United States
Environmental Protection
Agency
Office of Research and
Development
Washington, DC 20460
c/EPA
Of
Hazardous Wastes
EPA/600/R-92/126
August 1992
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EPA/600/R-92/126
August 1992
Bioremediation of Hazardous Wastes
by
Biosystems Technology Development Program
Office of Research and Development
U.S. Environmental Protection Agency
Ada, OK; Athens, GA; Cincinnati, OH;
Gulf Breeze, FL; and Research Triangle Park, NC
Printed on Recycled Paper
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Notice
The information in this document has been funded wholly or in part by the U.S. Environmental Protection
Agency. It has been subjected to the Agency's peer and administrative review and approved for publication as
an EPA document. Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
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Table of Contents
Page
Executive Summary
Introduction
.,
Section One: Site Characterization [[[ 5
Site Characterization at Pipeline Spill at Park City, Kansas:
Estimating Hydraulic and Geochemical Constraints on Bioremediation .................... 7
Field Demonstration of Innovative Bioremediation Strategies for
PCP and Creosote
Case Study on Site Characterization at a TCE Plume: St. Joseph,
Michigan, NPL Site [[[ 15
Approaches to the Characterization of Trichloroethylene Degradation by
Pseudomonas cepacia G4 [[[
Section Two: Bioremediation Field Initiative ............................................ 21
Progress in the Field Applications of Bioremediation ............................................... 23
Evaluation of Full-Scale In Situ and Ex Situ Bioremediation
of Creosote Wastes [[[ 25
Optimizing Bioventing in Shallow Vadose Zones and Cold Climates:
Eielson AFB Bioremediation of a JP-4 Spill [[[ 31
Optimizing Bioventing in Deep Vadose Zones and Moderate Climates:
Hill AFB Bioremediation of a JP-4 Spill [[[ 37
Use of Nitrate to Bioremediate a Pipeline Spill at Park City, Kansas:
Projecting from a Treatability Study to Full-Scale Remediation ............................... 41
Design and Treatability Study of In Situ Bioremediation of Chlorinated
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Table of Contents (continued)
Page
Characterization of Microorganisms, Microbial Consortia, and Microbial
Processes for the Reductive Dechlorination of Hazardous Wastes 53
Effects of Metals on the Reductive Dechlorination of Chlorophenols 55
Section Three: Performance Evaluation 57
Development of Comparative Genotoxicity Risk Methods for Evaluating
Alternative Hazardous Waste Control Technologies 59
Performance of Bioventing at Traverse City, Michigan 61
Bioventing of a Gasoline Spill at Traverse City, Michigan:
Practical Engineering Considerations 65
Clearance and Pulmonary Inflammatory Response after Intranasal
Exposure of C3h/HeJ Mice to Biotechnology Agents 67
Section Four: Process Research 69
Evaluation of Enhanced In Situ Aerobic Biodegradation of
cis- and trans-1-Trichloroethylene and cis- and trans-1,2-Dichloroethylene
by Phenol-Utilizing Bacteria 71
Fundamental Studies on the Treatment of VOCs in a Biofilter 75
Fungal Treatment of Pentachlorophenol 77
Anaerobic Degradation of Highly Chlorinated Dibenzo-p-Dioxins and
Dibenzofurans '"
Bacterial Degradation of Trichloroethylene in a Gas-Phase Bioreactor 83
Anaerobic Biodegradation of 5-Chlorovanillate as a Model Substrate
for the Bioremediation of Paper-Milling Waste 87
Methanogenic Degradation of Heterocyclic Aromatic Compounds by
Aquifer-Derived Microcosms 89
Chemical Interactions and pH Profiles in Microbial Biofilms 93
Decontamination of PCB-Contaminated Sediments through the Use of
Bioremediation Technologies 95
IV
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Table of Contents (continued)
Page
Bioremediation of Soils and Sediments Contaminated with
Aromatic Amines 97
Development of Small-Scale Evaluation Techniques for
Fungal Treatment of Soils 99
Development of Aerobic Biofilter Design Criteria for Treating VOCs 101
Sequential Anaerobic-Aerobic Treatment of Contaminated
Soils and Sediments 103
Influence of Low Levels of Nonionic Surfactants on the
Anaerobic Dechlorination of Hexachlorobenzene 105
Anaerobic Transformation and Degradation of Chlorobenzoates
and Chlorophenols under Four Reducing Conditions 107
Characterization of Biofilter Microbial Populations 109
Section Five: Modeling 111
Prediction of GAC Adsorptive Capacity with and without
Molecular Oxygen 113
Development of Tools for Monitoring Biofilm Processes 117
Evaluation of Anaerobic Respirometry to Quantify Intrinsic Anaerobic
Biodegradation Kinetics of Recalcitrant Organic Compounds 119
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Executive Summary
In 1987, the U.S. Environmental Protection Agency's (EPA) Office of Research and Development (ORD) initiated
the Biosystems Technology Development Program to anticipate and address research needs in managing our nation's
hazardous waste. The Agency believes that bioremediation offers one attractive alternative to conventional methods of
cleaning up hazardous waste and has developed a strategic plan for its acceptance and use by the technical and regulatory
communities. The Agency's 5-year strategic plan is centered around site-directed bioremediation research to expedite
the development and use of relevant technology.
In May 1992, ORD hosted the fifth annual Symposium on Bioremediation of Hazardous Wastes' U S EPA's
Biosystems Technology Development Program, in Chicago, Illinois. This was the first time this symposium had been
held in a regional office, and it drew the largest audience in 5 years. At the conference, 28 papers and 9 poster exhibits
were presented on recent program achievements and research projects aimed at bringing bioremediation into more
widespread use.
These Proceedings comprise an Executive Summary, an Introduction, and papers and poster presentations in five
key research and program areas, which, taken as a whole, represent a comprehensive approach to bioremediation of
hazardous waste sites. The five research and program areas are:
1. Site Characterization. EPA's Biosystems Program recognizes the basic need for more complete site
characterization techniques as a cornerstone to the application of in situ technology. Site characterization
includes detailed descriptions of the surface and subsurface, including hydrology, geology, and the physics
and chemistry of a site. The four topic papers presented covered research on petroleum-spill cleanup the
efficacy of Pseudonwnas to remediate chemical contamination, and on constraints to the proposed use of
methane-oxidizing bacteria for a TCE plume.
2. Bioremediation Field Initiative. This initiative was instituted in 1990 in response to research needs to
expand the nation's field experience in bioremediation techniques and to collect and disseminate
performance data from field application experiences. The Agency assists the regions and states in
conducting field tests and in carrying out independent evaluations of site cleanups using bioremediation
Through this initiative, tests are under way at Superfund sites, RCRA corrective action facilities and
Underground Storage Tank sites. Eight papers and two poster presentations were devoted to this key
program area and covered field evaluations currently under way at sites utilizing bioventing, biochemical
techniques, and bioremediation under a variety of aerobic and anaerobic conditions.
3. Performance Evaluation. Performance evaluation of bioremediation technologies involves determining
the extent and rate of cleanup by a particular bioremediation method as well as the environmental fate and
effects of the parent compounds and their by-products. Because remediation efforts at a contaminated site
can produce additional compounds during the remediation process, a key component of performance
evaluation is evaluating potential health effects. The particular purview of this area is to develop
bioremediation approaches that protect public health. Four presentations discussed risks related to
bioremediation and potential genotoxicity.
4. Process Research (including laboratory, pilot-scale, and field research). Process research focuses
primarily on identifying microorganisms that could degrade contaminants and developing methods for
their effective delivery. The work involves isolating and identifying microorganisms that carry out
biodegradation processes and developing new biosystems for treatment of environmental pollutants in
surface waters, sediments, soils, and subsurface materials. Eleven papers and five poster presentations
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addressed this critical area. Although the majority of presentations dealt with laboratory-scale work, both
pilot-scale and field research also were discussed. Pilot-scale research will become increasingly important
in the near future to provide critical information on process operation and control and residuals/emissions
management. Field research, in turn, is essential for evaluating the performance of bioremediation
processes and for accelerated site-directed research of process concepts that lend themselves to accelerated
testing.
5. Modeling. Mathematical modeling assists researchers in extrapolating field results to predict and guide
the application of technologies at other sites. Models provide assessments of the opportunities for
successful site bioremediation, provide cost-effective designs for field bioremediation, design appropriate
sampling strategies to support performance evaluations, and provide effective guidance with which to
apply research results to other sites. Two presentations concerned the mechanisms by which granular-
activated carbon degrades hazardous waste. A third presentation dealt with biodegradation kinetics.
The Biosystems Technology Development Program draws on ORD scientists who possess unique skills and
expertise in biodegradation, toxicology, engineering, modeling, biological and analytical chemistry, and molecular
biology. Participating laboratories and organizations are:
Environmental Research Laboratory-Ada, Oklahoma
Environmental Research Laboratory-Athens, Georgia
Center for Environmental Research Information-Cincinnati, Ohio
Risk Reduction Engineering Laboratory-Cincinnati, Ohio
Environmental Research Laboratory-Gulf Breeze, Florida
Health Effects Research Laboratory-Research Triangle Park, North Carolina
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Introduction
The U.S. Environmental Protection Agency (EPA) is responsible for protecting public health and the environment
from the adverse effects of pollutants. EPA's authority to develop regulations and to conduct environmental health
research is derived from major federal laws passed over the last 20 years that mandate broad programs to protect public
health and the environment. The Clean Air Act; the Safe Drinking Water Act; the Clean Water Act; the Toxic Substances
Control Act; the Federal Insecticide, Fungicide, andRodenticide Act; the Resource Conservation and Recovery Acf and
the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA, known as Superfund) all
require that EPA develop regulatory programs to protect public health and the environment.
For the control and cleanup of hazardous wastes, the Superfund law gives EPA broad authority to respond directly
to releases of hazardous materials that endanger public health or the environment. Also, the Superfund Amendments and
Reauthonzation Act of 1980 (SARA) expands EPA's authority in research and development, training health assess-
ments, community nght-to-know, and public participation. EPA's Office of Research and Development (ORD) conducts
basic and applied research in health and ecological effects, hazardous wastes, and remediation development and
demonstration of control technologies. Technologies are designed to provide efficient, cost-effective alternatives for
cleaning up the complex mixtures of pollutants found at Superfund sites or at other locations, such as oil spills As the
technologies advance, ORD transfers information on their use to groups that apply technologies at specific sites.
Some of the most promising new technologies for solving hazardous waste problems involve the use of
bioremediation, an engineered process that relies on microorganisms, such as bacteria or fungi, to transform hazardous
chemicals into less toxic or nontoxic compounds. These microorganisms have a wide range of abilities to metabolize
different chemicals; scientists can tailor the technology to the pollutants at specific sites and in specific media (e g
contaminated aquifers, waste lagoons, contaminated soils) by using organisms in the treatment system that break down
a particular pollutant under specific conditions.
Bioremediation is an attractive option because it is a natural process, and the residues from the biological processes
(such as carbon dioxide and water) are usually geochemically cycled in the environment as harmless products These
processes also are carefully monitored to reduce the possibility of a product or a process being more toxic than theoriginal
pollutant Another advantage of biological treatments-particularly in situ treatment of soils, sludges, and ground water-is
that they can be less expensive and less disruptive than options frequently used to remediate hazardous wastes such as
excavation followed by incineration or landfilling. Finally, bioremediation holds another clear advantage over many
technologies relying on physical or chemical processes: instead of merely transferring contaminants from one medium
to another, biological treatment can degrade the target chemical.
Until recently, the use of bioremediation was limited by lack of thorough understanding of biodegradation processes
their appropriate applications, their control and enhancement in environmental matrices, and the engineering techniques
required for broad application of the technology. The Agency recognized that, along with basic understanding of
biological processes, comprehensive mechanistic process control, engineering design, and cost data are necessary for
the acceptance and use of bioremediation by the technical and regulatory communities. In response to these needs ORD
developed an integrated Bioremediation Research Program, whose mission is to advance the understanding develop-
ment, and application of bioremediation solutions to hazardous waste problems threatening human health and the
environment. The program was designed to strike a balance between basic research activities leading to a fundamental
understanding of biological degradation processes and engineering activities leading to practical scientific applications
of the technology. The Bioremediation Research Program is made up of three major research components: the
Biosystems Technology Development Program, the In Situ Application Program, and the Bioremediation Field
initicitivc.
EPA's bioremediation research efforts to date have already produced significant results in the laboratory at the pilot
scale, and in the field. Accomplishments range from aquifer restoration to soil cleanup to process characterization to
tecnnolnav transfer
technology transfer.
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Section One
Site Characterization
The bioremediation community has focused major research on site characterization. EPA's Biosystems Program
recognizes the basic need for improved site characterization, which includes detailed descriptions of the surface and
subsurface, involving hydrology, geology, and the physics and chemistry of a site. These descriptions will clarify the
opportunities for and constraints to bioremediation.
Cleanup of petroleum spills is a subject of much current research. A recent project focused on the constraints to
bioremediation due to the hydraulic and geochemical nature of a site. A case study of a site contaminated with
tnchloroethylene also found constraints to bacterial remediation under the design condition originally planned Two
other studies described promising results from innovative strategies using Pseudomonas to facilitate bioremediation of
ground water contaminated with hydrocarbons. wouun ^
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Site Characterization at Pipeline Spill at Park City, Kansas:
Estimating Hydraulic and Geochemical Constraints on Bioremediation
Lonnie Kennedy
Coastal Remediation Company
Norman, OK
Engineering design for bioremediation of spills of petroleum
hydrocarbons is strongly influenced by (1) the vertical distribu-
tion of the oily-phase hydrocarbon, (2) the position of the water
table, (3) the hydraulic and pneumatic conductivity of the con-
taminated geological material, (4) the volumetric demand for
electron acceptor to remediate hazardous components of the
spill, (5) geochemical constraints on compatibility of nutrient
amendments in ground water, and (6) respiratory constraints on
the concentrations of amendments to ground water.
Geological Context
In the late 1970s,refined petroleum hydrocarbons were spilled
from a buried pipeline in a rural area in south central Kansas. The
pipeline carried a variety of refined products, including various
grades of gasoline and diesel. The resulting combination of fuels
resembles JP-4 jet fuel. The leak was detected when low concen-
trations of petroleum hydrocarbons were detected in a nearby
municipal water well in 1980.
The site is located within the floodplain of the Little Arkansas
River. Figure 1 is a schematic representation of the geology of the
site. The land surface was originally backswamp muds. Below
the clay is a layer of sand and gravel containing the water table
aquifer. Within the sands there is a general increase in sediment
grain size with depth. Below the aquifer is consolidated bedrock.
Petroleum hydrocarbons usually behave as LNAPLs, meaning
they drain through the unsaturated zone under the influence of
gravity, then accumulate near the water table. At the water table,
they spread laterally in the capillary fringe. The hydrocarbons
moved laterally over 400 ft to the water well, but the maximum
vertical extent of contamination near the spill was only 10 to 15
Emergency Response
The original amount of hydrocarbon released is unknown;
however, several thousand gallons of free-phase product were
pumped from two interceptor trenches installed after theproblem
was identified. The remaining hydrocarbon that collected in the
trenches was ignited and allowed to burn for about a year. By
1984 both trenches had been backfilled.
In an attempt to remove dissolved-phase hydrocarbon from the
aquifer, the municipal water well was pumped for 6 years
beginning in 1980. Pumping was discontinued when hydrocar-
bon could no longer be detected in the extracted water. Unfortu-
nately, when pumping ceased, concentrations of alkylbenzenes
(BTEX) rebounded in the water. Considerable amounts of dis-
solved- and oily-phase hydrocarbon remained in the aquifer;
these problems are common with the pump-and-treat approach to
aquifer remediation (3).
£££$^^
Figure 1. Geological setting of the site.
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Coastal Remediation Company (CRC) began a feasibility
study for in situ bioremediation. CRC joined with the U.S.
Environmental Protection Agency's Robert S. Kerr Environ-
mental Research Laboratory (Kerr ERL) in a Cooperative Re-
search Demonstration Agreement to evaluate nitrate as an elec-
tron acceptor for bioremediation of the spill.
Distribution of the Hydrocarbon
CRC began a site investigation in 1989. Eighteen monitoring
wells and two sets of five clustered piezometers were installed to
define the extent of contaminated ground water and the direction
of ground-water flow. In the spring of 1991, CRC and staff of the
Kerr ERL installed 12 additional monitoring wells and two
additional sets of clustered piezometers to define further the
distribution of oily-phase material. All the monitoring wells were
approximately 25 ft deep with well screens set 5 ft above and 5
ft below the water table. All wells were drilled using a hollow
stem auger drilling rig with continuous coring capabilities. Each
well was fully cored from surface to its greatest depth. The
lithologically distinct intervals were identified and described,
and the core extracts were analyzed for total petroleum hydrocar-
bons and alkylbenzenes (BTEX).
Figure 2 shows the relationship among the layers of geological
materials, the ambient water table, and the vertical distribution of
spilled hydrocarbon in an area adjacent to the spill. The hydrocar-
bon isroughly confined to an interval between the base of the clay
layer and the lowest water table since the spill. The hydrocarbon
penetrates the present water table only a few feet; most of the spill
lies above the present water table.
Delivery of Nutrients and Electron Acceptors
Wells were installed over the spill to distribute nutrients and
electron acceptors. Each well is screened in sandy, unsaturated
material below the clay layer, but above the water table. To
prevent failure (blowout) of the recharge well, the rule of thumb
requires that pressure in a recharge well should not exceed 20
percent of the pressure head between the land surface and the
water table (2). For the demonstration site, this constitutes a value
no greater than 2.0 per square in. (psi) as measured in the
discharge manifolds to the injection wells. Many injection wells
on a fairly tight spacing are needed to deliver the necessary
volume of water required to meet the nutrient and electron
acceptor demand of the spill within a reasonable time period.
More than 400 recharge wells have been installed on a 20-ft
grid spacing. Two large-capacity production wells have been
installed to recirculate ground water through distribution pipe-
lines back to the recharge wells. The production wells, pipelines,
and recharge wells are designed to recharge 125 gallons per
minute (gpm) per acre of surface area.
Aquifer tests indicate that the hydraulic conductivity ranges
from 0.17 to 0.36 cm/sec. Ground-water flow models were used
to estimate the variations in elevation of the water table during
pumping. The predicted rise in the water table, assuming uniform
conditions, is less than 1 ft, which is inadequate to flood the
Total petroleum
hydrocarbon
Depth (mg/kg)
— i (feet) 0 3000 6000
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|Sand[
| Sand j
[c
i •'.'.'•'.'•'•'.'. /. '• /. '« /. '« ;
'::': Clay /£
'/O^^O^V
[siitl
" r
Water
table
lay]
'x'x' Sand and gravel |^x.
r$j3fifiyif-jrfffif!f:J'!f-i'f
££K#
Bedrock £>££•>£
— 5 —
— 10 —
— 15 —
— 20 —
— 25 —
— 30 —
— 35 —
40
Vtf%MW/A
Figure 2. Relationship among spilled hydrocarbons, layers of
geological materials, the water table, and monitoring
wells.
contaminated interval totally (compare Figure 2). The local
conductivity at the depth of the recharge wells may be less than
average, resulting in a higher rise in the water table. Limitations
on water supply and the position of the water table, however,
probably make it impossible to reach all the spill with nitrate-
amended ground water. Remedial technology using air, such as
bioventing, may be required to reach the oily-phase hydrocarbon
in the unsaturated zone.
Supply of Electron Acceptor and Nutrients
Ground water in contact with the spill has a high concentration
of ferrous iron. To minimize problems with oxidation and pre-
cipitation of iron, nitrate was selected as the primary electron
acceptor. The rate of supply of nitrate (N) is limited by the flow
of water and the concentration of nitrate allowed by the state of
Kansas in its permit (10 mg/L as N).
The volumetric electron acceptor demand has been estimated
from the alkylbenzene content of the oily-phase material in the
spill. The relationship between the supply of nitrate and the
average demand, the demand of the most contaminated depth
interval, the rate of expression of nitrate demand, and the rate of
depletion of alkylbenzenes is discussed by Hutchins (1991). In
general, the rate of cleanup is controlled by the rate of supply of
nitrate, which is controlled, in turn, by hydrologic, geochemical,
and regulatory considerations.
A batch desorption isotherm demonstrated that the aquifer
solids could maintain solution concentrations of orthophosphate
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above 0.1 mg/L against very strong phosphate demands. Phos-
phate should not be limiting for biodegradation of hydrocarbons.
Because phosphate precipitates impose a risk of plugging the
recharge wells, no phosphate salts will be added to the recharged
ground water.
References
1. Hutchins, S.R. 1992. Use of nitrate to bioremediate a
pipeline spill at Park City, Kansas: projecting from a
treatability study to full-scale remediation. U.S. EPA's
1992 Symposium on Bioremediation of Hazardous
Wastes: U.S. EPA's Biosystems Technology Develop-
ment Program.
2. Olsthoorn, T.N. 1982. The clogging of recharge wells.
Assembled within the research program of the VEWIN
association, Rijswijk, the Netherlands. August 1982
pp. 15-17.
