic 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 methanogenic aquifer
material: A microcosms study. Environ. 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.
1 7. Berry-Spark, K.L., J.F. Barker, D. Major, and C.I. Mayfield. 1 986. Remediation of
gasoline-contaminated ground-waters: A controlled experiment. In: Proceedings of
petroleum hydrocarbons and organic chemicals in ground water: Prevention, detection,
and restoration, NWWA/API. Dublin, OH: Water Well Journal Publishing.
1 8. Vogel, T.M., and D. Grbic-Galic. 1 986. 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 MJ. Baedecker. 1990. Transformation of
monoaromatic hydrocarbons to organic acids in anoxic ground-water environment.
Environ. Geol. Water Sci. 16:135-141.
40
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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 contaminants and the
geochemical processes that result from their bresence in the subsurface. A large number of
contaminated sites contain petroleum-derived hydrocarbons. Many of the hydrocarbons are
biodegraded in ground-water environments, and the extent of their removal by natural processes
has been evaluated in field and laboratory investigations. Both types of investigations
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 .a re 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 contaminants is petroleum-derived hydrocarbons
from pipeline breaks, leaking storage tanks, spills, and disposal of wastes. Many of these sites,
such as those contaminated 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 remediation strategies to deal with petroleum-derived hydrocarbons. 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 bacteria were found and quantified in soil and ground water at
a fuel-oil contaminated site (11). Degradation of petroleum-derived hydrocarbons is generally
considered to occur more rapidly in aerobic or suboxic environments, where oxygen or nitrate
is available as an electron acceptor (7,12-1 3)j but anaerobic biodegradation also may remove
significant amounts of hydrocarbons from ground water (14-17).
An investigation of the effects of a crude-oil s
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 carrying crude oil ru Dtured and sprayed oil over land surface. Part of
the oil was removed during remediation, but part of it infiltrated through the unsaturated zone
and accumulated.as an oil body floating on tne ground water. A more detailed description of
the processes that occurred at the site is avai
Eganhouseetal. (9), Baedecker and Cozzarell
etal. (21).
able in Baedecker et al. (5), Bennett et al. (1 8),
et al. (19), Cozzarelli et al. (20), and Eganhouse
Symposium on Intrinsic Bioremediation of Ground Water
41
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The concentrations of benzene, toluene, ethyibenzene 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
decreased by several orders of magnitude (56 m to 1 37 m). The mass of BTEX lost near the oil
body (26 m to 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 transported to the water table. Degradation of
hydrocarbon vapors in the unsaturated zone most likely consumed the oxygen by aerobic
respiration (23). The downgradient movement of the BTEX plume averaged about 8 m/yrfrom
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 fartherfrom the oil body, additional processes such as dispersion, mixing (24), and
sorption become more important.
ea
a
a.
O)
.g
£
c
0>
o
o
O
100
10
1.0
0.1
0.01
0.001
0.0001
BTEX
50 100 150
meters downgradient from oil body
200
Figure 1. Concentrations of benzene, toluene, ethyibenzene, and o-, m-, and p-xylene (BTEX)
in the upper 1.5-m thickness of an aquifer contaminated with crude oil (1 9).
42
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Laboratory experiments demonstrated that these hydrocarbons can degrade under controlled
laboratory conditions. 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 concentration by 98 p srcent in 1 25 d, and toluene decreased by 99
percent in 45 d (5). No loss of naphthalene
controls, no loss of benzene, toluene, or
experiments and field results provide strong e
was observed during the same period. For the
naphthalene was observed. These laboratory
/idence 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 eliminated as possible
explanations for the loss of benzene and toluene.
Field and laboratory studies of anaerobic bictransformation or aromatic hydrocarbons were
reviewed by Barker and Wilson (25) who found evidence at five methanogenic sites for
biodegradation of benzene,toluene, ethylbenzene,and thexylenes. 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 concentrations of benzene are biodegraded in the subsurface (27), yet in
some ground-water environments benzene is the most persistent hydrocarbon among the
monoaromatics.
Natural processes can remove significant cc ncentrations of hydrocarbons and prevent the
spreading of a plume. At sites where the rates of solubilization, volatilization, and
biodegradation of hydrocarbons are such thai 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 a re attenuating contaminants provides information that can be used to accelerate
biodegradation processes.
Note
The data and interpretation in this report were previously published in Baedeckerand Cozzarelli
(1 9) and Baedecker et al. (5).
References
1. Baedecker, M.J., and W. Back. 1
reactions at a landfill. Ground Water
2. Nicholson, R.V., J.A. Cherry, and E.J
ground water at a landfill: A case study
979
. Hydrogeological processes and chemical
17:429-437.
Reardon. 1983. Migration of contaminants in
6. Hydrogeochemistry. J. Hydrol 63:1 31 -1 76.
Lesage, S., R.E. Jackson, M.W. Priddle, and P.G. Riemann. 1 990. Occurrence and fate
of organic solvent residues in anoxic ground water at the Gloucester landfill, Canada.
Environ. Sci. Technol. 24:559-566.
43
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4. Godsy, E.M., D.F. Goerlitz, and D. Grbic-Galic. 1 992. Methanogenic biodegradation
of creosote contaminants in natural and simulated ground-water ecosystems. Ground
Water 30:232-242.
5. Baedecker, M.J., I.M. Cozzarelli, R.P. Eganhouse, D.I. Siegel, and P.C. Bennett. 1993.
Crude oil in a shallow sand and gravel aquifer, III. Biogeochemical reactions and mass
balance modeling in anoxic ground water. App. Geochem. 8:569-586.
6. Barker, J.F., J.S.Tessman, P.E. Plotz,andM. Reinhard. 1 986. The organic geochemistry
of a sanitary landfill leachate plume. Contam. Hydrol. 1:171-1 89.
7. Major, D.W., C.I. Mayfield, and J.F. Barker. 1 988. 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. 1 993. Crude oil inashallowsand 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. 1 991. Microbiological 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. Schwarzenback. 1 988. 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. 1 991. Biodegradation of
aromatic hydrocarbons by aquifer microorganisms underdenitrifying conditions. Environ.
Sci. Technol. 25:68-76.
14. Wilson, B.H., G.B. Smith, and J.F. Rees. 1986. Biotransformations of selected
alkylbenzenes and halogenated aliphatic hydrocarbons in methanogenic aquifer
material: A microcosm study. Environ. Sci. Technol. 20:997-1,002.
15. Grbic-Galic, D., and T.M. Vogel. 1987. Transformation of toluene and benzene by
mixed methanogenic cultures. Appl. Environ. Microbiol. 53:254-260.
16. Lovley, D.R., M.J. Baedecker, DJ. Lonergan, I.M. Cozzarelli, E.P. Phillips, and D.I.
Siegel. 1 989. Oxidation of aromatic contaminants coupled to microbial iron reduction.
Nature 339:297-299.
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17. Haag, P.M., M. Reinhard, and P.L
18.
20.
21
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26.
McCarty. 1991. Degradation of toluene and
p-xylene in anaerobic microcosms: Evidence for sulfate as a terminal electron acceptor.
Environ. Toxicol. Chem. 10:1,379-1,390.
Bennett, P.C., D.I. Siegel, M.J. Baedec
sand and gravel aquifer, I. Hydrology
8:529-549.
cer, and M.F. Hult. 1 993. Crude oil in a shallow
and inorganic geochemistry. Appl. Geochem.
19. Baedecker, M.J., and I.M. Cozzarelli.
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substances and the hydrologic scieno
994. Biogeochemical processes and migration
voter contaminated with crude oil. In: Toxic
s. American Institute of Hydrology, pp. 69-79.
Cozzarelli, I.M., M.J. Baedecker, R.P. Eganhouse, and D.F. Goerlitz. 1994. The
geochemical evolution of low-molecular-weight organic acids derived from the
degradation of petroleum contaminan
58(2):863-877.
3 O
Eganhouse, R.P., M.J. Baedecker, and
in an aquifer contaminated by crud
Minnesota, research site. Presentee
Bioremediation of Ground Water, Denver,
. Aiken
Cozzarelli, I.M., M.J. Baedecker, G.
heterogeneities in a crude-oil-con
Morganwalp, D.W., and D.A. Aronsot
the U.S. Geological Survey Toxic Substances
(September 20-24, 1 993). Water Res
Hult, M.F., and R.R. Grabbe. 1988.
the unsaturated zone. In: Ragone, S.E.
water contamination study. Proceeding
Survey, Cape Cod, MA (October
C21-C26.
21
24. Essaid, H.I., M.J. Baedecker, and I .A/
s in ground water. Geochim. Cosmochim. Acta
M. Cozzarelli. 1 994. Biogeochemical processes
An overview of studies at the Bemidji,
at the U.S. EPA Symposium on Intrinsic
CO (August 30 to September 1).
, and C. Phinney. 1 994. Small-scale chemical
aminated aquifer, Bemidji, Minnesota. In:
i, eds. Proceedings of the Technical Meeting of
Hydrology Program, Colorado Springs, CO
Invest. Rep. 94-4014. In press.
pistribution of gases and hydrocarbon vapors in
ed. U.S. Geological Survey toxic waste/ground
; of the Technical Meeting of the U.S. Geological
-25, 1985). Open-File Report 86-481. pp.
. Cozzarelli. 1994. Use of simulation to study
field-scale solute transport and biodegiadation at the Bemidji, Minnesota, crude-oil spill
site. In: Morganwalp, D.W., and D.A. Aronson, eds. U.S. Geological Survey toxic
substances hydrology program. Proceedings of the Technical Meeting of the U.S.
Geological Survey Water Resources, Colorado Springs, CO (September 20-24,1 993).
Water Res. Invest. Rep. 94-4014. In press.
Barker, J.F., J.T.Wilson. 1992. Natural
under anaerobic conditions. Proceed!
Dallas, TX. pp. 57-58.
Grbic-Gdlic, D. 1 990. Anaerobic
and alicyclic compounds in soil, subsu
and G. Stotzky, eds. Soil Biochemistry
117-189.
biological attenuation of aromatic hydrocarbons
ngs of the Subsurface Restoration Conference,
micrbbial transformation of nonoxygenated aromatic
face, and freshwater sediments. In: Bollag, J.M.,
Vol. 6. New York, NY: Marcel Dekker, Inc. pp.
45
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27. Hadley, P.W., and R. Armstrong. 1991. Where's the benzene? Examining California
ground-water quality surveys: Ground Water 29(1):35-40.
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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,
00
sthod
In situ bioremediation is a cost-effective method for the remediation of soils and ground waters
contaminated with petroleum hydrocarbons. T iis paper presents a case study of the successful
application of in situ bioremediation at a pub
tie
The site was used as a truck maintenance
nonvolatile petroleum hydrocarbons from used motoroil
fluids were released to a used oil sump at
claystone and sandstone bedrock, covered with
gravels, as well as silty sands and clays.
interbedded alluvial sands and gravels.
ic utility site in Denver, Colorado.
The
acility for almost 30 years. During that time,
il, diesel, gasoline, and other automotive
site. The local geology of the site includes
recentdeposits of interbedded alluvial sands and
principal aquifer at the site lies within the
Soil and water sampling confirmed the presence of petroleum hydrocarbons in both the vadose
and saturated zones. Laboratory studies showed that the chemical, microbiological, and
hydrogeological characteristics of the site werj 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 bio-
degradation of contaminants and the mass tra
indicated that the primary contaminant at the
isport 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
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 soi
and sediment samples to determine the
biodegradability of contaminants under variou; nutrient loads and aerobic conditions. Aerobic
testing was performed to simulate optimal conditions for the bioremediation of hydrocarbons.
Column 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 ana 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
he site.
Symposium on Intrinsic Bioremediation of Ground Water
47
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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 nutrients. The primary functions of the remediation system included ground-water
recovery, treatment, and reinjection; 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.
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 system. Crews conducted field analyte tests, sampled ground water, and
gauged monitoring wells. The bioremediation 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 nutrients, and reinfiltrated. The site is currently undergoing closure with the
Colorado Department of Health.
Reference
For more information about the Denver site and supporting data, refer to:
Nelson, C.H., R.J. Hicks, and S.D. Andrews. 1 994. 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.
48
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The Role of intrinsic Bioremediotion 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 numbe r 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 correspondence generated at every site; instead, we prefer the
technical interpretation. Ourapproach to site evaluation is to reviewthis 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) occup
abridged version of this site presented a daunting task.
I needed to get up to speed quickly on the file
to verify if an approved corrective action plan
had been placed on this approval, if any. In a
been documented, we check to see if the poin'
ed 6 feet of shelf space. Even a review of the
o consider this site for closure. My first task was
was in place and to determine what conditions
iy case where ground-water contamination has
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
•elease. Based on the calculated ground-water
velocity, was the monitoring period long enoug h to detect 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 th
the next topic for examination prior to closure
and its potential for adversely affecting the environment,
identified receptors in the immediate vicinity an|d
by the remaining contamination.
3 closure borings had been drilled and analyzed,
was the extent of the remaining contamination
This information is considered with the
the potential for these receptors to be affected
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
remaining in the subsurface; 2) the closure
submitted; 3) similar to the owner's liability
rom any liability regarding the contamination
decision is based solely on the information
issue, the CDH wants to be informed of any
developments that could increase the potent al 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
proposing such technologies should be prepared
implemented in a manner that does not adve 'sely
Proposal of an innovative technology should b
innlovative remedial technologies. Those who are
to demonstrate that these technologies can be
affect human health and the environment.
3 approached from the viewpoint of a regulator:
Symposium on Intrinsic Bioremediation of Ground Water
49
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consider receptors and aspects of the technology that might be detrimental to the environment,
and address these concerns in a responsible, forthright manner.
50
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The Importance of Knowledge About Intr
Closure: The Client's Perspective
nsic Bioremediation for Cost-Effective Site
Harry E. Moseley
Public Service Company of Colorado, Denvei
, CO
Public Service Company of Colorado (PSCol is Colorado's largest electric and gas utility.
During an underground storage tank removal and replacement project in 1 987 at PSCo's
Seventh Avenue Service Center, a facility 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/n situ bibremediation 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 1 8 months of ground-water monitoring, the site received approval for final closure
in March 1 994, 7 years after discovery of the contamination.
PSCo, a business without extensive expertise in
with a fairly direct goal: to remediate the si
Intrinsic bioremediation was found to be a key
was known about intrinsic bioremediation, anc
additional knowledge of these fields might
For example, in investigating whether biorem
enhancing the activities of the bacteria that
nvironmental restoration, undertook the project
e as efficiently and cost-effectively as possible.
element of the project's success. In 1 988, little
some opportunities may have been lost; indeed,
have reduced costs.
sdiation was feasible, emphasis was placed on
vould perform the remediation, rather than on
determining how long the enhancement COL Id 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 halve been more effective for the sustained growth
of the bacteria than the continuous recirculation of the enriched ground water. If this fact had
been considered before the start of the project, operation and maintenance 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 monijtored the concentration of the injected chemicals,
or "nutrients"; these tests necessitated 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 reduc
bioremediation process would have reduced
ed to quarterly, but a better understanding of the
hese requirements earlier.
Symposium on Intrinsic Bioremediation of Ground Water
51
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The bioremediation process came in on budget and successfully removed all threat to human
health and the environment (see Table 1).
Table 1. Resources Used at the Seventh Avenue Bioremediation Project
1990
1991
1992
1993
Labor hours
Activities
Testing cost
Materials cost
Labor hours
Activities
Testing cost
Materials cost
Labor hours
Activities
Testing cost
Labor hours
Activities
Testing (quarterly)
2,024 hours total
1 68 hours per month average
O&M costs
Quarterly and monthly reporting
$12,778
$16,473
1 ,746 hours total
1 45 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
551 hours total
46 hours per month average
Quarterly reports
Monthly site visits
Closure report submittal
$11,539
94.25 hours total
7.85 hours per month average
Closure monitoring
$8,420
Information from past projects, sound scientific judgment, and risk analyses will provide an
understanding of intrinsic bioremediation that could increase the number of cleanups, speed up
the closure process, and reduce testing and reporting requirements, thereby substantially
reducing costs to industry.
52
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The Role of Intrinsic Bioremediation in C
In Situ Bioremediation: The Role of Mat
osure of Sites After Cleanup Through
lematical Models
Tissa H. Illangasekare and David C. Szlag
Department of Civil, Environmental, and Arch
University of Colorado, Boulder, CO
John T. Wilson
Robert S. Kerr Environmental Research Laborctory,
U.S. Environmental Protection Agency, Ada, OK
ectural Engineering,
Abstract
This pa per discusses the important processes ir
tools for the design and evaluation of
contaminated with nonaqueous phase orgc
application of a mathematical model that
methodology for retrospective evaluation of b
Introduction
developing mathematical models to be used as
grsund-water remediation schemes in aquifers
nic chemicals. The paper also presents the
considers these processes in developing a
oremediated sites.
Mathematical models of water flow and chemical transport have been extensively used in
ground-water quantity and quality managem
involve bioremediation. The basic processes
snt. Some of the applications of these models
of flow of water and transport of dissolved
substances are fairly well understood, and numerical models for the solution of the governing
equations have been developed. Field applications of models as prediction arid design tools,
however, have not been very successful for many reasons, including the complexities associated
with natural heterogeneities and the inadequacy of available field techniques for physical and
chemical characterization; this is especially
involving chemicals in the form of separate
true in ground-water contamination situations
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 adequatelj 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 bioremediation, pockets of NAPLs remained in the soil. These entrapped fluids have
Symposium on Intrinsic Bioremediation of Ground Water
53
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the potential to contribute to ground-water pollution after active remediation has been
discontinued.
In our research, we have identified two modes of entrapment: microscale and macroscale.
Microscale entrapment, 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, regardless 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 quantify mass transfer.
Laboratory experiments were conducted to obtain a fundamental 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 investigation 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 entrapment, 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 depends on
their ability to accurately represent the above processes. A case study conducted in Colorado
identifies some of these basic processes that are of importance 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 misleading. Monitoring wells may show very low or zero levels of contaminants after active
bioremediation, but levels may increase overtime. 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
bioremediation. This is a conservative assumption and generally leads to target levels of total
petroleum hydrocarbon (TPH) concentrations on the order of 10 to 1 00 mg/kg aquifer material.
Many bioremediation schemes, however, 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
answers 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 expected? 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.
54
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Problem Description
A temporary 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 mineral nutrients, and bioventing. An estimated 2,147 Ib of hydrocarbons have
been removed by pumping hydrocarbon emulsion from monitoring wells or by volatization
through soil aeration. In the treatment system,
recovery well was treated and amended with
he ground water pumped from a downgradient
organic nutrients and hydrogen peroxide. The
solution was then injected into the subsurface system upgradient of the contaminated site. The
system operated from October 1989 to March 1992. Table 1 compares the reduction in
concentration of benzene and total BTEX compounds in ground water achieved by this
remediation scheme (3, 4).
Table 1. Reduction in Concentration (ug/L) of Hydrocarbon Contaminants in Ground Water
Achieved by In Situ Bioremediation
Well
MW-1
MW-8
MW-2A
MW-3
RW-1
Benzene
Before
220
180
—
11
<1
During
<1
130
n
5
2
AfteV
<1
16
0.4
2
<1
Total BTEX
Before
2,030
1,800
—
1,200
<1
During
164
331
1,200
820
2
After
<6
34
13
46
<1
Water from the monitoring wells and the rec
contaminants by March 1 992. Active remediati
of postremediation monitoring. In June 1 992
determine the extent of hydrocarbon rema
contamination could return once active remedi
downgradient of the release. The cores we
hydrocarbons and for the concentration of in
boreholes with the highest concentrations of h
sampling, the elevation of the water table wa
rculation well contained low concentrations of
Jon was terminated, and the site entered a period
), core samples were taken from the aquifer to
ning and to determine whether a plume of
ition ceased. The site was cored along a transect
B extracted and analyzed for total petroleum
ividual BTEX compounds. Data from one of the
drocarbons are given in Table 2. At the time of
5,280.5 ft above mean sea level (AMSL).
55
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Table 2. Vertical Extent of Total BTEX and TPH (mg/kg) at a Borehole
Elevation (Feet AMSL)
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
TPH
<44
227
860
1,176
294
273
<34
<24
BTEX
<1
5.1
101
206
27
7.4
<1
<1
Benzene
<0.02
<0.2
<0.2
4.3
0.68
0.26
<0.2
<0.2
Color and Texture
Brown sand
Brown sand
Brown sand
Brown sand
Brown sand
Brown sand
Brown sand
Brown/yellow sand
To hydradically characterize the aquifer, selected soil cores that had previously been used to
measure hydrocarbon 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 hydrocarbon in the form of nonaqueous fluids
moved into preferential flow channels created by the local heterogeneities. After direct recovery,
the fluids remained entrapped 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 remediation scheme. The results of this study are reported in Szlag et al. (3).
Model Selection
The following observations at the site and in the laboratory indicated the need for three-
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.
56
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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 1 8 ft thick. The LNAPL organic contaminant is confined to the uppertwo layers.
Data for the period June 8, 1 989, to April 1,
standard deviation of the mean residual. The b<
990, were used for calibration. The goodness
of fit between the model and measured well cata was characterized by a mean residual and
st 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 cetermine 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 onl^ the upper two layers would see the highest
contaminant concentrations. A pumping well screened over the entire aquifer thickness is also
being considered; in this case, however, dilution will play a major role in reducing the maximum
concentrations. Two significant assumptions a B 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
MODFLOW, ground-water velocity through ea
BTEX concentration was then used to calculate t
residual NAPL. Using the heads generated by
ch source cell was calculated. The known NAPL
ie equilibrium concentration and, consequently,
the mass flux. Estimated benzene mass fluxes wsre 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 spec
of a NAPL spill, the source function will consi
quickly reached in spill scenarios if ground-wa
ification of the contaminant source. For the case
>t of a continuous mass flux of solute from the
residual NAPL phase to the aqueous phase. A/any researchers have shown that equilibrium is
er 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 conservative, as greater mass flu
tes cannot be achieved. The water flux can vary
significantly in the source zone, which often gives misleading indications that the contaminant
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 so
homogeneous. In general, the solute plume wil
ute plume emanating from a NAPL source is not
consist of subplumes at different depth intervals
and widely different concentrations, and moving at different velocities. A regulatory question
I I
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
57
-------
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 tremendous 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 existing monitoring wells at the site are screened overthe top
5 ft of the aquifer. The maximum concentration achieved in the well screened overa 5-ft interval
reaches a steady-state concentration of 26 ppb. If that well is screened overthe entire saturated
thickness, a concentration of 15 ppb is achieved. Even greater dilution 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-dimensional 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 formations in the unsaturated
zone. This hypothesis is supported 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 forthe 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 heterogeneities. Even though hydraulic conductivity
values determined in the laboratory on disturbed soil samples were used in this study, a more
appropriate characterization method would be well-designed bail tests (orslug 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.
References
1. Szlag, D., and T.H. Illangasekare. 1994. Quantification of residual entrapment of
nonaqueous phase organic fluids in soils. Ground Water. In review.
58
-------
2.
3.
Illangasekare, T.H., D. Szlag, J. Camp sell, J. Ramsey, M. Al-Sherida, and D.D. Reible.
1 991. 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: Kanscs State University.
Szlag, D.C., T.H. Illangasekare, and I.T. Wilson. 1993. Use of a three-dimensional
ground-water model for retrospectve evaluation of a bioremediated aquifer
contaminated with organic chemicals. Proceedings of the Ground-Water Modeling
Conference, Golden, CO (June 10).
4.
Wilson, J. 1 993. Retrospective perf
characterization. In: U.S. EPA.
Research, development, and field
Washington, DC (May).
orrnance evaluation on in situ bioremediation: Site
Symposium on bioremediation of hazardous wastes:
'evaluations (abstracts). EPA/600/R-93/054.
Acknowledgments
The support of the U.S. Environmental Protection
Research Center at Kansas State University (agreement
We would also like to thank Ms. Lisa Weers
assistance.