3. Wilson, J.L. and S.H. Conrad. 1984. Is physical
deplacement of residual hydrocarbons a realistic possi-
bility in aquifer restoration? Conference on Petroleum
Hydrocarbons and Organic Chemicals in Ground Wa-
ter: Prevention, Detection, and Restoration, NWWA/
API. Houston, Texas, pp. 274-298.
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Field Demonstration of Innovative Bioremediation
Strategies for PCP and Creosote
James G. Mueller, Cherie S. Heard, and Suzanne E. Lantz
SBP Technologies, Inc.
Stone Mountain, GA
and
Richard J. Colvin and Derek Ross
The ERM Group
Exton, PA
and
Douglas P. Middaugh and Parmely H. Pritchard
U.S. Environmental Protection Agency
Environmental Research Laboratory
Gulf Breeze, FL
In earlier papers, we had described the isolation and character-
ization of microorganisms capable of utilizing (mineralizing)
high-molecular-weightpolycyclic aromatic hydrocarbons (HMW
PAHs) and other recalcitrant creosote constituents as sole sources
of carbon and energy for growth (3,5,6). In addition, an axenic
culture ofPseudonwnas sp. strain SR3 was shown to mineralize
pentachlorophenol (PCP) when supplied as a sole carbon source
in liquid medium (12).
The object of the current study was to evaluate, at bench- and
pilot-scale levels, the ability of these specially selected microor-
ganisms to facilitate the bioremediation of ground water con-
taminated with creosote and PCP. The performance of the
bioremediation process was evaluated according to chemical
analyses of system influent, effluent, and bioreactor residues by
performing a mass balance evaluation and comparative biologi-
Ground-water
equilization/
feed tank
Ground-water
feed
Bioreactor I
CSTR
or
Membrane
Concentrate
indigenous microbes
or strains CRE1-13
Discharge
Bioreactor 2
(2a, b, and c)
BATCH
strains
EPA505
and SR3
= Sampling points
Figure 1. Ground-water bioremediation system.
11
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Table 1. Summary of Toxicological Analyses
Bioreactor Sample
Microtox
ECSO 5 min
Mysids
LCso96h
Ceriodaphnia
Embryonic Menidia
(1% solutions)
Specially selected microorganisms
Feed
Biorx 1
Effluent
Non-specific microorganisms
0.0185
2.3645
1.2956
0.01
0.34
0.75
0.007
0.710
0.530
embryotoxic
teratogenic
teratogenic
Feed
Biorx 1
Effluent
0.1456
25.057
54.107
0.02
0.32
0.16
0.079
0.292
0.230
embryotoxic
teratogenic
teratogenic
cal toxicity measurements. Results obtained upon the addition of
specialty microorganisms were compared with those obtained
when nonspecific microorganisms (i.e., obtained from soil at the
site and activated sludge) were employed for bioreactor opera-
tions.
The bioremediation process was field-tested with ground wa-
ter contaminated by creosote and PCP recovered from the Ameri-
can Creosote Works (ACW) Superfund Site, Pensacola, Florida.
A two-stage, continuous-flow, sequential inoculation system
was developed for the efficient biodegradation of creosote and
PCP (Figure 1).
The first stage of the pilot-scale system used a 120-gallon
EIMCO reactor (bioreactor 1). Ground water was delivered to
this bioreactor for 8 days at an average rate of 30 gallons per day
(4-day hydraulic retention time). Effluent from bioreactor 1 was
delivered to one of three, 60-gallon, completely mixed batch
reactors that served as the second bioreactor (bioreactors 2a, 2b,
and 2c). Once each chamber of bioreactor 2 was filled with
effluent from bioreactor 1, material was held in the batch mode
for 4 days. At the end of the 4-day batch reaction, the effluent was
clarified by settling biomass, and was placed into a holding tank
for final analysis and testing prior to discharge.
100
so
S 60H
n
T5
C
ffl
g
C
'to
a>
DC
40 —
20 —
31 ppm
539 ppm
0.1 ppm
368 ppm
1.6 ppm
F~| Input
I23 Output
5.2 ppm
Bi-cyclic compounds
3-ring PAHs
4 or more ring PAHs
Figure 2. Reduction of PAH contaminants in creosote-contaminated ground water.
12
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When specially selected bacteria were added to the system a
one-time inoculum of a 13-member bacterial community with
known degradative abilities was reconstituted from frozen stock
cultures (strains CRE1 through CRE13) and added to bioreactor
1 at a concentration of approximately 1 x 106 cells of each strain/
mL. Daily additions of Pseudomonas paucimobilis strain EPA 2
505 and Pseudomonas sp. strain SR3 (approximately 1 x 106
cells/mL) were added to each filled chamber of bioreactor 2
Daily measurements were taken of TCOD, SCOD, TSS VSS
DO, pH, oxygen uptake, temperature, NH3-N, reac'tive-P total
creosote, and PCP (GC-FID, GC-ECD analyses), and select
biological toxicity assays to monitor bioreactor activity and 3
performance.
With the two-stage, continuous-flow, sequential inoculation
design, the concentration of creosote constituents was reduced
from about 1,000 ppm in the ground-water feed to <7 ppm in the 4
system effluent (removal efficiency of >99 percent) Also the
toxicity and teratogenicity of the bioreactor effluent was signifi-
cantly reduced as determined by Microtox™, embryonic Menidla
rfu'/'T; Mysid°Psis bahia< and Ceriodaphnia dubia assays 5
(lable 1). Notably, the cumulative concentration of 9 HMW
PAHs (each containing 4 or more fused rings) was reduced from
368 ppm in the ground-water feed to 5.2 ppm in the system
effluent (Figure 2). However, mass balance evaluation found
approximately 20 percent of the HMW PAH associated with 6
residues in the flow lines, biomass, and reactor sludges.
When nonspecific (indigenous) microorganisms were used
bioreactor performance was much reduced. It was consistently
observed that indigenous microbes were unable to degrade PCP
Upon the addition of the PCP-degrading bacterium (Pseudomo-
nas sp. strain SR3) to the bioreactor system, however the
concentration of PCP in ground-water feed was reduced from
256 ppm to 31 ppm in the bioreactor effluent. Hence in the
absence of SR3 inoculation, recovery of PCP from the bioreactor
system could be used as a surrogate standard demonstrating an 85
to 90 percent recovery of the chemical tracer. g
Acknowledgements
Technical assistance was provided by Myke'lle Hertsgaard
Barbara Artelt, Beat Blattmann, Maureen Downy Sol Resnick
(Technical Resources, Inc.), Ellis Kline, Edward Kouba and
Arjan Van Buul (SBPTechnologies, Inc.). Assistance from Scott 9
Beckman, K.C. Mahesh (S AIC), Kim Kreiton (U.S. EPA, RREL
SITE Program), and Madolyn Streng (U.S. EPA, Region IV) is
also gratefully acknowledged.
Financial support for these studies was provided by the U S
EPA SITE Program, Cincinnati, Ohio. This work was performed
as part of a Cooperative Research and Development Agreement 10
between the Gulf Breeze Environmental Research Laboratory
and SBP Technologies, Inc. (Stone Mountain, Georgia) as de-
fined under the Federal Technology Transfer Act, 1986 (contract
no. FTTA-003).
References
1. Middaugh,D.P.,J.G.Mueller,R.L.Thomas,S.E.Lantz 11.
M.J. Hemmer, G.T. Brooks, and P.J. Chapman. 1991.
7.
Detoxification of creosote- and PCP-contaminated
ground water by physical extraction: chemical and
biological assessment. Arch. Environ. Contam. Toxicol.
21:233-244.
Middaugh, D.P., S.M. Resnick, S.E. Lantz, C.S. Heard
and J.G. Mueller. 1992. Toxicological assessment of
biodegraded pentachlorophenol: Microtox™ and em-
bryonic Menidia. Arch. Environ. Contam Toxicol
Submitted.
Mueller, J.G., PJ. Chapman, B.O. Blattmann, and P H
Pritchard. 1990. Isolation and characterization of a
fluoranthene-utilizing strain of Pseudomonas
paucimobilis. Appl. Environ. Microbiol. 56:1079-1086.
Mueller, J.G., P J. Chapman, and P.H. Pritchard. 1989
Creosote-contaminated sites: their potential for biore-
mediation. Environ. Sci. Technol. 23:1197-1201.
Mueller, J.G., PJ. Chapman, and P.H. Pritchard. 1989
Action of a fluoranthene-utilizing bacterial community
on polycyclic aromatic hydrocarbon components of
creosote. Appl. Environ. Microbiol. 55:3085-3090.
Mueller, J.G.,P.J. Chapman, S.E. Lantz, B.O. Blattman
and P.H. Pritchard. 1992. Mineralization of PAHs by
Pseudomonas paucimobilis strain EPA505 and identi-
fication of fluoranthene metabolites. Appl Environ
Microbiol. Submitted.
Mueller, J.G., RJ. Colvin, C.S. Heard, S.E. Lantz D P
Middaugh, D. Ross, and P.H. Pritchard. 1992. Biore-
mediation strategies for the treatment of ground water
contaminated with creosote and pentachlorophenol
Envron. Sci. Technol. Submitted.
Mueller, J.G., S.E. Lantz, B.O. Blattmann, and PJ
Chapman. 1990. Alternative biological treatment pro-
cesses for remediation of creosote-contaminated mate-
rials: bench-scale treatability studies. EPA/006/9-90/
04:7. 0:7 pp.
Mueller, J.G., S.E. Lantz, B.O. Blattmann, and PJ
Chapman. 1991. Bench-scale evaluation of alternative
biological treatment processes for remediation of pen-
tachlorophenol- and creosote-contaminated materials-
solid-phase bioremediation. Environ. Sci Technol
25:1045-1055.
Mueller, J.G., S.E. Lantz, B.O. Blattmann, and PJ
Chapman. 1991. Bench-scale evaluation of alternative
biological treatment processes for remediation of pen-
tachlorophenol- and creosote-contaminated materials-
slurry-phase bioremediation. Environ. Sci. Technol!
25:1055-1061.
Mueller, J.G., D.P. Middaugh, S.E. Lantz, and PJ
Chapman. 1991. Biodegradation of creosote and pen-
13
-------
tachlorophenol in ground water: chemical and biologi-
cal assessment. Appl. Environ. Microbiol. 57:1277-
1285.
12. Resnick, S.M. and P.J. Chapman. 1990. Isolation and
characterization of a pen tachlorophenol (PCP)-degrad-
ing gram-negative bacterium. Abstr. 90th Ann. Meeting
Am. Soc. Microbiol., Q70, p. 300.
14
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Case Study on Site Characterization at a TCE Plume:
St. Joseph, Michigan, NPL Site
John Kuhns
U.S. Environmental Protection Agency
Chicago, IL
and
Peter Kitanidis
Stanford University
Stanford, CA
A plume of TCE in a sandy water table aquifer originates from
an industrial facility in St. Joseph, Michigan. The origin of the
plume is on a ground-water divide. The plume bifurcates; part
drains to the east and part to the west. Based on concentrations
of compounds in conventional monitoring wells, bioremediation
through co-oxidation by methane-oxidizing bacteria was pro-
posed as the remedy for the west plume. To calibrate a design
model for bioremediation, it was necessary to estimate the
average concentration of TCE and its dechlorination products in
the ground water, and the flux of TCE and dechlorination
products along the plume.
Ground-Water Sampling
Estimates of contaminant concentrations were obtained from
two transects extending across the plume roughly perpendicular
to the flow of ground water. To sample ground water, an auger
slotted over the first 5 ft of its length was drilled into the earth.
Starting at the water table about 40 ft below land surface, the
auger was advanced 5 ft, water was pumped until conductivity
and redox stabilized, and samples were taken for chemical
analysis; then the auger was advanced another 5 ft to take the next
sample. This depth-discrete sampling was continued to the
bottom of the aquifer 80 ft below the land surface. Borings were
spaced along the transects at intervals of 20 to 40 ft. Both
transects extended all the way across the plume.
Estimates of Contaminant Flux
Concentrations in the plume were estimated by averaging the
individual depth-discrete samples. More than 50 such samples
were included in each transect. The hydraulic gradient was
multiplied by the hydraulic conductivity to determine the flow of
water in the plume. The flow was multiplied by the concentra-
tions of contaminants to estimate the flux. The flux of TCE off
the industrial facility was 60 kg/year. The flux of cis DCE and
vinyl chloride was 71 and 11 kg/year, respectively.
Implications for Bioremediation
The design model strongly indicated that bioremediation could
not reach acceptable concentrations for TCE or cis DCE when
methane was used as the primary feedstock for aerobic biotrans-
formation. As a result, plans for a pilot-scale evaluation of
bioremediation were abandoned. This case study illustrates the
importance of site characterization for proper implementation of
bioremediation. If data from the conventional monitoring wells
had been used to design the remedy, it would not have performed
as expected.
15
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-------
Approaches to the Characterization of TVichloroethylene
Degradation by Pseudomonas Cepada G4
Charles Somerville and Michael Reagin
Technical Resources, Inc.
Gulf Breeze, FL
and
Robert Gerger
U.S. Environmental Protection Agency
Gulf Breeze, FL
and
Malcolm Shields
University of West Florida Center for
Environmental Diagnostics and Bioremediation
Gulf Breeze, FL
Introduction
inthepresenceofaiomiAKCompomds, Pseudomonas cepada
G4 expresses a novel pathway for the metabolism of toluene (4)
The first two steps of the pathway are mediated by toluene ortho-
monooxygenase (Tom, Figure 1), an enzyme that also attacks
tnchloroethylene (TCE). To determine the genetic location of the
Tom gene for further study, aseriesof Tn5 mutants was made
b.
Tom
Tom
Tom
TFMP
in P. cepacia G4 (Table 1). One of the mutants, P. cepada G4
5223, lost the ability to degrade TCE, but gave rise to spontane-
ous revertants" (e.g., strain Phel), which express Tom activity
regardless of the presence of aromatics (2,3). While the use of P
cepacia G4 in the bioremediation of TCE-contaminated sites is
tanned by its requirement for aromatic compounds to induce
Tom activity, P. cepada G4 5223 Phel has no such require-
C230
(coo
N^ OH
C230
OH
TFMC
TFHA
17
-------
Xmnl
Seal
Ppal- Gsul
A1wNi
Afllll
Aval
Xmnl
derived from pLV59 are unshaded.
Aval Hindi
-t Seal
(EcoRI)
Pstl
Sail
•Xbal
3 //-MluT
Ndel
lindlll
. :Bg1M
Pstl
Aval
P cepacia DNA are shaded. Portions of the plasmid derived from pLV59 are unshaded.
ment Phel, therefore, provides significant advantages over
other bacteria for the bioremediation of TCE-contaminated aqui-
fers (2). Further study of the genetic basis for TCE degradation
in P cepacia G4 and the mechanism of Tom expression in Phel
is necessary to optimize their potential for application to biore-
mediation.
Results and Discussion
P cepacia G4 and its derivatives carry two cryptic plasmids ot
approximately 50 and 150 kb. In an effort to enhance the fitness
of Phel for use in bioremediation, a third plasmid, pROlOl, was
introduced via conjugation. Plasmid pROlOl, which carries the
genes of the 2,4-D degradative pathway, was intended to enable
Phel to eliminate potentially toxic intermediates from the me-
tabolism of halogenated aromatics that it would be likely to
encounter in polluted aquifers. Subsequent plasmid assays, how-
ever, indicated that some recipients of pROlOl had lost the large
resident plasmid (pPhelL). Those organisms, designated Phel
cure, lost the ability to degrade TCE and to transform tnfluoro-
methylphenol (TFMP) to trifluoroheptadienoic acid (TFHA), a
colorimetric reaction used to detect Tom activity (Figure 1). Phel
cure remained resistant to kanamycin, however, suggesting that
18
-------
Table 1. Relevant Characteristics of P. cepacia G4 and TnS-derivative Strains.
Designation
P. cepacia G4
P. cepacia G4 5220
P. cepacia G4 5223
P. cepacia G4 5227
P. cepacia G4 5223 Phe 1
TCE'
+C'
TFMP"
+C
Stability c
' Ability to degrade trichloroethylene
b Ability to produce hydroxytrifluoroheptadienoic acid from trifluoromethylphenol.
° Relative stability of phenotype after several passages on a nonselective medium
" Inducible activity.
' Constitute activity.
Phenotype
N/A
stable
yields TCE**
stable
stable
N/A
C230-
Tonr
Tonr
Tom+c
N/A
torn regulatory
torn
unknown
Tn5 does not reside on the large plasmid in Phel. Matings
between wild-type G4 and Phel cure yielded putative Phel
recipients of the large plasmid from G4 (pG4L), which had
regained wild-type (inducible) activity against TCE. Clearly the
large plasmid plays a role in toluene metabolism and TCE co-
metabolism in P. cepacia G4 and its derivatives. However there
may also be involvement of the chromosome and/or the smaller
plasmid, as indicated by loss of TCE activity in strain 5223
(parent of Phel), which does not seem to carry the Tn5 insertion
on the large plasmid.
While Phel is currently being tested for its ability to remove
TCE from contaminated water, efforts are also being directed
toward cloning genes involved in Tom activity from plasmid and
chromosomal DNAs. The advantages of cloned Tom activity
include the ability to alter expression of the gene, to place the gene
in a well-defined genetic background and the potential to produce
purified enzyme(s) for specialized field applications. To facili-
tate screening for Tom activity in potential clones, a novel host-
vector system has been constructed. The vector, pUCLVl (Fig-
ure 2), carries the£coRI endonuclease and methylase genes from
P^n !?' 3nd the 6-lactamase gene and replication origin of
PUL19 (5). The EcoSI methylase gene is temperature sensitive
and functions only at temperatures below 28°C. At high tempera-
tures, therefore, the methylase is inactive, and the cell is rapidly
killed due to the expression of EcoRl. EcoKl expression can be
inactivated by cloning into one of three unique restriction sites
within the EcoW gene. The result is a very tightly controlled
selection against nonrecombinant transformants on a high copy-
number plasmid. High copy number was considered desirable in
the cloning vector to increase gene dosage and, therefore to
increase the chances of detecting a weakly expressed gene
product.
The host cell for cloning is an E. coli strain that carries the gene
forcatechol-2,3-dioxygenase(C230,seeFigurel)onalowcopy-
numberplasmid (pCDOS). PlasmidpCDOS (Figure 3) is an over-
producer of C230, which carries the Cmr gene and origin of
replication from pLV59. The high level of expression of theC230
gene is fortuitous in that it will facilitate the detection of Tom
activity carried on the primary cloning vector.
In this system, Tom activity carried on the cloning vector can
bedetectedbytheconversionofTFMPtotrifluoromethylcatechol
VWT)- TFMC 1S> *"turn'a substrate for the C230 expressed by
pCDOS, producing the bright yellow-ring cleavage product
TFHA. Thus, the combination of plasmids pUCL V1 and pCDO5
allow for a rapid and sensitive detection of recombinant Tom
activity. While/', cepacia G4 5223 Phel has strong potential for
immediate use in bioremediation, the use of this new host-vector
system to isolate and characterize the gene involved in TCE
removal may provide even more powerful ways to address the
problem in the future.
References
1. O'Connor, C.D. and G.O. Humphreys. 1982. Gene
2.
3.
4.
5.
Shields, M.S. and M.J. Reagin. 1991. Abstract, 91st
General Meeting of the American Society for Microbi-
ology, Dallas, Texas.
Shields, M.S., S.O. Montgomery, S.M. Cuskey PJ
Chapman, and P.H. Pritchard. 1991. Appl Environ
Microbiol. 57:1935.
Shields, M.S., S.O. Montgomery, PJ. Chapman S M
Cuskey, and P.H. Pritchard. 1989. Appl. Environ
Microbiol. 55:1624.
Yanisch-Perron, C., J. Vieira, and J. Messing 1985
Gene 33:103.
19
-------
-------
Section Two
Bioremediation Field Initiative
e* ^__i_.... lvl.!es a. more tnan 130 Superfund sites, RCRA corrective action facilities, and TinHf>rorr.i,nH
^^
21
-------
-------
Progress in the Field Applications of Bioremediation
Fran V. Kremer
U.S. Environmental Protection Agency
Cincinnati, OH
and
Nancy Dean
U.S. Environmental Protection Agency
Washington, DC
The U.S. Environmental Protection Agency (EPA) has devel-
oped an initiative to expand the nation's field experience in the
use of bioremediation. The EPA's Office of Solid Waste and
Emergency Response and the Office of Research and Develop-
ment have instituted the Bioremediation Field Initiative to pro-
vide solid performance data to sufficiently define the capabilities
of this technology. The initiative provides assistance to the
Regions and states in conducting field tests and carrying out
independent evaluations of site cleanups using bioremediation.
Sites presently considered in this initiative include Superfund,
RCRA corrective action facilities, and Underground Storage
Tank (UST) sites. This initiative has three objectives. The first is
to more fully document performance on full-scale applications of
bioremediation. Seven field evaluations are being conducted on
petroleum, wood preserving, and solvent wastes contaminating
soil and ground water. The second is to provide technical assis-
tance to site managers for sites in a feasibility or design stage to
facilitate conducting treatability studies, field pilot studies, and
other studies through the Agency's Technical Support Centers.