Agency through the Hazardous Substance
R-81 5709) is gratefully acknowledged.
of the Colorado Department of Health for her
59
-------
Intrinsic Bioremediation of JP-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 discharges
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 apparent attenuation of contamination in water from monitoring wells that are distant
from the spill. These plans were often criticized because it was impossible to distinguish
attenuation due to contaminant destruction from attenuation 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 difficult. 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 attenuation 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
dilution and sorption of hydrocarbons in ground water. Cozzarelli et al. (1) used 1,2,3,4-
tetramethylbenzeneto 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 etal. (2) used 2,3-dimethylpentaneto normalize the concentrations of BTEX compounds
in ground water contaminated with gasoline from an underground 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 compounds and for total
petroleum hydrocarbons. Partitioning theory was used to predict the concentration of individual
60
Symposium on Intrinsic Bioremediation of Ground Water
-------
hydrocarbons in ground water in contact with
were compared with concentrations in water
water in the plume was in sorptive equilibriurr
To prove that the attenuation was due to bio
fhe core material. The predicted concentrations
from monitoring wells to determine if the ground
with the spilled fuel.
ogical activity, the geochemistry of the ground
3olism of petroleum hydrocarbons has
water was examined. Microbial meta
predictablegeochemical consequences. The hydrocarbons can be respired, resulting in the
consumption of oxygen, nitrate, sulfate, or irsn II minerals in the aquifer matrix and in the
production of water, dinitrogen, sulfide, or iron
microbial respiration as electron acceptors. A
production of methane. Simple stoichiometry
I. Microbiologist often referto the substrates for
kylbenzenes can be fermented, resulting in the
can be used to predict the quantity of electron
acceptors consumed or the quantity of methane produced during biotransformation of a given
concentration of petroleum-derived hydrocarbons.
Intrinsic Bioremediation of Ground Watei
POL is located over sands and silty peats chan
of the study, a plume of contaminated grounc
discharged to a small creek approximately 300
water table is approximately 8.4 ft above sea
table in the creek is 1.4 ft above mean sea leve
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
cteristic of a barrier island complex. At the time
water moved away from the residual JP-4 and
ftdowngradient (Figure 1). The elevation of the
level in the area with residual JP-4. The water
. Hydraulic conductivity determined by pumping
tests in monitoring wells varied from 48 ft to 1 02 ft per day. Based on these data, and assuming
an effective porosity of 30 percent, the residerce time along the flow path from the spill to the
creek is on the order of 1 0 wk. Water samples
screen. Seasonally, the temperature of ground
the pH varies from 5.6 to 6.7. Samples prod
1 993, when ground-water temperatures varied from 24°C to 28°C.
were acquired with a geoprobe, using an 1 8-in.
water at the site varies from 1 9°C to 28°C, and
jcing the data in Table 1 were taken in August
Correcting for Dilution at Eglin AFB
Table 1 presents the changes in conceri
trimethylbenzenes (TMB) along the flow path
rations of BTEX compounds and the three
from the spill to the creek (Figures 1 and 2).
Sample 80H-3 is from a location just outside 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 1 50 ft and 300 ft downgradienlj 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 .concentration of these compounds was found to be remarkably uniform.
Concentrations under the creek were 36, 27, and 46 percent of the concentrations near the
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.
61
-------
®PL-2
.
/
A - A'
SOIL BORING
2" MONITORING WELL
LINE OF EQUAL
WATER TABLE ELEVATION
(FEET ABOVE MEAN SEA LEVEL)
UNE °F HYDROGEOLOGIC SECTION
SCALE
rm
, T
FIGURE 1
Water Table Elevation Map
Showing Sampling Locatlona and
Flow Path of Ground Water
from JP-4 Spill to Point
of Surface Water Discharge.
EGLIN AIR FORCE BASE
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.
A
North
-20
LEGEND
X Sample Location
-sr_ Water Table
A'
South
r '5
. 0 E
-20
FIGURE 2
Hydrogeologlc
Section A-A'
EGLIN AIR FORCE BASE
Figure 2. Flow path of ground water from a JP-4 spill at Eglin AFB to the point of discharge to
surface water: cross section.
62
-------
Table 1. Bioattenuation in Methanogenic
Along a Flow Path From a JP-4 Spill
Ground Water: Changes in Ground-Water Chemistry
to the Point of Discharge to Surface Water
Compound
Benzene
Toluene
Ethyl benzene
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
Location Ale
80H-3
83Z-2
TOO
5,150
1,700
3,120
6,750
5,480
327
1,090
406
153
18.3
227
594
1,270
<1
114
420
182
24.1
0.4
0.14
1.6
3.7
2.3
2.97
0.2
0.12
0.62
16.8
7.8
>ng Flow Path to Surface Water
83U-3
83U-2
83U-1
fcg/L)
198
1.1
1.4
13.8
13.5
<1
70.9
172
115
6.9
<1
1.4
25.1
39.8
<1
119
299
187
<1
<1
<1
<1
<1
<1
2.3
<1
<1
(mg/L)
0.59
0.3
<0.05
5.6
12.5
3.3
0.68
0.6
<0.05
1.76
14.2
2.8
0.002
3.8
<0.05
<0.5
0.7
<0.5
Apparently, concentrations were reduced from one-half to one-fourth of the initial concentration
due to dilution, with further reductions due to biological activity. Benzene 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 bf 1,2,4-trimethylbenzene can be used to correct
forthe dilution of benzene and determine the true removal due to biodegradaWon. The corrected
concentration of a biologically transformed compound in a downgradient well would be the
measured concentration in the downgradient well, multiplied by the measured concentration of
1,2,4-trimethylbenzene in the upgradient well
1,2,4-trimethylbenzene in the downgradient v
and divided by the measured concentration of
63
-------
The concentration of benzene measured in sample 83U-2 was 6.9/ig/L Compared with sample
83Z-2, the concentration corrected for dilution would be 9.7/ig/L. Compared with 1 53 /tg/L
in 83Z-2, there was a 1 5-fold attenuation in benzene concentration between 83Z-2 and 83U-2.
The benzene attenuation at 83U-2 compared 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
contaminated ground water. Benzene may have been degraded aerobically. In any case,
benzene and the other BTEX compounds did not discharge to the stream at detectable
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 = Infcorrected cone, downgradient/conc. 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 temperatures. There seemed to be preferential removal of toluene and o-xylehe
in the segment close to the spill. Rates of removal of ethylbenzene and m+p-xylene increased
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
Benzene
Toluene
Ethylbenzene
p-Xvlene
m-Xylene
o-Xylene
80H-3 to 83Z-2
83Z-2 to 83U-3
(per week)
None
-0.94
-0.21
-0.14
-0.14
-1.5
None
-0.38
-0.38
-0.57
-0.73
Cannot calculate
64
-------
Stoichiometry of Bioattenuation at Eglin
There was very little oxygen, nitrate, orsulfate
in the plume of contaminated ground water
AFB
available for respiration of the BTEX compounds
(see Table 1). There was an increase in the
concentration of methane and iron and a corre; ponding decrease in aromatic fuel hydrocarbons
along the flow path, however. Assuming the following stoichiometry for methanogenesis from
BTEX compounds,
2 CH + 3/2 H2O -> 5/4 CH4 + 3/4 C(D2,
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 olj methane at 83Z-2, corrected 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 1 6.4
mg/L compared with 80H-3. This decrease in aromatic petroleum-derived hydrocarbons would
be expected to produce 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 Wate
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 approximately
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 conductivity 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
(82D), at the lower edge of the spill (82C), a
of the spill (821 in Table 3), near the midpoint
nd just beyond the edge of the spill (82B). The
continuous cores started in clean material above the spill and extended through the spill into
clean material underneath the interval conta
subdivided into core samples representing 0.3
appeared in the capillary fringe (Table 3). The
to 5.2 ft below the watertable, with TPH conce
ning hydrocarbon. The continuous cores were
vertical ft of the subsurface. Near the spill, JP-4
concentration maximum of 14,800 mg/kg was
located 0.5 ft above the watertable. A second interval that contained JP-4 extended from 3.8
itrations 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 watertable. 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 watertable. At the midpoint and lower edge of the
spill, hydrocarbons were not detected in core
reported in Table 3 (detection limit 1 0 mg/kg).
and lower edge of the spill were less than 70
Monitoring wells were installed in the borehol
monitoring wells were installed downgradient
material collected above or below the samples
Concentrations of hydrocarbons in the midpoint
D mg/kg.
3S used to acquire the core samples. Additional
of the lower edge of the spill.
65
-------
EXTENT OF JP-4 SPILL
SCALE
300
600
-STORM DRAIN
LEGEND
«MW-10 MONfTORING WELL LOCATION
B B' LINE OF HYDROGEOLOGIC SECTION
Map Showing Sample Locations
and Flow Path of Ground Water
from JP-4 Spill to Point of
Potential Discharge In Storm
Water Drain
HILL AIR FORCE BASE
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.
SW B
J. 4690-
-=.4680-
J.4670-
-J4660
Sj 4650-
3 4640-
V 4630-
| 4620
= 4610
| 4600
5 4590
z 4580
EPA-82-D
EPA-82-A
EPA-82-F-
LEGEND
Poorly Sorted Sands £g] Silt and Silly Clay
Silty or Clayey Sands 0 Storm Sewer
Ground Water Level 0150300
FEET
B' NE
EPA-82-1
Hydrogoologlc
Section B-B'
HILL AIR FORCE BASE
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.
66
-------
Table 3. Vertical Distribution of Oily Phase Hydrocarbons at Hill AFB
Core
Elevation
(ft)
TPH
Befnzene
821, near the spill area, water table elevatic
821-5
821-4
821-3
821-2
821-1
821-27 to
821-19
821-17
821-37
821-32
4,665.7
4,665.4
4,665.0
4,664.6
4,664.3
4,663.9 to
4,661.0
4,660.7
4,659.3
4,659.0
4,330
3,770
14,800
5,870
398
<300
1,290
1,370
3,830
Toluene
Ethyl-
benzene
1,2,4-
TMB
(mg/kg)
n 4,664.53 ft
0.0326
0.517
4.55
OJ.401
^0.01
c|.653
0.712
0,
0136
0.0266
0.235
2.73
12.6
0.142
0.591
0.182
0.032
14.5
4.8
47.7
17.5
0.556
3.39
2.72
1.24
49.9
42.7
167
69.8
4.43
2.34
5.38
8.37
82D, downgradient of spill, water table eleyation 4,631 .7 ft
82D-24
82D-23
4,631.0
4,630.7
77.1
572.0
q.271
<0.01
1.48
3.11
82C, lower end of spill, water table elevation 4,603.4 ft
82C-20
82C-19 '
4,603.5
4,603.2
593.0
638.0
<|o.oi
0
0062
0.0176
<0.01
0.00618
0.0180
1.03
1.16
82B, below the oil spill, water table elevation 4,608.3 ft
82B-20
82B-19
4,608.0
4,607.7
0.7
1.0
<0.01
<|o.oi
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
jnd water from the area containing oily phase
Correcting for Dilution at Hill AFB
The TMBs were remarkably persistent in gro
hydrocarbons (Table 4). Concentrations of the TMBs in water from the lower edge of the spill
varied from 50 percent to 147 percent off 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 concentration. The saturated
thickness of the channel sands was at most a fdw feet, and, therefore, there was little opportunity
67
-------
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.
Concentrations of all the aromatic hydrocarbons are low in well 82B just past the lower edge
of the spill.
Table 4. Changes in Ground-Water Chemistry Along a Flow Path Through a JP-4 Spill
Undergoing Sulfate Reduction
Compound
Benzene
Toluene
Ethyl benzene
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
Location Along Flow Path to Surface Water
821
MW-11
82D
82C
82B
(«g/L)
2,740
327
486
784
1,370
1,140
162
495
240
336
90
139
230
635
204
71
165
69
96
10
147
149
383
103
129
183
89
4.9
3.1
27
43
47
2.6
238
324
120
<1
4.3
<1
<1
<1
<1
1.1
1.4
<1
(mg/L)
7.7
0.68
2.1
0.1
0.4
98
0.022
0.05
1.3
1.3
0.5
193
<0.001
1.7
2.1
0.5
0.1
50
0.002
0.84
0.001
1.2
0.4
64
<0.001
0.11
Stoichiometry of Bioattenuation at Hill AFB
Similar to the plume at Eglin AFB, neither oxygen nor nitrate is available for respiration of the
BTEX compounds within the spill area (Table 4). Oxygen concentrations as high as 8.0 mg/L
occur outside the spill. Unlike the case at Eglin, there was little accumulation of iron II and
68
-------
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
H20,
4.6 mg of sulfate would be required to degra Je 1.0 mg of BTEX and 1MB. 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 of the theorstical demand remain in the water after it has
moved away from the spill. There is an adecuate supply of sulfate to remediate the plume
through sulfate reduction alone.
grou id
Concentrations of sulfate were higher in
lower concentrations of sulfate near the lower
exerted. It is just as possible that they represert
Hydrogen sulfide did not accumulate in ground
AFB. Unimpacted sediments at the site are tan
are black. Iron minerals in the aquifer matrix
precipitated as iron II sulfide.
water undergoing intrinsic remediation at Hill
or brown colored, while contaminated samples
Must act to precipitate sulfide. Sulfide very likely
The concentrations of electron acceptors were
from the lower edge of the spill. Well 82B is
82E are 300, 400, and 500 ft downgradient o
table aquifer (Figure 4); well 82A is adjacentt
underneath the water table aquifer.
compared in ground water at various distances
jist outside the lower edge; wells 82F, 82H, and
the spill. These wells are screened in the water-
82F but is screened in the first confined aquifer
None of the wells outside the spill had signif
(Table 5). Water adjacent to the spill (82B)
was
Intrinsic Bioremediation of Oily Phase JP
Two samples of floating oil from monitoring
analyzed by gas chromatography/mass spectra
mean molecular weight of the weathered fuel.
on duplicate samples.
The concentration of individual hydrocarbons
was estimated using Raoult's Law. With data f
petroleum hydrocarbon (mg/kg) was divided
divided by 1 60 to express TPH in moles/kg o
water near the source. It is possible that the
edge of the spill represent the sulfate demand
natural variations in sulfate concentrations.
cant concentrations of aromatic hydrocarbons
depleted in oxygen and nitrate.
•4 Jet Fuel at Hill AFB
Well MW-1 0, near the midpoint of the spill, were
netry(GC/MS) to determine the number-average
Values of 1 56 and 1 60 daltons were determined
in ground water in contact with oily phase JP-4
•om Table 3, the concentration of an individual
its molecular weight to express its concentration
in moles/kg core material. The concentration of total petroleum hydrocarbons (TPHs) was
re material. The mole fraction of the individual
hydrocarbon was calculated by dividing moles pf individual hydrocarbon per kilogram by moles
of TPH per kilogram. Then the mole fraction was multiplied by the water solubility of the
individual petroleum hydrocarbons to estimate the equilibrium concentration in ground water.
69
-------
TableS. Changes in Ground-Water Chemistry Along a Flow Path Downgradient of a JP-4 Spill
Compound
BTEX
TMB
Oxygen
Nitrate +
Nitrite - N
Sulfate
Methane
Iron II
Location Along Flow Path to Surface Water
82B
82F
82E
82H
82A
(«g/L)
-------
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
Elevation
(ft)
Benzene
Toluene
Ethyl-
benzene
1,2,4-
TMB
Measured Concentration («g/L)*
Predicted Concentration («g/L)
821, near the spill area, water table elevatio
11/93
821-5
821-4
821-3
821-2
821-1
821-27 to
821-19
821-17
821-37
821-32
4,665.7
4,665.4
4,665.0
4,664.6
4,664.3
4,663.9
to
4,661.0
4,660.7
4,659.3
4,659.0
2,740*
22.8
513
1,150
254
103
<300
1,870
1,940
13.2
n 4,664.53 ft
372*
5.5
55.6
165
1,923
320
. 410
119
7.5
486*
769
289
740
684
321
603
456
74.3
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
4,631.7
4,631.7
4,630.7
96*
174*
1,769
10*
4.6*
15.6
147*
30.8
594
183*
119*
460
82C, lower end of spill, water table elevaticjn 4,603.4 ft
7/93
11/93
82C-20
82C-19
4,603.4
4,603.0
4,603.5
4,603.2
4.9*
<1*
<63
36.3
3.2*
8.4*
26.6
13.9
27*
6.8*
2.4
6.5
324*
68.9*
135
142
71
-------
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. For the 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 contaminants. 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 overtime, a comparison of the mole fraction of benzene in 831-3 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 -0.1 I/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. 1 6(2) = 135-141.
2. Wilson, J.T., D.H. Kampbell, and J. Armstrong. 1 994. 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.
72
-------
A Natural Gradient Tracer Experiment in
In Situ Biodegradation Rates: A Case for
a Heterogeneous Aquifer With Measured
Natural Attenuation
Thomas B. Stauffer and Christopher P. Antwor h
Armstrong Laboratory, Tyndall AFB, FL
J. Mark Boggs
Engineering Laboratory, Tennessee Valley Lab
William G. Maclntyre
School of Marine Science, College of William
oratory, Morris, TN
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, Mississippi, during a pulse
release experiment. Reaction rate calculations were based on a kinetic model that includes the
hydrologic characteristics of the aquifer. Degracation kinetics were approximately first order with
the following rate constants: benzene, 0.0066
d'1; o-dichlorobenzene, 0.0059 d'1.
J'1; p-xylene, 0.0141 d'1; naphthalene, 0.0063
transport models. Field experiments for dete
Introduction
Biodegradation rates of organic contaminants in aquifers are needed for use in fate and
rmination of in situ biodegradation rates are
desirable because laboratory measurements may not relate to conditions in an aquifer. Madsen
(1) notes in a recent review "that determination of microbial activity in disturbed, displaced
environmental samples incubated in the laboratpry is likely to be quantitatively, even qualitatively
different from the same determination in situ." Reliable estimation of an in situ biodegradation
rate requires introduction of a known mass of contaminant at a defined time zero, and
observation of contaminant concentration variation in space and time. Confirmation of
biodegradation requires maintenance of mass b alances and determination of organic compound
degradation products. To our knowledge, no ex
rates and confirm biodegradation in an aquifc
>eriments to accurately measure biodegradation
r 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 heterogeneous unconfined aquifer formed by fluvial
I I .L'
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 contrast to the generally lower (10~4 to TO"3 cm/s) values and less
variable Kh field at the Borden Canadian Forces Base, Ontario (2).
Asummary of hydrogeologic properties of the Columbus aquiferand a description of the MADE
site are given by Boggs etal. (3). Rehfeldtetal. 4) measured the spatial distribution of hydraulic
Symposium on Intrinsic Bioremediation of Ground Water
73
-------
conductivity at the MADE site using borehole flow meters and other techniques. Figure 1 shows
the Kh distribution over a vertical section directed 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 Kh is relatively low, with values near TO"3 cm/s from the phreatic surface to the
lower confining layer and extending about 40 m downgradient from the injection location.
Immediately beyond this region, Kh increases to values near 1 O'1 cm/s in the upper 3 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 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 radioactive decay and biodegradation was developed
for use with field data to obtain in s/fu biodegradation rates for organic compounds in the
Columbus aquifer. Radiocarbon measurements using 14C-labeled p-xylene were done to confirm
biodegradation of p-xylene to its degradation 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 organic 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.
65
60
|=
J 55
03
LLJ
50
10° 10'1
10-2
10-" (cm/s)
FLOW
4-
25 50 75 100 125 150 175 200 225 250
Distance (m)
275
Figure 1. Distribution of hydraulic conductivity over a vertical section containing the center line
of the plume.
-------
300
250
200
i?150
100
50
-50
N
Injectlor
point
-100
-50
i-
X
Figure 2. Plan view of the injection and sampling wells at the MADE2 site.
barcad A
multilevel sampler •
Injection well a
inset
15
10
5
0
A A
A A
A6&
-15 -10 -5 0 5 10 15
50
On)
100
150
200
75
-------
Experimental parameters for the MADE2 test were: injection 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
compounds in water were determined by solvent extraction into pentane containing toluene as
an internal standard, and analysis of the extract was made by gas chromatography (GC) with
flame ionization detection. Most sampling wells contained multilevel samplers, but BARCAD
water samplers were used in a few wells. Water was collected from the sampling wells and
analyzed to provide three-dimensional snapshots of solute concentrations at 27,1 32,224,328,
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 overthe 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 for 3H2O. Selected
wells from high and low solute concentration regions of the plume were analyzed for p-xylene
degradation by 14C counting methods 421 d after injection. Each watersample was subsampled
in the field. A 2-mL subsample was mixed with cocktail and counted to obtain total 14C in the
water as either p-xylene or its degradation products. A 20-mL subsample was extracted with
2 mLof 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
carbonate produced by complete degradation of 14C p-xylene and the amount of 14C in water-
soluble organic intermediate products, respectively.
Leaky Reactor Model
Organic solutes remaining in the low Kh region near the injection wells may be regarded as
semiconfined in a 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 Kh 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
compounds, assumed to be first order with rate constants kb, kx, kn, and kc 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, k|, for all solutes.
Sorption-desorption processes at aquifer material surfaces 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 equations with initial conditions: att = t0 =
0, [Tr]0 = 0.539 Ci, [B]0 = 660 g, [X]0 = 402 g, [N]0 = 70 g, [C]0 = 31 8 g, where square
brackets represent solute mass and t is time following injection.
76
-------
INJECTION
WELL
GROUND
WATER
FLOW
/
f
f
/ ^
V
,„„„„„„,, J
REAC1
KB[B] '
KX[X]
KN[N]
KC[C]
T 1
ronz(
^DEG
DNE f
RADATION
IATES f
I
^
CHANNEL
K (LEAK RATE CONSTANT)
^
Figure 3. Diagram of the leaky reactor model.
Analytic solutions of equations 1 to 5 with these conditions are the integrated rate laws
(equations 6 to 1 0) describing solute concentrcitions as functions of time. The leak rate constant
k| is calculated by solving equation 6, using the known kd of tritium. Constants kb, k,,, kn/ and k,.
are then determined by substitution of k( and solving equations 7 through 1 0.
dt
= kb[B] -, kt[B]
dt
dt
= kn[N\ + k^N]
= kc(C] + k$C\
-k.
(1)
(2)
(3)
(4)
(5)
(6)
77
-------
In V'*l + lnU?T
(7)
*,-
,
(8)
InV1'*/ .in,--, .*
(9)
/ p>]|
, In l17^/ In U^
fc = + + k,
(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 sorption coefficients on a composite sample of Columbus
aquifer 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 remained near the injection
wells after 224 d. Biodegradation of organic solutes occurred primarily in a reactor zone
approximately delineated by the p-xylene distribution 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.
The spatial distributions of Kh in Figure 1 and of 3H2O in Figure 4a are similar, which is
expected because hydraulic conductivity is the controlling parameter for transport of an
unretained solute. High 3H2O concentrations 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 1 75 m depth. This distribution indicates that solutes were transported
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
78
-------
in the lower portion of the aquifer. The observed transport of 3H2O during this experiment was
consistent with the leaky reactor model.
65
"E" 60
J 55
"a
J> 50
a?
45
tritium
(a)
25 50 75 100 125 150 175 200 225 250 275
E 60
J 55
"a
.£ 50
us
p-xy!
ene
10° 10"'
C/Co
10
-4
(b)
0 25 50 75 100 25 150 175 200 225 250 275
Figure 4. Distribution of (a) normalized tritium and
over a vertical section containing tf
(b) normalized p-xylene concentration zone
e center line of plume motion.
79
-------
Leakage of solutes was determined by the loss of tritium from the reactor zone, defined as the
portion of the plume volume within 10m downstream from the injection 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 solute leakage from the reactor is first
order, with leakage rate constant, k, = 5.45 x 1 O'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.
Biodegradation rates within the reactor zone are approximately 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 degradation
rates are low. This lag period is followed by degradation 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(corrj(reactor)) in the second column have been corrected fororganic solute leakage
from the reactor zone. Maximum biodegradation rate constants for each organic solute were
also calculated by spatial integration over the entire well field, and are included in Table 1 as
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 and 14CO2 (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-labeIed organic compounds. 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 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 conversionfigures 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.
80
-------
-3
100 200
DAYS AFTER
Figure 5. Tritiated water content in the reactc r organic as a function of time
-1 -
-2-
-4-
-5'
300 400
INJECTION
500
D BENZENE
A NAPHTHALENE
O p-XYLENE
A o-DICHLOROBENZENE
0 100 200
DAYS AFTER
Figure 6. Degradation curves for the compo
300 400
NJECTION
500
jnds in the reactor zone.
81
-------
Table 1. Maximum Biodegradation Rate Constants From the MADE2 Site
Benzene
p-Xylene
Naphthalene
o-Dichlorobenzene
I^Jreactor)
(d-1)
0.0120
0.0196
0.0118
0.0114
kmaxicor^reactor)
(d-1)
0.0066
0.0141
0.0063
0.0059
k^whole field)
(d-1)
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
[xylene]>l ppm
(n-8)
[xylene]>l ppm
(n=10)
% 14C p-Xylene
Converted to All
Products
85.1 (± 6.3)
82.6 (± 6.1)
% 14C p-Xylene
Converted to
C02
73.3 (± 11.1)
74.2 (± 6.2)
Mass Balance
Based % p-Xylene
Converted
98
98
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 degradation 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
converted to CO2 may imply that some intermediate products 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.
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 expected, 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
82
-------
concentration can often be expressed as a firs
in the MADE2 test were quite low, so this conclusion is consistent with experimental results.
Larson (10) notes that the rate of decreas 3 in concentration as a function of substrate
-order equation, and that first-order kinetics are
generally expected for biodegradation at low organic substrate 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 methods of calculating biodegradation rates
gave similar results, thus providing confidence
hat these rates are useful for predictive purposes.
Organic solutes in MADE2 biodegraded under aerobic conditions in the aquifer. The organic
solute concentrations 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 biodegradation rates reported here were not affected by oxygen limitation. For very large
releases of similar organic compounds, the conditions near the source might rapidly become
reducing due to the biological oxygen demand. 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 remediation activities should be restricted to the source region, which might include
pumping to remove nonaqueous-phase liquids or excavation of contaminated aquifer material.
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 reduced source would reach
a steady state, with the boundary determined by the hydrology of the site, sorption, in situ
biodegradation, and oxygen and nutrient sup Dly.
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 experiments and
obtained in situ degradation rates forfour 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. 1 991. Determining in situ biodegradation: Facts and challenges. Environ. Sci.
Technol. 25(10):!,663-1,673.
2. Robin, M.J.L., E.A. Sudicky, R.W. Gillham, and R.G. Kachanoski. 1 991. Spatial variability
of strontium distribution coefficients and their correlation with hydraulic conductivity in the
Canadian Forces Base Borden aquifer. Water Resour. Res. 27(1 0):2,61 9-2,632.
83
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3. Boggs, J.M., S.C. Young, and L.M. Beard. 1992. Field study of dispersion in a
heterogeneous aquifer, 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 aquifer, 3. Geostatistical analysis of hydraulic conductivity. Water Resour.
Res. In press.
5. Mackay, D.M., D.L Freyberg, and P.V. Roberts. 1 986. A natural gradient experiment on
solute transport in a sand aquifer, 1. Approach and overview of plume movement. Water
Resour. Res. 22(13):2,01 7-2,029.
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, Massachusetts, 2. Analysis of spatial
moments for a nonreactive tracer. Water Resour. Res. 27(5):911 -924.
7. Adams, E.E., and L.W. Gelhar. 1 992. 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.P. Antworth. 1991. A comparison of sorption
coefficients determined by batch, column, and box methods on a low carbon aquifer
material. Ground Water 29(6):908-913.
9. Simkins, S., and M. Alexander. 1 984. Models for mineralization kinetics with the variables
of substrate concentration and population density. Appl. Environ. Microbiol. 47(6) :1,299-
1,306.
10. Larson, RJ. 1979. Role of biodegradation kinetics in predicting environmental fate. In:
Maki, A.W., et al., eds. Biotransformation and fate of chemicals in the aquatic
environment. Washington DC: American Society for Microbiology.
84
-------
Traverse City: Distribution of the Avgas Spill
David W. Ostendorf
Civil and Environmental Engineering Departm
ent, University of Massachusetts, Amherst, MA
Abstract
The capillary tension/liquid saturation equati
Parker (2) provide a reasonably accurate desc
phase aviation gasoline (avgas) in solid core
Station in Traverse City, Michigan.
Introduction
The depth of a continuous liquid below the c
tension in the absence of vertical velocity, sin
forms a vertically uniform hydraulic head unde
idns
of Parker and Lenhard (1) and Lenhard and
iption of the vertical distribution of the separate
samples taken from the U.S. Coast Guard Air
round surface is directly related to its capillary
:e the sum of the potential and pressure heads
static conditions. Thus, the theory (1) describing
capillary tension/saturation relations for immi: cible, continuous fluids also finds expression as
vertical profiles of water and avgas in the absence of infiltration. Lenhard and Parker (2)
recognize this equivalence and propose profi
es for total liquid and avgas saturation that are
reasonably borne out by Ostendorf et al. (3) in rheir analysis of solid core data from five stations
in the avgas plume at Traverse City (Figure 1).
implications of the distribution.
We summarize the theory here and discuss some
2. The water is more strongly attracted to the
Vertical Distribution of Free Avgas
The vertical distribution of water and free avgas in the contaminated soil is idealized in Figure
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 irreducible water 0WR contents combine to
define an effective apparent water saturation Sw and comprise the total saturation S when they
are added to the free avgas content 0LF (1).
Symposium on Intrinsic Bioremediation of Ground Water
85
-------
Interdiction'
Well Line
X
/'. Ni
/Flow
Direction
/
Smith
Building
606
• SOBT
/SUBS / Xpl"me
50CL* 50CE Boundary
50 m , /
Administration
Building
Spill
Origin
Figure 1. Site plan, U.S. Coast Guard Air Station.
Ground
Surface
Volumetric Liquid Content
Air
Water
Monitoring
Well
Figure 2. Vertical distribution of liquids in soil and a monitoring well.
86
-------
ri - (9,
S =
n
WR
" WR
(la)
(Ib)
with porosity n. The total saturation is contrail sd 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 = {1 + rA(bL-b)]<*}<]
S = 1
distribution to separate phase contamination
elevation and extends above the avgas table
bj
(2a)
(2b)
(2c)
Lenhard and Parker (2) derive equation 2 as an application of van Genuchten's (4) pore size
n soil. Figure 2 displays the minimum depth
JM
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
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. Th 3 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
/?L appearing in equation 2b. The latter parameter may be expressed in terms of the mean pore
radius 7 by noting that the avgas/air surface tension a^ relates fluid tension to interfacial radius
(3)
with gravitational acceleration g, avgas
1 lists parameter values calibrating the data
(circles) and theory (curve) for a typical total
pore radius of 5.4 x 10~5 m is implied by equal
based model of 7 checks this estimate. We
pore size flowing full, and equate this to a mo
2a,
LA
density pL, and an assumed zero contact angle. Table
from five stations at Traverse City (3). Observations
saturation profile are sketched in Figure 3. A mean
ion 3. A simple extension of a classical grain size
n 3te that 7/2 is the hydraulic radius of the mean
Jified Fair and Hatch (5) estimate of the quantity
87
-------
r
"2
V.
voros
^SOLIDS
(4a)
r =
(4b)
Table 1. Total and Avgas Saturation Profile Parameters at Traverse City
Symbol
n
C'WR
PL
a
A.
Aw
A
Y
Parameter
Porosity
Irreducible moisture content
Avgas density
Pore size uniformity
Avgas scaling factor
Water scaling factor
Water table amplitude
Trapping factor
Value
0.367
0.059
7.07 kg/m3
3.00
8.20 rrr1
1 .53 m'1
0.35 m
40
The observed mean grain size of 3.8 x 10'4 m at Traverse City leads to a mean pore size value
of 5.7 x 10"5 m, in excellent agreement with the equation 3 value.
The water/avgas interface controls the water saturation in the presence of the free avgas
(b
w
bM)
(5a)
Sw-1
bw)
(5b)
•'LA
''WL
(5c)
with water density p and scaling factor/?w predicated on the water/avgas surface tension <7WL.
The upper extent bM of free avgas may be estimated by equating the total saturation (equation
2b) and the water saturation (equation 5a) at this elevation. A water/air interface with a surface
tension <7WA governs the water saturation above this depth
-------
_
(_J —
sw =
''WA
The free avgas saturation S^, in view of equa
LF
S - S
6 -
0.0
-6
L f^
(6a)
(6b)
ion 1, is simply given by
w
(7)
Figure 4 shows a typical free avgas profile atth 3 site, based upon the scaling factors of Table 1,
1.0
Figure 3. Total saturation, Core 50BT.
5.5
0.0
0.2
Figure 4. Free avgas, Core 50CE.
89
-------
Vertical Distribution of Residual Avgas
Residual avgas may be attributed to hysteretical trapping 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 tensions as the water table falls due to their higher position
above this reference level. After the table attains a maximum 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 saturations SWMAX, SWM,N are established by
corresponding minimum (bWM|N) and maximum (bWMAX) water table depths, respectively
WMIN
(8a)
(8b)
(8c)
Ostendorf et al. (3) infer the historical water table excursion amplitude A cited in Table 1 from
water level variations in nearby Lake Michigan. Parker and Lenhard (1) suggest that the extreme
saturations induced by extreme water positions trap residual avgas saturation SLR that is given
by
1-S,
WMIN
(bWMAX>b>bWMIN)
(9)
with empirical trapping factor Y. Ostendorf et al. (3) calibrate residual saturation profiles from
their five stations with the trapping factor cited in Table 1, with the typical results sketched in
Figure 5.
E
.a
0.00 0.01
0.02 0.03
LR
Figure 5. Residual avgas, Core 50CL.
90
-------
Discussion
The vertical distribution of free and residual avjas has important implications forthe 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, and 7 from
the water table to the upper limit of free avgqs occurrence, with the result
MO,=
Mu^-?1
'w
K»-[
bw
S^db
(10a)
(lOb)
ft.
(IOC)
The integral function l(y) varies with the uniforrrpity exponent of the pore sizes, while its argument
is determined by the scaling factors and the p -oduct thickness in the monitoring well. Figure 6
displays the equilibrium variation of the depth
function of the monitoring well avgas thickness
this estimate, since the trapped avgas is not cor tinuously connected to product in the monitoring
well. Fora point of reference, Figure 5 implies
5.3 kg/m2; if this mass were free avgas, we
monitoring well.
integrated mass predicted at Traverse City as a
We note the exclusion of the residual mass from
a depth integrated residual avgas mass of about
would observe a 0.3-m product thickness in a
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 lot 12°C). This fraction becomes more important
in the lower elevations due presumedly to decreased volatilization 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 stripping away of the lighter fraction.
91
-------
CNJ
<
£
JP
b —b , m
W L
Figure 6. Depth integrated free avgas mass.
8- 5.5
Q
O High
• Medium
V Low
_! i
10 20 30 40
Avgas Fraction, %
Figure 7. Avgas composition, SOBS.
-------
The partitioning of avgas into free and resic
relatively high-permeability, saturated soil
ual fractions has important implications on its
downgradient fate and transport as well. Figure 2 suggests that the free avgas occupies a
region
il I r •
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 vertical 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 mechanisms is that the avgas travels with
the underlying ground water, but at a retarded velocity.
the soil due to its direct interface with air. The
remediate, since it is surrounded by essentially
initial period of relatively rapid remediation in
owed by an asymptotically lower removal rate
The free avgas is relatively easy to strip out o
residual avgas is much more difficult to
immiscible water. We accordingly expect an
response to soil venting or air sparging, fol
exacerbated by avgas distillation.
References
1. Parker, J.C., and R.J. Lenhard. 1 987. A model for hysteretic constitutive relations
governing multiphase flow, 1. Satu "ation pressure relations. Water Resour. Res.
23:2,187-2,196.
Lenhard, R.J., and J.C. Parker. 1 990. Estimation of free hydrocarbon volume from fluid
levels in monitoring wells. Ground Water 28:57-67.
2.
3. Ostendorf, D.W., R.J. Richards, and
Ground Water 31:285-292.
:.P. Beck. 1 993. LNAPL retention in sandy soil.
conductivity of unsaturated soils. Soil
of water through sand. J. Am. Waterv
residual aviation gasoline in sandy s
4. van Genuchten, /vi.T. 1980. A clossd form equation for predicting the hydraulic
Sci. Soc. Am. J. 44:892-898.
5. Fair, G.M., and L.P. Hatch. 1 933. Fu idamental factors governing the streamline flow
•orks Assoc. 25:1,551 -1,565.
6. Ostendorf, D.W., LE. Leach, E.S. Hnlein, and Y.F. Xie. 1991. Field sampling of
7.
il. Ground Water Monitor. Rev. 11 =107-120.
Ostendorf, D.W. 1 990. Long-term fate and transport of immiscible aviation gasoline in
the subsurface environment. Water Sci. Tech. 22:37-44.
93
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Traverse City: Geochemistry and Intrinsic Bioremediafion of BTX 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 majorsource of contamination of unsaturated soils, aquifer solids, and ground water. Volatile
aromatics such as benzene, toluene, 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 occurthrough reaeration from soil gases, ground-water
recharge, and usually inefficient mixing with surrounding oxygenated ground waters (4,5).
Although ground waters nearthe perimeter of the contaminant plume may be reoxygenated, the
interior of the plume will remain anoxic for a distance downgradient. Anaerobic biological
processes can account for most of the removal of BTEX from the plume (6,7). The
biogeochemical mechanisms that contribute to anaerobic processes in the subsurface, however,
are not well understood.
The impact of anaerobic microbial processes on the fate of monoaromatics, substituted
aromatics, and chlorinated 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 otheralkylbenzenes present (1 3). Methanogenesis has been observed at two sites
with ground water contaminated 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 contaminated with creosote. Intermediate products of methane fermentation,
formate and acetate, were found in the ground water at the second site.
In 1 969 the unsaturated soil and ground water underlying the U.S. Coast Guard Air Station at
Traverse City, Michigan, were contaminated with an.estimated 25,000 gal of aviation gasoline
when a flange in an underground storage 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 aromatics, and gaseous hydrocarbons resulting from volatilization (16-1 8). 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
94
Symposium on Intrinsic Bioremediation of Ground Water
-------
operations during aircraft maintenance. A smc II plume of chlorinated solvents lies adjacent to
the gasoline plume.
Geochemical Characferization
Geochemical analyses of the water samples collected at Traverse City revealed waters of four
distinct geochemistries: 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 concentre tions of the alkylbenzenes. The perimeter of the
plume was surrounded by a renovated or prist
detectable alkylbenzenes, and no methane.
ne zone with high concentrations of oxygen, no
Gas chromatography/mass spectrometry (GC/MS) analyses of the waters confirm the presence
of BTEX at Traverse 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 la) identified phenols and aromatic acids indicative
of anaerobic microbial action on the soluble aromatic constituents of petroleum products
(7,19,20). These compounds are found in portions of the plume with substantial concentrations
of methane and no detectable oxygen, and are probably precursors of the methane. Complete
information on the geochemical characterization of the ground waters at Traverse City,
collection of aquifer material, analytical methods used, and microcosm construction may be
found in Wilson et al. (21).
Laboratory Studies
To confirm field evidence of intrinsic biorernediation, laboratory microcosm studies were
conducted on aquifer 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 la). Aquifer
material from site A (1 1.7 m to 1 2.3 m below
was used to construct the microcosms for the
land surface), the zone of anaerobic treatment,
anaerobic fate study. Aquifer material from site
B (9.6 m to 1 0.2 m below land surface), the aerobic zone of active biological treatment, and
from site C (6.6 m to 7.2 m below land surfa
:e), 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
and chlorobenzene.
The initial compound concentrations and resu
are shown in Table 1. Benzene, toluene, p-:
aerobic and anaerobic aquifer material. The
each of the three geochemical zones studied,
aerobic zone of treatment, or the renovated
of the compounds in the anaerobic aquifer
p-xylene, and o-xylene had been reduced one
benzene, toluene, p-xylene, o-xylene, TCA, TCE,
ts of the fate studies at various incubation times
•x/lene, and o-xylene were biodegraded in both
•emovals were quite rapid for all compounds in
whether in the anaerobic zone of treatment, the
(pristine) material. By the end of 8 wk of incubation
rraterial, the concentrations of benzene, toluene,
order of magnitude. The biotransformation of the
95
-------
four compounds in the aerobic zone of biological treatment occurred even more rapidly. By the
end of 2 wk 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.
Table 1. Behavior of Benzene, Alkylbenzenes, TCAa, TCEb, and Chlorobenzene in Aquifer
Material From an Aviation Gasoline Plume
Subsurface
Material
A,
anaerobic
treatment,
11.7 mto 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
Compound lftg/\- Pore Water)
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
Chloro-
benzene
500
66
34
3
500
53
50
30
500
100
106
67 '
500
300
210
°1,1,1 -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 fig/L
At the end of 4 wk of incubation, the respective concentrations 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 concentrations. Due to the removals of BTX in the controls, a second set
of autoclaved samples was prepared. The concentrations after 1 6 wk of incubation in the
duplicated controls were benzene, 83 percent; toluene, 66 percent; m- + p-xylene, 51 percent;
o-xylene, 50 percent; TCA, 86 percent; TCE, 73 percent; and chlorobenzene, 58 percent. The
96
-------
cause for the removal of organics in the contrc Is has not been determined; however, sorption
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 of
incubation. Similar removal was seen in the renovated/pristine material after 14 wk of incuba-
tion. In the autoclaved samples, 44 percent of the chlorobenzene remained after 4 wk of
incubation.
No significant biotransformation of TCA or TC
the end of 8 wk of incubation. Evidence of
:, compared with the controls, was observed at
reductive dechlorination of both compounds,
however, was indicated by GC/MS analyses of anaerobic microcosms at 95 wk of incubation
by the identification of 1,1-dichloroethylene
(22,23).
1,1-DCE) and 1,1-dichloroethane (1,1-DCA)
Headspace concentrations of methane were measured immediately before sampling to
determine the maintenance of methanogenic conditions in the microcosms. Methane was found
m all the anaerobic samples, with concentrations ranging from 50 ppm to TOO 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
concentrations 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 1 0 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
1 984 through 1 987. 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 operation, the water behind the field near site
A (Figure la) became stagnant, and concentrations of BTX began to drop. Solution
concentrations of BTX dropped to low values by late 1 985 (data not shown). Because the water
was not moving, the decline in concentration overtime could be used to estimate the first-order
rate constant for anaerobic BTX biotransforma
3).
Ion in the part of the aquifer near site A (Table
97
-------
Figure 1a
U.S. COAST GUARD
AIR STATION BOUNDARY
PURGE WELL FIELD
Figure I a. Locations of wells in an aviation gasoline plume atthe 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.
Rgure 1fo
TOLUENE BENZENE
XYLENES
ua/L
30,000
10,000
3,000
1,000
300
100
30
10
P R
Figure 1 b. Concentration of BTX in monitoring wells along a flow path down the central axis
of the plume in the first quarter of 1 986.
98
-------
Table 2. First-Order Rates of Anaerobic Biolransformation (per Week) of Total BTX Along
Segments in the Aquifer
Year
1984
1985
1986
1987
StoQ,
Quarter of Year
First
0.11
0.02
0.43
0.34
Third
0
0
17
27
oll4
olio
S to A,
Quarter of Year
First
0.10
0.07
0.20
0.33
Third
0.09
cnca
cnc
cnc
°Cannot calculate: the plume was interceptec
Table 3. Comparison of the First-Order Rates
by a purge well field and did not reach site A.
per Week) of Anaerobic Biotransformation of BTX
Compound
Benzene
Toluene
m + p-Xylene
o-Xylene
All xylenes
Microcosms'"
0.5
0.3
0.4
0.5
Aquifer Segment
StoQ
0.05
1.3
0.03
Change Over Time at
Ac, 6/85 to 1 0/85
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
Sediment samples for the microcosm study were acquired from site A in late 1 985, and this
study was conducted in the first quarter of 1 986. 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 segment S to Q during the first quarter of
1 986 and the depletion in the static water a site A in the last half of 1 985 with the rates of
anaerobic BTX biotransformation in the microcosm study.
The rate of disappearance of BTX compouncs
compare well with actual rates of aerobic BTX
in the field is controlled by mass-transport lim
are limited by reaction rates. The anaerobic fa1
as measured in aerobic microcosms does not
degradation in the field. The rate of degradation
tations for oxygen (4,5), while laboratory studies
e study compared quite favorably with those rates
99
-------
measured by field data. Methanogenesis and other anaerobic processes are not limited by the
availability of oxygen in either microcosms or subsurface materials. Mass-transport
considerations are therefore not as critical to the comparison of microcosm and field data.
Anaerobic microcosms might prove to be a valuable tool to evaluate intrinsic biorestoration of
aquifers contaminated with petroleum products.
Conclusions
The results of the laboratory study confirm field evidence of both aerobic and anaerobic
transformation of alkylbenzenes and suggest that intrinsic aerobic and anaerobic in s/fu
biorestoration of ground water contaminated with petroleum products can occur. The anaerobic
transformations seen at this site and confirmed by the laboratory study provide an attractive
alternative to aerobic restoration. The removals of the alkylbenzenes in the anaerobic material
were quite rapid and compared favorably with removals seen in the aerobic zone of treatment.
Comparison of first-order rates of disappearance in anaerobic microcosms with those calculated
from field data show acceptable agreement. Anaerobic processes in the subsurface are probably
limited by in situ reaction rates ratherthan by mass-transport limitations for nutrients. Potentially,
anaerobic microcosm studies could be useful in the evaluation of intrinsic bioremediation 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 occurring anaerobic biological processes can potentially remediate
ground water contaminated with petroleum products and significantly increase the reliability of
existing remediation technologies.
References
1. Coleman, W.E., J.W. Munch, R.P. Streicher, H.P. Ringhand, and F.C. Kopfler. 1 984.
The identification and measurement of components in gasoline, kerosene, and no. 2
fuel oil that partition into the aqueous phase after mixing. Arch. Environ. Contam.
Toxicol. 13:171-178.
2. Atlas, R.M. 1 981. Microbial degradation of petroleum hydrocarbons: An environmental
perspective. Microbiol. Rev. 45(1 ):1 80-209.
3. Gibson, D.T., and V. Subramanian. 1984. Microbial degradation of aromatic
hydrocarbons. In: D.T. Gibson, ed. Microbial degradation of organic compounds. New
York: Marcel Dekker. pp. 181-252.
4. Borden, R.C., and P.B. Bedient. 1 986. Transport of dissolved hydrocarbons influenced
by oxygen-limited biodegradation, 1. Theoretical development. Water Resour. Res.
22(131:1,973-1,982.
100
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5. Borden, R.C., P.B. Bedient, M.D. Lee,
dissolved hydrocarbons influenced
application. Water Resour. Res. 22(13)
9.
10.
11.
13.
14.
of crude oil in a shallow aquifer,
geochemical facies. In: Mallard, G.
C.H. Ward, and J.T. Wilson. 1986. Transport of
by oxygen-limited biodegradation, 2. Field
:1,983-1,990.
6. Baedecker, M.J., D.I. Siegel, P. Bennett, and I.M. Cozzarelli. 1 988. The fate and effects
1. The distribution of chemical species and
:., and S.E. Ragone, eds. Proceedings of the
Technical Meeting of the U.S. Geological Survey Toxic Substances Hydrology Program,
Phoenix, AZ (September 26-30). Water Res. Invest. Rep. 88-4220. pp. 13-20.
7. Cozzarelli, I.M., R.P. Eganhouse, and M.J. Baedecker. 1 988. The fate and effects of
crude oil in a shallow aquifer, II. Evide nee of anaerobic degradation of monoaromatic
hydrocarbons. In: Mallard, G.E., and S.E. Ragone, eds. Proceedings of the Technical
Meeting of the U.S. Geological Survey Toxic Substances Hydrology Program, Phoenix,
AZ (September 26-30, 1988), Water
8. Kuhn, E.P., PJ. Colberg, J.L. Schnoor,
Res. Invest. Rep. 88-4220. pp. 21-33.
0. Wanner, A.J.B. Zehnder, and R.P. Schwarzen-
bach. 1 985. Microbial transformations of substituted benzenes during infiltration of river
water to ground water: Laboratory column studies. Environ. Sci. Technol.
19(10):961-968.
Lovley, D.R., M.J. Baedecker, D.J. Lonergan, I.M. Cozzarelli, E.J.P. Phillips, and D.I.
Siegei. 1 989. Oxidation of aromatic contaminants coupled to microbial iron reduction.
Nature 339:297-300.
Lovley, D.R. and D.J. Lonergan. 1993. Anaerobic oxidation of toluene, phenol, and
p-cresol by the dissimilatory iron-reducing organism, GS-1 5. Appl. Environ. Microbiol.
56(6):1,858-1,864.
Suflita, J.M., S.A. Gibson, and R.E. B
pollutant chemicals in aquifers. J. Ind
Beman. 1 988. Anaerobic biotransformations of
Microbiol. 3:179-194.
Rees. 1986. Biotransformations of selected
12. Wilson, B.H., G.B. Smith, and IF
alkylbenzenes and halogenated ali hatic hydrocarbons in methanogenic aquifer
material: A microcosm study. Environ
Reinhard, M., N.L. Goodman, and J.
organic chemicals in two landfi
18(12):953-961.
Sci. Technol. 20(10):997-1,002.
Barker. 1 984. Occurrence and distribution of
leachate plumes. Environ. Sci. Technol.
Ehrlich, G.G., D.F. Goerlitz, E.M. Godsy/and M.F. Hult. 1982. Degradation of
phenolic contaminants in ground water by anaerobic bacteria: St. Louis Park,
Minnesota. Ground Water 20(6):703-710.
1 5. Goerlitz, D.F., D.E. Troutman, E.M. Gpdsy, and B.J. Franks. 1 985. Migration of wood-
preserving chemicals in contaminated
Florida. Environ. Sci. Technol. 19(10
ground water from a sand aquifer in Pensacola,
=955-961.