The third objective is to provide information regularly on treat-
ability studies, design, and full-scale operations of bioremedia-
tion projects. Current cost and performance data will be available
on the operation of biological treatment systems for a variety of
wastes and contaminated matrices.
23
-------
-------
Evaluation of Full-Scale In Situ and Ex Situ Bioremediation of Creosote Wastes
John E. Matthews
U.S. Environmental Protection Agency
Ada, OK
The bioremediation field initiative has three main objectives:
(1) to document more fully the performance of full-scale biore-
mediation field applications in terms of treatment effectiveness,
operational reliability, and cost; (2) to provide technical assis-
tance to EPA and state agency site managers that are overseeing
or considering the use of bioremediation; and (3) to develop a
treatability data base that will be available through EPA's Alter-
native Treatment Technology Information Center (ATTIC). The
performance evaluation project in Libby, Montana, focuses
primarily on the first objective. Champion International Super-
fund Site in Libby, Montana, was nominated by the Robert S.
Kerr Environmental Research Laboratory as a candidate site for
performance evaluation.
The first phase of the bioremediation performance evaluation
(which has been completed) is to summarize previous and current
remediation activities. This summary is to be followed by iden-
tification of critical site characterization and treatment param-
eters that are important to evaluate bioremediation for each of the
treatment units identified. The final objective is to evaluate
bioremediation performance based on the information obtained.
Three biological treatment processes wiO be addressed: (1)
surface soil bioremediation in a prepared-bed, lined land treat-
ment unit (LTU); (2) oil/water separation of extracted ground
water and treatment of the aqueous phase in an above-ground
fixed-film bioreactor; and (3) in situ bioremediation of the upper
aquifer at the site. Each biological treatment process will be
addressed with regard to design, performance, and monitoring
activities.
Two forms of wood preservative were used at the site: creosote
and pentachlorophenol. Polycyclic aromatic hydrocarbons are
the primary components of concern at the site and are associated
with the soil phase primarily by adsorption. Contaminated soil
from three primary source areas (tank farm, butt dip, and waste
pit) has been excavated and moved to one central location (waste
pit). The soil is pretreated in the waste pit area and is further
treated in a prepared-bed, lined LTU. Planned activities associ-
ated with the field initiative are (1) statistical sampling of the soil
treatment unit, (2) field-scale treatment kinetics, (3) toxicity
reduction, (4) clean-up levels achieved, (5) influence of moisture
and soil structure, and (6) mass balance of contaminants by soil
and leachate analyses.
The upper aquifer ground-level treatment unit is provided for
the separation of LNAPL and DNAPL from ground water ex-
tracted for subsequent biological treatment prior to reinjection
via the infiltration trench. The subsequent biological treatment is
two, fixed-film reactors that will be operated in series. The first
reactor is for roughing purposes, while the second is for polishing
and reoxygenation of the effluent prior to reinjection. Planned
activities include (1) flow composited sampling, (2) evaluation
and prediction of reactor performance, (3) analysis of biofilm
dynamics, (4) mass balance of contaminants, and (5) treatment
optimization.
The upper aquifer in situ bioremediation system involves the
addition of hydrogen peroxide and inorganic nutrients to stimu-
late the growth of contaminant-specific microbes (Figure 1).
Planned activities include (1) dissolved oxygen profiles, (2)
aquifer material sampling to distinguish abiotic and biotic ef-
fects, (3) dissolved oxygen uptake evaluation and correlation to
the rate of biodegradation, and (4) toxicity reduction.
The current status of the three biological treatment processes
is as follows.
Above-Ground Bioreactor
The preliminary performance analysis of the existing bioreactors
indicates the following: (1) removal of PAHs andPCPis strongly
influenced by flow rate through thereactor(10gallonsperminute
versus 15 gallons per minute); (2) areas within the reactor may be
anaerobic as indicated by dissolved oxygen measurements taken-
(3) removal of PAHs (80 to 90 percent) is more efficient than
removal of PCP (40 to 80 percent); (4) nutrients in the form of
nitrogen and phosphorus, although added to the reactor, are not
proportioned either to flow rate or to mass loading of contami-
nants, and are generally insufficient; (5) it was possible to
increase the removal of PCP within the bioreactor by lowering
the flow rate through the reactor from 15 gpm to 10 gpm.
A pilot-scale bioreactor has been designed and constructed for
physical placement beside the full-scale bioreactor at the site.
The pilot-scale reactor will simulate the behavior of the full-scale
reactor and allow modification of nutrients, flow rate, and dis-
solved oxygen to evaluate the effect of process parameters,
without interfering with the operation of the full-scale reactor.
The pilot-scale bioreactor has been hydraulically tested in the
25
-------
Proposed
boundary
injection
system
N
• Contaminant
plume
Waste pit area
Infiltration trench
Bioreactor
housing
Rock
filter
LTU
LTU
Scale (feet)
0 300 600
Regional
ground-water
flow
<^p Monitoring well
O Extraction well
D Injection well
Figure 1. In situ bioremediation of upper aquifer (from Woodward
Clyde).
laboratory and will be transported to the site in the spring of 1992.
In Situ Ground Water
Results of both sampling trips indicate that the aquifer materi-
als below the site are in a chemically reduced condition (presence
of reduced iron and manganese); therefore, the site has an abiotic
oxygen demand in addition to any biological oxygen demand.
This situation provides challenges to designing in situ evalua-
tions for determining how much uptake of oxygen is due to
microorganisms. Therefore, the absence of hydrogen peroxide in
one well and the presence of oxygen in a downstream well that
has received hydrogen peroxide upgradient may indicate that the
abiotic oxygen demand, rather than the biological oxygen de-
mand, has been met. It is likely that both abiotic and biotic
demands for oxygen are occurring simultaneously.
No additional ground-water samples have been taken since
October 1991 because of the onset of winter. Present plans are to
take aquifer material in the spring (if possible) for PAH analysis
and for reducing potential. A single-well, "push-pull" test, as
well as a multiple-well test, will be performed to determine
oxygen consumption at contaminated and background wells at
the site.
Land Treatment Units
Total estimated contaminated soil volume for treatment is
45,000 yd3 (uncompacted). Contaminated soil clean-up goals
(dry-weight basis) are (1) 88 mg/kg total (sum of 10) carcino-
genic PAHs, (2) 8 mg/kg naphthalene, (3) 8 mg/kg phenanthrene,
(4) 7.3 mg/kg pyrene, (5) 37 mg/kg PCP, and (6) < 0.001 mg/kg
dioxin equivalency.
The LTU comprises two adjacent 1-acre cells. Each cell is
lined with low-permeability materials to minimize leachate infil-
tration from the unit. When reduction of contaminant concentra-
tions in all lifts placed in the LTU has reached an acceptable level,
a protective cover will be installed over the total 2-acre unit and
maintained in such a way as to minimize surface infiltration,
erosion, and direct contact.
Contaminated soil is applied to and treated in lifts (approxi-
mately 9 in. thick) in the designated LTU cell until target soil
contaminant levels are achieved for a given lift. Degradation
rates, volume of soil to be treated, initial contaminant concentra-
tion, yearly degradation period duration, and LTU cell size
determine time required for remediation of a given lift. Based on
an estimated 45-day timeframe for remediation of each applied
lift to acceptable contaminant levels, an estimated 45,000 yd3 of
contaminated soil, and a 2-acre total LTU surface area, projected
time to complete soil remediation is 8 to 10 years.
The primary purpose of the LTU soil sampling programs being
carried out in this project is to determine statistical significance,
confidence of rate, and extent of contaminated soil biodegrada-
tion at this site. A quantitative expression for data variability is
necessary to determine an accurate estimate of biodegradation of
these contaminants at field scale. Such an expression will allow
data generated to be used by others to help estimate biodegrada-
tion potential of similar types of wastes under similar conditions
at other sites. A critical element in design of this project involves
obtaining a good representation of mass balance for those com-
pounds being monitored. Correspondingly, assessment of degra-
dation intermediates, humification, leachate, soils, and air qual-
ity is planned. If Phase II sampli ng and analytical program design
will allow tracking of target contaminants through the system,
investigators can determine an accurate fate and transport profile
that denotes biodegradation performance. Completion of other
project activities, such as degradation kinetics and toxicity reduc-
tion studies, will generate data that can be used to help assess
overall bioremediation effectiveness and, subsequently, help
predict performance of bioremediation processes at other, simi-
lar sites.
Three field sampling events directed toward Phase II perfor-
mance evaluation studies were completed during 1991. The first
event, in early May, consisted of collecting soil core samples for
analyses immediately following placement of treatment lift #3
(first 1991 lift;lifts#l and #2 were applied in 1990) in LTU cell
#1. A second sampling event in late June involved collection of
32 soil core samples during treatment process for lift #3. In late
July, treatment lift #4 (second 1991 lift) was placed in LTU cell
#l,andlift#l was placed in cell #2. A third sampling event in late
26
-------
September involved collection of 32 core samples from both
LTU cell #1 and LTU cell #2. All 32 core samples collected for
each sampling event penetrated to the bottom of treatment lift # 1.
Performance Evaluation Data-Year 1
Data on the concentrations of pentachlorophenol are presented
in Table 1, on phenanthrene in Table 2, on pyrene in Table 3, and
on total carcinogenic PAHs in Table 4. Data currently available
from the three 1991 LTU sampling events are presented. Full
results from these sampling events will be available after data
quality checks are completed for the remaining data.
These data indicate that the variability inherent in lift applica-
tion and sampling is high enough to account for any differences
between discrete sampling regimes, as in the performance evalu-
ation, and composite sampling regimes, as in the operations
monitoring.
It may be noted that the pentachlorophenol concentrations
have a lower variability in lifts #1 and #2 than in lift #3, applied
just before sampling. Lifts #1 and #2 had been tilled many times
since application, and perhaps this tilling had mixed the soil
sufficiently to reduce the spatial heterogeneity of the PCP con-
centrations. No doubt some of the lower variability is due to the
lower concentrations of PCP in lifts #1 and #2 at the sampling
date being considered. If it is assumed that the initial concentra-
tions of PCP at application were similar for all lifts, then there is
some indication that the PCP concentrations are being reduced.
Phenanthrene concentrations are lower than treatment levels,
except for a small number of the discrete samples. One high value
in lift #1 (151 mg/kg) markedly increases the variability of the
data in this lift. Otherwise, phenanthrene levels were low even
immediately after application of the lift (lift #3).
Pyrene concentrations in the LTU are somewhat more variable
than pentachlorophenol and phenanthrene concentrations. This
may be partly due to the tendency of the larger molecule to be
more strongly adsorbed on the soil, causing the measured con-
centrations to be more sensitive to variations in the efficiency of
the extraction step in the analytical process. The concentrations
in lifts # 1 and #2 are clustered near the treatment level, with many
just above or just below the treatment level. The reduction in
concentration is apparently good, especially in lifts #1 and #3.
Since lift #2 concentrations were low at application of the lift, the
reduction is not as apparent as in lifts #1 and #3.
The variability of the total carcinogenic PAHs concentration,
being the sum of the variability of 10 individual components, is
high. The data variability precludes any definitive conclusion
concerning concentration reductions for this parameter. Most of
the concentrations are around the 88 mg/kg treatment level.
Table 1. Pentachlorophenol Concentrations In
The Libby Soil
Treatment Facility (mg/kg)
Sampling Date 5/6/91
*^*gMri:>rjii,1r
East' o l '-
§;;
ioA
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Depth
Lift 3
Lift 2
Liftl
Lift3
Lift 2
Liftl
Lift 3
Lift 2
Liftl
LiftS
Lift 2
Liftl
% " i •> Ji ^ - «
f W ^. :
229
30
76
257
4
31
189
0
42
275
226
156
k ;% ^ ^ * ^ ' i
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313
155
9
217
15
28
321
9
17
891
26
13
x,'* -;
56$;
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33
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42
21
38
367
21
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798
177
165
°"*.
843-
49
12
33
167
0
17
204
17
28
710
18
176
*&£**
•^ % *
0
12
30
303
0
29
440
11
64
212
24
136
' 'till
43
144
147
325
58
380
336
11
27
55
12.5
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====c
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435
79
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761
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94
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667
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% f V" i -J^ •• t
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169
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275
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56
27
-------
ro
oo
Table 2. Phenanthrene In The Libby Soil Treatment Facility (mg/kg)
Sampling Date 9/18/91
Sampling Location
Feet Sample
East Depth
Lift 4
S44> Lift 3
Lift 2
Lift 1
Lift 4
56.6 Lift 3
Lift 2
Lift 1
Lift 4
283 Lift 3
Lift 2
Lift 1
Lift 4
0 Lift 3
Lift 2
Liftl
Feet South
U 283 564 S4J 1
32 31 30 29
5.6 6.6 6.4 4.4
1.7 2.6 0.9 1.4
0.2 4.1 1.0 3.6
0.3 1.9 ND 0.8
17 18 19 20
3.1 5.6 6J 12
2.8 4.3 0.9 3.9
1.6 1.6 2.7 8.1
1.2 2.6 1.0 3.0
16 15 14 13
12.0 4.8 3.8 2.4
3.3 0.8 3.7 ND
13.0 1.6 ND ND
2.7 1.4 3.5 3.1
1234
1.4 3.0 2.8 ND
0.2 0.9 2.9 ND
6.1 2.9 4.6 0.2
ND 2.5 3.3 5.5
!&£ 14L5 10*8 mi
28 27 26 25
7.4 7.5 10.9
2.0 ND 4.0 3.1
0.6 ND 2.7 7.0
0.6 ND 3.8 ND
21 22 23 24
10.2 8.7 10.5 4.5
2.8 6.8 2.8 7.0
4.7 2.1 12 4.8
0.2 1.6 42 1.6
12 11 10 9
4.9 3.3 ND 4.3
ND 3.0 ND 2.5
1.7 ND 1.1
2.8 3.7 1.4 1.9
5678
ND 2.7 4.7 4.0
ND 0.8 2.3 ND
2.0 ND 22 ND
3.6 3.5 3.5 1.6
-------
ro
CO
Table 3. Pyrene In The Libby Soil Treatment Facility (mg/kg)
Sampling Date
Sampling Location
Feet
East
84.9
$6J$
28 .3
0
Sample
Depth
Lift 4
Lift 3
Lift 2
Liftl
Lift 4
Lift 3
Lift 2
Liftl
Lift 4
Lift3
Lift 2
Liftl
Lift 4
Lift3
Lift 2
Lift 1
» 283
32 31
5/6/91
FeetSoutb
56.6
30
84.9 im 14LS 169.8 198.1
29 28 27 26 25
17 18
19
20 21 22 23 24
16 15
79.4 159.6
ND 23.7
266.6 26.4
1 2
8.9 62.7
1L3 13.3
23.3 3.3
14
75.9
17.1
3.8
3
95.9
15.4
8.1
13 12 11 10 9
92.7 118.0 82.0 105.3 110.0
13.4 10.1 7.9 ND 10.8
14.4 ND 18.8 ND 9.1
4 5 6 78
66.1 39.6 80.1 28.4 44.9
ND 19.8 ND 172 6.1
282 143 133.6 30.7 ND
-------
CO
o
Table 4. Total Carcinogenic PAHs
In The Libby Soil Treatment Facility (mg/kg)
Sampling Date
Sampling Location
Feet
East
84.9
56.6
283
0
Sample
Depth
Lift 4
Lift 3
Lift 2
Lift 1
Lift 4
Lift3
Lift 2
Lift 1
Lift 4
Lift 3
Lift 2
Liftl
Lift 4
Lift3
Lift 2
Liftl
9/18/91
Feet South
0 283
32 31
352 194
53 134
47 81
38 74
17 18
165 352
165 217
94 94
35 232
16 15
223 254
108 370
357 86
116 92
1 2
44 187
81 120
143 154
23 116
56.6
30
187
101
84
86
19
250
42
117
43
14
150
106
51
71
3
128
136
274
188
84.9
29
131
83
93
53
20
317
160
90
102
13
130
159
124
131
4
151
63
59
119
113.2
28
241
51
59
35
21
201
115
114
22
12
223
193
125
5
17
19
88
151
141.5
27
78
66
19
22
192
187
129
116
11
172
149
72
113
6
ill
131
172
192
169.S
26
282
79
60
105
23
241
88
75
104
10
185
229
220
68
7
211
150
113
212
198.1
25
247
228
206
23
24
110
172
174
117
9
244
151
71
114
8
134
113
71
68
-------
Optimizing Bioventing in Shallow Vadose Zones and Cold Climates:
Eielson AFB Bioremediation of a JP-4 Spill
Gregory D. Sayles and Richard C. Brenner
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
Robert E. Hinchee
Battelle Laboratories, Columbus Division
Columbus, OH
and
Catherine M. Vogel
U.S. Air Force Engineering Services Center
Tyndall Air Force Base, FL
and
Ross N. Miller
U.S. Air Force Center for Environmental Excellence
Brooks Air Force Base, TX
Introduction
Bioventing is the process of supplying oxygen in situ to
oxygen-deprived soil microbes by forcing air through unsatur-
ated contaminated soil at low flowrates (1). Unlike soil venting
or soil vacuum extraction technologies, bioventing attempts to
stimulate biodegradative activity while minimizing stripping of
volatile organics, thereby destroying the toxic compounds in the
ground. Bioventing technology is especially valuable for treating
contaminated soils in areas where structures and utilities cannot
be disturbed because the bioventing equipment (air injection/
withdrawal wells, air blower, and soil gas monitoring wells) is
relatively noninvasive.
The U.S. EPA Risk Reduction Engineering Laboratory, with
resources from the U.S. EPA Bioremediation Field Initiative,
began a 2-year field study of in situ bioventing in the summer of
1991 in collaboration with the U.S. AirForceatEielson Air Force
Base (AFB) near Fairbanks, Alaska. The site has JP-4 jet fuel-
contaminated unsaturated soil where a spill occurred in associa-
tion with a fuel distribution network. The contractor operating the
project is Battelle Laboratories, Columbus, Ohio. With the pilot-
scale experience gained in these studies and others, bioventing
should be available in the very near future as a reliable, inexpen-
sive, and unobtrusive means of treating large quantities of organi-
cally contaminated soils. The following is a report on progress
through January 1992.
Methodology
At Eielson AFB, we are studying bioventing in shallow soils in
a cold climate in conjunction with soil-warming methods to
enhance the average biodegradation rate during the year. Roughly
1 acre of soil is contaminated with JP-4 from a depth of roughly
2 ft to the water table at 6 to 7 ft. Initial (prebioventing) soil gas
measurements taken in July 1991 ranged from 600 to 40,000 ppm
total hydrocarbons, 0 to 13 percent O2, and 10 to 18 percent CO2,
indicating oxygen-limited biological activity and a high degree
of contamination. Thus, addition of oxygen as air to the site
would be expected to increase the rate of biodegradation. In
comparison, dry atmospheric air composition contains 20.9
percent O2 and 0.03 percent CO2.
31
-------
The test area was established by laying down a relatively
uniform distribution of air injection/withdrawal wells and con-
structing three 50-ft square test plots within this test area (see
Figure 1). Air is injected from 2 to 6 ft below grade at a rate of 2.5
ft3/min/well, spaced 30 ft apart. Thus, the test plots should receive
relatively uniform aeration. One plot is a control, i.e., bioventing
only (no heating). Two plots are being used to evaluate the
following two strategies of combining bioventing with warming
of the soil above ambient temperature to increase the rate of
biodegradation year-round:
• Passive warming: enhanced solar warming in late spring,
summer, and early fall using plastic covering (mulch) over
the plot and passive heat retention the remainder of the year
by applying insulation on the surface of the plot.
• Active warming: warming by applying heated water from
soaker hoses 2 ft below the surface. Water is applied at
roughly 35°C and at an overall rate to the plot of roughly 1
gal/min. Five parallel hoses 10 ft apart deliver the warm
water. The surface is covered with insulation year-round.
In addition to the network of air injection/withdrawal wells,
three-level soil gas monitoring wells at 2.5,4.5, and 6 ft below
grade, and three-level temperature probes at 2.5,4, and 5 ft below
grade were installed to provide independent measurement at
three depths throughout the site. The venting of air and the
trickling of unheated water to the actively wanned plot began in
September 1991. Warming of the water began in October 1991.
A plan view of the installation is presented in Figure 1. A cross-
section of aportion of theactively warmed plot is shown inFigure
2. An air injection well, a soil gas monitoring well, and a
temperature probe were installed in a nearby uncontaminated
area of similar geologic structure east of the site shown in Figure
1 for a background control.
Periodically, in situ respirometry tests (2) are conducted to
measure the in situ oxygen uptake rates by the microorganisms.
These tests involve temporarily (4 to 8 days) shutting off the air
and monitoring the soil gas oxygen concentration with time.
Oxygen uptake due to oxygen demands other than biological
activity is calculated by conducting a parallel shutdown test in the
background (uncontaminated) area. The rate of decrease in
oxygen concentration with time, relative to the rate observed in
the background area, indicates a relative biodegradation rate.