101
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16. Kampbell, D.H., J.T. Wilson, and D.W. Ostendorf. 1990. Simplified soil gas sensing
techniques for plume mapping and remediation monitoring. In: Kostecki, P.T., and EJ.
Calabrese, eds. Petroleum Contaminated Soils. Chelsea, Ml: Lewis Publishers, pp.
125-139.
1 7. Ostendorf, D.W. 1 990. Long-term fate and transport of immiscible aviation gasoline in
the subsurface environment. Water Sci. Technol. 22:37-44.
18. Rifai, M.S., P.B. 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.
1 9. Grbic-Galic, D. 1 989. Microbial degradation of homocyclic and heterocyclic aromatic
hydrocarbons under anaerobic conditions. Develop. Ind. Microbiol. 30:237-253.
20. Grbi<5-Galic, D., and T.M. Vogel. 1 987. Transformation of toluene and benzene by
mixed methanogenic cultures. Appl. Environ. Microbiol. 53(2):254-260.
21. Wilson, B.H., J.T. Wilson, D.H. Kampbell, B.E. Bledsoe, and J.M. Armstrong. 1990.
Biotransformation of monoaromatic and chlorinated hydrocarbons at an aviation
gasoline spill site. Geomicrobiol. J. 8:225-240.
22. Barrio-Lage, G., F.Z. Parsons, R.S. Nassar, and P.A. Lorenzo. 1986. Sequential
dehalogenation of chlorinated ethenes. Environ. Sci. Technol. 20(l):96-99.
23. Klecka, G.M., SJ. Gonsior, and D.A. Markham. 1990. Biological transformations of
1,1,1-trichloroethane in subsurface soils and ground water. Environ. Toxicol. Chem.
9:1,437-1,451.
24. Sammons, J. 1994. Personal communication between John Sammons, The Traverse
Group, Inc., and the authors.
25. Lee, M.D., J.M. Thomas, R.C. Borden, P.B. Bedient, C.H. Ward, and J.T. Wilson. 1 988.
Biorestoration of aquifers contaminated with organic compounds. CRC Crit. Rev.
Environ. Control 1 8:29-89.
26. Raymond, R.L. 1 974. Reclamation of hydrocarbon contaminated waters. U.S. Patent
Office 3,846,290.
102
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Mathematical Modeling of Intrinsic Bioremediation at Field Sites
Hanadi S. Rifai
Energy and Environmental Systems Institute, R
ce University, Houston, TX
Introduction
Intrinsic bioremediation is an important attenuation mechanism at contaminated field sites
because it limits pollutant migration and reduces contaminant mass in the subsurface.
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
bioremediation 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 presented, as
well as the difficulties in validating and verify ng these models. Finally, a few case studies of
modeling intrinsic bioremediation are reviewed.
Intrinsic Bioremediation Processes
Intrinsic bioremediation refers to the reduction of contaminant mass at a field site due to
biodegradation. This reduction 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 (anox/c conditions), microorganisms can use organic chemicals or
inorganic onions as alternate electron acceptors. Anaerobic biodegradation refers to
fermentative, denitrifying, iron-reducing, su/f^te-rec/uc/ng, or mefhanogen/c processes. To
quantify the impact of intrinsic bioremediation on contaminant concentrations at a field site, one
needs to develop an accurate picture of the c istribution 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
biodegradation may be occurring.
in
contaminated areas provides an indication that
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 expression that represents
the chemical reaction can be written to account for the effect that the rate of the reaction has
on biodegradation. This mathematical expression can then be combined with the transport
equation to account for the electron acceptor limitation effect on the biodegradation process
in the subsurface.
Many biodegradation models have been
kinetic expression for biodegradation (see Tc
number of aerobic and anaerobic biodegrad
developed in recent years, most of which utilize some
ble 1). The models listed in Table 1 simulate a
altion processes subjectto specified conditions and
Symposium on Intrinsic Bioremediation of Ground Water
103
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assumptions. The difficulties involved in applying these models include 1) the data required as
input to the model is lacking, such as the kinetic rate parameters 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 certain
level of expertise.
Table 1. Biodegradation Models
Name
-
BIOPLUME
-
BI01D
-
-
-
BIOPLUME II
-
BIOPLUS
ULTRA
-
-
Description
1 -D, aerobic, microcolony, Monod
1 -D, Monod
1 -D, analytical first-order
1-D, aerobic and anaerobic,
Monod
1 -D, co-metabolic, Monod
1 -D, aerobic, anaerobic, nutrient
limitations, microcolony, Monod
1-D, aerobic, co-metabolic,
multiple substrates, fermentative,
Monod
2-D, aerobic, instantaneous
2-D, Monod
2-D, aerobic, Monod
2-D, first-order
2-D, denitrification
2-D, Monod, Biofilm
Author(s)
Molzetal. (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. (1 3)
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 laboratory investigation programs
with modeling studies.
104
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at the Conroe Superfund site in Texas. The
operated as a wood-preserving facility from 1
ponds were composed of predominant
Conroe Superfund Site, Texas
Borden et al. (2) applied the first version of the BIOPLUME model to simulate biodegradation
United Creosoting Company (UCC) site was
946 to 1 972. Wastes disposed in two unlined
y polycyclic 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. The ground-water velocity at the UCC site is approximately 5 m/yr.
Oxygen exchange with the unsaturated zone was simulated 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 generating 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 Guc
in the northwestern portion of the lower penin: ula 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
rd Air Station located in Grand Traverse County
from 1 50 ft to 400 ft wide and is about 4,000
ft long. A pumping wellfield system was installed
at the downgradient end of the dissolved plune to control offsite migration.
A modeling effort of natural attenuation at ths 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 turned on. The model predict
concentrations at the monitoring wells reason ably well except in the vicinity of well M31. Rifai
et al. (8) indicated that this was because he simulation did not account for anaerobic
biodegradation, which was occurring in the interior of the plume.
ons by Rifai et al. (8) matched the observed
105
-------
a Obaervad Data
OBIOPUUME n Model
M28
U26 TP4 TP3
W30
M31
IN2W2
IN4W4
Figure 1. BIOPLUME II model predictions for the Traverse City field site (8).
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 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) concluded 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 etal. (14) evaluated a first-order decay biodegradation approach and the BIOPLUME
II model for simulating biodegradation at the gas plant facility. Using the model and assuming
first-order decay, several simulations were made to match the observed benzene concentration
distribution of 1/22/85 by setting the observed concentration distribution of 1 1/1/84 as the
initial condition. 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 distribution of February 1 987 as an initial condition. The data in Figure 2 show
the comparison between the measured and the simulated soluble BTX concentrations of July
1 987. As can be seen from Figure 2, the model predictions for BTX were reasonable. The model
predictions for oxygen concentrations, however, were not as similar. The authors attributed the
106
-------
differences to the fact that the BIOPLUME II
for 1 ppm of benzene, whereas the actual rec
Cliffs-Dow Superfund Site
model assumes a requirement of 3 ppm of oxygen
uirement is in the range of 1 ppm to 3 ppm.
Ground water at the Cliffs-Dow site is contam
compounds. The aquifer sediments at the site
hydraulic conductivity ranges between 3.5
contaminants found at the site near the source
phenols, and naphthalene at concentrations re
noted with low levels of phenolic and polycyclic
consist of mostly coarse sands and gravels. The
x 10"3 to 4.6 x 10"2 cm/sec. The principal
area include phenol, several methyl-substituted
nging from 220/fg/Lto 860/fg/L. Based on the
analysis of samples obtained from monitoring wells, Klecka et al. (4) found that the levels of
organic contaminants are reduced to near ojr below the detection limit within a distance of
1 00 m downgradientfrom the source. Further analyses of the ground-water chemistry were used
to verify that biodegradation was occurring a
contaminants.
• the site and causing the disappearance of the
The migration of organic constituents in th
e ac
uiferwas simulated using the BIO1 D model and
assuming a first-order decay expression. Half-lives for the contaminants at the site were
estimated from the results of soil 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 re nge of ground-water flow rates at the site.
Figure 3 illustrates the impact of biodegradction on contaminant concentrations at the site.
Model simulations performed using a half-li
components were reduced by greater than 99
of 2 d indicated that levels of the phenolic
percent within a distance of 30 m downgradient
of the source. When the half-life was increased by a factor of 1 0, the concentrations were
reduced to a similar extent within 75 m. Because of the dominance of biodegradation, 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.
developed overthe last 1 0 years. These models
Conclusions
A number of biodegradation models have bee
are generally similar in that they simulate the transport and biodegradation of a number of
components in the ground water. The models differ in the mathematical biodegradation
expressions that they use and in the numerical procedures used to solve the complicated system
of equations. Application of these models to field sites has proven to be complicated due to the
lack of biodegradation 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.
107
-------
= 2 days
t1/2 = 10 days
t1/2 = 20 days
Retardation factor = 5
Ground water velocity = 0.2 m/day
i i i i I i i i i
100 150
Distance (m)
(a)
200
250
t1/2 = 2 days
t1/2 = 10 days
t1/2 = 20 days
Retardation factor = 5
Ground water velocity = 0.46 m/day
J I I I I I I L_l I I I I !_
100 150
Distance (m)
(b)
200
250
Figure 2. BIOID model predictions for the Cliffs-Dow Superfund site (15).
108
-------
000
000
000
000
000
000
000
000
000
000
0
0
158
63
[el
|o I
0
0
0
0
0
0
[3"
15.
LP_
[4"
1°
82
0
0
0
5
532
923
1503
0 0
0 0
0 0
0 0
0 0
0 0
0
0
fo]
[oj
5317
3421
10690
10544
lol
[OJ
0
0
0
0
0
0
0
0
0
38
0
0
0
0
0
fo
o
o"
c«
0
0
0
!2
0
Uj
[0]
0
0
0
0
0
0
0
0
0
1387
1573
0
0
0
0
0
0
0
0
0
5876
5024
1301
1398
0
fol
lol
0
0
0
0
0
0
0
0
0
0
0 0
0 0
0 0
0 0
0 0
0 0
0 0
0 0
Flare pit
0 0
0 0
000
0 0
000
000
000
000
000
000
000
000
FigureS. BIOPLUME II model predictions fcr the gas plant facility in Michigan (14)—top
number: observed data; bottom number: simulated data.
References
i.
2.
3.
4.
5.
Molz, F.J., MA Widdowson, and LD.
dynamics coupled to nutrient and oxyg
22(8):1/207-1,216.
Borden, R.C., P.B. Bedient, M.D. Lee,
dissolved hydrocarbons influenced
application. Water Resour. Res. 13:1,
Benefield. 1 986. Simulation of microbial growth
n transport in porous media. Water Resour. Res.
C.H. Ward, and J.T. Wilson. 1 986. Transport of
by oxygen-limited biodegradation, 2. Field
983-1,990.
Domenico, PA 1 987. An analytical madel for multidimensional transport of a decaying
contaminant species. J. Hydrol. 91:49-58.
Srinivasan, P., and J.W. Mercer. 1 988. Simulation of biodegradation and sorption
processes in ground water. Ground Water 26(4):475-487.
Semprini, L., and P.L. McCarty. 1 991. Comparison between model simulations and field
results for in situ biorestoration of chlorinated aliphatics, 1. Biostimulation of
methanotrophic bacteria. Ground Water 29(3):365-374.
109
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7.
8.
9.
10.
6. Widdowson, M.A., F.J. Molz, and LD. Benefield. 1 988. A numerical transport model
for oxygen- and nitrate-based respiration linked to substrate and nutrient availability in
porous media. Water Resour. Res. 24(9):1,553-1,565.
Celia, M.A., J.S. Kindred, and I. Herrera. 1989. Contaminant transport and
biodegradation, 1. A numerical model for reactive transport in porous media. Water
Resour. Res. 25(6): 1,141-1,148.
Rifai, H.S., P.B. Bedient, J.T. Wilson, K.M. Miller, and J.M. Armstrong. 1988.
Biodegradation modeling at aviation fuel spill site. J. Environ. Eng. 1 14(5):1,007-1,029.
MacQuarrie, K.T.B., E.A. Sudicky, and E.O. Frind. 1 990. Simulation of biodegradable
organic contaminants in ground water, 1. Numerical formulation in principal directions.
Water Resour. Res. 26(2):207-222.
Wheeler, M.F., C.N. Dawson, P.B. Bedient, C.Y. Chiang, R.C. Borden, and H.S. Rifai.
1 987. Numerical simulation of microbial biodegradation of hydrocarbons in ground
water. Proceedings of the Solving Ground Water Problems With Models Conference,
Denver, CO (February 1 0-12). Dublin, OH: National Water Well Association (NWWA).
11. Tucker, W.A., C.T. Huang, J.M. Bral, and R.E. Dickinson. 1986. Development and
validation of the underground leak transport assessment model (ULTRA). Proceedings
of Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention,
Detection, and Restoration, Houston, TX (October-November). Dublin, OH: National
Water Well Association (NWWA). pp. 53-75.
12. Kinzelbach, W., W. Schafer, and J. Herzer. 1 991. Numerical modeling of natural and
enhanced denitrification processes in aquifers. Water Resour. Res. 27(6):1,1 23-1,1 35.
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
Restoration, 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.
1 5. Klecka, G.M., J.W. Davis, D.R. Gray, and S.S. Madsen. 1 990. Natural bioremediation
of organic contaminants in ground water: Cliffs-Dow Superfund site. Ground Water
4:534-543.
110
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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, an
U.S. Geological Survey, Reston, VA
Isabelle M. Cozzarelli
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. Upgradientfrom
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 reduction, 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 anoxia zone is indicated by the presence of structurally
related oxygenated intermediates (for example, alkylated benzenecarboxylic acids) and
differences in the removal rates of isomeric a Iky I 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 attiis site.
Introduction
Contamination of ground water by intentional or inadvertent 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 efforts. The efficacy of current engineering approaches is subject
to considerable debate. One thing is clear, however. For future remediation efforts to be
effective and useful, we must improve our understanding of how nature responds to such
impacts. In principle, this should permit the manipulation of natural processes for purposes of
contaminant removal. Investigations that have
oil spill site by the U.S. Geological Survey
understanding. The present paper discusses s
seen carried out at the Bemidji, Minnesota, crude
since 1 984 are aimed at developing such an
ome of the major findings of those studies.
Symposium on Intrinsic Bioremediation of Ground Water
111
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Methods
Sampling and Analysis
Wells were installed at the study site using a hollow stem auger without lubricants or grease 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 stainless steel screens (1.5 m long) intersected
the water table at the approximate midpoint of the screen; locations 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 samples to be used for determination of inorganic constituents and dissolved
organic carbon were collected with submersible pumps, whereas a Teflon bailer was used to
collect water for determination of organic constituents. The analyses included 1) water—
dissolved oxygen (DO), volatile dissolved organic carbon (VDOC), nonvolatile dissolved organic
carbon (NVDOC), methane, pH, Eh, major cations (Ca, Mg, Na, K), alkalinity, NH4+, NO3", Ch,
SO/', sulfide, iron (total 6 Fe2+), Mn2+, Si, Ba, Al, Sr, <513Cmethane, c513CTIC(TIC=total dissolved
inorganic carbon), volatile hydrocarbons (VHCs), extractable hydrocarbons, XAD resin isolates,
and low-molecular-weight organic acids; 2) oil—<313C, elemental analysis (C,H,N,S),
hydrocarbons, 13C-NMR, and heavy metals; and 3) sediment—(513C, hydrocarbons, and
elemental 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) forXAD resin isolates; by Baedeckerand Cozzarelli (7), Baedecker
et al. (8), and Bennett et al. (9) for inorganic constituents, methane, and elemental analyses;
and by Cozzarelli et al. (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 permeable till at about 24 m below
land surface. The outwash sediments are heterogeneous and composed of moderately
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 interbedded with gravel deposits and clay lenses. The water table is 6 m to
10m below land surface, and ground-water flow 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 nearthe watertable range from 0.05 meters/day (m/d) to 0.5 m/d (9) forfine-grained
and coarse-grained sediment, respectively.
112
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Sampling Sites
• Ground water
O Sediment
X Oil
X Oil and sediment
X Oil and ground water
Figure 1. Map of the field site near Bemidji showing location of the ruptured pipeline,
approximate location of the oil bodv, area over which oil was sprayed, and sampling
locations (2).
A pipeline rupture occurred in August 1 979, spilling 1,670 m3 of crude oil. A portion of the 410
m3 of crude oil unaccounted for after the clea
the water table. The oil body described here is
interval of unconsolidated sediment above the
iup effort is present in a body of oil floating on
irregularly distributed over a 7-m to 8-m vertical
water table; by 1 990, it had spread to a length
of 70 m to 80 m in the direction of ground-wa
received oil spray during the pipeline rupture,
This area, extending 140 m to 1 80 m to the w
approximately 6,500 m2, is hereafter referrec
erflow (13). An area upgradientof this oil body
and crude oil coated only the surface sediment.
est-southwest of the pipeline and encompassing
to as the "spray area" (see Figure 1).
113
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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
forthese 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 1 980s, 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 of ground-water flow (Figure 2), correspond to:
I—uncontaminated native ground water, II—ground water upgradient of the oil body but within
the "spray area," HI—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 downgradient, 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.
IB
f
435,-
430
425
420
415
Water table
Oil body
(( ((
-200
-150
-100 -50 0 50
Distance from center of oil body, meters
100
150
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). Fora detailed description
of zonation, see Baedecker et al. (8).
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
composition 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).
114
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Ground water within zone II is influenced by th
quantities were insufficient, however, for forma
e aerobic degradation of soluble oil constituents
originally deposited 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
ion of a discrete oil phase. Ground water in zone
II exhibits elevated concentrations of TDOC, Ga, Mg and HCO3"; lower pH (about 0.5 to 1 pH
unit); and a slightly reduced DO concentration. Carbon dioxide partial pressures (PCO2)
calculated by Bennett et al. (9) are as high as TO"1'5 atm (as compared with ambient
atmospheric values of 10"3-5atm). 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 ex ensive biodegradation. Presumably, some of the
degradation/oxidation occurred in the unsakirated 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 degradation is likely to be
occurring within the ground water as well.-If rremoval of the NVDOC is occurring via aerobic
respiration, however, the rate of supply of NVDOC must exceed its rate of removal because
NVDOC concentrations tend to increase with approach to zone 111. Bennett and others (9) have
suggested that the large increases in Ca, MgJ and HCO3" with increasing PCO2/ as we'l as the
similarity of the [Ca +Mg]/[HCO3-] mole ratio of ground water in this zone to that of the native
ground water, indicate thatthe principal reaction responsible forgenerating the plumes of alkali-
earth solutes is dissolution of carbonates By reaction with carbonic acid. Because these
processes are occurring under oxic (as opposed to anoxic) conditions, the primary chemical
signatures of the contaminant plumes reflect tie products of organic remineralization (HCO3~)
and transformation (TDOC) and those result
reactions.
ing from acidification (pH, Ca, Mg), not redox
Zone III encompasses ground water near, dovmgradientfrom (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
metabolism, leading to complete consumption of oxygen and the dominance of anoxic
degradation reactions, including iron and manganese reduction and methanogenesis. Sulfate
reduction and denitrification are not impo-tant in this system because of the very low
concentrations of SO42" and NO3" found in the native ground water. In this zone, TDOC
concentrations rise to a maximum of 48 mg/., 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 measurements of Eh indicate a strongly reducing
environment. In 1 987 the average stable carbon isotope ratios forTDOC ((513CT|C) and methane
((513Cmethane) were -8.23 and -55.45 per mil,
native ground water is -12.55 per mil. The
respectively. The average 513C value for TIC in
heavier ratio for TIC in zone III, thus, reflects
fractionation resulting from methanogenesis. The most dramatic changes in inorganic chemistry
are seen as large increases in the concentrati
>ns of dissolved iron and manganese that result
from mobilization of these redox species in response to the microbially mediated oxidation of
hydrocarbons. Silica concentrations also increase due to enhanced dissolution of silicate
minerals. Other compounds present in the ground water of zone III, but not found in the oil or
(except in trace amounts) ground water in zor es I and II, are a complex mixture of oxygenated
products of hydrocarbon degradation, including low-molecular-weight organic acids (10, 1 1]
The acids are structurally related to coexisting
monoaromatic hydrocarbons from the oil. Those
115
-------
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 metabolites.
Zone IV is a transition zone characterized by small but.detectable quantities of oxygen. Dissolved
iron and silica decrease to below detection limits at the boundary of this zone due to
precipitation reactions. Low-molecular-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 constituents
is conservative.
Evolution of the Contaminant Plume. Figure 3 depicts the concentrations of dissolved Fe2+,
Mn2+, methane, and <513CnC from ground water taken from zone III nearthe downgradient edge
of the oil body for the years 1 984 to 1 992 (8, 14). At this site, concentrations of methane and
Fe2+ increased by factors of 100 and 25, respectively, during the first 5 years. Thereafter, the
concentrations of methane and Fe2+ have virtually leveled off. Over the same period,
manganese concentrations first increased and then declined, whereas (513CTK; has increased
continuously. These variations in the contaminant plume chemistry reflect evolutionary changes
in the biology and geochemistry of this perturbed system with continued supply of hydrocarbons
and depletion of terminal electron acceptors.
i 1 1 T T 1
0.001
1984 1986 1988 1990 1992
Figure 3. Concentrations of dissolved ferrous iron, manganese, methane (in millimoles), and
<513C in per mil of TDOC for the years 1984 to 1992 (14).
116
-------
In particular, it would appear that manganese
•eduction was an important degradative process
at the earliest stages of plume evolution, b jt as time has progressed iron reduction and
methanogenesis have become dominant, presumably due to depletion of the supply of reducible
manganese. As of 1 992, the dominant process at this location appears to be methanogenesis.
This is supported by the steady increase in (513OT|C . During this same period, the distribution and
size of the contaminant plume, as measured by the concentrations of monoaromatic
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 bioc eochemistry 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 111 and IV. Benzene is the dominan individual VHC in anoxic ground water near the
leading edge of the oil body (70 percent), reacning concentrations in excess of 9,300 ^ag/L. The
C, .4 alkylated benzenes are second in abundarce, representing approximately 14 percent of the
total VHCs. Alkanes, consisting of a complex m xture of normal and branched hydrocarbons with
four to seven carbon atoms, account for T2 percent of the VHCs. Cyclic hydrocarbons
(cyclopentanes and cyclohexanes), cycloaromatics (e.g., alkylated indans), and heteroatomic
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: hydrocarbons with two to four a I kyl 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-butylpenzene, which was not detected in the crude oil,
are present in the contaminated ground water within zone 111. The composition of the
monoaromatic hydrocarbons changes systematically with distance downgradient from the oil
body and with increasing depth in the saturated zone (15, 16). Most importantly, isomeric
alkylbenzenes show dramatically different ap
parent removal rates downgradient from the oil
body. Because these isomers have similar physical properties, the attenuation of VHCs (and
therefore VDOC) is attributable to biological,
rather than physical, processes.
117
-------
1
§
1
u
u
10s
104
105
102
10°
10''
io-2
Toluene
20 40 60 80 100 120 140
Distance from Center of Oil Body (meters)
160
Figure 4. Concentrations of benzene, toluene, and ethylbenzene (ag/L) along the flowpath of
the contaminant plume. Also shown is the concentration of a conservative solute with
an assumed starting concentration of 10,000 fig/L. Model assumptions discussed
in Baedecker et al. (8); figure from Mallard and Baedecker (1 7).
This hypothesis is reinforced by data illustrated in Figure 4. Here concentrations of benzene,
toluene, ethylbenzene, 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 features of
these data are readily apparent. First, the concentration 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 homologous series with
increasing Kows (octanol-water partition 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
contaminants 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 methods
in the future.
118
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References
i.
2.
3.
4.
5.
6.
7.
8.
9.
10.
Hult, M.F. 1984. Ground-water contamination by crude oil at the Bemidji,
Minnesota, research site: An introduction. In: Proceedings of the U.S. Geological
Survey toxic-waste technical meeting, Tucson, AZ (March 20-22). Water Res. Invest.
Rep. 94-4188. pp. 1-15.
Eganhouse, R.P., M.J. Baedecker,
Dorsey. 1993. Crude oil in a s
geochemistry. Appl. Geochem. 8:551-567.
M. Cozzarelli, G.R. Aiken, K.A. Thorn, and T.F.
hallow sand and gravel aquifer, II. Organic
Eganhouse, R.P., T. Dorsey, C. Ph
C6-C10 aromatic hydrocarbons
chromatography. J. Chromatog. 6
Leenheer, J., and E.W.D. Huffman
nney, and S. Westcott. 1 993. Determination of
in water by purge-and-trap capillary gas
28:81-92.
1 979. Analytical method for dissolved organic
carbon fractionation. Water Res. Invest. Rep. No. 79-4.