Thus, these tests allow estimation of the biodegradation rate as a
function of time, and therefore, as a function of ambient tempera-
ture and soil-warming technique. Quarterly comprehensive and
monthly abbreviated in situ respiration tests are planned. Final
soil hydrocarbon analyses will be conducted in summer 1993 and
o-
o
Passive
Active
Active
warming
system
Site trailer
N
•O •
rtXA Ground-water monitoring well
O Air injection/withdrawal well
^ Three-level soil gas probe
I Three-level thermocouple probe
Scale
25'
50'
Figure 1. Plan view of the joint U.S. EPA and U.S. Air Force bioventing activities at Eielson AFB, near Fairbanks, Ak.
32
-------
Ground surface
Plastic sheeting
— Styrofoam Insulation
— Plywood
Ground-water
monitoring
well (MW-3)
Air injection/withdrawal well
/ Three-level
* soil gas probe
TlaT AB r A B _ B
Valve A,,| C (S4) A ,, C (S5) A°|C (S6)
Ox
««««'':':^'^'. Gray sand and gravel ;X
fc«
-------
Table 1. Results of an In Situ Respiration Test in Early December
1991 at EielsonAFB
Oxygen uptake rate
Biodegredation rate
Plot
Actively warmed
Passively warmed
Control
Background
(% O2/hr)
0.15
0.039
0.071
0
(mg/kg-day)
2.9
0.74
1.4
0
Biodegradation rate is calculated from the oxygen uptake rate assuming the
reaction stoichiometry in Equation 1.
Conclusions
This paper summarizes the first 6 months of a 2-year joint U.S.
EPA and U.S. Air Force study of in situ bioventing. Already, the
work at Eielson AFB has shown that the active soil-warming
techniques are successful at maintaining soil at summer-like
temperatures during cold (winter) ambient temperatures. The
elevated temperatures have allowed at least a doubling of the
biodegradation rate in the heated plot relative to the other test
plots. The most efficient means of delivering the warm water to
avoid blockage of the buried hoses and the optimal water and air
flowrates that provide adequate warming and aeration continue
to be investigated. Additional data over the next 1-1/2 years are
required to fully evaluate the soil warming strategies and
bioventing at Eielson AFB. Operational costs of the soil warming
techniques will be evaluated at the completion of the study.
These bioventing studies are generating valuable pilot-scale
performance data and operational experience for a technology
that in the near future could provide a very economical means of
in situ cleanup of organically contaminated unsaturated soils. In
addition, the soil warming techniques investigated here will be
applicable to enhancing biological treatment rates of unsaturated
soils for any bioremediation technology at any location where a
significant portion of the year is too cold to allow satisfactory
biological activity.
We thank Dr. Andrea Leeson for preparing the graphical
portions of the paper.
14
8.
0)
12 -
10 -
8 -
6 -
4 -
2 -
0 -
Pressure pulse
on soaker hose
Active
Background
Control
Passive
-2
T~
9/1
"I—
10/1
1
11/1
Time (days)
12/1
1/1
Figure 3. Average temperature of each test plot and background as a function of time at EielsonAFB. Time scale is date beginning in late
August 1991 through early January 1992.
34
-------
References
1. R.E. Hoeppel, R.E. Hinchee, and M.F. Arthur. 1991.
Bioventing soils contaminated with petroleum hydro-
carbons. J. Indust. Microbiol. 8:141.
2. S.K. Ong, R.E. Hinchee, R. Hoeppel, and R. Schultz.
1991. In situ respirometry for determining aerobic deg-
radation rates. In In Situ Bioreclamation, R.E. Hinchee
and R. F. Olfenbuttel, eds. Butterworth-Heinemann,
Boston, pp. 541-545.
35
-------
-------
Optimizing Bioventing in Deep Vadose Zones and Moderate Climates:
Hill AFB Bioremediation of a JP-4 Spill
Gregory D. Sayles and Richard C. Brenner
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
Robert E. Hinchee
Battelle Laboratories, Columbus Division
Columbus, OH
and
Catherine M. Vogel
U.S. Air Force Engineering Services Center
Tyndall Air Force Base, FL
and
Ross N. Miller
U.S. Air Force Center for Environmental Excellence
Brooks Air Force Base, TX
Introduction
Bioventing is the process of supplying oxygen in situ to oxygen
deprived soil microbes by forcing air through unsaturated con-
taminated soil at low flow rates (1). Unlike soil venting or soil
vacuum extraction technologies, bioventing attempts to stimu-
late biodegradative activity while minimizing stripping of vola-
tile organics, thus destroying the toxic compounds in the ground.
Bioventing technology is especially valuable for treating con-
taminated soils in areas where structures and utilities cannot be
disturbed because bioventing equipment(air injection/withdrawal
wells, air blower, and soil gas monitoring wells) is relatively
noninvasive.
The U.S. EPA Risk Reduction Engineering Laboratory, with
resources from the U.S. EPA Bioremediation Field Initiative,
began a 2-year field study of in situ bioventing in the summer of
1991 in collaboration with the U.S. Air Force at Hill Air Force
Base (AFB) near Salt Lake City, Utah. The site has JP-4 jet fuel-
contaminated unsaturated soil where a spill has occurred in
association with a fuel distribution network. The contractor
operating the project is Battelle Laboratories, Columbus, Ohio.
With the pilot-scale experience gained in these studies and
others, bioventing should be available in the very near future as
a reliable, inexpensive, and unobtrusive means of treating large
quantities of organically contaminated soils.
The objectives of this project are to increase our understanding
of bioventing large volumes of soil and to determine the influence
of air flow rate on biodegradation and volatilization rates. The
following is a summary of progress through January 1992.
Methodology
The site is contaminated with JP-4 from depths of approxi-
mately 35 ft to perched water at roughly 95 ft. Here, bioventing,
if successful, will stimulate biodegradation of the fuel plume
under roads, underground utilities, and buildings without dis-
turbing these structures. A plan view of the installation is shown
in Figure 1. The air injection well is indicated; "CW" wells are
soil gas cluster wells where independent soil-gas samples can be
taken at 10-ft intervals from 10 to 90 ft deep, and "WW" wells are
ground-water wells. A cross-sectional view along the path AA'
in Figure 1 is shown in Figure 2. Air is currently being injected
from one well into the plume at a rate of 40 ft3/min depths from
30 to 95 ft.
37
-------
Ground-water monitoring well
CW = Soil vapor cluster well
(1.5) = TPH in ground water (mg/L)(9/91)
A-A' = Cross section trace
Figure 1 Plan view of the joint U.S. EPA and U.S. Air Force bioventing activities at HillAFB, near Salt Lake City, Utah. CWare cluster soil-gas
monitoring wells, WW are ground-water monitoring wells, and the air injection well is indicated. The path AA' indicates the location of
the cross-sectional view shown in Figure 2.
An inert gas tracer study, regular soil gas measurements at
several locations and depths, and semiannual in situ respiration
tests (2) are planned to demonstrate the effectiveness of deliver-
ing oxygen and stimulating biodegradation in a large volume of
soil of substantial depth. The in situ respirometry tests involve
temporarily (4 to 8 days) shutting the air off to the injection well
and monitoring the soil-gas oxygen concentration with time.
Oxygen uptake due to oxygen demands other than biological
activity is calculated by conducting a parallel shut-down test in
the background (uncontaminated) area. The rate of decrease in
38
-------
4790
co
4670 - bd
4790
- 4770
- 4750
— - 4730
- 4710
- 4690
- 4670
Sand with gravel and clay
Silty sand
Sand
CD Screened interval
•^ Perched water (approx. surface)
CW = Soil gas cluster well
890 = TPHinsoil(mg/kg)(9/91)
1.5 mg/L = TPH in ground water
Figure 2. Cross-sectional view at HHIAFB along the path AA' shown in Figure 1, showing the relative locations of the air injection well, soil-gas
cluster wells, ground-water wells, and some geological features of the site. Vertical axis indicates elevation above sea level in feet.
Numbers along wells indicate soil total petroleum hydrocarbon concentrations (TPH) (mg/kg) from soil samples taken at the depth
shown during drilling in September 1991.
oxygen concentration with time, relative to the rate observed in
the background area, indicates a relative biodegradation rate at
that time during the study. Thus, these tests allow estimation of
the biodegradation rate as a function of time.
The inert gas tracer study involves temporarily replacing the
injection of air with the injection of helium and observing the
transport of gas in the soils by monitoring for the inert gas at the
various soil-gas wells with time. The spatial distribution of tracer
with time will indicate the uniformity (or lack thereof) of aeration
of the site. The air injection rate will be decreased semiannually
to evaluate the tradeoff between the loss in area of influence of the
injected air for bioremediation and the decrease in volatilization
of volatile organics to the atmosphere at the soil surface.
Final soil hydrocarbon analyses will be conducted in summer
1993 and compared with the initial soil analysis to document
actual hydrocarbon loss due to bioventing.
Results
Progress to date at Hill AFB includes completion of the
installation of the wells shown in Figures 1 and 2, initial soil
sampling for total hydrocarbon levels as a function of depth, and
commencement of air injection on September 5, 1991. Total
petroleum hydrocarbon (TPH) soil analysis results also are
shown in Figure 2. The jet fuel contamination is most highly
concentrated in a 50-ft radius around the injection well.
Thus, the cleanup strategy being investigated here is (1) to
aerate the highly contaminated zone near the injection well to
stimulate biodegradation of the jet fuel in this zone, and (2) to
allow the organics that are volatilized, due to aeration of the
plume, to travel through a significant volume of soil outside the
plume to provide biodegradation before the air stream reaches the
surface.
Budgetary constraints have not allowed significant measure-
ments to date. However, the inert gas tracer study and the first in
situ respiration test and air flow rate change are scheduled for
spring 1992.
We thank Chris Perry for preparing the graphic portions of the
paper.
References
1. R.E. Hoeppel, R.E. Hinchee and M.F. Arthur, 1990.
Bioventing soils contaminated with petroleum hydro-
carbons. J. Indust. Microbiol. 8: 141.
2. S.K. Ong, R.E. Hinchee, R. Hoeppel, and R. Schultz,
1991. In situ respirometry for determining aerobic
degradation rates. In In Situ Bioreclamation; R.E.
Hinchee and R.F. Olfenbuttel, eds. Butterworth-
Heinemann, Boston, pp 541-545.
39
-------
-------
Use of Nitrate to Bioremediate a Pipeline Spill at Park City, Kansas:
Projecting From a Treatability Study to Full-Scale Remediation
Stephen R. Hutchins
U.S. Environmental Protection Agency
Robert S. Kerr Environmental Research Laboratory
Ada, OK
Microcosm tests were conducted to determine whether an
aquifer contaminated with petroleum hydrocarbons at Park City,
Kansas, could be remediated in situ under denitrifying condi-
tions, and to determine the rate and extent of nitrate consumption
and alkylbenzene biodegradation. These rates, in conjunction
with detailed core analyses from the site, were then used to
estimate nitrate demand and assess the current design of the
remediation system.
Discrete 15-cm cores were obtained over continuous lengths
from 3 to 10 m below land surface to define the vertical extent of
contamination, using methods described previously (1). These
cores were extracted and analyzed for benzene, toluene,
ethylbenzene, and xylenes (BTEX) by GC/MS and for total
petroleum by GC using JP-4 jet fuel as a reference standard.
Aquifer material was then obtained for microcosm preparation
from both contaminated and uncontaminated intervals at the site
using aseptic sampling techniques under anaerobic conditions.
Microcosms were prepared in an anaerobic glovebox using 13-
mL headspace vials and sterile diluted spring water, and were
amended with ammonium and phosphorus nutrients. Sets pre-
pared with both contaminated and uncontaminated aquifer mate-
rial received either nitrate alone, BTEX spike alone, nitrate plus
BTEX spike, or nitrate plus BTEX spike with biocides. Micro-
cosms were sealed without headspace using Teflon-lined septa
and incubated in an anaerobic glovebox at 20°C. Periodically,
three replicates were sacrificed for each set and analyzed for
aqueous BTEX, nitrate, nitrite, and nutrients.
Toluene, ethylbenzene, m-xylene, p-xylene, 1,3,5-
trimethylbenzene, and 1,2,4-trimethylbenzene were degraded to
less than 5 jig/L within 20 days in the clean-aquifer microcosms
amended with nitrate and BTEX. About half of the o-xylene was
removed, whereas benzene and 1,2,3-trimethylbenzene were
recalcitrant. There was no degradation of the compounds in the
microcosms without nitrate addition during the same time period,
relative to controls. However, nitrate consumption occurred
without BTEX addition, indicating a background nitrate demand
in the uncontaminated aquifer material. Alkylbenzene biodegra-
dation also occurred in the contaminated-aquifer microcosms,
although sorption and leaching effects mediated by the residual
petroleum hydrocarbons precluded a direct assessment of the
rates of removal. Although the measured nitrate-consumption
rate was equivalent to that observed in the clean-aquifer micro-
cosms, the actual nitrate demand was at least four times as high
in the contaminated-aquifer microcosms. Zero-order rate con-
stants were obtained for removal of the alkylbenzenes (BTEX)
and nitrate and were expressed on a dry-weight basis (Table 1).
It is important to note that this method of rate expression is
preferable, because rates based on solution analyses only, where
unrealistic water/solid ratios are employed, as is the case with
most microcosm studies, can underestimate the absolute rates by
as much as an order of magnitude. Rates expressed on a dry-
weight basis can be extrapolated easily to the natural water
content of an aquifer.
Although benzene was recalcitrant in the microcosms, previ-
ous work had shown benzene to be degraded at field scale,
presumably due to the presence of residual oxygen in the system
(1). It is expected to be degraded at Park City as well, and hence
these rates were used to calculate the penetration zone for nitrate
application and to estimate the time required for remediation.
Calculations were made for Cell #2, a 3,400-m2 area that is to
receive ground-water recharge containing 10 mg/L NO3-N at an
application rate of 680 m3/d through a network of injection wells.
Continuous cores were obtained from two locations within this
cell; the summarized data are shown in Table 2. Based on these
analyses, Cell #2 contains 1,300 kg BTEX and 58,000 kg total
petroleum hydrocarbons (TPH). The nitrate demand of the con-
taminated zone can then be calculated using stoichiometric
relationships discussed elsewhere (2), and the time required for
remediation can be estimated based on the projected application
rate. These estimates were compared to those obtained using the
intrinsic microbial rates of BTEX degradation and nitrate re-
moval, considering either the average values for the entire depth
interval or the most contaminated interval within the aquifer
profile.
Table 1. Zero-Order Rate Constants for BTEX and Nitrate-Nitrogen
Removal in Microcosms Prepared with Uncontaminated
and Contaminated Park City Aquifer Material (all units in
mg/kg dry weight/day)
Parameter
BTEX
N/trate-N
Treatment Uncontaminated
Nitrate only
Nitrate plus BTEX
Nitrate only
Nitrate plus BTEX
1.65
2.80
2.80
Contaminated
0.248
0.275
2.92
3.06
41
-------
Table 2. Summary Core Data for Locations 60S and 60H, Cell #2
Parameter
Units
60B
60H
Depth interval m BLS* 2.4-8.5
Contaminated zone(s) m BLS 2.4-6.9
Uncontaminatedzone(s) m BLS 6.9-8.5
Most contaminated interval" m BLS 6.4-6.6
2.7-8.9
6.0-7.2
2.7-6.0
7.2-8.9
6.7-6.8
Mean BTEX concentration
Mean TPH concentration
Peak BTEX concentration
Peak TPH concentration
mg/kg dry wt
mg/kg dry wt
mg/kg dry wt
mg/kg dry wt
42.6
2,360
353
10,500
24.3
686
412
13,400
* Meters below land surface
" For BTEX, not necessarily Total Petroleum Hydrocarbon (TPH).
Based on projected nitrate demand alone, 210 days would be
required to supply enough nitrate to remediate the aquifer. This
assumes that BTEX constitutes the bulk of the nitrate demand,
and that desorption and reaction rates are instantaneous relative
to the application rate. The reaction rates, at least, should be fast.
Only 13 days would be required to satisfy the average theoretical
nitrate demand if nitrate were continuously present, based on the
microbial reaction rate for nitrate removal observed in the micro-
cosm tests. Similarly, only 21 days would be required on the
average to degrade all of the BTEX, based on the observed
microbial reaction rate for BTEX biodegradation. This observa-
tion illustrates that the main factor affecting remediation time
will be the rate of application of nitrate.
This result can also be shown by calculating the penetration
zone of nitrate once the infiltrate reaches the water table. Based
on the application rate, the design nitrate concentration, and the
microbial rate of nitrate removal, the penetration zones would be
5.8 to 6.1 m for the Core 60B location and 6.3 to 6.6 m for the Core
60H location. As shown in Table 2, the actual contaminated
interval extends to below these zones in both locations, and hence
nitrate breakthrough is not expected until substantial remediation
has occurred.
Although remediation will be controlled by the nitrate applica-
tion rate, microbial reaction rates could become the important
process parameters, if the rate of remediation is controlled by
contributions from the most contaminated interval. For example,
it would take 270 days to remediate the most contaminated
interval in Cell #2, even given the fast rates of TEX removal by
the microorganisms. Of course, if this interval is small relative to
the entire volume being treated, it may not contribute signifi-
cantly to the final average BTEX concentration in the ground
water after remediation.
In summary, these microcosm tests have shown that many of
the aklylbenzenes can be degraded under denitrifying conditions
in this aquifer, although benzene may be recalcitrant unless some
oxygen is available. These data can then be used to predict the
limiting factors for remediation, and indicate that regulatory
constraints on the permissible nitrate loading will govern the
length of time required for remediation, since the in situ micro-
bial processes are sufficiently rapid not to be a limitation.
References
1. Hutchins, S.R., W.C. Downs, J.T. Wilson, G.B. Smith,
D. A. Kovacs, D.D.Fine,R.H. Douglas, and D.J. Hendrix.
1991. Effect of nitrate addition on biorestoration of
fuel-contaminated aquifer: field demonstration. Ground
Water 29:571-580.
2. Hutchins, S.R. 1991. Optimizing BTEX biodegrada-
tion under denitrifying conditions. Environ. Toxicol.
Chem. 10:1437-1448.
42
-------
Design and Treatability Study of In Situ Bioremediation of Chlorinated
Aliphatics by Methanotrophs at St. Joseph, Michigan
Lewis Semprini, Perry McCarty, Mark Dolan, Margaret Lang, Thomas McDonald,
Jaeho Bae, and Peter Kitanidis
Western Region Hazardous Substance Research Center
Stanford University
Stanford, CA
Introduction
The St. Joseph, Michigan, National Priority List industrial site
is a relatively homogeneous fine-sand aquifer contaminated with
mg/L concentrations of trichloroethylene (TCE), cis- and trans-
1,2-dichloroethylene (c-DCE, t-DCE), and vinyl chloride (VC).
Laboratory, field, and modeling studies indicate that conditions
were ideal for evaluating in situ bioremediation at this site using
methanotrophic bacteria (1). To develop an appropriate treat-
ment system design, additional site characterization is required,
alternative technologies need to be researched, and modeling
studies for alternatives need to be evaluated. These tasks are
being carried out by a team of researchers at Stanford University
in cooperation with Allied Signal Corporation, Engineering
Science, and Region 5 and the Kerr Laboratory of the U.S.
Environmental Protection Agency. This abstract presents a sum-
mary of the status of this study.
Background
Methanotrophic bacteria have the potential for cometabolically
oxidizing chlorinated aliphatic compounds to nontoxic end prod-
ucts. In the Moffett Field pilot-scale study, methanotrophs
biostimulated in a 2-m zone through oxygen and methane addi-
tion achieved the following percentages of biodegradation: TCE,
20 percent; c-DCE, 50 percent; t-DCE, 85 percent; and VC, 95
percent (2). The Saint Joseph site appeared ideal for studying the
potential of this process at full scale since the aquifer is relatively
homogeneous and the major contaminants present from the
source of contamination are c-DCE, t-DCE, and VC (1).
Results of a preliminary feasibility study of the potential for
methanotrophic treatment at the site were presented by McCarty
et al. (1). Soil microcosm studies in columns packed with St.
Joseph aquifer solids showed methanotrophs were present. Upon
stimulation through the batch addition of 3.5 mg/L of methane
and 25 mg/L of oxygen, the following extents of degradation
were achieved: VC, 95 percent; t-DCE, 80 percent; and TCE, 20
percent, at contaminant concentrations of 100 ng/L, 160|ag/L,
and 70 |ig/L, respectively. Sorption studies indicated the con-
taminants were weakly sorbed onto the aquifer solids with
retardation factors ranging from 1.2 to 2.9 for VC, to 2.7 to 5.2 for
TCE.
Simulations performed of a potential in situ treatment scheme
using laboratory-derived parameters indicated that in situ treat-
ment of VC and t-DCE compared favorably to pump-and-treat,
while the more slowly degraded c-DCE and TCE were more
rapidly removed by pump-and-treat. The feasibility study indi-
cated the methanotrophic process has potential for treating the
downgradient portion of the plume that contained VC and the
DCE isomers. Thus, the design of an in situ treatment system and
more detailed characterization of the site were initiated.
Mixing System for Methane and Oxygen Addition
One of the challenges in in situ bioremediation is the delivery
and mixing of growth substrate and nutrients, such as oxygen and
methane needed for the development of the methanotrophic
bacterial population. The mixing method being investigated uses
a recirculation well with two screens and a pump, which induces
flow through the bore and flow recirculation through the porous
formation (Figure 1). Methane and oxygen would be introduced
directly into the recirculating ground water. The method elimi-
nates pumping the contaminated ground water to the surface,
Figure 1. Recirculation system for in situ remediation at the St.