Huffman, E.W.D., and H.A. Stube'. 1985. Analytical methodology for elemental
analysis of humic substances. In: Aiken, G.R., D.M. McKnight, R.L. Wershaw, and
P. MacCarthy, eds. Humic substances in soil, sediment and water. New York, NY:
John Wiley and Sons. pp. 433-456.
Thorn, K.A., and G.R. Aiken. 1 989. Characterization of nonvolatile organic acids
resulting from the biodegradation of crude oil by nuclear magnetic resonance
spectrometry. In: Mallard, G.E., and S.E. Ragone, eds. Proceedings of the technical
meeting of the U.S. Geological jSurvey Toxic Substances Hydrology Program,
Phoenix, AZ (September 26-30, 1 988). pp. 41-52.
Baedecker, M.J., and I.M. Cozzare
constituents in contaminated ground
Ground-water quality and analysis
Dekker. pp. 425-461.
illi. 1 992. The determination and fate of unstable
water. In: Lesage, S., and R.E. Jackson, eds.
at hazardous waste sites. New York, NY: Marcel
Baedecker, M.J., I.M. Cozzarelli, D.I. Siegel, P.C. Bennett, and R.P. Eganhouse.
1 993. Crude oil in a shallow sand and gravel aquifer, III. Biogeochemical reactions
and mass balance modeling in anoxic ground water. Appl. Geochem. 8:569-586.
Bennett, P.C. D.I. Siegel, M.J. Baedecker, I.M. Cozzarelli, and M.F. Hult. 1 993.
Crude oil in a shallow sand and |gravel aquifer, I. Hydrogeology and inorganic
geochemistry. Appl. Geochem. 8:529-549.
Cozzarelli, I.M., M.J. Baedecker, R.P. Eganhouse, and D.F. Goerlitz. 1994. The
geochemical evolution of low-molecular-weight organic acids derived from the
degradation of petroleum contaminants in ground water. Geochim. Cosmochim.
Acta 58:863-877.
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Cozzarelli, I.M., R.P. Eganhouse, and M.J. Baedecker. 1 990. Transformation of
monoaromatic hydrocarbons to organic acids in anoxic ground-water environment.
Environ. Geol. Water Sci. 16:135-141.
Franzi, D.A. 1 985. The surficial and subsurface distribution of aquifer sediments at
the Bemidji research site, Bemidji, Minnesota. Presented at the Second Toxic-Waste
Technical Meeting, Cape Cod, MA (October 21 -25).
Essaid, H.I., W.N. Herkelrath, and K.M. Hess. 1 991. Air, oil, and water distributions
at a crude-oil spill site, Bemidji, Minnesota. 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). pp. 614-620.
Baedecker, M.J., and I.M. Cozzarelli. 1994. Biogeochemical processes and
migration of aqueous constituents in ground water contaminated with crude oil. In:
Button, A.R., ed. Toxic substances and the hydrologic sciences. Minneapolis, MN:
American Institute of Hydrology, pp. 69-70.
Cozzarelli, I.M., R.P. Eganhouse, and M.J. Baedecker. 1 989. The fate and effects
of crude oil in a shallow aquifer, II. Evidence of anaerobic degradation of
monoaromatic hydrocarbons. In: Mallard, G.E., and S.E. Ragone, eds. In:
Proceedings of the technical meeting of the U.S. Geological Survey Toxic Substances
Hydrology Program, Phoenix, AZ (September 26-30, 1 988). Water Res. Invest. Rep.
88-4220. pp. 21-34.
Eganhouse, R.P., T.F. Dorsey, C.S. Phinney, M.J. Baedecker, and I.M. Cozzarelli.
1 987. 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.
Mallard, G.E., and M.J. Baedecker. 1 993. Hydrocarbon transport and degradation
in ground water: U.S. Geological Survey investigations. In: Pare, K.M., ed.
Proceedings of the Air Combat Command environmental quality 1 993 symposium,
Langley AFB, VA (March 1 -5) pp. 102-108.
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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 opportunity 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 transport processes are useful tools for integrating
information collected in the field, fortesting hypotheses, and forstudying the relative importance
of simultaneously occurring processes in complex, real-world systems. Numerical modeling of
multiphase flow iat the Bemidji site 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 degradation processes is being used as a tool to study the field-scale,
naturally occurring solute-transport and degrcdation processes occurring at the site.
Introduction
On August 20, 1 979, a buried ojl pipeline
(1 1,000 barrels) of crude oil (Figure 1). The
outwash plain. Depth to the water table rangi
lear Bemidji broke, spilling about 1.7 x TO6 L
site is located in a pitted and dissected glacial
s from 0 m to 8 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 pipe ine rupture. An estimated 1.2 x 1 O6 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 hydrocarbor
with the flowing ground water.
components that dissolve in and are transported
Numerical models that simulate flow and transport processes are useful tools for integrating
information collected in the field, fortesting hypotheses, and forstudying the relative importance
of simultaneously occurring processes in compl sx, 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 multiphase flow and
of ground-water transport and biodegradation at the site.
Symposium on Intrinsic Bioremediation of Ground Water
121
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95 5'40"
95 5'4"
47 34'32"
47 34'14"
Direction of
ground-water flow
Trace of north pool
simulated section
SCALE
0 50 100 150 Meters
I I I
Figure 1. The north oil pool at the Bemidji crude-oil spill site. Plus symbols are locations of
boreholes where samples were collected forsaturation analyses, and circles are wells
sampled for ground-water concentrations. The line A-A' is the trace of the multiphase
flow simulation section, and B-B' is the trace of the transport simulation section.
The Observed Subsurface Oil Distribution
The Bemidji field site has provided an opportunity to study in detail the subsurface oil distribution
following a spill. Field cores were collected for the purpose of determining the oil-saturation
distribution (the fraction of the pore space that is occupied by oil) in the subsurface.
Determination of the subsurface fluid-saturation distributions required the implementation of a
sampling technique 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 barrel 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 polyethylene (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 gravimetrically. Air saturation can then be
calculated by subtracting the sum of oil and water saturation from unity. Following the saturation
analysis, each sample was sieved to obtain the particle-size distribution.
122
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To date, samples have been collected and analyzed from 1 0 boreholes (550 subsamples) at the
site of the north oil pool (Figure 1). The three-dimensional distribution 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 locations
where oil infiltrated from a trench excavated
following 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 watertable is not lens-sha'ped. 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 simulatior
difficulty of measuring sediment hydraulic
particle-size distribution data were used to es
have been used to assess the effect of spatial
variability of hydraulic properties on the oil-saturation distribution at the site. Because of the
properties for oil-contaminated samples, the
imate the permeabilities (k) of the field samples
and the retention curves (7). The mean and irariance of the log(k) distribution (k in m2) were
-1 1.2 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)>-l 1.64) and a fine fraction
Kriged North Pool Oil Saturations
430
Elevation
(m)
422
0 Oil Saturation .75
Figure 2. Three-dimensional oil saturation
is at an elevation of about 423.5
di stribution at the north oil pool site. The watertable
n above sea level.
123
-------
99.99
ra
•§
.1
3
o
0.01
-15.
-14.
-13.
-12.
Log (k)
-10.
-9.0
Figure 3. Cumulative probability plot of log (permeability) at the north pool site.
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 realization was
obtained that reproduced the geometry of the fine and coarse fractions and also reproduced
the variability structure within these fractions. The details of this process are explained by Dillard
(9) and Dillard et al. (10).
Multiphase Flow Modeling
Atwo-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 equation for the oil and water phases, assuming that the air phase is
maintained at atmospheric pressure. An important feature of the model is that it incorporates
hysteretic relations between capillary pressures and fluid saturations. 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 1 979 until
the samples were collected in June 1990. The.simulated section was 120 m long and 10m
deep. The initial condition for the simulation was a hydrostatic water pressure distribution
corresponding to the measured water-table elevation at the time of sampling. The lateral and
bottom boundaries were assumed to be hydrostatic water-pressure boundaries. Oil was assumed
to infiltrate through five constant oil pressure nodes at the top boundary representing the trench
124
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that was excavated following the spill. Oil infiltration 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
01
and Dillard et al. (10).
Data from the first sample transect (Figure 4a) were compared 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/saturation fur ction parameters were used. A uniform k resulted
in a symmetric, lens-shaped oil body (Figure 4 D) that did not reflect the features of the observed
distribution. In the second simulation, spatial variability of hydraulic properties was
introduced, resulting in variability in oil saturations within the lens (Figure 4c). Very little oil was
trapped in the unsaturated zone. The shape o
complex and irregular, with zones of low oil s
silt layers. In the third simulation, hysteresis
the oil body floating on the water table became
aturation corresponding to the low permeability
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 distribution in the subsurface.
The Observed Ground-Water Plume
Numerous researchers working at the Bemidji
microbial activity and degradation of petrolei
:rude-oil spill site have documented evidence for
m hydrocarbons in the field (1 1 -1 8). As a result
of these studies, five geochemical zones in ground water (Figure 5) have been identified at the
site (1 1,1 2,1 9). Zone 1 consists of oxygenated, uncontaminated native ground water. Zone 2,
which is belowthe area where the land surface was sprayed by oil following the pipeline rupture,
characterized by reduced oxygen conce itrations and the presence of refractory high
is
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 conditions, 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 1 984 has shown that, near the water
table, the concentration of total dissolved organic carbon (TDOC) and dissolved oxygen (DO)
downgradient from the oil body has remained relatively stable with time. In the anoxic zone
(Zone 3), concentrations of reduced manganese (Mn) and iron (Fe) and of methane have
increased with time, indicating a sequence of Mn reduction followed by Fe reduction and
methanogenesis.
125
-------
430
Elevation
(m)
420
b 430
Elevation
(m)
420
C 430
Elevation
(m)
420
d 430
Elevation
(m)
420
Distance (m)
100
Distance (m)
120
Distance (m)
120
Distance (m)
120
0 Oil Saturation 1'°
Figure 4. Oil saturation distributions along the sample transect: a) observed oil saturations, b)
simulated oil saturations with uniform mean properties, c) simulated oil saturation
with spatially variable properties, d) simulated oil saturation with spatially variable
properties and hysteresis.
126
-------
100
RECHARGE ZONES
— < > — < > — *•
to 424
re
1U
8
S_ 420
LU
Q
t
x\ x i A \ ^
* V! Zone 2 \ \c Zc
Zono 1 ^ ^ "•'•^
X ^^-^
EXPLANATION j* ^^
X midpoint o( well x j
screen (
Datum is soa lovol /
Oil Body ^ "^
^UUIJIU^ x x x \ x 'X x x
ne 3 x *is * \. X
X /
v Zone 4 .. /
~-~-— _Ji— -^ x
x Dicoclion ol Ground-Water Flow
I I
DIST/
200
, IN METERS
300
400
Figure 5. The simulated cross section showing the five geochemical zones in ground water.
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 factors
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 multiple salutes and to include biodegradation terms. The
approach of Kindred and Celia (22) was used to represent 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 approximately parallel to the direction of
ground-water flow was simulated using the trc ns.port model for the period from the time of the
spill in ~\979 until September 1990 (Figure 5). Ground-water samples from numerous wells
along this section have been analyzed overtims (11,12,1 5). Steady-state flow, no sorption, and
isothermal conditions were assumed. Foursolu
TDOC was split into two fractions: degradable
tes and two microbial populations were modeled.
dissolved organic carbon (DDOC) and refractory
To represent the aerobic and anaerobic deg
populations of bacteria were included in the
dissolved organic carbon (RDOC). The remair ing two solutes modeled were DO and methane.
radation processes, aerobic and methanogenic
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 dissolution 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 recharge
water. The estimated values of initial concentrations 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 biodegradation
parameters under natural field conditions. Therefore, 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).
Table 1. Initial and Recharge Water Concentration (mg/L)"
Solute
DDOC
RDOC
TDOC
DO
Methane
Initial
Concentration
0.0
2.0
2.0
9.0
0.0
Recharge Zone
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.
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 1 986 and 1 990 concentration profiles. There is an increase in TDOC concentration
in the upgradient spray zone, followed by a rapid increase in TDOC concentration in the zone
of the oil body. Downgradient from the oil body, the TDOC concentration decreases gradually
to the background concentration. This decrease is a result of microbial and physical processes.
There is anaerobic degradation of DDOC within the anoxic zone near the oil body and aerobic
degradation of DDOC at the margins of the plume, where oxygenated 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 displacement and mixing of flowing ground water with
recharge water.
128
-------
TDOC
E 80
_J
DC
UJ
1 60
CD
| 40
1
5 20
111
o
100 200 300
DISTANCE, IN METERS
METHANE
tr
UJ
K
EC
UJ
Q-
30
400
S 20
Q 10
I
UJ
o
O
O
100 200 300
DISTANCE, IN METERS
Figure 6. Graphs of simulated and observed
b) DO, c) methane.
100 200 300 400
DISTANCE, IN METERS
EXPLANATION
— 1986 Simulated * 1987 Observed
— 1990 Simulated -i- 1888 Observed
A 1S86 Observed x 1990 Observed
400
concentrations at the water table: a) TDOC,
The simulated DO concentration (Figure 6b) decreases in the spray zone because of the
assumed decrease in DO concentration in recharge water, caused by degradation of
hydrocarbons in the unsaturated zone and by the consumption of oxygen by degradation in the
ground water. Nearthe oil body and immediately downgradientfrom it, the DO of the recharge
water is assumed to have been completely co isumed in the unsaturated zone. An anoxic zone
develops in this area. Farther downgradient, DO begins to increase as oxygenated recharge
water enters the system.
129
-------
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 1 986 to 1 990 as
the population of methanogens increases. This increase in methanogenesis results in a slight
decrease in TDOC concentrations from 1986 to 1990 (Figure 6a). The rate of increase in
methane production was quite sensitive to the biodegradation parameters used in the simulation.
To examine the effect of degradation on DDOC in the aquifer, two-dimensional distributions of
DDOC for three 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 hydraulic conductivity distribution is heterogeneous
(Figure 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.
RECHARGE ZONES
C z
-. 420
< 41G
EXPLANATION
X midpoint of wallscroon
zono boundary
_ 1 Isoconconlralion lino
Datum is soa lovol
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.
130
-------
In the second simulation, the anaerobic a
contaminant plume that is narrower than
id aerobic degradation processes result in a
the plume in the first simulation and whose
concentration gradients are comparatively she rp at the edges (Figure 7b). In this simulation, 46
percent of the total DDOC mass entering the aquifer is degraded: 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. (1 (D) 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 dispersion of ground water. This,
in turn, results in increased biodegradation. In this simulation, of the total DDOC mass entering
the aquifer, 60 percent is degraded: 21 percent by anaerobic degradation and 39 percent by
aerobic degradation.
The simulations represent a highly simplified representation of the true field conditions and
neglect Fe and Mn reduction. Also, the parameters used in the simulations are highly uncertain.
Nevertheless, the results do reproduce the general features of the observed contaminated
ground-water plume. In addition to the kinetics of the biodegradation processes, important
factors that affect 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 con
Juctivity.
opportunity to study in detail the processes that
Summary
The Bemidji crude-oil spill site has provided a
occur following a spill of an organic immiscible fluid that is slightly soluble in water. Detailed
field studies have characterized the subsurface oil distribution and the characteristics 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 aerob
used as a tool to study the field-scale solute-t
c and anaerobic degradation processes is being
•ansport and degradation processes. In addition
to the kinetics of the biodegradation processes, important factors that affect the distribution of
the solutes in the field are the recharge influx
ground-water system.
References
Hult, M.F. 1984. Ground-water
Minnesota, research site: An in
contamination by crude oil at the
Invest. Rep. 84-4188. pp. 1-15.
Hess, K.M., W.N. Herkelrath, and
fluid contents at a crude-oil spill site
and the degree of dispersion and mixing in the
contamination by crude oil at the Bemidji,
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9. Dillard, L.A. 1993. Multiphase flow modeling of a crude-oil spill site using
geostatistical simulation of soil hydraulic properties. M.S. thesis. Stanford, CA:
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10. Dillard, L.A., H.I. Essaid, and W.N. Herkelrath. 1 994. Multiphase flow modeling at
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94-4014. In press.
11. Baedecker, M.J., I.M. Cozzarelli, D.I. Siegel, P.C. Bennett, and R.P. Eganhouse.
1 993. Crude oil in a shallow sand and gravel aquifer, 3. Biogeochemical reactions
and mass balance modeling in anoxic ground water. Appl. Geochem. 8:569-586.
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Chang, F.H., H. Wang, B. Denz
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Geological Survey Toxic Substancjes Hydrology Program, Monterey, CA (March
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and M.J. Baedecker. 1 990. Transformation of
monoaromatic hydrocarbons to organic acids in anoxic ground-water environment.
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Dorsey. 1993. Crude oil in a
geochemistry. App. Geochem. 8:5
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organic rich ground water. Science 258:278-281.
Hult, M.F., M.K. London, and H.O.
models of mobilization and tran
unsaturated zone near Bemidji, Min
Proceedings of the Technical Meetir
Hydrology Program, Monterey, CA
4034. pp. 621-626.
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sport of volatile petroleum derivatives in the
lesota. In: Mallard, G.E., and D.A. Aronson, eds.
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March 1 1 -15,1 991). Water Res. Invest. Rep. 91 -
Lovley, D.R., M.J. Baedecker, D.J. lonergan, I.M. Cozzarelli, E.J.P. Phillips, and D.I.
Siegel. 1989. Oxidation of aronatic contaminants coupled to microbial iron
reduction. Nature 339:297-299.
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effects of crude oil in a shallow aquifer, I. The distribution of chemical species and
geochemical facies. In: Mallard, G.E., and S.E. Ragone, eds. Proceedings of the
Technical Meeting of the U.S. Geological Survey Toxic Substances Hydrology
Program, Phoenix, AZ (September 26-30, 1 988). Water Res. Invest. Rep. 88-4220.
pp. 13-20.
Goode, D.J., and LF. Konikow. 1 989. Modification of a method-of-characteristics
solute-transport model to incorporate decay and equilibrium-controlled sorption or
ion exchange. Water Res. Invest. Rep. 89-4030. p. 65.
Konikow, L.F., and J.D. Bredehoeft. 1 978. Computer model of two-dimensional
solute transport and dispersion in ground water. U.S. Geological Survey Techniques
of Water Resources Investigations.
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Kindred, J.S., and M.A. Celia. 1989. Contaminant transport and biodegradation, 2.
Conceptual model and test simulations. Water Resour. Res. 26(6):1,149-1,1 60.
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23. Essaid, H.I., MJ. Baedecker, and I.M. Cozzarelli. 1 994. Use of simulation to study
field-scale solute transport and biodegradation at the Bemidji, Minnesota, crude-oil
spill site. In: Morganwalp, D.W., and DA Aronson, eds. Proceedings of the
Technical Meeting of the U.S. Geological Survey Toxic Substances Hydrology
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94-4014 (in press).
134
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An Overview of Anaerobic Transformation of Chlorinated Solvents
Perry L. McCarty
Department of Civil Engineering, Stanford Uqiversity, Stanford, CA
Abstract
Intrinsic cometabolic transformation of chlorinated solvents commonly occurs at sites where
co-contaminants are present as primary substi
activities of transforming bacteria. The extent
relative concentration of primary substrates
conditions present. Reduction of tetrachloroeth
ates to support the energy needs and metabolic
of transformation that occurs depends upon the
and the microorganisms and environmental
sne (PCE) and trichloroethene (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 presence of
CAH transformation products and indicators of anaerobic biogical activity, such as the
disappearance of dissolved oxygen, nitrates, qnd sulfates, and the production of methane and
soluble iron (II).
Introduction
Chlorinated solvents and their natural transformation products represent the most prevalent
organic ground-water contaminants in the cou itry. These solvents, consisting primarily of CAHs,
have been used widely for degreasing of aircraft engines, automobile parts, electronic
components, and clothing. Only during the pcist 15 years has it become recognized that CAHs
can be transformed biologically (1). Such transformations sometimes occur under the
environmental conditions present in an aquifer in the absence of planned human intervention,
a process called intrinsic biotransformation (2)1 Conditions underwhich this is likely to occurwith
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 transformations. The
transformations are brought about through co-metabolism or through interactions of the CAHs
with enzymes or cofactors produced by tne microorganisms for other purposes. In co-
metabolism, other organic chemicals must be present to serve as primary substrates to satisfy
the energy needs of the microorganisms. 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.
Symposium on Intrinsic Bioremediation of Ground Water
135
-------
Chemical Transformation
TCA is the only major chlorinated solvent that can be transformed chemically in ground water
under all conditions likely to be found and within the one- to two-decade time span of general
interest, although chemical transformation of CT through reductive processes is a possibility.
TCA chemical transformation occurs by two different pathways, leading to the formation of
1,1-DCE and acetic acid (HAc):
CH3CC13
TCA
CH2=cci2 + H+ + cr
1,1-DCE
(elimination) (1)
CH3COOH + 3H+ + 3CT
HAc
(hydrolysis) (2)
The rate of each chemical transformation is given by the first-order reaction:
C = C,e
-kt
(3)
where C is the concentration of TCA at any time t, C0 represents the initial concentration at
t = 0, and k is a transformation rate constant. The overall rate constant for TCA transformation
(kTCA) is equal to the sum of the individual rate constants (kDCE + kHAc). The transformation 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 forAand E for TCA abiotic transformation by various investigators, and
calculated values for the TCA transformation rate constant for 1 0°C, 1 5°C, and 20°C using
equation 4. Also given is the average calculated TCA half-life based upon tl/2 = 0.69/k. The
temperature effect on TCA half-life is quite significant.
Table 1. Reported First-Order TCA Abiotic Transformation Rates (kTCA)
A
yri
3.47 (10)20
6.31 (10)20
1.56 (10)20
Average half-life (yr)
E
Id
118.0
119.3
116.1
10
°C
0.058
0.060
0.058
12
I
-------
Ciine and Delfino (4) found that kDCE equaled
found it to be 22 percent. This means that a
7 fig/L compared with 200,Mg/L forTCA. Wh<
about 21 percent of KTCA/ and Haag and Mill (3)
most 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 significance
because 1,1-DCE is considered more toxic than TCA, with a maximum contaminant level of
neverTCAis present as a contaminant, 1,1 -DCE
can also be expected. In general, TCA is probably the main source of 1,1 -DCE contamination
found in aquifers.
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 nor harmful products (6).
Biological Transformations
CAHs can be oxidized or reduced, generally
.ground waters, intrinsic reductive transformat
presence of intermediate products that are
through co-metabolism, as noted in Table 2. In
ons are most often noted, perhaps because the
formed provide strong evidence that reductive
transformations are taking place. Intrinsic aerobic transformation of TCE is also possible,
although if it did occur, the intermediate products are unstable and more difficult, analytically,
to measure. Thus, convincing evidence for the latter is difficult to obtain. Also, aerobic co-
metabolism of TCE would only occur if sufficient dissolved oxygen and a suitable electron donor,
such as methane, ammonia, or phenol, were present. Since circumstances under which the
proper environmental conditions for significant aerobic co-metabolism are unlikely to occur
often, intrinsic aerobic co-metabolism of TCE is probably of little significance. Evidence is ample,
however, that anaerobic reductived transformations of CAHs occur frequently, and this process
is important to the transformation of all chlorinated solvents and their transformation products.
The major environmental requirement is the presence of sufficient concentrations of other
organics that can serve as electron donors for energy metabolism, which 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
but much more is actually required because
be required to convert 1 g of PCE to ethene (7),
f the co-metabolic nature of the transformation.
Figure 1 illustrates the potential chemical and biological transformation pathways for the four
majorchlorinated solvents underanaerobicenyironmental conditions (6). Freedmanand 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 inorganic end
products and of PCE and TCE to ethene generally requires conditions suitable for methane
fermentation. Extensive reduction, although perhaps not complete, can also occur under sulfate-
reducing conditions. For methane fermentation to occur in an aquifer, the presence of sufficient
organic co-contaminant is required to reduce the oxygen, nitrate, nitrite, and sulfate present.
Some organics will be required to reduce the CAHs, and perhaps Fe(ll) as well, if present in
significant amounts. If the potential for intrinsic transformation of CAHs is to be evaluated, then
the concentrations of nitrate, nitrite, sulfate, Fe(ll), and methane, and of organics (as indicated
by COD ortotal organic carbon [TOC]) should be determined. Unfortunately, such analyses are
not considered essential in remedial investigations, but it is evident that they should be.
137
-------
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
Carbon
Tetrachloride
(CT)
Tetrachloro-
ethylene
(PCE)
Trichloro-
ethylene
(TCE)
1 ,1 ,1 -Tri-
chloroethane
(TCA)
Aerobic Biotransformation
Primary substrate
Co-metabolism
No
No
No
No
No
Yes
No
Perhaps
Anaerobic Biotransformation
Primary substrate
Co-metabolism
Hazardous intermediates
Chemical Transformation
No
Yes
Yes
Perhaps
Perhaps
Yes
Yes
No
Perhaps
Yes
Yes
No
No
Yes
Yes
Yes
CCI2 =CCI2 PCE
CT
CH 3 CH3
ANAEROBIC TRANSFORMATIONS OF CHLORINATED SOLVENTS
Figure 1. Anaerobic chemical and biological transformation pathways for chlorinated solvents.
138
-------
Table 3. Environmental Conditions Genera
Chlorinated Solvents
ly Associated With Reductive Transformations of
Chlorinated Solvent
Carbon tetrachloride
1 ,1 ,1 -Trichloro-
ethane
Tetrachloroethylene
Trichloroethylene
Redox Environment
All
TCA-*1,1-DCE
+ CH3COOH
E
c
(
)enitrifi-
ation
T-»CF
Sulfate
Reduction
CT->C02+C|-
TCA^1,1-DCA
PCE^1,2-DCE
TCE^1,2-DCE
Metha no-
genesis
TCA-*C02+d-
PCE-s-ethene
TCE->ethene
Case Studies
Major et al. (10) reported field evidence for intrinsic bioremediation of PCE to ethene and
ethane at a chemical transfer facility in North Toronto. PCE was stored at the site 1 0 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 forthe transforming organisms. Where
high PCE was found, TCE (1.7 mg/L), cis-DJCE (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 (0.42 mg/L) were present,
suggesting thatsignificantdehalogenation had occurred. Otherdichloroethylenes (1,1 -DCEand
trans-DCE) were not significant in concentration,, indicating that cis-DCE was the major
transformation intermediate. Microcosm studies also supported that biotransformation was
occurring at the site, with complete disappearance of PCE, TCE, and cis-DCE and production
of both VC and ethene. The conversions were accompanied by significant methane production,
indicating that suitable redox conditions were present for the transformation.
Fiorenza et al. (1 1) reported on PCE, TCE, T
3A, and dichloromethane (DCM) contamination
of ground water at two separate locations at a carpet-backing manufacturing plant in
Hawkesbury, Ontario. The waste lagoon was pie 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 forthe dehalogenation
reactions. Here, the sulfate concentration was nondetected, but the concentration in native
ground water was about 1 5 to 18 mg/L. Total dissolved iron was quite high (1 9.5 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 suitable for
intrinsic biodegradation of the chlorinated solvents. Whilesome chemical transformation of TCA
was indicated (0.4 mg/L), biotransformation was quite extensive, as indicated by a 1,1-DCA
concentration of 7.2 mg/L compared with the TCA concentration of 5.5 mg/L. Some CA was
also present (0.1 9 mg/L). Transformation was also indicated for PCE and TCE, which remained
at concentrations of only 0.01 6 mg/L and 1.5 mg/L, respectively, while the cis-DCE, VC, and
139
-------
ethene concentrations were 56, 4.2, and 0.076 mg/L, respectively. Only traces of ethane were
found. Trans-DCE concentration was only 0.57 mg/L, again providing evidence that cis-DCE
is the most common transformation intermediate from TCE and PCE. Downgradient from the
lagoon, the dominant products were cis-DCE (4.5 mg/L), VC (5.2 mg/L), and 1,1 -DCA (2.1
mg/L). While good evidence for 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 chlorinated solvent hazard at the site, although it was
producing compounds that may be more susceptible to aerobic co-metabolism.
Evidence for intrinsic biotransformation of chlorinated solvents has also been provided from
analyses of gas from municipal refuse landfills where active methane fermenation exists. A
summary by McCarty and Reinhard (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. ForTCA, gaseous concentrations were: TCA, 0.1 7; 1,1 -DCE,
0.10; 1,1 -DCA, 2.5; and CA, 0.37. These data indicate that TCA biotransformation was quite
extensive, with the transformation intermediate, 1,1 -DCA, present at quite significant levels, as
is frequently found in ground water.
Perhaps the most extensively studied and reported intrinsic chlorinated solvent biodegradation
is that at the St. Joseph, Michigan, Superfund site (7, 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. (1 7) 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
sulfate 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. 87-11 6.
2. Council, N.R. 1 993. 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 epoxides. Environ. Sci. Technol. 22:658-663.
140
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4.
5.
6.
7.
8.
9.
10.
n.
12.
13.
14.
15.
Cline, P.V., and J.J. Delfino. 1 989.
to the stable product 1,1 -dichloroe
water treatment. Chelsea, Ml: Lew^s
Jeffers, P., L. Ward, L. Woytowitch
Transformation kinetics of 1,1,1 -trichloroethane
hene. In: R.A. Larson, ed. Biohazards of drinking
Publishers, Inc. pp. 47-56.
and L. Wolfe. 1 989. Homogeneous hydrolysis
rate constants for selected chlorine ted methanes, ethanes, ethenes, and propones.
Environ. Sci. Technol. 23(8): 965-969.
Vogel, T.M., C.S. Griddle, and P.L
aliphatic compounds. Environ. Sci,
vlcCarty. 1 987. Transformations of halogenated
Technol. 21:722-736.
McCarty, P.L, and J.T. Wilson. 1 992. Natural anaerobic treatment of a TCE plume,
St. Joseph, Michigan, NPL site. In: U.S. EPA. Bioremediation of hazardous wastes.
EPA/600/R-92/126. Cincinnati, OH. pp. 47-50.
Freedman, D.L., and J.M. Gossett.
tetrachloroethylene and trichloroeth yl
Appl. Environ. Microbiol. 55(9):2,
1 989. Biological reductive dechlorination of
ene to ethylene under methanogenic conditions.
144-2,151.
de Bruin, W.P., et al. 1992. Complete biological reductive fransformation of
tetrachloroethene to ethane. Appl.
Major, D.W., W.W. Hodgins, and
Environ. Microbiol. 58(6):1,996-2,000.
3.J. Butler. 1 991. Field and laboratory evidence
of in situ biotransformation of tetrachloroethene to ethene and ethane at a chemical
transfer facility in North Toronto. In: Hinchee, R.E., and R.F. Olfenbuttel, eds. Onsite
bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 147-171.
Fiorenza, S., et al. 1 994. Natural anaerobic degradation of chlorinated solvents at
a Canadian manufacturing plant. In: Hinchee, R.E., A. Leeson, L. Semprini, and S.K.
Ong, eds. Bioremediation of chliorinated and polycyclic aromatic hydrocarbon
pounds. Boca Raton, FL: Lewis Publishers, Inc. pp. 277-286.
com
McCarty, P.L., and M. Reinhard. 1993. Biological and chemical transformations of
halogenated aliphatic compounds in aquatic and terrestrial environments. In:
Oremland, R.S., ed. The biogeocnemistry of global change: Radiative trace gases.
New York, NY: Chapman & Hall, Inc.
Charnley, G., E.A.C. Crouch, L.G. Green, and T.L. Lash. 1988. Municipal solid
waste landfilling: A review of environmental effects. No. Meta Systems, Inc.
Hasten, Z.C., P.K. Sharma, J.N. Black, and P.L. McCarty. 1 994. Enhanced reductive
dechlorination of chlorinated ethenes. In: U.S. EPA. Bioremediation of hazardous
wastes. San Francisco, CA.
Kitanidis, P.K., L. Semprini, D.rj. Kampbell, and J.T. Wilson. 1993. Natural
anaerobic bioremediation of TCE ctthe St. Joseph, Michigan, Superfund site. In: U.S.
EPA. Symposium on bioremediation of hazardous wastes: Research, development,
and field evaluations (abstracts). EPA/600/R-93/054. Washington, DC (May).
Cincinnati, OH. pp. 47-50.
141
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1 6. McCarfy, P.L., et al. 1 991. In situ methanotrophic bioremediation for contaminated
ground water at St. Joseph, Michigan. In: Hinchee, R.E., and R.G. Olfenbuttel, eels.
Onsite bioreclamation processes forxenobiotic and hydrocarbon treatment. Boston,
MA: Butterworth-Heinemann. pp. 1 6-40.
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. Presented at the U.S.
EPA Symposium on Intrinsic Bioremediation of Ground Water, Denver, CO (August
30 to September 1).
142
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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 Pic
jtinny Arsenal in Morris County, New Jersey, is
contaminated withtrichloroethylene (TCE). Estimated average linearground-waterflowvelocities
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 1 30,000-m2 plume area is estimated to be
970 kg. About 65 percent of the mass is in areas where TCE concentrations exceed 10,000
ftg/t, whereas about 30 percent is in areas vhere TCE concentrations are 1,000 to 10,000
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 unconfined aquifer and measured
base-flow discharge in Green Pond Brook, s 1 to 2 mg/sec. Biotransformation is the most
important mechanism by which dissolved TC: leaves the ground-water system.
Introduction
Picatinny Arsenal, a U.S. Army armament re
glaciated valley in north-central New Jersey (
and a new metal-plating facility and industria
1 960 to 1 981, the wastewater treatment s
search and development center, is located in a
Figure 1). In 1960, Building 24 was remodeled,
wastewater treatment plant were installed. From
stem discharged tens of thousands of liters of
wastewater daily into two 2.5-m deep, sand-bottomed settling lagoons behind the building (1).
The metal-plating wastewater contained trace metals, such as cadmium, chromium, copper,
lead, nickel, tin, vanadium, and zinc, and othpr ions used in plating solutions, such as sodium,
potassium, sulfate, chloride, and cyanide (2). From 1 973 to 1 985, 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 1 983, 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
porosities, 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
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
of these estimated velocities, the residence time
Symposium on Intrinsic Bioremediation of Ground Water
143
-------
A. Areal extent of trichloroethylene
contaminant plume
EXPLANATION
Ky*:|:'x) Area in which trichloroethylene concentration
l^ili'a exceeds 10 micrognuns per liter
10 LINE OF EQUAL TRICHLOROETHYLENE
CONCENT'RATION-Shows trichloroethylene
ennCCTTff tjon, in nucrognu&s per liter.
Dashed where approximate
A "A' Line of section
*41 -9 Sampling site location and local identifier
BOO 400 200 0 200 400 600 800 1.000 1.200 1.40O 1.6O01.800 2.000
DISTANCE FROM BUILDING 24 CB-24). IN FEET
B. Vertical distribution of trichloroethylene
concentrations
EXPLANATION
-100 LINE OF EQUAL TRICHLOROETHYLENE
CONCENTRATION-Shows trichloroeibyiene
conccxumion, in nucrogninu per liter.
Dashed where approximate
Weil sacea ani tricblorodhylene conccnmtlion.
Utoo
NS
Not sampled
Less than
Localioa of well and local identifier
Figure 1. (A) Location of Building 24 study area at Picatinny Arsenal, and areal extent of TCE
plume and altitude of water table, January 1 993, and (B) vertical distribution of TCE
concentrations, October to November 1 991. (Location of section A-A' is shown in
Figure 1A.)
144
-------
Ground-water contamination in the unconfinec
et al. (3), Fusillo et al. (2), and Imbrigiotta et c
aquifer has been discussed previously by Sargent
I. (4,5). Results of water-quality sampling during
1 986 to 1 991 confirm that TCE remains the dominant contaminant in the unconfined aquifer
and that the extent of the plume of TCE-contaminated ground water has changed little since
September 1 986. These water-quality measurements also show that 1) Building 24 is the source
of the TCE plume; 2) the highest TCE concen
less than 6 m and within 10 m downgradienl
from Building 24 to Green Pond Brook and fo
rations near Building 24 are found at depths of
from the dry well; 3) the plume extends 500 m
lows the ground-water head gradients as it flows
downward 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 1 30,000 m2; 5) the highest concentrations of TCE are
found at the base of the unconfined aquifer
Brook; and 6) the highest TCE concentrations
than 6 m (Figure 1).
midway between Building 24 and Green Pond
at Green Pond Brook are found at depths of less
ard
The U.S. Geological Survey is conducting an
contamination by the chlorinated solvents
Picatinny Arsenal. The objectives of the
biological processes that affect the movement
in the subsurface; 2) determine the relative
predictive models of contaminant transport.
interdisciplinary research study of ground-water
other contaminants at the Building 24 site at
are to 1) describe the chemical, physical, and
and fate of these contaminants, particularly TCE,
mportance of these processes; and 3) develop
This paper describes a conceptual model o
chlorinated solvents at Picatinny Arsenal resea
in the unconfined aquifer at the site
solute-transport simulations.
I tie
processes that affect the fate and transport of
•ch site. A preliminary solute mass balance of TCE
is presented and compared with results of numerical
Conceptual Model
A conceptual model of the physical, chemical,
and mass balance of TCE within the plume
(Figure 2). TCE is present at the site in severe
phase (soil gas), and matrix phase (sorbed o
also may be present as a dense nonaqueous
of TCE from Building 24 to the base of the
of high concentrations of TCE at the base of
Results of chemical analyses of samples of
the plume area indicate that the fate and
chemical, and biological processes at the site
affect the movement of dissolved TCE and
discharge to Green Pond Brook. TCE als
biotransformation (reductive dehalogenation
saturated-zone sediments. Volatilization of TCJE
by Smith and others (7). Transport in the
principally by molecular diffusion. The
cis-1 -2-dichloroethylene (cis-DCE) and vinyl c
and biological processes that affect the transport
was developed by Imbrigiotta and Martin (6)
forms: solute phase (dissolved in water), vapor
ito solid surfaces or associated with biota). TCE
-phase liquid (DNAPL). Nonaqueous-phase flow
ur confined aquifer is hypothesized to be the cause
the unconfined aquifer (5).
ground
water, soil gas, and aquifer sediments from
fransport of TCE are affected by the physical,
. Advection and dispersion in the saturated zone
cause it to be removed from the system in the
removed from the system by anaerobic
o is
, volatilization at the water table, and sorption to
to the unsaturated-zone soil gas was measured
imsaturated zone was determined to be driven
detection of the biotransformation products
hloride (VC) in waterfrom 75 percent of the wells
145
-------
VOLATILIZATION
Bnildini 24 (0.1 mg/s)
\ W«er table
LONG-TERM
DE3ORPTION
(1S-8S mg/s)
SHORT-TERM
DESOHPTION
(Not estimated)
ADVECTIVE
TRANSPORT TO
QREEN PONO BROOK
(1-2 mg/s)
Land surface
TH>'C.H:LORO T.«YUEJ*E
DISSOLUTION
(Not estimated)
ANAEROBIC
BIOTRANSFORMATION
(1-3O mg/s)
Estimated top of confining unit
NOT TO SCALE
GAINS
TRICHLOROETHYLENE MASS BALANCE COMPONENTS
[mg/s. milligrams par second; <. lass than]
LOSSES
LONG-TERM DESORPTION
INFILTRATION
DISSOLUTION
SHORT-TERM DESORPTION
15-85 mg/s
< 0.1 mg/s
Not estimated
Not estimated
ANAEROBIC BIOTRANSFORMATION 1-30 mg/s
ADVECTIVE TRANSPORT TO
GREEN POND BROOK 1-2 mg/s
VOLATILIZATION 0.1 mg/s
SHORT-TERM ADSORPTION 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.
screened within the plume is indicative of biologically mediated, reductive dehalogenation of
TCE. The occurrence of methanogenesis is indicated by the detection of dissolved methane in
water from 85 percent of the wells in which cis-DCE was detected and in water from 94 percent
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 concentrations measured near Building 24 and at the base of the
unconfined aquiferfrom 1986 to 1991 indicate that slow desorption of TCE from contaminated
146
-------
aquifer sediments, dissolution of DNAPL TCE n 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
concentration is assumed to represent the TCE concentration of a volume of ground water
extending half the horizontal and vertical d stance to the adjacent sampling points. These
volumes are limited vertically by the extent OT the unconfined aquifer and horizontally by the
1 0-jug/L maximum TCE concentration line. Aquifer porosity is assumed to be 30 percent. The
total volume of dissolved TCE within the ground water outside the 10-^g/L maximum TCE
concentration line is estimated to be about 1 L.
The estimated total volume of TCE appears to be unrelated to the total number of samples
collected. Four sets of data plotted in Figure 8 (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
1 0,000 jug/L, and about 30 percent is founc
1,000 to 1 0,000 pg/L The estimate of the
mostly on the number of samples that conta
volume of water each sample is assumed to
The relative amounts of TCE in the dissolved
in areas where TCE concentrations range from
volume of TCE in the water appears to depend
n relatively high concentrations of TCE and the
represent.
vapor, and matrix phases in a block of aquifer
and unsaturated zone immediately downgradientfrom Building 24 were estimated on the basis
of measured TCE concentrations in samples cf all three phases (8) (Figure 4). Most of the 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 representative sample of each TCE phas 5 in the area of Building 24. The ratio of the mass
of TCE sorbed to the soil to the mass of disso ved 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 unusual
TCE at the site has not been estimated.
y high ratio of about 20:1. The amount of DNAPL
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
because the direct release of TCE from Build
DNAPL TCE has not been estimated, the rate of TCE dissolution cannot be estimated. Three
first-order rate constants of TCE desorption
calculated by Koller et al. (8) ranged from 0.
measured in laboratory flowthrough columns
which most TCE enters the ground-water system,
ng 24 stopped in 1 985. Because the amount of
from shallow aquifer sediments at the arsenal
003 to 0.01 5 per week. The rate constants were
using uncontaminated water as the influent fluid.
Desorption rates in the field, where ground water containing 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 estimated fluxof TCE
into the ground-water system through desorpti
on is 15 to 85 mg/sec. This flux estimate is made
147
-------
900
£ 800
1U
700 -
UJ
g 600 |-
UJ
O
cc
3
X
O
£
i-
500
400
300
O
111 ZOO
13
_!
o 100
Sample set A
Number of samples 96
B
39
C
40
D
59
e
50
F
42
EXPLANATION
(jig;L, mlcrograms pur llt«r: s, less than or oqual to: >, groator than)
Sample
sal
Volume of Irichloroathylana calculated
using samples with Irichloroathylana
concentrations of:
m 1 to £100u.g/L
gjjgg >100 to s1.000ng/L
>1,000 to 510.000|ig/L
>10,000|ig/L
Data of collection...
1|300
1,200
1.100
1,000
900
800
700
600
500
^.
O
400 K
1-
300 £
200 CO
100
0
A April and August-Soptambor 19S6
B August 1987
C June 1989
0 November-December 1989
E April-March 1S90
F October-November 1991
Figure 3. Estimated volume of dissolved TCE in ground water at the Building 24 site at Picatinny
Arsenal, 1986-91.
24 Mass distril3ution of tricnoloroethylene near building 24
[kg, kilogram; <, less than]
NOT TO SCALE
Phase
Percent
Mass (kg) of total
10 feet
10 faat
Unsaturated
zone
Saturated
zone
Vapor
Matrix
Solute
Matrix
0.001
3
1
4
99.9
20
80
Figure 4. Mass distribution of TCE in the saturated and unsaturated zones immediately
downgradient from Building 24 at Picatinny Arsenal.
148
-------
by assuming that, over long periods (years), the short-term desorption rate (weeks and months)
is equal to the short-term adsorption rate, anc
adsorption, which is no longer occurring.
that soils previously have undergone long-term
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 wateirand measured base-flow discharge in the brook,
is 1 to 2 mg/sec. The flux of TCE volatilized fj-om the water table is estimated to be about 0.1
mg/sec 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 constan
0.001 to 0.02 per week were estimated by Wi
microcosm studies of soil from five sites within
s for TCE transformation ranging from less than
son et al. (9) on the basis of results of laboratory
he 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 calculated to be about 1 to 30 mg/sec. 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 actual flux
greater than that shown in Figure 2.
A reactive multispecies transport model of a two-dimensional vertical section along the central
axis of the plume is being used to analyze the
chemical transport characteristics and th
calibration, and sensitivity analysis have been
volatilization, and microbial degradation of T
of TCE lost through biotransformation may be
aboratory and field estimates of the physical and
timated TCE mass balance. The model design,
described by Martin (1 1). Transport, desorption,
IE are simulated. The formation and transport of
the degradation products cis-DCE and VC also are simulated. Specified-flux boundary conditions
are used to represent ground-water recharge and flows across the horizontal 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 ordissolution 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 cverage steady-state flow conditions and virtually
steady-state transportconditions after 1 985. S
the calibrated mode! are shown in Figure 5.
imulated concentrations of TCE and cis-DCE from
Most concentrations were simulated to within an
order of magnitude of average concentrations 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 estimate of the magnitudes of the various
mass-balance components of the plume of TCE-contaminated ground water at the Building 24
site; however, sensitivity simulations were use
ground water. Results of a series of sensitivity
general conceptual model as defined by the estimated solute mass balance presented above.
to test hypotheses concerning the fate of TCE in
simulations discussed by Martin (11) support the
149
-------
(A) Trichloroethylene
185 • —
1OO 0 100 200 300 400 SOO
DISTANCE FROM BUILDING 24 (B-24). IN METERS
(B) cis-l-2-Dichlordethylene
IBS -
DISTANCE FROM BUILDING 24 (B-24). IN METERS
EXPLANATION
—100 LINE OF EQUAL SOLUTE CONCENTRATION-
-Shows solute concentration, in micrograms per liter.
Q 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.)
150
-------
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-be
an important mechanism for removing solutes
the water table. The overall flux of TCE into
lance component, this process was shown to be
and thereby affecting solute concentrations near
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 urconfined aquifer might result in a reasonable
simulated TCE mass balance.
chlorinated solvents from the dry well at Building
Summary
Infiltration of wastewaterfrom the lagoons anc
24 at Picatinny Arsenal, New Jersey, has created a plume of contaminated ground water
downgradient from the building. TCE is the predominant contaminant in the 1 30,000 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 highest 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 wa
solid surfaces or associated with biota. TCE a
of dissolved TCE below the water table is est
of TCE in the system is associated with the se
A conceptual model of the physical, chemical,
and mass balance of TCE within the plume ir
the ground-water system by discharge to
volatilization; and 3) gain of dissolved TCE
sediments and possibly from dissolution of D
er, as a vapor in the soil gas, and sorbed onto
so may be present as a DNAPL. The total volume
mated to be about 660 L, but most of the mass
diments in the saturated and unsaturated zones.
and biological processes that affect the transport
eludes 1) transport of TCE from the Building 24
source area to Green Pond Brook by advection and dispersion; 2) loss of dissolved TCE from
Green Pond Brook, biotransformation, and
by slow desorption from contaminated aquifer
IAPLTCE in the aquifer. Preliminary estimates of
the flux of dissolved TCE discharged to Gredn Pond Brook from the plume is 1 to 2 mg/sec.
Most dissolved TCE leaves the ground-water system by means of biotransformation. The
estimated flux of TCE out of the system by th s 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 estimated to be about an order of
magnitude less than the flux of TCE discharge to Green Pond Brook, volatilization is an
important mechanism for removing solutes and thereby affecting solute concentrations near the
water table.
151
-------
References
1. Benioff, PA, M.H. Bhattacharyya, C. Biang, S.Y. Chiu, S. Miller, T. Fatten, D. Pearl,
A. Yonk, and C.R. Yuen. 1 990. Remedial investigation concept plan for Picatinny
Arsenal, Vol. 2. Descriptions of and sampling plans for remedial investigation sites.
Argonne, IL: Argonne National Laboratory, Environmental Assessment and Information
Sciences Division, pp. 22-1 to 22-24.
2. Fusillo, T.V., T.A. Ehlke, M. Martin, and B.P. Sargent. 1 987. Movement and fate of
chlorinated solvents in ground water: Preliminary results and future research plans. In:
Franks, B.J., ed. Proceedings of the U.S. Geological Survey Program on Toxic
Waste—Ground-Water Contamination, Pensacola, 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. Vowinkel. 1 986. Ground-water-quality
data for Picatinny Arsenal, New Jersey, 1 958-85. U.S. Geological Survey Open File
Rep. 86-58. 66 pp. ;
4. Imbrigiotta, I.E., M. Martin, B.P. Sargent, and L.M. Voronin. 1 989. Preliminary results
of a study of the chemistry of ground water at the Building 24 research site, Picatinny
Arsenal, New Jersey. In: Mallard, G.E., and S.E. Ragone, eds. Proceedings of the U.S.
Geological Survey Toxic Substances Hydrology 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 Hydrology Program, Monterey, CA (March 1 1 -1 5).
Water Res. Invest. Rep. 91-4034. pp. 681-688.
6. Imbrigiotta, T.E., and M.Martin. 1 991. 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, Monterey, CA (March 1 1 -1 5). Water Res. Invest. Rep.
91-4034. pp. 673-680.
7. Smith, J.A., C.T. Chiou, J.A. Kammer, and D.E. Kile. 1 990. Effect of soil moisture on
sorption of trichloroethene 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. Proceedings of the U.S. Geological Survey
Toxic Substances Hydrology Program, Colorado Springs, CO (September 20-24,1 993).