Joseph site.
43
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surface treatment, and the subsequent reinjection. Our design
work for the recirculation system has focused on the mathemati-
cal modeling of the flow, modeling of the biostimulation and
biodegradation process, and developing a mass transfer device
for dissolving methane and oxygen in the recirculating ground
water.
The flow modeling has been performed using an axisymmetric
boundary element method, which for simple cases (no regional
flow, homogeneous aquifer properties) reduces a three-dimen-
sional problem to a one-dimensional problem. This procedure
enables accurate representation of the dimensions of the well, the
free surface, and the effects of uncertainty in the hydraulic
conductivity. The results of a simulation with streamtubes from
the recirculation unit are shown in Figure 2. Much of the recircu-
lating flow occurs near the well, and the sphere of influence is
approximately equal to the thickness of the aquifer. A three-
dimensional boundary element code is being formulated for a
more detailed analysis, including the effects of regional flow,
multiple wells, and irregular-shaped boundaries.
Biostimulation and biotransformation modeling of the recircu-
lation system has been performed by adapting the code devel-
oped in the Moffett Field study (3,4) for the flow geometry
shown in Figure 2. Initial simulations indicate that an alternate
pulse cycle of 2 days oxygen and 1 day methane provides the
most favorable biological growth pattern and degradation rate.
However, there is a potential for bioclogging of the aquifer near
the well screens due to the recirculation of nutrients.
A Venturi mass-transfer device has been developed and stud-
ied in the laboratory as a means of dissolving methane and
oxygen in the recirculating ground water. The tests indicate the
device could achieve the desired methane and oxygen addition
rates at the St. Joseph site. The study showed that the presence of
other gases in the ground water (e.g., nitrogen) decrease the
efficiency of oxygen and methane addition. Thus, measurements
of trace gas composition of the ground water are required.
Contaminant Characterization Results
A detailed characterization of the ground-water contamination
was conducted at the location proposed for the in situ evaluation.
Details of this characterization are provided by Wilson et al. (5).
The distribution of the contamination is significantly different
than we had anticipated from the data previously collected from
the nearby monitoring wells. The concentrations of the chlori-
nated aliphatic compounds varied significantly with depth. Rela-
tively high concentrations (several mg/L of TCE, c-DCE, and
VC) exist at all locations within 20 m of the center of the plume.
The maximum concentrations of the contaminants were much
higher than expected (TCE, 133 mg/L; c-DCE, 128 mg/L; vinyl
chloride, 56 mg/L). Ethylene, as well as methane, is present in
relatively high concentrations. Higher ethylene concentrations
tend to be associated with higher VC concentrations, suggesting
the ethylene is a product of anaerobic VC transformation. The
aquifer is anoxic at depth with a low redox potential, which also
suggests that conditions are suitable for anaerobic methane-
forming processes. These data, discussed in detail by McCarty
and Wilson (6), strongly suggest that anaerobic microbial trans-
Z(m)
Figure 2 Streamlines calculated for the recirculation unit. Each streamtube represents 10 percent of the total flow. Half of a symmetric profile
is shown.
44
-------
formation processes have occurred in the aquifer, and not only
are responsible for the presence of c-DCE and VC, but also for the
complete dechlorination of a significant amount of TCE to form
ethylene, a relatively nonharmful chemical.
Microcosm Experiments
The high concentrations of the chlorinated organics observed
in the characterization studies might inhibit the biostimulation of
a methanotrophic population because of product toxicity shown
by Alvarez-Cohen and McCarty (7), and others. Microcosm
studies with St. Joseph aquifer solids were initiated to investigate
this possibility. VC and TCE degradation was studied at concen-
trations ranging from 0.8 mg/L to 8 mg/L. The results showed
complete methane utilization in all of the enriched columns
regardless of VC or TCE concentration. In the methane-stimu-
lated columns, approximately 25 percent and 80 percent VC
removal was observed in the columns fed 8 mg/L and 0.8 mg/L
of VC, respectively. The TCE concentrations in all of the en-
riched columns remained approximately the same as in the
control columns. Studies performed at 12°C, the aquifer tem-
perature, indicated that enrichment of a mixed methanotrophic
culture is possible and occurs within about 40 days and that the
extent of VC transformation approaches than found at room
temperature.
The results of the microcosm studies indicate that high TCE
concentrations and lower temperatures do not inhibit
methanotrophic growth, and that the extent of VC transformation
appears to be the same with or without TCE present. TCE in mg/
L concentrations, however, is not effectively transformed by the
methanotrophic process; thus, efforts should be made to locate
the proposed recirculation unit in an area where TCE concentra-
tions are low and VC is the downgradient contaminant.
Summary and Conclusions
Studies leading to the design of a system for the methanotrophic
in situ treatment of VC and DCE isomers have been initiated. A
recirculation system has been chosen for study that would elimi-
nate the need for pumping contaminated ground water to the
surface. Initial hydrodynamic and biological modeling studies
indicate that effective remediation of VC can be achieved with
the system. A method for adding methane and oxygen directly to
the recirculating ground water has been developed that could be
implemented at the site.
A detailed characterization of ground-water contamination
showed zones of high concentrations of CACs. The presence of
ethylene indicated that anaerobic transformation to a nontoxic
end product has occurred. Thus, enhanced anaerobic transforma-
tion should be explored. Microcosm studies indicated that
methanotrophs would grow in the presence of mg/L concentra-
tions of VC and TCE and would effectively degrade VC. The
results indicate that a promising strategy would be to promote
anaerobic transformation of TCE upgradient and use aerobic
methanotrophic treatment downgradient as a final polishing step.
This scenario is currently being investigated.
References
1.
2.
5.
6.
7.
McCarty, P.L. et al. 1991. In situ methanotrophic biore-
mediation for contaminated ground water at St. Joseph,
Michigan. In R.E. Hinchee and R.G. Olfenbuttel, eds.
Onsite Bioreclamation Processes for Xenobiotic and
Hydrocarbon Treatment. Butterworth-Heinemann, Bos-
ton, pp. 16-40.
Semprini, L. et al. 1991. A field evaluation of in situ
bioremediation of chlorinated ethenes. Part 2: Results
of Biostimulation and Biotransformation Experiments.
Ground Water 28(5):715-727.
Semprini, L. and P.L. McCarty. 1991. Comparison
between model simulations and field results for in situ
biorestoration of chlorinated aliphatics. Part 1:
Biostimulation of methanotrophic bacteria. Ground
Water 29(3):365-374.
Semprini, L. and P.L. McCarty. 1992. Comparison
between model simulations and field results for in situ
biorestoration of chlorinated aliphatics. Part 2:
Cometabolic transformations. Ground Water 30(1):37-
45.
Wilson et al. 1992. Characterization of contamination at
the St. Joseph, Michigan, NPL site. U.S. EPA's 1992
Symposium on Bioremediation of Hazardous Wastes:
U.S. EPA's Biosystems Technology DevelopmentPro-
gram.
McCarty, P.L. and J.L. Wilson. 1992. Natural anaerobic
treatment of a TCE plume, St. Joseph, Michigan, NPL
site. U.S. EPA's 1992 Symposium on Bioremediation
of Hazardous Wastes: U.S. EPA's Biosystems Tech-
nology Development Program.
Alvarez-Cohen and P.L. McCarty. Effects of toxicity,
aeration, and reductant supply on trichloroethylene
transformation by a mixed methanotrophic culture.
Applied and Environmental Microbiology 57(1): 228-
234.
45
-------
-------
Natural Anaerobic Treatment of a TCE Plume
St. Joseph, Michigan, NPL Site
Perry L. McCarty
Western Region Hazardous Substance Research Center
Stanford University
Stanford, CA
and
John T. Wilson
U.S. Environmental Protection Agency
Robert S. Kerr Environmental Research Laboratory
Ada, OK
Introduction
A ground-water plume containing trichloroethylene (TCE)
and other chlorinated aliphatic compounds (CACs) is present at
a National Priority List industrial site in St. Joseph, Michigan,
approximately 750 m east of Lake Michigan. Lagoons covering
about 2 hectares were used at the site between 1968 and 1976 for
disposal of industrial wastes. Between 1976 and 1978, the
lagoons had liquids removed from them, and were then backfilled
and capped. Whether the TCE in ground water resulted from
discharge into the lagoon or from spills elsewhere on the site is
not known. Surface drainage from the site is controlled by both
Lake Michigan to the west and Hickory Creek, which lies about
350 meters to the east, with a ground-water divide running north
and south through the plant site. The water table lies about 10m
below the ground surface in a fine sand aquifer that is about 20
mdeep.
TCE and other chlorinated aliphatic compounds were found in
ground-water plumes emanating from the site in 1982. Subse-
quent studies indicated that the ground-water plume is divided,
with a segment moving toward Hickory Creek and another
toward Lake Michigan. Both segments contained cis- and trans-
1,2-dichloroethylene (cDCE, tDCE), vinyl chloride (VC), and a
few other CACs in addition to TCE. The possibility of in situ
treatment of the VC portion of the plume by methanotrophic
bacteria was proposed through injection of methane and oxygen
(1). To provide additional information for the pilot study, a
detailed site characterization was undertaken at one location
through a cooperative effort with Allied-Signal Corporation,
Engineering Science, Region 5 and the Kerr Laboratory of EPA,
and Stanford University. This study indicated that the ground
water contained significant levels of ethene, a likely end product
of anaerobic TCE transformation, plus high concentrations of
methane. The extent of anaerobic transformation of TCE and the
potential for enhanced anaerobic rather than aerobic biological
transformation are being explored. A summary of information
supporting in situ reduction of TCE follows.
Background
The anaerobic reduction of TCE to DCE and VC was reported
in theearly 1980s (2,3,4). Furtherreduction of tetrachloroethylene
(PCE) to VC and ethene under methanogenic conditions was
reported recently by Freedman and Gossett (5). In subsequent
studies, DiStefano et al. (6) found that the reduction of PCE (as
high as 55 mg/L) to ethene could take place under anaerobic
conditions with methanol as the growth substrate, even in the
absence of methanogenesis. In their study, about 70percentof the
methanol was converted to acetate, while about 30 percent was
associated with the dechlorination reactions. They suggested this
efficient reductive dechlorination might be exploited for biore-
mediation of PCE-contaminated sites. Major et al. (7) provided
field evidence for the reductive transformation of low mg/L
concentrations of PCE to VC and ethene at a chemical-transfer
site in North Toronto where methanol and acetate were found as
ground-water contaminants. Transformation of PCE to ethene
was successful in methane-producing laboratory microcosms
using site aquifer material supplemented with methanol and
acetate. Since TCE is the first reduction product from PCE, these
studies support the potential for TCE transformation to the
environmentally benign ethene. Remaining VC might then be
treated by the originally proposed methanotrophic treatment.
Charecterization Results
As discussed above, reductive dechlorination of CACs re-
quires an electron donor, which may be native organic matter, or
an added organic compound such as methanol or acetate. The
capacity of the donor to supply electrons can be analytically
determined as its chemical oxygen demand (COD). Figure 1
provides contours of COD in ground water at the St. Joseph site
based upon monitoring data reported in 1986 (8). Figure 2 is a
47
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Chemical Oxygen Demand
(mg/l)
Figure 1. COD contours for the St. Joseph site (after Keck (8)).
composite of 10 mg/L contours for chlorinated aliphatic com-
pounds, also reported previously. A correlation between COD
decrease in Figure 1 and transformations of TCE to VC in Figure
2 is apparent. Also shown in Figure 1 are the locations of the three
transects taken for the detailed characterization, which consisted
of 17 boreholes with ground-water sampling from each at 5-ft
vertical intervals down to the aquitard at about 30 m.
Table 1 summarizes the electrons released by TCE reduction
to different products, along with the equivalent amount of COD
decrease required for the various transformations. COD decrease
associated with methane production is indicated as well. An
electron equivalent of reduction requires one-fourth mole mo-
lecular oxygen, or 8 g COD decrease. Since the reduction of one
mole or 131 g TCE to ethene releases six electrons, an equivalent
decrease of 48 g of COD is necessary. Table 2 summarizes the
CACs, ethene, and methane found at selected sampling loca-
tions, and the equivalent COD decrease that would be associated
with these products. The average concentration of methane found
at all sampling depths of 25 m or more was 6 mg/L, which
corresponds with an average COD conversion of 24 mg/L. Figure
1 indicates a COD decrease as high as 200 mg/L may have
occurred at the site, which is four to six times the equivalent COD
change to form products at the selected Table 2 locations. These
observations may be consistent as the amount of COD dilution
between the lagoon and the detailed characterization location is
unknown, other possible electron acceptors such as nitrate and
sulfate were not considered, the effect of sorption of chemicals on
their distribution is unknown, and some methane may have
escaped by volatilization or other processes. At the sampling
locations where the greatest degrees of dehalogenation were
found (Table 1), reductive dehalogenation itself was associated
with 8 to 36 mg/L of the total equivalent COD decrease. This too
represents a fairly high percentage of the COD reduction sug-
gested by Figure 1 contours.
Summary and Conclusions
Relatively high concentrations of TCE, DCE, VC, ethene, and
methane were found in ground water at the St. Joseph site.
Organic chemicals represented by ground-water COD appear to
have undergone biological removal, a significant portion being
associated with methane production (as high as 25 percent) and
dechlorination of TCE (as high as 18 percent). An active anaero-
bic dehalogenating population appears to be present in the
aquifer. It appears that enhanced reduction of remaining CACs to
ethene may be possible if additional organic compounds, such as
methanol or acetate, were added to the aquifer. Microcosm
studies to determine chemical requirements, reaction rates, and
transformation end products are proposed.
48
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Table 1.
Half Reactions Indicating Electron Equivalents of Change and Associated Equivalent COD Decrease Associated with
Change
Transform
product
Methane
DCE
VC
Ethene
' From [ O? + 4H* + 4e- -» 2
Mol.
wt.
16
97
62.5
28
!H O ], the half-reaction for o:
CCV
CHCI=CCI2
CHCI=CCI2-t
CHCI=CCI2
Half reaction
8H+ + 8e- -» CH4 + 2Hp
+ H* + 2e--> CHCI=CHCI -t
-c/-
-2H* + 4e--+ CH^CHCI + 2CI •
+ 3H+ + 6e~ -> CH -CH + ;
Equiv. COD
decrease
g COD/g
product1
4
0.16
0.51
1.71
References
1.
2.
McCarty, P.L. et al. 1991. In situ methanotropohic
bioremediation for contaminated ground water at St.
Joseph, Michigan. In On-SiteBioreclamation Processes
for Xenobiotic and Hydrocarbon Treatment, R.E.
Hinchee and R.G. Olfenbuttel, eds. Butterworth-
Heinemann, Boston, pp. 16-40.
Vogel, T.M. and P.L. McCarty. 1985. Biotransforma-
tion of tetrachloroethylene to trichloroethylene, di-
chloroethylene, vinyl chloride, and carbon dioxide un-
der methanogenic conditions. Appl. and Environ
Microbiol. 49:1080-1083.
Parsons, F. and G.B. Lage. 1985. Chlorinated organics
in simulated ground-water environments. Amer Water
Works Assoc. 77:52-59.
Barrio-Lage, G. et al. 1986. Sequential dehalogenation
of chlorinated ethenes. Environ. Sci. Technol. 20(1):96-
Figure 2. Chlorinated aliphatic compound 10 mg/L contours at St. Joseph, Michigan.
49
-------
5. Freedman, D.L. and J.M. Gossett. 1989. Biological
reductive dechlorination of tetrachloroethylene and
trichloroethylene to ethylene under methanogenic con-
ditions. Appl. and Environ. Microbiol. 55(9):2144-
2151.
6. DiStefano, T.D., J.M. Gossett, and S.H. Zinder. 1991.
Reductive dechlorination of high concentrations of
tetrachloroethene to ethene by an anaerobic enrichment
culture in the absence of methanogenesis. Appl. Environ.
Microbiol. 57(8):2287-2292.
7. Major, D.W., W.W. Hodgins, and BJ. Butler. 1991.
Field and laboratory evidence of in situ biotransforma-
tion of tetrachloroethene to ethene and ethane at a
chemical transfer facility in North Toronto. In On-Site
Bioreclamation, R.E. Hinchee and R.F. Olfenbuttel,
eds. Butterworth-Heinemann, Stoneham, p. 147-171.
8. Keck Consulting Services, Inc. Hydrogeological Study
Report. Section 10, Lincoln Township, Berrien County,
Michigan, June 1986.
Table 2 Concentrations of CACs, Ethene, and Methane Found at Selected Sampling Locations Along Detailed Characterization Transects, and
the Equivalent COD Decrease Associated with the Products
Sample @
1-2-70'
1-3-75'
2-1-75'
2-6-65'
3-2-80'
mg/l
Equiv. COD
Decrease, mg/l
% of Equiv. COD
mg/l
Equiv. COD
Decrease, mg/l
% of Equiv. COD
mg/l
Equiv. COD
Decrease, mg/l
% of Equiv. COD
mg/l
Equiv. COD
Decrease, mg/l
% of Equiv. COD
mg/l
Equiv. COD
Decrease, mg/l
% of Equiv. COD
TCE
4.03
0.00
0.0
12.80
0.00
0.0
0.44
0.00
0.0
0.51
0.00
0.0
2.53
0.00
0.0
1, 1DCE
0.09
0.01
0.0
0.27
0.04
0.1
0.09
0.01
0.0
0.00
0.0
0.01
0.00
0.0
cDCE
4.14
0.66
2.1
16.90
2.70
4.4
13.40
2.14
5.6
4.70
0.75
1.4
0.90
0.14
0.3
tDCE
0.81
0.13
0.4
0.67
0.11
0.2
0.21
0.03
0.1
0.03
0.00
0.0
0.03
0.00
0.0
VC
3.19
0.41
1.3
56.40
28.80
46.8
1.46
0.74
1.9
2.66
1.36
2.5
0.27
0.14
0.3
Ethene
6.62
11.32
36.6
2.25
3.84
6.2
3.15
5.39
14.2
4.27
5.98
10.9
4.87
8.32
15.1
CH4
4.61
18.44
59.5
6.62
26.00
42.2
7.43
29.72
78.1
11.72
46.88
85.3
11.59
46.36
84.4
Total
30.97
61.5
38.03
54.97
54.96
*> First value is transect number, second value is borehole number, and third value is depth of sample below ground surface in
50
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Proposed PCB Biodegradation Study at
The Sheboygan River Confined Treatment Facility
John Rogers, Jack Jones, and Eric Weber
U.S. Environmental Protection Agency
Athens, GA
and
Rebecca Adams
TAI
Athens, GA
The primary objective of the proposed field demonstration
study at the Sheboygan River Confined TreatmentFacility (CTF)
is to examine the effects of alternating aerobic and anaerobic
conditions, with and without the addition of inorganic nutrients,
on the rate and extent of PCB biotransformation under simulated
field conditions. Collecting data under conditions that closely
resemble a large-scale treatment of PCBs will provide valuable
information regarding the correlation of laboratory and field data
and the development of scenarios for enhanced bioremediation.
The selection of treatment designs for the Sheboygan River
CTF has been based, in part, on the outcome of laboratory bench-
scale studies in the context of engineering feasibility. Two
potential treatments, inoculation with acclimated organisms and
augmentation with organic amendments, have not been demon-
strated to enhance significantly the rate of dechlorination of
PCBs in bench-scale laboratory studies. Addition of inorganic
nutrients, however, is considered beneficial for enhancing micro-
bial activity. Other strategies still under consideration are those
designed to increase the physical-chemical availability of PCBs
bound to sediments, including the addition of surfactants, and the
application of aerobic/anaerobic cycling. Surfactants may in-
crease the availability of "bound" PCBs to microorganisms, but
identifying a surfactant that is nontoxic and not a preferred
substrate for microbial growth is a challenging problem. Thus,
surfactant treatment is not considered to be a viable option at this
time.
Aerobic/anaerobic cycling is currently the most promising
treatment option and merits trial at the field scale. During the
aerobic phase, the bioavailability of PCBs may be enhanced
through the degradation of naturally occurring organic com-
pounds as well as contaminants such as oil and grease, all of
which are known to bind PCBs. Further, biphenyl and lower
chlorinated PCB congeners, primarily mono- and di-chlorinated
PCBs, are degraded preferentially under aerobic conditions (1,2).
Thus, an aerobic treatment phase contributes to more complete
biodegradation. Anaerobic conditions favor the biotransforma-
tion of the more highly chlorinated PCB congeners, e.g., 6,7, and
8 chlorines, through reductive dechlorination reactions (3,4),
producing lower chlorinated products amenable to aerobic at-
tack. Taken together, the alternating aerobic/anaerobic scheme
seems promising for more extensive bioremediation of PCBs (5).
This scenario is supported by preliminary laboratory studies
indicating that aerobic/anaerobic cycling reduces the concentra-
tion of both higher and lower chlorinatedPCBs. Furthermore, we
have recent evidence of ort/w-dechlorination of lower chlori-
nated congeners in lab-scale, aerobic/anaerobic cycling experi-
ments, an observation not often cited.