Water Res. Invest. Rep. 94-4014. In press.
152
-------
9.
Wilson, B.H., T.A. Ehlke, I.E.
dechlorination of trichloroethyiene in
New Jersey. In: Mallard, G.E., and
Geological Survey Toxic Substances H
Water Res. Invest. Rep. 91-4034. pp.
10.
Imbrigiotta, and J.T. Wilson. 1991. Reductive
anoxic aquifer material from Picatinny Arsenal,
D.A. Aronson, eds. Proceedings of the U.S.
'drology Program, Monterey, CA (March 1 1-15).
704-707.
Ehlke, T.A., B.H. Wilson, J.T. Wilson, and T.E. Imbrigiotta. 1994. In situ
biotransformation of trichloroethyiene 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
(September 20-24, 1 993). Water Res
11.
Hydrology Program, Colorado Springs, CO
. Invest. Rep. 94-4014. In press.
Martin, M. 1994. Simulation of transport, desorption, volatilization, and microbial
degradation of trichloroethyiene in ground water at Picatinny Arsenal, New Jersey. In:
Morganwalp, D.W., and D.A. Aronson|, eds. Proceedings of the U.S. Geological Survey
Toxic Substances Hydrology Program, Colorado Springs, CO (September20-24,1 993).
Water Res. Invest. Rep. 94-4014. In press.
153
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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 determined 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, because 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 sampling of the
western plume near Lake Michigan, to estimate 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 1 7 borings with a slotted auger. In 1 992,
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 concentration. Three of the transects (2, 4, and 5) are roughly perpendicular
to the contaminant plume. Of the remaining transects, transect 1 crosses the plume at an angle
and transects lies along the length of the plume. The perpendiculartransects form logical units
for study of TCE biotransformation.
154
Symposium on Intrinsic Bioremediation of Ground Weter
-------
Approx. Extent of TCE Plume
II
St Joseph, Michigan
NPL Site
measurement point. The blocks are defined
Figure 1. Site plan, St. Joseph, Michigan, NPL site.
The site data from the transects are visualized as sets of blocks that are centered around the
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-dimensional version of the fully three-
dimensional field-data analysis program calle<
S. Kerr Environmental Research Laboratory.
The mass of each chemical per unit thickness
calculated by summing over the blocks. By f
concentrations are not extrapolated into the c
d SITE-3Dthat is under development at the Robert
and the advective mass flux of each chemical are
llowing this procedure, the measured chemical
ay layer under the site, nor are they extrapolated
beyond a short distance from the measurement locations (5 ft vertically and 50 ft to TOO ft
horizontally). Other interpolation schemes such as inverse distance weighting or kriging could
also be used to estimate the concentration fie
3 show the distributions of the VC and TCE a
"color" scale. Notably, the maximum VC con
transect 5 was 205,Mg/L. The maximum TCE
d and perform the mass estimates. Figures 2 and
transect 5 using a logarithmic, black-and-white
centration at transect 4 was 1,660 jug/L and at
concentration at transect 4 was 8,720/ig/L and
at transect 5 was 163 ^g/L. As noted previously for other portions of the site (2, 4), the
contamination is found near the bottom of the aquifer. The highest concentrations of VC and
TCE do not appear to be co-located. In
perpendicular transects ordered from farthesn
(transect 5). The data in Table 1 represent tre 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.
able 1, mass estimates are presented for the
upgradient (transect 2) to farthest downgradient
155
-------
St. Joseph, Michigan
Vinyl Chloride
transect: 5
mass: 0.4811 E-01 Kg/m
151 tS2
tS4 tS3 tSS
Ground surfica
10faet,
approx. N
100 feet
Concentration
ug/L
250000.1
25000.
2500.
250.0
25.00
2.500
0.2500
0.0250
Figure 2. VC distribution at transect 5.
St. Joseph, Michigan
Trichloroethene
transect: 5
mass: 0.2821 E-01 Kg/m
tS1 152
Ground surface
10 feet
approx. N
154 153 tSS
100 feet
Concentration
ug/L
250000.
25000.
2500.
250.0
25.00
2.500
0.2500
0.0250
Figure 3. TCE distribution at transect 5.
156
-------
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
2
1.523
0.2377
0.566
12.32
10.67
5.855
0.6847
No data
No data
129.9
37.05
2.904
1.835
2.987
1
Transect
1
8969
o|()816
0
5
5
5
5059
1127
5804
4826
OJ8925
N|D data
N
148
34
2
> data
8
376
471
2J5609
3
8357
4
0.4868
0.01451
0.03628
1 .890
1.397
4.620
0.1747
0.2085
12.63
213.1
95.78
4.421
0.4562
0.6353
5
0.04811
0.001047
0.007041
0.2832
0.02821
1.373
0.004901
0.001689
8.314
156.2
66.19
8.247
0.2256
0.3646
Advective Mass Flux Estimates
Results from the calibrated MODFLOW mocel of Tiedeman 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 declining hydraulic conductivity toward
the west (i.e., towards Lake Michigan and tramsects 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 decline 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 1 3 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 transec 12 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
157
-------
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 1 7 kg/y at transect 5. The range of fluxes at transect 5 is an upper bound on, and
best estimate of, the flux into Lake Michigan.
Table 2. Mass Flux (kg/y) 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
Transect
2
18.81
2.934
6.995
152.1
131.7
72.29
8.453
No data
No data
1604
457.4
35.85
22.66
36.88
1
36.03
1.551
9.609
97.11
106.0
104.1
16.95
No data
No data
2826
652.9
46.93
48.64
72.85
4
10.69
0.3185
0.7963
41.48
30.67
101.4
3.836
4.577
277.2
4678
2102
97.05
10.01
13.95
5
1.676
0.03648
0.2453
9.868
0.9829
47.86
0.1708
0.05885
289.7
5444 •
2306
287.4
7.861
12.70
Apparent Degradation Constants
The mass per unit thickness of TCE at transects 2, 4, and 5 was used to estimate apparent first-
order degradation constants. The constants are estimated by applying the first-order rate
equation
In
= XAt
(1)
to the site data, where cf is the average concentration in the transect j, c)+1 is the average
concentration in the downgradient transect j+1, At is the advective travel time for TCE to move
158
-------
between the transects, and A is the apparent cegradation constant. The mass per unit thickness
data forTCE and the cross-sectional area were used to determine the average concentrations
G| and ci+1 in the up- and downgradient transects. The porosity, bulk density, fraction organic
carbon, organic carbon partition coefficient (5), ground-water gradient, and distance 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.
Table 3. Chemical and Hydraulic Values Us« d in Estimating Apparent Degradation Rates
Tran-
sect
2
4
5
Area with
nonzero
TCE
concen-
tration
(m2)
1592
2774
1943
Mass per
unit thickness
from SITE-3D
(kg/m)
10.67
1.397
0.0282
Average TCI
concentratio
in the transe
(kg/m3)
Cj and Cj+1 in
equation 1
6.70e-3
5.04e-4
1 .44e-5
i
ct
Distance
between
transects
(m)
260
160
Gradient
estimated
from
Tiedeman
and
Gorelick
(1993)
7.3e-3
l.le-2
Retarded
seepage
velocity
for TCE°
(m/d)
0.11
0.156
Estimated
travel time
between
transects
(weeks)
At in
equation 1
340
145
°Constants used in seepage velocity calculation:
Hydraulic conductivity: 7.5 m/d
Retardation factor for TCE: 1.78 = 1 + K.xiocpb/e
Porosity, 0: 0.30
Bulk density pb: 1.86 g/cm3
K^: 126 mL/g, foc: 0.001
References
i.
2.
Engineering Science, Inc. 1 990. Reme
Ml, phase 1 technical memorandum.
dial investigation and feasibility study, St. Joseph,
.iverpool, NY.
Kitanidis, P.K., L Semprini, D.H. Kampbell, and J.T. Wilson. 1993. Natural anaerobic
bioremediation of TCE at the St. Joseph, Michigan, Superfund site. In: U.S. EPA.
Symposium on bioremediation of hazardous wastes: Research, development, and field
evaluations. EPA/600/R-93/054. Washington, DC (May), pp. 57-60.
3.
McCarty, P.L., and J.T. Wilson. 1 992.
Joseph, Michigan, NPL site. In: U.
EPA/600/R-92/126. pp. 47-50.
Natural anaerobic treatment of a TCE plume, St.
S. EPA. Bioremediation of hazardous wastes.
159
-------
4. Tiedeman, C., and S. Gorelick. 1 993. Analysis of uncertainty in optimal ground-water
contaminant capture design. Water Resour. Res. 29(7):2,139-2,1 53.
5. U.S. EPA. 1 990. Subsurface remediation guidance table 3. EPA/540/2-90/01 1 b.
160
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Poster
Session
-------
-------
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, Rober
Ada, OK
Ross N. Miller and Jerry E. Hansen
U.S. Air Force Center for Environmental Exce
Brooks AFB, TX
S. Kerr Environmental Research Laboratory,
lence, Technology Transfer Division,
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
contaminated ground water (1). The material
effort of AFCEE, the
Agency's Robert S.
Engineering-Science, Inc., to facilitate
remediation (natural attenuation) with long-term monitoring for restoration of fuel-hydrocarbon
presented herein was prepared through the joint
Bioremediation Research Team at the U.S. Environmental Protection
Kerr Environmental Research Laboratory in Ada, Oklahoma, and
of dissolved-phase fuel hydrocarbons having
The intended audience for this document
consultants, regulatory personnel, and othe
contaminated ground water at USAF facilitie:
mplementation of intrinsic remediation at
fuel-hydrocarbon contaminated sites. Specifically, this protocol is designed to evaluate the fate
regulatory maximum contaminant levels (MCLs).
s U.S. Air Force (USAF) personnel, scientists,
rs charged with remediating fuel-hydrocarbon
Intrinsic remediation is achieved when
biodegradation (aerobic and anaerobic),
contaminantdissolved in ground water. During
transformed to innocuous byproducts (e.g.,
another phase or location within the
integration of several subsurface attenuation
or nondestructive. Destructive processes
hydrolysis. Nondestructive attenuation
dispersion and infiltration), and volatilization
b ing
In some cases, intrinsic remediation reduces
below MCLs before the contaminant plume
naturally occurring attenuation mechanisms, such as
about a reduction in the total mass of a
intrinsic remediation, contaminants are ultimately
carbon dioxide and water), not just transferred to
environment. Intrinsic remediation results from the
mechanisms that are classified as either destructive
include biodegradation, abiotic oxidation, and
include sorption, dilution (caused by
mechanisms
dissolved-phase contaminant concentrations to
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.
that intrinsic remediation will result in a continual reduction in contaminant concentrations over
time such that calculated risk values are reduced.
This is especially likely if it can be demonstrated
Symposium on Intrinsic Bioremediation of Ground Water
163
-------
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 induced 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 requires the quantification of ground-water
flow and solute transport and transformation processes, including rates of natural attenuation.
Quantification of contaminant migration and attenuation rates, and successful implementation
of the intrinsic remediation option, requires completion of the following steps, each of which is
discussed 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 remediation.
3. Perform site characterization in support of intrinsic remediation.
4. Refine the conceptual model based on site characterization 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 biodegradation term (e.g., Bioplume II or Bioplume 111).
6. Conduct an exposure assessment.
7. Prepare a long-term monitoring plan, long-term monitoring wells atthe site, and
point-of compliance 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 hydraulic conductivity, hydraulic gradients, and potential
contaminant migration pathwaysto human orecological receptors. Contaminant sources include
nonaqueous-phase liquid (NAPL) hydrocarbons present as mobile NAPL (NAPL occurring at
sufficiently high saturations to drain, under the 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).
164
-------
Review Available >
Site Data
\
f
Develop Preliminary
Conceptual Model
Jr
Make Preliminary
Assessment of Potential
For Intrinsic Remediation
Based on Existing Site
Characterization Data
- Contaminant Type
and Distribution
- Hydrogaology
- Location of Receptors
'^
f
Perform Site Characterization
in Support of Intrinsic Remediation
>
f
Refine Conceptual Model and
Complete Pre-Modeling
Calculations
\
r
Document Occurrence of
Intrinsic Remediation and
Model Intrinsic Remediation
Using Numerical Models
>
f
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
^'"unacceptable Risk Tn^YES
Ev<
(
S Con
Intrins
Free Product
Recovery j
/
Pump
&
Treat
iuat
erR
)ptic
unc
cR
' h
[Bioslurpingj '
t 15-°! Assess Potential For
emfdial ^ Intrinsic RfimfiHiatinn
;i'3n"',. ' With Remediation
£«&, s^m lnstalled
f^V >
r
I \ | Bioventing | Refine Conceptual Model and
1 \ Complete Pre-Modeling
1 \ Calculations
I 1 Sparging!
I1- ' >
1 Bairiei-6 I Model Intrins
bprSlL .OptionSe,
>
r
c Remediation
irith Remedial
acted Above
3rical Models
r
Use Results of Modeling and
Site-Specific Information in
an Exposure Assessment
^<
YES ^x^UnacceDt
^xsPotential
NO
r
rhere"*-^
able Risk To""^.
Receptors'x^^
Site Point-pf-Compliance
Monitoring Wells and
Prepare, Long-Term
Monitorinq Plan
u,
Present Findings
and Lbng-Term FIGURE 2. 1
Monitoring Plan To
Regulatory Agencies . . ,. .
and RearJh Agreement Intrinsic Remediation
on Monitdring Strategy Flow Chart
Figure 1. Intrinsic remediation flow chart.
165
-------
The following analytical protocol should be used for analysis of soil and ground-water samples.
This analytical 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 samples 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, sulfide, ferrous iron, carbon dioxide,
methane, chloride, total petroleum hydrocarbons, and aromatic hydrocarbons. The extent and
distribution (vertical and horizontal) of contamination and electron acceptor and metabolic
byproduct concentrations and distributions are of paramount importance in documenting the
occurrence 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 hydrocarbon biodegradation. Similarly, nitrate and
sulfate concentrations below background in an area with fuel-hydrocarbon contamination are
indicative of anaerobic biodegradation through denitrification and sulfanogenesis. Contour maps
can be used to provide visible evidence of these relationships. Elevated concentrations of
metabolic byproducts in areas with fuel-hydrocarbon contamination are indicative of
hydrocarbon biodegradation. As iron II and methane concentrations increase during iron (III)
reduction and methanogenesis (anaerobic processes), BTEX concentrations should be seen to
decrease. 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 degradation of site contaminants is occurring at rates sufficientto be protective
of human health and the environment. Three lines of evidence can be used to support intrinsic
remediation: 1) documented loss of contaminants at the field scale, 2) the use of chemical
analytical data in mass balance calculations of microbial metabolism, and 3) laboratory
microcosm studies using aquifer samples collected from the site.
The first line of evidence involves using measured dissolved-phase concentrations of biologically
recalcitrant tracers found in fuels in conjunction with aquifer hydrogeologic parameters such as
seepage velocity and dilution to show that a reduction in the total mass of contaminants is
occurring at the site. The second line of evidence 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 byproduct 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
laboratory 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 investigation is to determine if natural
processes of degradation 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 extentand concentration of the contaminant plume in time and space
is made based on governing physical, chemical, and biological processes. This projection
should be based on historic variations in—and the current extent and concentration of—the
contaminant plume, as well as on the measured rates of contaminant attenuation.
166
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modeling the combined effects of advection,
assess the possible risk to potential downgrad
The data collected during site characterization can be used to model the fate and transport of
contaminants in the subsurface. Such mode ing 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 o 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 concentre tion of a dissolved-phase contaminant plume by
dispersion, sorption, and biodegradation; 2) to
ent receptors; and 3) to provide technical support
for the natural attenuation remedial option at postmodeling regulatory negotiations.
Microorganisms generally utilize dissolved oxygen and nitrate in areas with dissolved-phase
fuel-hydrocarbon contamination at rates that a re instantaneous relative to the average advective
transport velocity of ground water. This results in the consumption of these compounds 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 biodegradation 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-hydrocarbon
contamination at rates that are slow relative to the advective transport velocity of ground water.
This results in the consumption of these comppunds 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-limited
process that can be approximated by first-order kinetics. The Bioplume II model utilizes a
first-order rate constant to model such biocegradation. First-order decay constants can be
determined by simple calculations based on ground-water chemistry or through the use of
laboratory microcosm studies. In addition, tre 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
data input into the model and the model itse
modeling effort are only as good as the original
f. Because of the inherent uncertainty associated
with such predictions, it is the responsibility cf the proponent to provide sufficient evidence to
demonstrate thatthe mechanisms of intrinsic remediation will reduce contaminant concentrations
to acceptable levels before potential receptors
input parameters and numerous sensitivity anc
contaminant migration scenarios. When poss
used to provide information that collectively a
are reached. This requires the use of conservative
lyses so that consideration is given to all plausible
ble, both historical data and modeling should be
id consistently supports the natural reduction and
of the dissolved-phase contaminant plume. In some cases, simple calculations of
contaminant attenuation rates are all thai are required to successfully support intrinsic
remediation.
removal
Upon completion of the fate and transport rrodeling 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 negotiating this
option with regulators. The exposure assessrient allows the proponent to show that potential
exposure pathways will not be completed.
167
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The long-term monitoring plan consists of locating ground-water monitoring wells and
developing a ground-water sampling and analysis strategy. This plan is used to monitor plume
migration over time and to verify that intrinsic remediation is occurring at rates sufficient to
protect potential downgradient receptors. The long-term monitoring 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 contaminant 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 velocity 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 monitoring 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
numberand location of long-term monitoring wells depends 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 to help site both the long-term and POC monitoring wells.
Anaerobic Treatment Zone
Plume Migration
Extent of Dissolved-
Phase BTEX Plume
Aerobic Treatment
Zone
LEGEND
8 Point-of-Compliance Monitoring Well
O Long-Term Monitoring Well
Not To Scale
FIGURE 2
Hypothetical Long-Term
Monitoring Strategy
Figure 2. Hypothetical long-term monitoring strategy.
168
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References
i.
2.
3.
4.
5.
6.
7.
8.
"OCQ
Wiedemeier, T.H., D.C. Downey, J.T
Hansen. 1994. Draft technical prot<
long-term monitoring option for
contamination in ground water. U.S.
Technology Transfer Division.
Wilson, D.H. Kampbell, R.N. Miller, and J.E.
for implementing the intrinsic remediation with
natural attenuation of dissolved-phase fuel
Air Force Center for Environmental Excellence,
Klecka, G.M., J.W. Davis, D.R. Gray, and S.S. Madsen. 1 990. Natural bioremediation
of organic contaminants in ground water—Cliffs-Dow Superfund site. Ground Water
28(4):534-543.
Downey, D.C., and M.J. Gier. 1991
hydrocarbon spill site. In: Proceec
Technology Symposium, San Antonio
dt
Wiedemeier, T.H., P.R. Guest, R.L. H«
to support regulatory negotiations
Proceedings of the Petroleum
Prevention, Detection, and Restoratio i
nry, and C.B. Keith. 1 993. The use of Bioplume
a fuel spill site near Denver, Colorado. In:
Hydroccrbons and Organic Chemicals in Ground Water:
Conference. NWWA/API. pp. 445-459.
Wiedemeier, T.H., B. Blicker, and
bioremediation of fuel hydrocarbons
Federal Environmental Restoration
Exhibition, New Orleans, LA.
60.
P.R. Guest. 1994. Risk-based approach to
at a major airport. In: Proceedings of the 1994
and Waste Minimization Conference and
Hazardous Materials Control Resources Institute, pp. 51 -
Rifai, H.S., P.B. Bedient, J.T. Wils
Biodegradation modeling at aviatior
1,0029.
Borden, R.C., and P.B. Bedient. 1 986
by oxygen limited biodegradation
22(13):1973-1 982.
Wilson, J.T., J.F. McNabb, J.W. Coch
1 985. Influence of microbial adapta
water. Environ. Toxicol. Chem. 4:72
Supporting the no-action alternative at a
ings of the USAF Environmental Restoration
, TX (May 7-8). Section U. pp. 1 -1 1.
>n, K.M. Miller, and J.M. Armstrong. 1988.
fuel spill site. J. Environ. Eng. 1 14(5):1,007-
. Transport of dissolved hydrocarbons influenced
•theoretical development. Water Resour. Res.
ran, T.H. Want, M.B. Tomson, and P.B. Bedient.
ion on the fate of organic pollutants in ground
-726.
169
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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 1 993, the Wisconsin Department of Natural Resources issued an interim guidance
on naturally occurring biodegradation as a remedial action option for contaminated sites. The
focus of this guidance was primarily on soil contamination by petroleum hydrocarbons and on
the requirements for site characterization and monitoring necessary to use this approach.
Subsequent 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 concentrations to acceptable levels within a reasonable period with
respect to potential contaminant migration or impacts 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 requirements are essentially the same. Characterization involves
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
representthe 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 overtime, ensures that no
unexpected contaminant migration is occurring, and provides information regarding the nature
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 responsible parties are unlikely to select this
option due to the long time frame involved. This suggests that naturally occurring
biodegradation is more viable as an option 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
information that can be used in the evaluation and design of enhanced bioremediation systems.
170
Symposium on Intrinsic Bioremediation of Ground Water
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Assessing the Efficiency of Intrinsic Bioremediation
Francis H. Chapelle
U.S. Geological Survey, Columbia, SC
of biodegradation. If transport rates are fasf i
can migrate freely with ground-water flow anc
wildlife populations. Conversely, if transport
Abstract
The efficiency of intrinsic bioremediation to contain contaminant migration in ground-water
systems can be quantitatively assessed by conparing rates of contaminant transport with rates
elative to rates of biodegradation, contaminants
possibly reach a point of contact with human or
rates are slow relative 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 bioremediation can be assessed by evaluating the presence or
absence of contaminant transport to predetermined points of contact. Thus, this assessment
includes hydrologic (rates of ground-water flow), microbiologic (rates of biodegradation), and
sociopolitical (points of contact) considerations.
The U.S. Geological Survey, in cooperation wi
h the Naval Facilities Engineering Command, has
developed a framework for assessing the efficiency of intrinsic bioremediation that is based on
these three considerations. In this framework, hydrologic and microbiologic information is
synthesized using a solute-transport code (SUJTRA) 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 contaminants are effectively contained and are not transported to adjacent points of
contact. At the Hanahan site, biodegradation rates are similarly slow under the ambient
anaerobic conditions (Kbio ~0.01 d'1), but beqause rates of ground-water flow are relatively high
(~ 1.0 ft/d), contaminants are transported to multiple points of contact with humans. These
examples illustrate the complex interplay that develops between hydrologic, microbiologic, and
sociopolitical considerations, and show that
be assessed on a site-by-site basis.
he efficiency of intrinsic bioremediation can only
Symposium on Intrinsic Bioremediation of Ground Water
171
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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 consideration 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 provide
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) levels in ground water and plume characteristics. Background DO levels greater than 1
to 2 mg/L and inverse correlation between DO and soluble hydrocarbon concentration 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 balances, laboratory microcosm studies, and detailed ground-water modeling)
can be applied to demonstrate natural attenuation as well. The recommended approach 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 indicators should be applied to evaluate natural attenuation.
172
Symposium on Intrinsic Bioremediation of Ground Water
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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, N.
Daniel P. Shine
Sun Company, Inc., Aston, PA
Intrinsic bioremediation as discussed in the recent report of the National Academy of Sciences
Committee on Bioremediation (1) can reduce the concentrations of some common contaminants
to levels generally considered 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-
preserving site (2), many other sites have been
observed to have undergone natural attenuation
at sufficient rates to limit the size of contaminart plumes; in several instances, contaminant levels
decreased to below cleanup levels.
The use of intrinsic bioremediation, while generally attractive from a cost perspective, may
actually be desirable environmentally because secondary effects of active remediation are
avoided. Intrinsic bioremediation, however, las several costs associated with its use. These
include some level of documentation that bio ogical degradation of constituents of concern is
taking place and costs associated with monitoring and management 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
management costs have been approximately $1 1,900 per year, consisting of site visits and
sampling ($1,000), chemical analysis ($7,200), reporting ($2,400), and consulting ($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 electron 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
management burdens.
Addition of appropriate electron acceptors would accelerate reduction in constituents such as
monoaromatic hydrocarbons. In some cases where intrinsic bioremediation is technically
feasible, it may not be the most cost-effectivd approach. To evaluate the concept of applying
limited engineering solutions at sites where intrinsic 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 shortened by 1 yr or more,
cost that would have been incurred by only nonitoring and managing the site.
the cost of treatment will have been less than the
Symposium on Intrinsic Bioremediation of Ground Water
173
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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 peroxide 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 along the downgradientedge 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 perpendicular to the ground-water gradient near the
downgradientedge 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 Stanford University
in tests conducted at the Borden Landfill in Canada, this approach can successfully prevent
migration of monoaromatic hydrocarbons.
Intrinsic bioremediation should be.viewed as one approach in a continuum of methods of
utilizing biodegradation processes to remediate soil and ground water. It should be used alone,
in combination with other approaches, or as a polishing step based on evaluation of the site
conditions, regulatory issues, technical feasibility, implementability, and cost.
This poster session will present cost and analytical data from three New Jersey service stations
during the monitoring-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.
Additionally, the costs and relative advantages of the use of slow-release oxygen compounds
and air sparging will be presented.
References
i.
2.
3.
4.
5.
National Research Council. 1993. In situ bioremediation: When does it work?
Washington, DC: National Academy Press.
Wilson, J.T., J.F. McNabb, J. Cochran, T.H. Wang, M.B. Tomson, and P.B. Bedient.
1 985. Influence of microbial adaptation on the fate of organic pollutants in ground
water. Environ. Toxicol. Chem. 4:721-726.
Norris, R.D., J.C. Dey, and D.P. Shine. 1993. The advantages of concerted
bioremediation of lightly contaminated sites compared to intrinsic bioremediation.
Presented at the American Chemical Society I&E Special Symposium, Atlanta, GA.
Kao, C.-M., and R.C. Borden. 1 994. Enhanced aerobic bioremediation of a gasoline
contaminated aquifer by oxygen-releasing barriers. In: Hydrocarbon bioremediation.
Boca Raton, FL: Lewis Publishers.
Bianchi-Mosquera, G.C., R.M. Allen-King, and D.D. Mackay.
degradation of dissolved benzene and toluene using a solid
compound. Ground Water Monitor. Rev. pp. 120-128.
1994. Enhanced
oxygen-releasing
174
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Natural Attenuation of Jet Fuel in Ground Wafer
Greg Doyle, Dwayne Graves, and Kandi Brovn
International Technologies Corporation, San
Bernardino, CA
Natural attenuation is a minimum action remedial strategy that permits the biodegradation of
organic contaminants under natural, in si u conditions. Mechanisms that act to affect
contaminant biodegradation include aerobic b
odegradation atthe plume boundary and various
anaerobic processes within the plume. Nitra e, iron, manganese, and sulfate reduction and
linant biodegradation. Naturally occurring levels
mariganese, and carbonates support anaerobic
biodegradation.
methanogenesis are thought to support contami
of nitrate, sulfate, oxidized iron and ma ic
Natural attenuation was proposed as a feasible remedial alternative for the dissolved jet fuel
plume at George Air Force Base in Victorvillel California. The plume, covering 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 approximately 20 tt/yr, and the edge of the plume
was about 2,000 ft from the property line. The ground water in the affected aquifer was not
being used. Because of subsurface conditions,
natural attenuation represented the most effic
dissolved jet fuel at this site.
A separate-phase layer of jet fuel floated on
aggressive skimming operation was employed
from the subsurface. Bioventing will be used to
in areas where jet fuel was detected and
slow migration, poor water yield, and lack of use,
ent and cost-effective approach for remediating
he water table near the center of the plume. An
to remove all recoverable separate-phase jet fuel
further remediate contaminated vadose zone soil
d.
remove^
Natural attenuation for aerobic and aero Die/anaerobic conditions 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,
various 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 remediate the plume before it migrated off site.
Based on this prediction and the cost savings
performance period was established prior to tr
efforts by the Air Force, the U.S. Environmental
Research Laboratory, and IT will provide data
realized by applying natural attenuation, a 5-yr
ie issuance of a finalized record of decision. Joint
Protection Agency's Robert S. Kerr Environmental
verifying the accuracy of the BIOPLUME II model
predictions and demonstrating the level of natural biological activity occurring in the ground
water. These efforts are expected to lead to regulatory acceptance of natural attenuation for the
full-scale remediation of the jet fuel plume on site.
Symposium on Intrinsic Bioremediation of Ground Water
175
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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 Research 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 description 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 1 8,000
gal UST, removed in 1 989, were the probable sources of observed ground-water contamination.
Total dissolved petroleum hydrocarbon contaminant mass at the beginning 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 concentrations at a number of
locations in unconta mi noted areas surrounding the plume were over 2 mg/L.
Initial site assessment activities utilized cone penetrometry for rapid collection of subsurface soil
stratigraphicdata and forthe placement of more than 60small-diameterground~watersarnpling
points at the two shallow field sites. Ambient temperature headspace analyses were conducted
using a polyethylene bag method developed by the Utah Water Research Laboratory in addition
to a commercial Lag-in-a-Bag apparatus to provide rapid field-determined measurements of
ground-watertotal 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 techniques (soil gas survey, collection of limited soil core samples, and
placement of more limited numbers of large-diameter ground-water monitoring wejls), was
greatly in error. The model could be improved significantly with these cost-effective field
techniques, which are now widely available to the consulting community.
Seven field sampling events, beginning in April 1 992, were conducted overthe 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 transport of contaminants, along with the
predominant microbial reactions taking place within the contaminant plume. Ground-water
quality data collected included field measurements 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 benzene,
toluene, ethylbenzene,p-xylene, naphthalene, and methylnaphthalene), and contaminant boiling
point split concentrations.
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Symposium on Intrinsic Bioremediation of Ground Water
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Ground-water data were used to generate tote11 and specific compound dissolved contaminant
and dissolved electron acceptor mass data for each sampling period. The location of the center
of the mass of contaminants and electron acceptors were evaluated, and changes in these
parameters overtime, with respect to bulkgro
magnitude of natural degradation processes
jnd-waterflow, were used to assess the rate and
aking place at each site.
constituents displayed zero-order decay rates.
mg (>99.9 percent removal) by the end of
Over the 2-year study period, the mass of TPH showed exponential decay, while all specific
Dissolved TPH mass declined to less than 1,000
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 -0.01 3/d
(p = 0.03, r2 = 0.933), -11.2 mg benzene/d jp = 0.01, r2 = 0.976), and -14.2 mg toluene/d
(p = 0.01, r2 = 0.998), respectively. 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 attenuated significantly, with measured center of mass velocities 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 electrDn acceptors occurred across the plume.
Intrinsic bioremediation of the dissolved plum
removed the residual hydrocarbon mass exis
at the Hill AFB site successfully attenuated and
ing at the site at the beginning of the study. In
January 1 994, only 1 of 34 monitoring well/piezometer ground-water samples contained a
benzene concentration (20.7 /Jg/L) above regulatory concern, and the site is expected to be
eligible for closure at the next routine, semiannual sampling event.
This postersession will detail the physical/chemical characteristics of the field sites and the rapid
site assessment techniques and typical results collected, as well as highlight the data
collection/reduction/interpretation methodology developed in this study. Finally, more complete
results demonstrating natural degradation o
presented, along with a summary of a natu
hydrocarbon contaminants at this site will be
ral attenuation decision support system to aid
investigators in assessing the viability of intrnsic bioremediation for the selection of a "no
action'/natural attenuation monitoring alternctive at their sites.
177
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Case Studies of Field Sites To Demonstrate Natural Attenuation of BTEX
Compounds in Ground Water
Chen Y. Chiang and Paul M. McAllister
Shell Development Company, Houston, TX
The most definitive indicptors 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
attenuation 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 oxidation under varying levels of DO. The
results are remarkably consistent with field data on the presence of high or low levels of BTEEX
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 overtime. 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 downgradient. 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 relatively constant out to
the leading edge, where a sharp front would be observed. It is emphasized that natural
attenuation also includes other mechanisms than biodegradation: dispersion, sorption,
volatilization, and chemical transformation.
178
Symposium on Intrinsic Bioremediation of Ground Water
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Demonstrating Intrinsic Bioremediation of BTEX at a Natural Gas Plant
Keith Piontek and Tom Sale
CH2M Hill, St. Louis, MO, and Denver, CO
Steve de Albuquerque and John Cruze
Phillips Petroleum Company, Bellaire, TX, an
J Bartlesville, OK
Intrinsic bioremediation is being characterizqd
plant operations resulted in the release of
deposits beneath the site. Ground water
ethylbenzene, and xylene (BTEX) constituen
biodegradation that occurs under natural con
realistic assessment of potential risks posed
management decision-making.
at the site of a former natural gas plant. Gas
nonaqueous-phase liquid (NAPL) to the eolian
beneath the site contains benzene, toluene,
•s of the NAPL. The rate and extent of BTEX
Jitions is being characterized to provide for more
by the site, and to support long-term site
The characterization of intrinsic bioremediaticn is being conducted in general accordance with
the protocol proposed by Dr. John Wilson of the U.S. Environmental Protection Agency's Robert
S. Kerr Environmental Research Laboratory. Key parameters being assessed include seasonal
variations in ground-water flow direction and velocity, changes in hydrocarbon concentrations,
and changes in the concentrations of electror
of the stoichiometry of hydrocarbon biodegrcdation under various redox conditions, are used
to confirm that BTEX biodegradation is occurring at the site and to estimate the BTEX
biodegradation rate.
The characterization of intrinsic bioremediation
are collected and analyzed on a quarterly bas
performed, and two additional sampling
following:
acceptors. These data, together with knowledge
at this site is underway. Ground-water samples
s. Two ground-water sampling events have been
are planned. To date, findings include the
events
The extent of the dissolved hyd.rocarbon plume is smaller than would be
expected if the hydrocarbons were not being biodegraded. BTEX concentrations
decrease by five orders of magnitude within a lateral, downgradient distance of
300 ft from the NAPL zone.
While some oxygen flux into th 3 plume occurs, the majority of contaminant mass
removal occurs underanoxic conditions. Naturally high concentrations ofsulfate
in the ground water provide ar
biodegradation undersulfate-
essentially infinite supply of electron acceptor for
"educing conditions. Site data suggest that sulfate
is the most significant electron acceptor in terms of hydrocarbon mass removal,
with over 90 percent of the h/drocarbon mass removal attributable to sulfate
reduction.
Additional hydrocarbon mass removal is attributable to reduction of nitrate, iron,
carbon dioxide, and oxygen.
Symposium on Intrinsic Bioremediation of Ground Water
179
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• Elevated concentrations of bicarbonate downgradient of the plume, associated
with hydrocarbon mineralization, confirm the role of intrinsic bioremediation in
limiting plume migration and provide evidence that the plume has reached its
steady-state extent.
The poster paper will present information on the site setting, monitoring and data evaluation
methodology, evidence of intrinsic bioremediation mechanisms and rates, and impact of
observed biodegradation on plume migration.
180
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Demonstrating the Feasibility of Intrinsic
Manufactured Gas Plant
Bioremediation at a Former
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)
investigation to assess natural, in situ biodeg
decisions on intrinsic bioremediation and
in Baltimore, Maryland, is the subject of an
adation for the purposes of basing remediation
sngineered, 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 (DNAP.L) that contains
monocyclic aromatic hydrocarbons (MAHs),
aromatic hydrocarbons (PAHs), such as naph
as a long-term source of aromatic hydrocarbo ns to ground water, and concentrations in ground
water over much of the 70-acre site are clos
Laboratory and field data were collected to
such as benzene and toluene, and polycyclic
halene and benzo(a)pyrene. The tar DNAPL acts
5 to theoretical effective solubilities.
evaluate biodegradation. The first phase of the
laboratory investigation, consisting of tests c n 1 6 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 established between hydrogeology, contaminant
distribution, and microbiological characteristics. Aerobic and anaerobic enumerations were
performed, followed by fatty acid analyses to identify the microbes. Viable aerobic bacteria were
detected in all subsurface samples. Bacteria wjere grown in the samples plated under anaerobic
conditions, 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 Unive -sity of Wyoming (UW). JHU used 49 soil samples
from five boreholes with site ground-water and individual radiolabeled target substrates
(benzene, naphthalene, phenanthrene, and acetic 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 perforjmed by the acridine orange direct-count method.
Similar microcosm studies are being performed by UW under sulfidogenic, iron-reducing, and
methanogenic conditions using phenol, benzsne, toluene, naphthalene, and phenanthrene as
the targeted radiolabeled compounds. Benzere, 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
naphthalene mineralization (7 percent to 13
over a 4-week incubation period. Anaerobic
oercent) was observed in two JHU samples in the
presence of NO3. Half-lives calculated from f
tens to hundreds of days, with a lower half-li
rst-order degradation rates typically ranged from
e in the initial stages of incubation followed by a
slower rate presumably indicative of oxygen- or nutrient-limiting conditions. Under sulfate-
Symposium on Intrinsic Bioremediation of Ground Water
181
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reducihg conditions, phenol mineralized 1 3 percent to 1 8 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 mineralization 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 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 atmospheric in 1 1 of the 16 wells
measured.
Apparent degradation rates were calculated using the first-order model by regressing the natural
log of constituent concentrations (adjusted for dilution) by estimated in situ travel time. Because
no nonreactive, conservative tracers are unique to the tar sources, dilution 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 anthropogenic carbon, resulting in
overestimates of dilution and underestimates of degradation rates. Half-lives for benzene,
toluene, ethylbenzene, xylenes, and naphthalene were calculated to be 729, 660, 877, 855,
and 2,1 66 d, respectively. Half-lives forthe three-ring and greater PAHs were not calculated due
to the poor regression correlation coefficients. Degradation estimates showed thattoluene is the
most preferred aromatic substrate studied and that naphthalene appears to degrade the slowest
as predictable from the literature. Surprisingly, 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 aerobic
and anaerobic conditions. Field evidence showing various redox conditions and biogenic
products of organic 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 estimates of in situ constituent decay are based on numerous
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
availableto 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 sorption. The attenuation assessment
demonstrated that contaminant loss was observed overtime along the aqueous plume travel
path (i.e., travel time, rather than atone point over time) and degradation, probably biotic, was
measured for target contaminants.
The combined laboratory and field evidence point to natural in situ biodegradation as an active
process in the site's low oxygen, subsurface environment. Intrinsic bioremediation, albeit slow
due to mass transfer limitations from the tar to the aqueous phase/may be technically viable
for controlling aqueous aromatic hydrocarbon contamination emanating from MGP-tar sources.
Laboratory studies are continuing to explore compound-specific biodegradation under various
conditions, and plans are being formulated now for an in situ biodegradation pilot study.
182
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Natural and Enhanced Bioremediation o
California: Laboratory and Field Investigations
Aromatic. Hydrocarbons at Seal Beach,
Harold A. Ball, Gary D. Hopkins, Eva Orwin,
Western Region Hazardous Substance Resear
and Martin Reinhard
ch Center, Stanford, CA
Introduction
The objective of this study was to develop our
important for intrinsic anaerobic biodegrada
ground-water aquifers, and to determine me
understanding of environmental factors that are
ion of aromatic hydrocarbons in contaminated
hods to enhance this process. The focus of the
investigation was a site at the Seal Beach Naval Weapons Station in southern California, where
a significant gasoline spill resulted in contamiration of the ground-water aquifer (1). Inthe field,
nitrate was present at about 5 mg/L in background wells and approached detection levels (<0.1
mg/L) in the contaminated wells. There was a
ligh natural background sulfate concentration of
about 85 mg/L in the ground water, and methane was detected in the contaminated well
headspace. The distribution of aromatics present in the contaminated ground water differs from
that expected from dissolution of pure gasoline (2). This suggests that natural biotransformation
of several organic species is occurring at the site. The project was divided into laboratory and
field components, 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 anaerobic conditions and to understand the
effect of alternate environmental conditions on the transformation processes. Field experiments
were carried out on site at Seal Beach.
biotransformation under natural (presumed si
have been carried out.
At the field site, experimental monitoring of
Ifate-reducing) and nitrate-enhanced conditions
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-J 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 pare meter) were biotransformed in the unamended
ground-water samples under presumed sulfa e-reducing conditions. Addition of nitrate to the
ground water increased rates of toluene biotransformation coupled to nitrate reduction,
stimulated biotransformation of ethylbenzene, and inhibited the complete loss of m+p-xylene
that was observed when nitrate was not added and sulfate-reducing conditions prevailed.
Addition of the nutrients ammonia and phosphate had no effect on either the rate of aromatics
transformation or the distribution of aromatics transformed. When Seal Beach sediment was
placed into nitrate-reducing media, ethylbenzene was transformed first, followed by toluene.
When the sediment was placed into sulfatejreducing media, lag times were increased, but
toluene and m-xylene were ultimately transformed just as in the microcosms with ground water
Symposium on Intrinsic Bioremediation of Ground Water
183
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no
alone. Although methane had been detected in the field, there appeared to be .,~
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 constructed 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 capability to control
and monitor both hydrodynamic flow and supplements to the composition of the native ground-
water influent. Initial operation of the three anoxic/anaerobic reactors focused on evaluating
anaerobic bioremediation strategies foraromatic hydrocarbons under natural (presumed sulfate-
reducing) and enhanced denitrifying conditions.
Bioreactor results were consistent with the laboratory microcosm experiments. Toluene arid m+p-
xylene were degraded in both the unamended and nitrate-amended bioreactors. Degradation
of ethylbenzene was stimulated 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 aromatic biotransformation in a test zone
within the contaminated aquifer at the Seal Beach site. Initial work focused on evaluation of
intrinsic bioremediation as evidenced by the distribution of aromatic species in background wells.
Subsequent experiments to determine our ability to enhance this biotransformation have been
conducted using a slug test experimental design in which a single well was used forthe 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 deoxygenated
injection stream, and the stream was metered into the injection well. Sodium bromide was added
as a conservative tracer. Under this scenario, the'different 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 bioremediation studies, benzene, toluene, ethylbenzene, m-
xylene, and o-xylene were added with the injection pulse at a concentration of approximately
200 fj.g/1 each.
The initial bromide tracer data showed stable tracer concentrations and indicated no substantial
encroachment 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 1 03-d
period. During the tracer test, the equilibrium desorption concentrations for the aromatic
hydrocarbons when the electron acceptors nitrate and sulfate were absent from the ground water
were evaluated. Benzene, ethylbenzene, and o-xylene concentrations remained relatively stable
and thus appeared to be at an equilibrium. The toluene and m+p-xylene concentrations had
a downward trend relative to benzene once the native ground water encroached after
approximately 0.4 pore volumes, suggesting thatthe nitrate and sulfate concentrations available
184
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in the native ground water supported some i
experiment for toluene and m+p-xylene remc
In a nitrate augmentation experiment, nitrate c
irtrinsic biological activity in the latter part of the
val.
nd aromatics were added to the injection pulse,
resulting in complete consumption of toluene cind 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 anaerobic denitrifying population. This
would confirm that there was already an active
activity was enhanced by the addition of nitrate. With the exception of o-xylene transformation,
these results were comparable with those from
nitrate-reducing population in the aquifer whose
the nitrate-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. Environmental Protection Agency's Office of
Research and Development, under agreement R-81 5738-01 through the Western Region
Hazardous Substance Research Center. The content of this study does not necessarily represent
the views of the Agency. Additional funding
Technology Company, Richmond, California.
was obtained from the Chevron Research and
References
1. Schroeder, R.A. 1991. Delineation o
shallow deposits at the U.S. Naval
Invest. Rep. 89-4203.
a hydrocarbon (weathered gasoline) plume in
Weapons Station, Seal Beach, California. Water Res.
2. Cline, P.V.,J.J. Delfino, and P.S.C. Rao. 1 991. Partitioning of aromatic constituents into
water from gasoline and
25(5):914-920.
other complex solvent mixtures. Environ. Sci. Technol.
3. Ball, H.A., and M. Reinhard. 1 994. laboratory study of monoaromatic hydrocarbon
degradation under anaerobic conditions at Seal Beach, California. In preparation.
4. Ball, H.A., and M. Reinhard, M. 1994
degradation under anaerobic conditi
, Pilot-scale study of monoaromatic hydrocarbon
ns at Seal Beach, California. In preparation.
5. Huxley, M.P., C. Lebron, M. Reinhard, H. Ball, H.F; Ridgway, and D. Phipps. 1 992.
Anaerobic and aerobic degradation of aromatic hydrocarbons using in situ bioreactors
at an unleaded gasoline spill site. Presented at the 18th Environmental Symposium of
the American Defense Preparedness /Association, Alexandria, VA.
185
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6. Reinhard, M., LE. Wills, H.A. Ball, T. Harmon, D.W. Phipps, H.F. Ridgway, and M.P.
Eisman. 1991. A field experiment for the anaerobic biotransformation of aromatic
hydrocarbon compounds at Seal Beach, California. In: Hinchee, R.E., and R.F.
Olfenbuttel, eds. In situ bioreclamation: Applications and investigations for hydrocarbon
and contaminated site remediation. Boston, MA: Butterworth-Heinemann.
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The Complete Dechiorination of Trichloroefhene 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 Envircnmental Programs, Albany, NY
Introduction
In anaerobic environments, chlorinated ethenes can act as electron acceptors in a process called
reductive dehalogenation (specifically, reduct
however, has been shown to vary in anaerobi
:ve dechlorination). The extent of dechlorination,
: dechlorination studies depending upon the flow
and availability of electrons within the anaerobic microbial community. For example,
dechlorinated intermediates such as dichloro<
ithene (DCE) and vinyl chloride (VC) were found
to accumulate during the dechlorination of tetrachloroethene (PCE) and trichloroethene (TCE)
(1 -3). Freedman and Gossett (4), however, were the first to observe the complete dechlorination
of PCE to ethene in a methanogenic enrichme it culture, and DeBruin et al. (5) showed that PCE
could be reduced to ethane. The first field observation of the complete dechlorination of PCE
to ethene was documented by Major et al. (5). Their laboratory and field study showed that
anaerobic microorganisms in a low-permeabil
tyaquiferwere capable of naturally dechlorinating
PCE in the presence of methanol. This paper documents that microorganisms in a bedrock
aquifer unit are also capable of completely cechlorinating TCE to ethene.
Study Site Conditions
The study site is located in the Finger Lakes region of central New York. The property was used
for a variety of electrical components, including
high-voltage semiconductors. In the early to mid-1 960s waste solvents were disposed of in an
unlined evaporation pit. TCE, which was often mixed with acetone or methanol, was among the
solvents disposed of in the unlined evaporation pit. As a result, these chemicals are now found
Lll X
in the overburden and bedrock units beneath the study site.
Results
Our study involved collecting representative
ground-water samples from 21 existing ground-
water monitoring wells, mostly in the shallow bedrock unit, for geochemical and microbiological
analyses. 1,2-DCE and VC were detected in around-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
Symposium on Intrinsic Bioremediation of Ground Water
187
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unit is being controlled by biodegradation. First, the distribution 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 distributions, 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
acetone 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 onions, methane and
methane isotopes, and volatile fatty acids (e.g., acetate) were used as indirect measures of the
activity of functional groups of 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
thatthe 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
phospholipid fatty acids (PLFAs) and respiratory quinones from microorganisms thatwere trapped
onto 0.22-^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 1.6x 102 cells/mLto4.2 x TO4 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 methanol. The total biomass was orders of magnitude higher near VOC
source areas, as well as in areas with measurable concentrations of acetone and methanol, than
at downgradient or background (upgradient) locations. The microbial biomass distribution
suggested that a biologically active zone (BAZ) has developed in response to the presence of
VOCs, acetone, and methanol. The microbial populations in the samples generally
demonstrated 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 support ideal
growth. Cluster analysis of the PLFA data showed that three population groups exist at the site.
The population groups appearto 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 anaerobically dechlorinate TCE
to ethene when suitable substrates are present to support their growth. At this study site, an
active and diverse anaerobic microbial community, 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. :
188
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References
i.
2.
3.
Bouwer, E.J., and P.L McCarty. 1 983.
aliphatic organic compounds
45(4):1,286-1,294.
Transformation of 1 - and 2-carbon halogenated
under methanogenic conditions. Appl. Environ. Microbiol.
Parsons, F., P.R. Wood, and J. DeMarco
and trichloroethylene in microcosms cm-
Wilson, B.H., G.B. Smith, and J.J.
alkylbenzenes and halogenated alif
material: A microcosm study. Environ
4. Freedman, D.L, and J.M. Gossett.
. 1984. Transformations of tetrachloroethylene
id ground water. J. Am. Water. Assoc. 76:56-59.
Rees. 1986. Biotransformations of selected
hatic hydrocarbons in methanogenic aquifer
Sci. Technol. 20(10):997-1,002.
1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions.
Appl. Environ. Microbiol. 55(9):2,14-
5. De Bruin, W.P., M.J.J. Kotterman, M,
-2,151.
A. Posthumus, G. Schraa, and A.J.B. Zehnder.
1 992. Complete biological 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. 1 991. Field and laboratory evidence of in
situ biotransformation of tetrachloroet
facility in North Toronto. In: Hind
bioremediation, pp. 147-171. Bostor
nene to ethene and ethane at a chemical transfer
ee, R.E., and R.F. Olfenbuttel, eds. On-site
, MA: Butterworth-Heinemann.
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