The aerobic/anaerobic field experiments will be conducted for
a minimum of 1 year and perhaps for 2 years to provide sufficient
time for potentially significant/rtttfu bioremediation. Anaerobic
conditions will be established by natural means; that is, no
externally added source of oxygen will be provided to the CTF
during this phase of the treatment. Sediments will be aerated by
pumping soluble oxidants, such as peroxides or oxygen-satu-
rated water, into the CTF. Oxidant addition can be made either
directly through the existing distribution piping system or to the
CTF outflow liquid in a recycle fashion.
The effectiveness of either strategy will depend on the nature
of dispersion of the added oxidant and its consumption within the
sediments. Tracer studies have been proposed to test dispersion
from the upflow nutrient distribution piping system. Results
from that study are necessary to determine the adequacy of the
existing distribution system for oxygen addition in the aerobic
cycle. Oxidants under consideration include hydrogen peroxide
and oxygen-saturated cold water. In the event the currently
existing plumbing should be inadequate for the distribution of
oxidant, flow and distribution of the oxidant could be augmented
by sinking wells at points of low flow and pumping or withdraw-
ing interstitial fluids.
51
-------
Determination of the optimum cycle times for the system to be
operated under aerobic and anaerobic conditions will have two
components: the time required for the transition from anaerobic
to aerobic conditions and vice versa, and the time required within
each regime for the desired extent of PCB transformation. Indi-
cators of aerobic status may include direct measurement of
oxygen levels in liquid outflows and interstitial sediment fluids
or such indirect measures as chemical oxygen demand (COD) or
total soluble carbon of sediment samples. Adequate aeration
would be expected to lower the soluble organic content of
sediments. Sampling procedures will be designed to detect flow
problems such as channeling/short-circuiting.
References
1. Bedard, D.L. and ML. Habewrl. 1990. Microb. Ecol.
20:87-102.
2. Furukawa, K., N. Tomizuka, and Kamibayashi. 1979.
Appl. Environ. Microbiol. 38:301-310.
3. Brown, J.F., R.E. Wagner, D.L. Bedard, M.J. Brennan,
J.C. Carnahan, R.J. May, and T.J. Tofflemire. 1984.
Northeast Environ. Sci. 3:167-179.
4. Quensen, J.F., J.M. Tiedje, and S.A. Boyd. 1988. Sci-
ence 242:752-754.
5. Unterman,R. 1990. Anaerobic biodegradation of PCBs.
In Biological Remediation of Contaminated Sediments,
with Special Emphasis on the Great Lakes, C.T. Jafvert
and J.E. Rogers, eds. U.S. Environmental Protection
Agency, Athens, Georgia. Publication no. EPA/600/9-
91/001.
52
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Characterization of Microorganisms, Microbial Consortia, and
Microbial Processes for the Reductive Dechlorination of Hazardous Wastes
John E. Rogers and Jack Jones
U.S. Environmental Protection Agency
Athens, GA
and
Dorothy Hale and Wren Howard
Technology Applications, Inc.
Athens, GA
The purpose of this study is to determine the nutritional,
physiological, and environmental factors that enhance the activ-
ity, growth, and enrichment of microbial populations capable of
the transformation of chlorophenols. The ultimate goal is to gain
sufficient knowledge for the development and characterization
of enriched microbial consortia and pure microbial cultures
responsible for the reductive dechlorination of chlorinated aro-
matic compounds.
To date, only one pure microbial culture has been described
that is capable of the reductive dechlorination of a chlorinated
aromatic compound. This isolate (Desulfomonile tiedje) is re-
stricted to the dechlorination of meta-halogenated benzoates and
some chlorophenols (1,2,3). Few studies have been reported that
describe the physiological or nutritional conditions affecting
reductive dechlorination of chloroaromatics. Gibson and Suflita
(4) and Kuhn et al. (5) have demonstrated enhanced dehalogenation
rates of 2,4,5-T and chloroanilines, respectively, by addition of
organic carbon supplements to methanogenic aquifer slurries.
Further, Nies and Vogel (6) reported that addition of organic
substrates, such as methanol, acetone, glucose, and acetate,
accelerated the dechlorination of PCBs in river sediments. The
enhanced dechlorination rate is theorized to be related to the
supply of potential electron donors to the dechlorinating popula-
tion.
In this study, we report the effects of pH, nutrient supplements,
and heat (pasteurization) on enriched microbial cultures affect-
ing the dechlorination of mono- and di-chlorophenols (CP)
primarily under methanogenic conditions. Results will be used in
future studies to characterize microbes and microbial consortia
responsible for reductive dechlorination.
In initial studies, both unadapted and 2,4-DCP-adapted fresh-
water sediment slurries were used as model systems to assess the
effects of added nutrients on the onset and rate of reductive
dechlorination. Additions of individual substrates, including
formate, butyrate, propionate, acetate, and a vitamin mix, did not
substantially enhance dechlorination in chlorophenol-adapted
cultures or in adapted cultures with recently depleted activity.
However, addition of yeast extract enhanced dechlorination
activity as well as the onset of dechlorination in most experi-
ments. Addition of acetate also stimulated dechlorination activ-
ity in fresh, unadapted sediment slurries. However, in some
experiments, addition of organics, such as formate and propi-
onate, extended the time before the onset of dechlorination.
Studies are in progress to determine the source of additional
(potential) electron donors.
The effects of heat and pH on the dechlorination of selected
mono- and di-chlorophenols (DCP) were studied using CP-
adapted freshwater sediment slurries and CP-adapted consortia.
CP-adapted sediment cultures were enriched by the repeated
addition of the mono- or di-CP over a period of 6 months and
subsequent dilution into sterile, diluted (10 percent solids) sedi-
ment. Some cultures subsequently were transferred to filter-
sterilized site water supplemented with 0.1 percent yeast extract
and amended with the appropriate chlorophenol. For heat-treat-
ment studies, most cultures were heated to 85°C for 15 minutes
and diluted 10-fold into sterile sediment slurry (10 percent dry
weight). Dechlorination was followed over time. In many experi-
ments, the heat treatment was repeated without further dilution.
The initial heat treatment of DCP-adapted slurries did not
significantly decrease the rate or extent of dechlorination activ-
ity. A second heat treatment had a more pronounced inhibitory
effect. Dechlorination of the first addition of 2,3-DCP, for
example, was complete by day 6 in control slurries (non-heated)
and by day 9 in slurries heated for 15 and 30 minutes to 85°C.
Additional DCP added to these cultures subsequently was dechlo-
rinated within 2 days in control slurries, whereas the rate of
dechlorination was decreased by 50 percent following the second
53
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(identical) heat treatment. Cultures enriched with 2,4-DCP and
3,4-DCP were not affected as severely by the second heat
treatment compared to 2,3-DCP heat-treated cultures.
Sediment slurries adapted to dechlorinate 2-, 3-, and 4-
chlorophenol were completely inhibited by the initial heat treat-
ment. Further, nonheated cultures (controls), which were also
diluted into sterile 10 percent sediment, exhibited a long lag
period before the onset of chlorophenol dechlorination. It is
evident that physiological and environmental factors have differ-
ing effects on the various dechlorinating populations.
The effect of pH on the onset and rate of dechlorination of
dichlorophenols was examined in sterile sediment slurries (10
percent w/v) adjusted to pH 5 to 9 and inoculated with either heat-
treated, DCP-adapted sediment slurries or DCP-enriched cul-
tures. In general, the rate of dechlorination of each of the DCPs
was greatest at pH 7 over the range of pH tested (5 to 9).
Dechlorination of all DCPs tested was complete within 6 days of
incubation in experiments adjusted to pH 7. However, longer lag
times and decreased dechlorination rates were observed for
experiments incubated at the extremes of pH tested (pH 5 and 9).
These results indicate that physiological and environmental
factors are important considerations in the development of strat-
egies for remediation of hazardous wastes.
References
1. Shelton, D.R. and J.M. Tiedje. 1984. Appl. Environ.
Microbol. 48:840-848.
2. DeWeerd, K. A., L. Mandelco,R.S. Tanner, C.R. Woese,
and J.M. Suflita. 1990. Arch. Microbiol. 154:23-30.
3. Mohn, W.W. and K.J. Kennedy. 1992. Appl. Environ.
Microbiol. 58:1367-1370.
4. Gibson, S.A. and J.M. Suflita. 1990. Appl. Environ.
Microbiol. 56:1825-1832.
5. Kuhn, E.P., G.T. Townsend, and J.M. Suflita. 1990.
Appl. Environ. Microbiol. 56:2630-2637.
6. Nies,L.andT.M.Vogel. 1990. Appl. Environ. Microbiol.
56:2612-2617.
54
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Effects of Metals on the Reductive Dechlorination
of Chlorophenols
Jack Jones
U.S. Environmental Protection Agency
Athens, GA
and
In Chul Kong
Technology Applications, Jjic.
Athens, GA
Recent interest in the use of bioremediation technology for the
cleanup of contaminated soils and sediments has led to a greater
understanding of the fate of a variety of toxic organic compounds
in natural environments. Most of the research has focused on the
biological transformation of a single organic contaminant in
laboratory microcosms using naturally occurring microbial in-
ocula. However, most polluted ecosystems and hazardous waste
sites are most often contaminated with a mixture of organic
compounds and toxic inorganic wastes such as metals. The fate
and transport of organic compounds in natural environments
such as soils and sediments has received considerable interest in
recent years, but little is known about the potential toxicity of
metals to naturally occurring microorganisms and microbial
processes of importance to the biotransformation of organic
compounds.
Soils and sediments are quite varied in composition but gener-
ally consist of an array of mineral particles, organic matter,
microbial cells and debris, and inorganic solutes, all of which are
important in the formation of metal complexes. Further, recent
studies have indicated the importance of microbial cell surfaces
in the sorption and immobilization of metals in natural ecosys-
tems.
The purpose of this study is to investigate the effects of toxic
heavy metals on anaerobic microbial processes affecting the
biotransformation of organic compounds. As a model system, we
are studying metal toxicity on the reductive dechlorination of
chloroaromatic compounds in unadapted and substrate-adapted
microbial communities from freshwater sediments.
The onset, rate, and extent of biotransformation of several
mono-, di-, and trichlorophenols wereexamined using unadapted
and chlorophenol(CP)-adapted freshwater sediment slurries (pH
7.0) in the presence and absence of added metal salts (CuCl2,
CdCl2, K2Cr2O7). The time required for unadapted control sedi2-
ment slurries (no added metals) to dechlorinate the respective
chlorophenol (10 mg/L) was as follows: 2,4-DCP and 2,4,6-TCP
(16 to 17 days); 2,3-DCP (21 days); 2-CP (31 days); 2,4^5-TCP
(37 days); 3-CP (80 days). Addition of a20ppm concentration of
the chloride salts of Cu(II) and Cd(II) had little or no effect on the
onset, rate, or extent of dechlorination for most chlorophenols
tested. Addition of Cr(VI) at 20 to 40 ppm, however, increased
the lag time before initiation of dechlorination for most
chlorophenols tested. Addition of lOOppm (or greater) Cr(VI), as
K2Cr2°7' caused total inhibition of dechlorination of 2- and 3-CP,
2,3-DCP, and 2,4,5- and 2,4,6-TCP. Addition of 100 ppm Cr(VI)
increased the lag time of 2,4-DCP dechlorination from 13 days
(control) to 40 days. Higher concentrations of Cd(II) (100 to 200
ppm) also caused complete inhibition of dechlorination for all
chlorophenols with the exception of 2,4-DCP. Therate of dechlo-
rination of 2,4-DCP was reduced by 70 percent and 100 percent
at 100 and200ppmCd(II) concentrations, respectively. Addition
of CuCl2 at 100 ppm increased the lag time before the onset of
dechlorination for most chlorophenols tested, but dechlorination
of 2,4-DCP was still evident at 200 ppm Cu(II). The distribution
of the test metals between the aqueous and complexed (sediment)
phase of the experimental samples was determined by induc-
tively coupled plasma (ICP) spectrometry. In most experiments,
the aqueous phase concentrations of the added metals were low
and varied from 0 to 2 ppm for Cu(H), 0 to 30 ppm for Cd(II), and
0 to 28 ppm for Cr(VI) over the range of metal salts added (20 to
200 ppm).
The effects of metal salts (20 to 200 ppm of Cu(II), Cd(II)) on
the reductive dechlorination of 3,4-DCP were compared in two
distinct freshwater sediments of similar pH but of differing total
organic carbon content. Control experiments (no added metals)
for both sediment types differed slightly with regard to the onset
of dechlorination but rates of dechlorination were similar. How-
ever, sediment cultures containing the higher organic carbon
content (34 mg C/g) were, in general, more resistant to the
55
-------
inhibitory effects of the added metal salts. The onset and rate of
dechlorination of 3,4-DCP were only slightly affected in the
higher organic carbon sediment cultures amended with Cu and
Cd salts at total metal concentrations up to 150 ppm. Dechlorina-
tion in sediment cultures with lower organic carbon content (17
mg C/g) was completely inhibited (100 days incubation) in
cultures amended with a total metal (Cu or Cd) concentration of
50 ppm. Further, longer lag times (40 days compared to 14 days)
were observed in those cultures with the lower organic carbon
content at a lower (20 ppm) added metal concentration. Results
of recent experiments using 3,4-DCP-adapted sediment cultures
of low organic carbon content indicated that dechlorination was
more resistant to the inhibitory effects of the added metal salts
than those reported above for the unadapted sediment cultures.
Additional studies are in progress to determine the inhibitory
nature of metals on reductive dechlorination of higher chlori-
nated chloroaromatics. These preliminary data suggest that the
metal type, aqueous metal concentration, and organic carbon
content may affect the transformation of chlorophenols in anoxic
sediments.
56
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Section Three
Performance Evaluation
Performance evaluation of bioremediation technologies involves determining the extent and rate of cleanup by a
ue? SS±T fT n ^ * "* "^ ^^^ ^ and effectS °f the P3™1 comP°^ -SS by-
products. Remediation efforts at a contaminated site can produce intermediate compounds that can themselves be
detemine *e risk °f '
exDonPA1™11!6'1 *" T00™0"31 f11"^ (NCPB> TO aPOtentiaUy important source of hazardous
exposure. In an EPA project, researchers documented the fact that NCPs limit the ability of scientists to perform risk
characterizations through analytical chemistry and are developing bioassays from which to determiS Zn"S
geno ox,c,tyat remediation sites. TwootherprojectsevaluatedtheperformanceofbioventingsystemstoTtemSeir
effectiveness and to monitor possible health risks from surface emissions of hazardous compounds l°aetermuietlieir
fourth study concerned potential health effects associated with microorganisms genetically engineered for
*^^
57
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-------
Development of Comparative Genotoxicity Risk Methods for
Evaluating Alternative Hazardous Waste Control Technologies
Larry D. Claxton and S. Elizabeth George
U.S. Environmental Protection Agency
Health Effects Research Laboratory
Research Triangle Park, NC
Most environmental remediation situations present mixtures
of toxicants and other pollutants (1). Remediation efforts,
especially biological ones, produce additional compounds that
add to the complexity of evaluating the potential health effects of
a contaminated site during and after remediation efforts. Expo-
sures to toxicants from most hazardous waste sites through soil,
water, and air are typically low-level chronic exposures. There-
fore, one of the lexicological endpoints of greatest interest is
genotoxicity, which includes both carcinogenicity and mutage-
nicity endpoints. Due to the presence of multiple pollutants and
the potential production of other pollutants by remediation pro-
cesses, most of the actual toxicants within a typical remediation
site are not identified. The Committee on Environmental Epide-
miology of the National Research Council states, 'There is
evidence that NCPs [nonconventional pollutants] are a poten-
tially important source of hazardous exposure. Some prelimi-
nary toxicologic studies suggest that NCPs have important bio-
logic properties, environmental persistence, and mobility... In
the broadest sense, these unidentified substances represent risk of
unknown magnitude" (1).
NCPs, therefore, limit the ability to do risk characterizations of
remediation sites when only analytical chemistry is used for
exposure assessment studies. By incorporating biological tests
into assessment studies, it is possible to improve the estimations
of potential human toxicity before, during, and after remediation
efforts. When appropriately coupled with analytical chemistry,
bioassays can also be used to identify the major toxic pollutants.
In addition, during the development of remediation methods,
bioassays can be used for comparative assessments between
differing technological approaches. This presentation will dis-
cuss the use of bioassays for comparative assessments of com-
plex mixtures, for identifying individual toxins within complex
mixtures, and for enhancing exposure assessment.
In bioremediation research, the ideal situation is to avoid the
production of additional toxicants and to demonstrate the safety
of the expected products—even products for which analytical
chemistry monitoring is not done. For determining whether or
not bioremediation is an appropriate alternative for aremediation
effort prior to its application, two approaches are available. The
first approach is to test individual metabolites. Although this
approach appears straightforward, it may take extended periods
of time to elucidate the actual metabolic pathways and to test each
metabolite produced. An alternative approach for examining
health risks is to examine products produced in the culture
medium when the biodegrading organism(s) is grown on the
carbon source of interest. Complete knowledge of the degrada-
tion pathway is unnecessary for this evaluation. This method also
lessens the likelihood that post-treatment analysis of aremediated
site will identify new toxicants that were not a part of the original
monitoring scheme.
Recent efforts with 2,4,5-trichlorophenoxyacetic acid (2,4,5-
T) illustrate the application of these two approaches (2). 2,4,5-
T is well known because it is one of the principal components of
the defoliant Agent Orange (4). The herbicide 2,4,5-T is a
pollutant in the United States; it was used for weed control from
the 1940s until banned in 1969, after a study indicated that it is
teratogenic (3). Because 2,4,5-T is persistent in the environment,
it remains an environmental problem. Some soil microorgan-
isms, however, have been shown to degrade chlorinated
phenoxyherbicides (4,5). One study reported that, aerobically,
2,4,5-T is metabolized to 2,4,5-trichlorophenol (2,4,5-TCP), 2,5-
dichlorohydroquinone (DCHQ), and 5-chloro-2-hydroxy-
hydroquinone (CHHQ) before the arene ring is cleaved (6).
Another study indicated that the principal 2,4,5-T intermediates
are 3,5-dichlorocatechol and 2-chlorosuccinate (CS) (7). A
strain provided by Doctors A.M. Chakrabarty and U. Sangodkar,
Pseudomonas cepacia AC 1100, has the ability to degrade 2,4,5-
T. Dr. Sangodkar indicates that AC 1100 metabolizes 2,4,5-T
through the intermediate CHHQ and that catechol is a by-product
of the metabolism. This information is more than is typically
available for the degradation products of a pollutant.
To test the utility of these two approaches, two genotoxicity
bioassays (the Salmonella plate incorporation mutagenicity as-
say (8) and the prophage-induction bioassay (9)) were used to
detect mutagens and potential carcinogens. Initially, 2,4,5-T and
its commercially available metabolites were tested with both
assays. When tested up to 10 mg perplate using S. typhimurium
strains TA98, TA100, TA102, and TA104 in the presence and
absence of an exogenous metabolic activation system, none of
the compounds was mutagenic. Two of the compounds (2,4,5-
T and 2,4,5-TCP) were positive in the prophage-induction assay
when the metabolic activation system was present. DCHQ was
considered a weak positive in this assay when the metabolic
activation system was absent. These results indicate that one of
59
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the proposed pathways produced genotoxic metabolites and that
one of these metabolites (2,4,5-TCP) is more biologically active
than the parent compound. These results, in turn, bring into
question whether or not to apply this bacterial strain in remediation,
and they might make all bioremediation efforts suspect. The
question remains, do toxic metabolites persist?
In the second portion of the study, P. cepacia strain AC 1100
was used to degrade 2,4,5-T as described by Kilbaneetal. (4). At
selected times over a 60-hour period, aliquots were withdrawn,
cell growth was monitored, and the level of 2,4,5-T was mea-
sured. In addition, aliquots from the different time intervals were
bioassayed using both methods. In no case did the Salmonella
strains TA100 and TA98 indicate any mutagenicity. This would
indicate that mutagenic hydrocarbon metabolites were not pro-
duced. When the prophage-induction bioassay was used in the
presence of metabolic activation, the growth of AC 1100 was
inversely related to the genotoxicity of the 2,4,5-T concentration.
The mutagenicity of the uninoculated 2,4,5-T culture did not
decline over the same time period. This indicates that the 2,4,5-
T was metabolized by AC 1100 and that mutagenic metabolites,
if produced, did not accumulate within the media.
The ideal situation for evaluating a biodegradation process
prior to use would be to determine the toxicity associated with a
particular process by (1) understanding the metabolic pathways
of the process and the toxicity of each metabolite, and (2)
examining the toxicity of a process during a bench-scale opera-
tion. This is not always possible. The estimation of toxicity
through the use of bioassays, however, provides a powerful tool
for the development of bioremediation approaches that protect
public health.
References
1. National Research Council, Committee on Environ-
mental Epidemiology. 1991. Environmental Epidemi-
ology: Volume 1, Public Health and Hazardous Wastes.
National Academy Press, Washington, DC, pp. 282.
2.
6.
7.
8.
9.
George, S.E., D.A. Whitehouse, and L.D. Claxton.
1992. Genotoxicity of 2,4,5-trichlorophenoxyacetic
acid biodegradation products in the Salmonella rever-
sion and Lambda prophage-induction bioassays. Envi-
ronmental Toxicology and Chemistry 11:733-740.
Grant, W.F. 1979. The genotoxic effects of 2,4,5-T.
Mutation Research 65:83-119.
Kilbane, J.J., D.K. Chatterjee, J.S. Karns, S.T. Kellogg,
andA.M.Chakrabarty. 1982. Biodegradation of 2,4,5-
trichlorophenoxyacetic Acid by a pure culture of
Pseudomonas cepacia. Applied and Environmental
Microbiology 44:72-78.
Pfarl, C., G. Ditzelmuller, M. Loidl, and F. Strichsbier.
1990. Microbial degradation of xenobiotic compounds
in soil columns. FEMS Microbiol. Ecol. 73:255-262.
Haugland, R.A., D.J. Schlemm, R.P. Lyon III, P.R.
Sferra, and A.M. Chakrabarty. 1990. Anaerobic bio-
degradation of 2,4,5-trichlorophenoxyacetic acid in
samples from a methanogenic aquifer: Stimulation by
short-chain organic acidsandalcohols. AppliedEnviron.
Microbiol. 56:1357-1362.
Rosenberg, A. and M. Alexander. 1980. Microbial
metabolism of 2,4,5-trichlorophenoxyacetic acid in soil,
soil suspensions, and axenic culture. J. Agric. Food
Chem. 28: 297-302.
Maron,D. and B.N.Ames. 1983. Revised methods for
the Salmonella mutagenicity test. Mutat.Res. 113:173-
212.
Elespuru, R.K. 1984. Induction of bacteriophage
Lambda by DNA-interacting chemicals. In Chemical
Mutagens: Principles and Methods for Their Detection,
F.J. de Serres, ed., Vol. 9. Plenum, New York, New
York, pp. 213-231.
60
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Performance of Bioventing at Traverse City, Michigan
Don H. Kampbell and John T. Wilson
U.S. Environmental Protection Agency
Ada, OK
and
Christopher J. Griffin
The Traverse Group, Inc. (TGI)
Traverse City, MI
Two pilot-scale bioventing systems in adjacent plots were
evaluated during a 9-month operating period at an aviation
gasoline spill site. The unsaturated subsurface was bioremediated
to less than 100 mg TPH/kg soil within 5 months. Benzene was
less than 5 ppb in the underlying ground water when bioventing
was concluded after 9 months of operation. Minimal surface
emissions were indicative of a biological cleansing process near
the soil surface. Considerable oily-phase residue remained below
the water table. Biosparging by air injection below the depth of
contamination in the saturated subsurface presently is being
evaluated.
Objective
The objective of the project was to design, install, operate, and
evaluate two pilot-scale bioventing systems. Performance of the
two systems included demonstrating that surface emissions were
minimal, total fuel hydrocarbons in the remediated core material
would be less than 100 mg/kg, final benzene in the underlying
ground water would not exceed 5 ng/L, remediation would be
completed in a reasonable time, and the technique would be
applicable to full-scale reclamation.
Treatability Studies
Prior to design of the pilot-scale bioventing systems, labora-
tory soil microcosm studies were conducted using surface soil
from the spill site. Microcosms have been used to simulate
biodegradation of volatile gasoline components as reported by
Kampbell (2).
Aerobic microcosms of 160 mL glass serum bottles and 32 g
dry-weight basis Rubicon sand were used to determine biodeg-
radation of fuel vapors under variable parameters of moisture,
nutrients, vapor concentrations, and temperature. Degradation
under favorable conditions was rapid and complete, showing a
curve typical of first-order kinetics. An NPK nutrient response
was obtained. Differential responses occurred for temperature
ranges from 4°C to 37°C and moisture levels of 3.5 percent to 24
percent, but none greatly hindered the biodegradation process.
The rates indicated that at a vapor exposure concentration of 14.5
mg aviation gasoline/kg soil, consumption would be complete
during 8 hours of retention in the subsurface.
Introduction
Bioventing is an in situ bioremediation process that provides
an air flow to vaporize and transport volatile organic pollutants
upward from the subsurface to more amenable media for miner-
alization. The air also provides oxygen for microbial degradation
processes. Naturally occurring soil microbes, once acclimated,
can biodegrade the volatiles, such as fuel hydrocarbon contami-
nants.
An aviation gasoline spill of about 35,000 gallons 20 years
earlier had an 80 x 360 m surface area plume in 1989. Much of
the spill has remained as oily globules in subsurface capillary
pores near the ground-water table. Water table changes have
resulted in a vertical contaminant smear of nearly 1 m. The water
table depth has been near 5 m. The vertical profile was relatively
uniform beach sand from the surface to thick clay at 15 m.
Detailed descriptions of the surface and subsurface at the field
site are presented in Ostendorf et al. (1) and Twenter et al. (3).
Experimental Design
Turf was established on a 75 ft x 90 ft rectangular area
overlying the plume of contamination. A nutrient solution was
applied for dispersion throughout the unsaturated subsurface.
Two blowers in a building shown in Figure 1 were connected to
aeration transfer piping and to screened air-injection wells with
adjustable depth to emit air flow just above the water table. The
north plot venting system was injection only, and the south plot
was injection, extraction, and reinjection. The venting systems
began operating in October 1990. A blower rate of 5 cpm to each
plot was calculated to average an injected-air retention time of 24
hours.
61
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-^ —
V" " "
\
\ DG
\ +
\
\
\
\
\
\
\
\
\
\
\
\
V
o
South Plot
DF
_± B
I
I / ' '
9/\J>\ \
\ & A/
\y\ *{\ }
*'\ *\ ''
\»4t-0/^->
North Plot
-------
Table 1. Soil Gas Hydrocarbons (mg/L)
Date
North Plot
One Meter Three Meters
South Plot
One Meter
10/22/90
10/29/90
1 1/12/90
01/10/91
04/04/91'
04/18/91
05/02/91
08/08/91
10/07/91
1,940
860
430
180
400
310
78
32
16
2,060
910
490
200
990
820
52
40
16
4,220
1,960
760
200
600
480
66
20
14
2,960
1,210
550
260
1,580
1,160
38
20
14
Table 2. Fuel Carbon Mass in North Plot (mg/ft')
09/25/90
Rep1
Rep 2
Avove water table
Below water table
Avove water table
Below water table
43,640
221,630
Rep1
253
147,980
54,020
232,800
10/16/91
Rep 2
172
234,390
48,830
227,220
Rep 3
480
150,360
X
302
177,580
Table 3. Fuel Carbon Mass in South Plot and South Control (mg/ft *)
South Plot
Rep1
Rep 2
Avove water table
Below water table
37
105,335
27
62,502
32
83,920
South Control
Rep1
Avove water table
Below water table
12,090
54,616
4,510
103,910
8,300
79,260
causative factors for the reductions were possible. Adsorption of
air with oxygen by the ground water from overlying venting may
have stimulated vaporization and biodegradation of the water-
phase BTEX. Gasoline globules trapped in capillary pores re-
main as a liable pool for potential release of BTEX.
Discussion
Both bioventing systems performed satisfactorily to remediate
the unsaturated test area subsurface contamination of aviation
gasoline. Surface emissions, once acclimation was attained 1
month after system startup, did not exceed 10ngTPH/L and were
less than 1 pg TPH/L in most instances. Ground-water levels
were essentially the same in October 1990 and October 1991,
although variations near 0.4 m occurred during the bioventing
operating period. Considerable oily-phase residue remained be-
low theOctober 1991 water-table level.Finalbenzeneconcentra-
tions were less than 5 ng/L compared to initial levels up to 133
Hg/L in the underlying ground water near the water table. Large
63
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reductions in gasoline hydrocarbons between 1 m-depth soil gas
measurements and surface emissions indicated an active rhizo-
sphere cleansing action near the soil's surface.
Costs
Frequent statements in the literature state that bioremediation
is emerging as a cost-effective alternative to treat subsurface
contamination. It does reduce capital costs by eliminating trans-
portation and disposal of recovered contaminants. In situ biore-
mediation may be less costly, but the biological process can be
difficult to control and, in many instances, slow. The cost of soil
venting by a field-scale system has been reported to be roughly
$50 per ton (U.S. EPA Report 540/A5-89/003, 1989), while
incineration costs are more than 10 times this amount. A cost
estimate of about $ 15 per yd3 sandy soil for bioventing treatment
ataJP-4jetfuel contaminated site has been reported by Vogel (4).
Total costs incurred for the pilot-scale bioventing systems over
the 6-month period for construction, operation, field monitoring,
and sample analyses were near $147,000. Expansion of the
injection well grid from 3,000 ft2 to 60,000 ft2 for a field-scale
system would increase construction costs about $69,400. Less
extensive field monitoring and sample analyses could be insti-
gated to offset the increase in construction cost. Maintaining total
costs of $147,000 for the field-scale bioventing system with 2 ft
of vertical oily-phase residue above the water table would result
in a clean-up cost estimation of $33 per yd3 of contaminated
subsurface.
A grid of air injection biosparging wells was installed in the
north plot during the latter part of October 1991 to a depth of 3 m
below the water table. Evaluation of biosparging effectiveness
will be completed in May 1992.
References
1. Ostendorf, D.W., D.H. Kampbell, J.T. Wilson, and J.H.
Sammons. 1989. Mobilization of aviation gasoline from
a residual source. Research Journal WPCF, Vol. 61, pp.
1684-1690.
2. Kampbell, D.H. and J.T. Wilson. 1991. Bioventing to
treat fuel spills from underground storage tanks. Jour.
Haz. Materials 28, pp. 75-80.
3. Twenter, F.R., T.R. Cummings, and N.G. Grannemann.
1985. Ground-water contamination in East Bay Town-
ship, Michigan. U.S. Geological Survey Water-Re-
sources InvestigationsReport85-4064,Lansing,Michi-
gan.
4. Vogel, C. 1991. Enhanced in situ biodegradation of
petroleum hydrocarbons through soil venting. Techdata
RDV 91 -7, Air ForceE&S Center, Tyndall AFB, Florida.
Table 4. Ground- Water BTEX
Sample
Well depth
Benzene (mg/L)
Total BTEX (mg/L)
January 1991
MW2N
MW2N
MW2S
MW2S
Total
MW2N
MW2N
MW2S
MW2S
Total
MW2N
MW2N
MW2S
MW2S
Total
17
21
18
20
17
21
18
20
17
21
18
20
<5
<5
<5
43
<43
April 1991
<5
174
<5
7
<181
October 1991
<5
<5
<5
<5
<0
1,018
440
1,028
<2,495
149
492
39
444
< 1,1 24
51
159
19
31
<260
64
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Bioventing of a Gasoline Spill at TVaverse City, Michigan:
Practical Engineering Considerations
Christopher J. Griffin and John M. Armstrong
The Traverse Group, Inc. (TGI)
Traverse City, MI
The Bioventing Reclamation Pilot Study is designed to evalu-
ate the biodegradation of hydrocarbon-contaminated vapors within
the unsaturated zone during induced volatilization. This study is
being conducted at the U.S. Coast Guard Air Station in Traverse
City, Michigan, the site of a spill of about 35,000 gallons of
aviation gasoline that occurred in 1969. After 20 years, a major
portion of the spill still persists in the subsurface as a residual
plume that is about 1,100 ft long and 250 ft wide. This study is a
cooperative effort between the U.S. Coast Guard and the U.S.
Environmental Protection Agency's Robert S. Kerr Ground
Water Research Laboratory.
The subsurface conditions at the site consist of a uniform beach
sand extending to depths of about 50 ft, underlain by a gray,
glacial silly clay. The water table is located at a nominal depth of
about 15 ft below the ground surface. Over the past 6 years, the
water table elevation has fluctuated 6 to 8 ft.
The 90- x 75-ft study area has been divided into two equal areas
of 45 x 75 ft to evaluate the effects of different flows and
extraction patterns. A conceptual design is presented in Figure 1.
The northern area has an injection system, while the southern
area has an injection and extractionAeinjection system. The
South Plot
North Plot
Extraction well with tip
of well screen located
at a depth of approx.
16 ft below grade
Reinjection well with tip
of well screen located
at a depth of approx.
10 ft below grade
Pumps
Pump
Fresh air injection well with
tip of well screen located at
a depth of approx. 16 ft
below grade
. \
Fresh air injection well with tip -*
of well screen located at a depth
of approx. 15 ft below grade
Figure 1. Conceptual design.
65
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pneumatic properties of the unsaturated zone evaluated by the
performance of a pneumatic pump test resulted in a design radius
of influence of 10 ft. Ambient air is injected into both areas at an
initial rate that replaces the volume of calculated air-filled pore
space in 24 hours. The flow rate will be increased to a vapor
recharge rate of 2 hours or higher as the system becomes
acclimated. Based on the treatment area size, depth to water, and
air-filled porosity, the recharge rates correspond to flows ranging
from 5 cubic feet per minute (cfm) for the 24-hour recharge to 63
cfm for the 2-hour recharge.
The blower system, therefore, has to be capable of extracting
vapors in the south study area from the water table (depths of 15
to 18 ft, at the corresponding flow rates; ranging from 5 to 63 cfm)
and then reinjecting the vapors at the same rate at a depth of 10
ft. In addition, the system has to be able to inject ambient air at the
same flow rate within both the extraction/reinjection plot (south
area) and the air injection plot (north area). Accordingly, because
the injected ambient air will be placed in twice the area (both test
plots), the blower has to be able to inject air at flow rates ranging
from 10 cfm to 128 cfm.
As previously mentioned, the project is divided into two
different treatment areas. Construction consisted of installing 15
aeration injection points placed on 10-ft centers in the north area.
These wells were placed in a 3- x 5-grid pattern and screened just
above the water table. In the south area, eight sets of injection
points coupled with seven extraction points, 10 ft on centers,
were installed with screens placed just above the water table.
Eight reinjection wells were installed with the screen placed at a
depth of 10 ft.
Root 45 vacuum blowers with amaximum flow rate of 130 cfm
were used to inject or extract and reinject the contaminated
vapors. These blowers extract at vacuum pressures of 4 to 6 in. of
mercury and inject (or reinject) at pressures of 6 pounds per
square in. (psi). The extraction/reinjection blower is driven by a
10 HP 3-phase electric motor, while the injection blower is driven
by a 7-1/2 HP 3-phase electric motor. All of the equipment is
explosion-proof.
The monitoring requirements of the EPA Bioventing Work
Plan called for the installation of several different types and
depths of monitoring equipment and/or sampling points. To
monitor hydrocarbon vapor and oxygen concentrations, three 5-
point cluster wells were installed per plot. The cluster wells
consisted of 1/4-in. diameter copper tubing with a wire mesh
screen covering the tip. The five sampling points of each cluster
well were installed at 3.28-ft (1 m) depth increments throughout
the unsaturated zone. In addition, we installed three 14-point
cluster monitoring wells (well screens at 1.5-ft intervals from
ground surface to 21 ft—one per plot and one at an upgradient
location), and one set of moisture temperature probes per plot.
The monitoring wells are of the same construction as the cluster
wells. These wells were installed for water-quality sampling
below the water table and relative-humidity monitoring above
the water table. The moisture temperature probes are Soil Test
Series 300 moisture temperature cells consisting of thermistor
soil cells buried at depths of 5,10, and 15 ft below grade.
The development of a sufficient microbial population to de-
grade the hydrocarbon vapors requires adequate quantities of
nitrogen, phosphorus, and potassium. In accordance with the
EPA Work Plan, 64 pounds of nitrogen, 13 pounds of phospho-
rus, and 5 pounds of potassium were applied to each area prior to
startup. Also, during the growing season, 10 pounds of nitrogen,
2 pounds of phosphorus, and 1 pound of potassium are applied to
each area monthly. These nutrients were added as an aqueous
solution by sprinklers.
The Bioventing Project is sampled and/or monitored daily,
twice a week, and monthly. Daily monitoring consists of measur-
ing the blower's operating parameters, such as flow rate, pres-
sure, and vapor temperature. Combustible gas concentration
within the vapor reinjection line is determined daily using a
Bacharach ThresholdLimit Value (TLV) combustible gas meter.
Twice-a-week monitoring includes determining the combus-
tible gas and oxygen concentration within the three 5-point
cluster wells located in each plot. The combustible gas concen-
tration is determined using the TLV meter, while the oxygen
concentration is determined using a Bacharach oxygen meter.
The soil moisture content and temperature are measured twice a
week using the soil test moisture temperature probes.
Surface emissions are sampled twice a week at two locations
within each study area and at two upgradient locations weekly.
To measure surface emissions, a 19-in. diameter stainless steel
bowl, which has a volume of 4.3 gallons (16 liters), is inverted
and placed flush on the ground as a vapor collection chamber.
The sample is pulled from the chamber through teflon tubing that
is attached to the bowl by a 1/4-in. diameter steel ball valve
tapped into the top of the chamber. Any water that collects within
the emission chamber is removed by a water trap, consisting of
a flask containing a drying agent (Drierite), located upstream
from the sample trap. Samples are drawn over a 4-hour period,
using an Ismatec peristaltic pump set for a flow of approximately
1 L/hour.
Water quality samples analyzed for nutrients and BTEX are
obtained monthly by sampling at two depths in each of the 14-
point monitoring wells.
66
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Clearance and Pulmonary Inflammatory Response After
Intranasal Exposure of C3h/HeJ Mice to Biotechnology Agents
S. Elizabeth George, M.J. Kohan, and Larry D. Claxton
U.S. Environmental Protection Agency
Health Effects Research Laboratory
Research Triangle Park, NC
and
M.S. Taylor and H.G. Brooks
Environmental Health Research and Testing, Inc.
Research Triangle Park, NC
and
M.I. Gilmore
University of North Carolina
Chapel Hill, NC
Environmental application of engineered microorganisms has
initiated research into health effects potentially associated with
these organisms. Pulmonary exposure to biotechnology agents
may occur during their production or application. In this study,
survival of the dosed microorganisms and pulmonary inflamma-
tory response were explored in 30-day-old male C3h/HeJ mice
exposed intranasally (i.n.) to each agent. Mice were administered
10s colony-forming units of Pseudomonas maltophila BC6 or P.
aeruginosa strains BC16, BC17, BC18, or AC869. Strains BC6,
BC16, BC17, and BC18 were isolated from a commercial prod-
uct designed for environmental application and strain AC869
was engineered for the ability to degrade 3,5-dichlorobenzate.
Two clinical isolates, P. aeruginosa, strains PA01 and DG1,
were included as positive controls. None of the isolates were
detected in the lungs 14 days following treatment. However,
strains BC6 and BC16 were recovered from nasal washings. The
two clinical isolates and strain BC 16 induced a strong inflamma-
tory response that was evident throughout the duration of the
experiment (14 days). With the exception of strain BC17, all of
the isolates were recovered from the intestinal tract 14 days
following treatment. Translocation to the liver, spleen, and/or
mesenteric lymph nodes was observed, depending on the dosed
strain.
67
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Section Four
Process Research
Process research involves isolating and identifying microorganisms that carry out biodegradation processes and
developing new biosystems for treatment of en vironmental pollutants. Although the majority of conference presentations
dealt with laboratory-scale work, it is clear that both pilot-scale and field research are becoming increasingly important.
Pilot-scale research provides critical information on process operation and control and residuals/emissions management.
Field research is essential for evaluating the performance of full-scale bioremediation processes and for accelerated
testing on technologies that lend themselves to field testing.
In one project, researchers studied the ability to stimulate an indigenous aerobic microbial population in a shallow
aquifer to enhance its bioremediation characteristics. Another paper described three experimental and theoretical studies
on biodegrading volatile organic compounds (VOCs) in aerobic biofilters. Two additional studies focused on design
criteria for biofilter treatment of VOCs.
Researchers at a field project evaluated the effectiveness of a fungal treatment to deplete pentachlorophenol in the soil
at a former wood treating plant. Two groups studied the effectiveness of anaerobic bacteria to remediate river and lake
sediments contaminated with PCBs. River sediments were also used to investigate the effects of four reducing conditions
on the biodegradability of halogenated aromatic compounds.
The subject of another research presentation was the unique metabolic capabilities of Pseudomonas to remove
trichloroethylene from contaminated air and ground water using gas-phase bioreactors. Another study covered the
biodegradation of paper-milling effluents by anaerobic microorganisms. A Florida wood preserving plant was the site
for a study of bioremediation of contaminated ground water by indigenous microorganisms. One presentation discussed
the chemical interactions in microbial biofilms used in water quality control.
Researchers observed bioremediation of soils and sediments contaminated with aromatic amines, important environ-
mental contaminants that can have toxic effects on microbes as well as on animals. EPA presented areport on developing
small-scale evaluation techniques for bioremediation of soils by fungi. A new project has as its objective the development
of sequential anaerobic-aerobic technologies for biotreatment of soils or sediments contaminated with highly chlorinated
aromatic compounds. Researchers also evaluated the effects of surfactant action on microbial degradation.
69
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Evaluation of Enhanced In Situ Aerobic Biodegradation of CIS- and
Trans-l-TVichloroethylene and CIS- and Trans-l,2-Dichloroethylene by
Phenol-Utilizing Bacteria
Gary D. Hopkins, Lewis Semprini, and Perry L. McCarty
Western Region Hazardous Substance Research Center
Stanford University
Stanford, CA
Introduction
Recent research has demonstrated that aerobic microorgan-
isms grown on phenol or toluene can initiate the cometabolic
oxidation of chlorinated aliphatic compounds (CACs) to stable
nontoxic end products (1,2,3,4). Such microorganisms possess-
ing toluene oxygenase (TO) (2,3,4) have good potential for
bioremediating aquifers contaminated with CACs and their
anaerobic and abiotic transformation products. In this study the
ability to stimulate an indigenous phenol-utilizing population in
a shallow aquifer that can degrade cis- and trans-1,2-
dichloroethylene (c-DCE, t-DCE) and trichloroethylene (TCE)
was evaluated. In addition, microcosm studies were performed to
evaluate the ability of organisms with other oxygenase systems
to degrade CACs. Growth substrates and the assumed enzyme
systems studied include: phenol (TO); toluene (TO); methane
(methane monooxygenase, MMO) (5); and ammonia (ammonia
monooxygenase, AMO) (6).
Background
Previous evaluations at the Moffett test site of enhanced in situ
biodegradation of CACs focused on stimulating indigenous
methanotrophic bacteria through methane and oxygen addition
(7). In these tests the following percentage transformations were
achieved in a 2-m biostimulated zone: TCE, 20 percent; c-DCE,
50 percent; t-DCE, 90 percent; and vinyl chloride, 95 percent.
TCE and c-DCE degradation were limited, indicating other
oxygenase systems should be evaluated. The objective of this
study was to evaluate the TO system for in situ biodegradation of
1.0'
0.8 —
o
^ 0.6 —
re
O)
a>
Q
o 0.4 —
Phenol
Toluene
Methane
Ammonia
Figure 1. Relative fractions of CACs degraded in the ground-water microcosms following stimulations with different primary growth substrates.
71
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TCE, c-DCE, and t-DCE at the Moffett test site by phenol and
oxygen addition. Performing the evaluation at the same site
permits a direct comparison of the TO system with the MMO
system previously studied
Microcosm Studies
Microcosm studies were performed at the field site with
columns fed oxygenated ground water from the test zone. The 1-
m, 3-in. O.D. columns were filled with 5/8-in. Flexrings as a
biological support medium. The columns were inoculated with
microbes in ground water from the test zone. Five columns were
batch fed twice weekly. Each was fed a different substrate to
induce different enzyme systems: methane, MMO; phenol, TO;
toluene, TO; ammonia, AMO, and a control-fed dissolved oxy-
gen only. The ammonia addition was to stimulate nitrifiers that
possess AMO. The substrates were added at the following
concentrations: phenol, 11 mg/L; toluene, 8.5 mg/L; methane,
6.7 mg/L; and ammonia, 7.1 mg/L.
The DO consumption as a function of time indicated microbes
growing on phenol and toluene were most rapidly stimulated,
followed by methane oxidizers, and then ammonia oxidizers. DO
consumption in each column was approximately 20 mg/L, con-
sistent with the amount of substrate added. After 50 days of
primary substrate addition, a mixture of TCE, c-DCE, and t-DCE
at concentrations of 40 (xg/L, 20 |xg/L, and 20 \ig/L, respectively,
were added, while continuing to feed the primary substrates and
DO. After several exchanges pseudo-steady-state decreases in
concentrations of the CACs were observed in the stimulated
columns compared to the control. Figure 1 shows the estimates
of the percentage degradation observed, correcting for losses in
the control column. The columns fed phenol or toluene were most
effective in degrading TCE and c-DCE, while the methane-fed
column was more effective in degrading t-DCE. The ammonia-
fed column was the least effective.
The relative extent of transformation in the methane column—
(t-DCE > c-DCE > TCE) — is consistent with our earlier in situ
studies with methane-utilizing bacteria. These results indicate
that the ground-water-fed columns do mimic the in situ condi-
tions. Thus we anticipated that the in situ tests with phenol would
result in more effective transformation of TCE and c-DCE than
was observed previously with methane.
Results of In Situ Biostimulation with Phenol
The in situ tests were performed along SSE, a new experimen-
tal leg that was offset approximately 1 m from the SI leg used in
previous experiments with methanotrophic bacteria. The same
stimulus-response experimental methodology was used as in our
previous studies. The stimulus was the injection of ground water
amended with the chemicals of interest and the response was the
chemical concentration history at the monitoring locations. Ex-
periments were performed under the induced gradient conditions
created by the injection and extraction of ground water.
Bromide tracer tests indicated the new test zone had good
hydraulic characteristics, as the injected fluid completely perme-
ated the regions around the observation wells. The fluid transport
times were rapid, ranging from 6 hours to the SSE1 observation
well (1 m from the injection well), to 24 hours to the SSE3
observation well (4 m from the injection well).
The test zone was not contaminated with the CACs of interest.
TCE, c-DCE, and t-DCE were added to the reinjected ground
water at concentrations of 30 |.ig/L, 40 ng/L, and 40 ^g/L,
respectively, prior to biostimulation. The transport of the CACs
was retarded compared to the bromide tracer, due to sorption onto
the aquifer solids. c-DCE was the least retarded, while TCE was
the most, which is consistent with our previous observations. The
steady-state concentrations indicated essentially no transforma-
tion in the presence of dissolved oxygen alone.
Active biostimulation was initiated through phenol addition
after steady-state contaminant concentrations were achieved.
Phenol was pulse injected for 1 hour in an 8-hour pulse cycle at
a concentration of 50 mg/L (6 mg/L time averaged). The phenol
injection concentration was doubled after 520 hours, and then
was reduced to the original concentration after 840 hours.
The responses of DO, TCE, and c-DCE at the SSE2 well, 2 m
from the injection well, are shown in Figure 2. The biostimulation
with phenol is indicated by the DO decreases, which were small
during the periods of low phenol addition, but increased after
higher phenol concentrations were added. Decreases in c-DCE
and TCE concentrations were associated with decreases in DO,
indicating cometabolic transformations resulted from
biostimulation. Significant degradation of c-DCE and TCE were
observed, with c-DCE being more rapidly degraded than TCE.
The c-DCE concentration decreased by approximately 60 to 70
percent and TCE by 20 to 30 percent during the period of low
phenol addition. Doubling the phenol injection concentration
resulted in a greater transformation of both TCE and c-DCE, with
85 to 90 percent, and over 90 percent, transformed, respectively.
Least transformed of the three compounds studied was t-DCE
(not shown), consistent with the microcosm studies. Decreasing
the amount of phenol increased the TCE concentration, indicat-
ing that the extent of transformation was related to the amount of
phenol added.
The field results also showed evidence for competitive inhibi-
tion of cometabolic transformation by phenol. When phenol
concentrations were high due to its pulse addition, the concentra-
tions of c-DCE and t-DCE, which were not pulsed, also increased
in concentration. These results are consistent with those observed
for the methanotrophic study and indicate that competitive
inhibition affects the cometabolic transformation rates.
Summary
An indigenous phenol-utilizing population that effectively
degraded TCE and c-DCE was easily stimulated in situ. The
phenol-utilizing population degraded up to 90 percent of the TCE
in a 2-m biostimulated zone, compared to 20 to 30 percent
observed previously with methane-utilizing bacteria. The phe-
nol-utilizing population more effectively degraded c-DCE, but
was less effective in degrading t-DCE.
Microcosm studies performed under conditions similar to the
field tests agreed qualitatively with the in situ tests. The results
are promising, indicating that microcosm studies are of use in
evaluating the potential for cometabolic in situ treatment at
contaminated sites. Future studies at the site will explore TCE
concentration effects on degradation efficiency, and will deter-
mine whether a compound more environmentally acceptable
72
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1.2
1200
Figure 2. Response of DO, TCE, and c-DCE at the SSE2 well during in situ biostimulation with phenol at Moffett Field.
than phenol or toluene can be used to induce an indigenous TO 4.
population that effectively degrades TCE.
References
1. Nelson, M.J.K., S.O. Montgomery, E.J. O'Neill, and
P.H. Pritchard. 1986. Aerobic metabolism of trichloro- 5.
ethyleneby abacterial isolate. Appl. Environ. Microbiol.
52:383-384.
2. Nelson, M.J.K., S.O. Montgomery, W.R. Mahaffey,
and P.H. Pritchard. 1987. Biodegradation of trichloro-
ethyleneandinvolvementofanaromaticbiodegradative 6.
pathway. Appl. Environ. Microbiol. 53:949-954.
3. Nelson, M J.K., S.O. Montgomery, and P.H. Pritchard.
1988. Trichloroethylene metabolism by microorgan- 7.
isms that degrade aromatic compounds. Appl. Environ.
Microbiol. 54: 604-606.
Wackett, L.P. and D.T. Gibson. 1988. Degradation of
trichloroethylene by toluene dioxygenase in whole-cell
studies with Pseudomonas putida Fl. Appl. Environ.
Microbiol. 54:1703-1708.
Fox, E.G., J.G. Bourneman, L.P. Wackett, and J.D.
Lipscomb. 1990. Haloalkene oxidation by the soluble
methane monooxygenase from Methylosinus
trichosporium OB3b: mechanistic and environmental
implications. Biochemistry 29:6419-6427.
Arciero, D., T. Vannelli, M. Logan, and A.B. Hooper.
1989. Degradation of trichloroethylene by the ammo-
nia-oxidizing.
Semprini, L., P.V. Roberts, G.D. Hopkins, and P.L.
McCarty. 1990. A field evaluation of in situ biodegra-
dation of chlorinated ethenes: Part 2, Results of
biostimulation and biotransformation experiments.
Ground Water 28:715-727.
73
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Fundamental Studies on the Treatment of VOCs in a Biofllter
R. Govind, V. Utgikar, Y. Shan, W. Zhao
Department of Chemical Engineering
University of Cincinnati
Cincinnati, OH
and
D.F. Bishop, S.I. Safferman, G.D. Sayles
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
The Superfund Amendments andReauthorization Act (SARA)
emission summary for petroleum and chemical manufacturing
companies reveals that the largest environmental releases of
chemicals are volatile organic compounds (VOCs) into air.
Current technologies for treating VOCs include activated carbon
adsorption, wet scrubbing, and incineration. An emerging tech-
nology for treatment of VOCs involves the use of biofilters. This
paper describes three experimental and theoretical studies on the
biodegradation of toluene, methylene chloride, Irichloroethylene
(TCE), ethylbenzene, and chlorobenzene conducted in aerobic
biofilters.
The first study on three compounds describes the performance
of a pelletized, activated carbon biofilter (1). Complete degrada-
tion of toluene, methylene chloride, and TCE was demonstrated
at the bench-scale at 2 minutes gas retention time. After about 180
days of operation the biofilter flooded due to plugging by
biomass accumulation. Thereafter, the carbon pellets had to be
cleaned mechanically.
The second study on the above five compounds describes the
performance of a bench-scale biofllter packed with ceramic
pellets (celite, Manville Corporation, California). Complete deg-
radation of the five compounds was obtained in the experiments.
The TCE concentration was reduced by about 35 percent.
The third study on the five compounds describes the perfor-
mance of a biofllter constructed from corrugated plates of celite.
The straight corrugations form straight passages, which, unlike
the tortuous packed-bed system, offer an unobstructed means for
biomass release from the biofllter. The performance of the
straight passages system exhibited self-regulation of biomass
and provided essentially complete degradation of four com-
pounds. The TCE concentration was reduced by about 40 per-
cent.
Two models for a biofllter system were developed; the first
model quantitates biofilter performance at low biomass loadings
(biofilm regime), and the second at high biomass loadings (plug
flow regime) of biofilter operation. These models were applied to
the experimental data that had been obtained from the above three
studies. Insights into biofllter design and operation gained from
this analysis are presented. Preliminary design procedures for
biofilters and cost comparison with other technologies are in-
cluded in the presentation.
References
1. Govind, R., V. Utgikar, Y. Shan, W. Zhao, D.F. Bishop,
and S.I. Safferman. 1992. Studies on aerobic degrada-
tion of volatile organic compounds (VOCs) in an acti-
vated carbon packed-bed biofilter. Paper submitted to
ES&T.
75
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Fungal TVeatment of Pentachlorophenol
John A. Glaser
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
Richard T. Lamar, Mark W. Davis, and Diane M. Dietrich
U.S. Department of Agriculture
Forest Products Laboratory
Madison, WI
The investigation of fungal treatment systems has been mainly
confined to laboratory or bench-scale studies. A systematic series
of investigations (1,2,3) has permitted us to consider and engage
in a treatment effectiveness study under field conditions at an
abandoned wood-treating site in Mississippi. The study site is
located in Brookhaven, Mississippi, 60 miles south of Jackson,
and was identified as a removal action site for EPA Region 4.'
Ownership of the operation changed hands several times be-
tween 1980 and 1991, when the owners filed for bankruptcy. The
wood-treating operation was based on both creosote and pen-
tachlprophenol technology. As a result of a treatment chemical
spill in the facility's operations, surrounding soil was contami-
nated to some depth. The soil was excavated and mounded above
the ground surface in a RCRA hazardous waste treatment unit.
The field study was undertaken as a SITE Program demonstra-
tion involving two phases. The first phase was designed to
evaluate the ability of three different fungal species to deplete
pentachlorophenol in soil. Three inoculum loadings were used
with the required controls. The experimental design consisted of
a randomized complete block (RGB) without replication and a
balanced incomplete block (BIB) with treatment replicated four
times (Figure 1). Treatments (Table 1) were applied to screened
25 cm/deep soil in 3.05-m x 3.05-m plots. Six of the plots were
allocated to the RGB design, and the four plots for the BIB design
were subdivided into four 1.53- m x 1.53-m subplots.
After inoculation with fungi, as outlined by the design, each
plot was irrigated and tilled with agarden rototiller. The tiller was
cleaned as it was moved throughout the plots to prevent cross
contamination between treatments and controls. Woodchips were
added to the soil plots, in accordance with the study design, to
provide a substrate to sustain growth of the fungi. Soil moisture
was monitored daily throughout the study and maintained at a
BIB Study
RGB Study
G
7
J
9
H
2
1
10
Rx5
W
10
z
9
X
2
Y
8
Figure 1. Treatment plot arrangement.
77
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prescribed level. Both ambient and plot temperatures were re-
corded daily throughout the study. Weekly plot tilling was
scheduled for the duration of the study. A time series analysis of
treatment performance was accomplished by sampling the plots
before application of the treatments, immediately after treatment
application, and then after 1,2,4, and 8 weeks of operation.
The study was conducted over a2-month span from September
18 through November 13,1991. The analysis of treatment data
has identified the P. sordida treatment as the most effective to
biotransform pentachlorophenol (Figure 2). This fungal treat-
ment has been selected as the treatment for the SITE demonstra-
tion in 1992 when the period of operation is designed to be 5
months.
Table 1. Experimental Design for Fungal Treatment Effectiveness
Fungus/Control
References
1. Lamar, R.T., J. A. Glaser, and T.K. Kirk. 1990. Fate of
pentachlorophenol (PCP) in sterile soils inoculated
with white-rot basidiomycete Phanerochaete
chrysosporlum: Mineralization, volatilization, and
depletion of PCP. Soil Biol. Biochem. 22:433-440.
2. Lamar, R.T. and D. Dietrich. 1990. In situ depletion of
pentachlorophenol from contaminated soil by
Phanerochaete spp. App. Environ. Microbiol. 56:3093-
3100.
3. Lamar, R.T., M.J. Larsen, and T.K. Kirk. 1990. Sensi-
tivity to and degradation of pentachlorophenol by
Phanerochaete spp. App. Environ. Microbiol. 56:3519-
3526.
Inoculum Loading
Plots
1 P. chrysosporium
3 P. sordida
4 P. chrysosporium/
T. hirsuta
g Inoculum carrier control
e No treatment control
7 T. hirsuta
8 P. chrysosporium
9
•JQ Wood chip control
5%
10%
10%
5%
5%
10%
-
10%
13%
10% (Day O)
3% (Day 14)
-
B
C,H,K,V,X
E
F
A
D
G,L,P,S
N.R.T.Y
J,Q,U,Z
I,M,O,W
1000
800
600 -
400
200 -
No treatment control
10
Phanerochaete sordida
20 30
Time (days)
40
50
Figure 2. Pentachlorophenol depletion by Phanerochaete sordida.
78
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Anaerobic Degradation of Highly Chlorinated Dibenzo-p-Dioxins
and Dibenzofurans
P. Adriaens and D. Grbic-Galic
Department of Civil Engineering
Stanford University
Stanford, CA
Methanogenic Hudson River sediments contaminated with
100 mg/kg "weathered" Aroclor® 1242 were incubated anaerobi-
cally and spiked with 144 ± 14 ug/kg of the following polychlo-
nnated dioxin (PCDD) and dibenzofuran (PCDF) congeners-
1,2,3,4,7,8-H^CDD; 1,2,4,6,8,9/1,2,4,6,7,9-H CDo!
1,2,3,4,6,7,9-HCDD; 1,2,4,6,8-PentoCDF; and 1,2,3,4,6,7,8-
HpCDF. Previously, a hexachlorinated congener was identified
as a metabolite in HpCDD incubations. No further PCDD/F
metabolites were detected in final extracts of Hudson River
microcosms, yet the methanogenic population exhibited dechlo-
rinating activity on Aroclor® 1242, suggesting a concentration or
bioavailability threshold below which noreductive dehalogenation
occurs or at rates not measurable within the time of this project.
Background
Reductivedehalogenationofpolychlorinatedbiphenyls(PCBs)
with a high degree of chlorine substitution has been demonstrated
to occur in previously contaminated Hudson River sediments (1)
and in methanogenic microcosms containing pristine or adapted
sediments, spiked with mg/kg concentrations of either Aroclor®
mixtures (2,3) or individual PCS congeners (3,4). The PCBs
presumably serve as an alternative "electron sink" to CO for
methanogenic bacteria. Because of the structural similarity2 be-
tween dioxins or furans and PCBs, the potential for susceptibility
of the PCDD/F to reductive dehalogenation processes was inves-
tigated.
Substrate disappearance plots expressed as C/C of HexaCD-
D-, HexaCDDi-, and HeptaCDD-spiked microcosms indicated
that biotransformation at least contributed to substrate disappear-
ance in the live replicates, relative to the killed or chemical
controls (Figure 1, A-C). Previously, it was reported that after 2
months, approximately 9 ug/kg of hexa-chlorinated dioxin was
formedmHudsonRiver sediment-inoculated microcosms spiked
with HeptaCDD (5). The concentration of the intermediate
decreased to 6 ug/kg after 9 months (Figure 1, D), but no lesser
chlorinated congeners resulting from further dehalogenation
were detected (5). In aparallel series of experiments, partitioning
the PCDD/F in microcosms inoculated with aquifer material was
shown to influence the interpretation of long-term incubation
results, as 60 to 90 percent of the substrate was found to be
associated with the settled sediment, which was only a minor
fraction of the solids sampled at any given time.
In this paper we describe dechlorinating activity on
Aroclor®1242 and partitioning PCDD/F in Hudson River sedi-
ments during long-term incubations.
Results
The analyses of Hudson River extracts spiked with PCDD/F
for lower chlorinated congeners (penta- and tetrachlorinated)
were impeded by interferences in the PCDD/F window by the
higher chlorinated PCB congeners of Aroclor® 1242. Most of
these interferences can be eliminated by GC/MS analyses of M+
(M++2), and (M+-2) isotopic ratios (6). No products were conclu-
sively identified, however, in part because of the low intensities
obtained relative to background values. The recovery of PCDD/
F in Hudson River sediment-inoculated microcosms, as exempli-
fied for the dioxins, is given in Figure 2. All values are normal-
ized with respect to octachloronaphthalene recovery (52 to 65
percent) after 24-hour Soxhlet extraction with hexane:acetone
(1:1). The low PCDD/F recovery in both live and killed incuba-
tions follows the same pattern observed in the C/Co plots
indicating sorption to the high organic carbon river sediments.'
The lack of further dehalogenation of PCDD/F was not due to
inactivity of the methanogenic populations present. The distribu-
tion of PCB homologs in an Aroclor® 1242 standard in the
"weathered" Aroclor® 1242 present in Hudson River sediments
(time 0), and in the active microcosms after 16 months, is shown
in Figure 3. Since a mixture of fatty acids was added to the
microcosms to augment the indigenous methanogenic popula-
tions, the already dechlorinated Aroclor® 1242 (B) was further
dehalogenated, resulting in the accumulation of di- and
tnchlorobiphenyls at the expense of tetra- and pentachlorinated
congeners (C).
Conclusions
None of the PCDD/F congeners other than H CDD showed
accumulation of lesser chlorinated isomers, although all de-
creased over time relative to the killed controls. This may be
explained by the presence of "weathered" Aroclor® 1242 as an
alternative electron acceptor. Aroclor® 1242 was present at much
higherconcentrations(100mg/kg)andwas further dehalogenated
by the methanogenic population. Assuming the same popula-
tions are responsible for dechlorinating activity on PCDD/F and
PCBs, a threshold value might be invoked below which no
79
-------
1.1
0.9 r
0.8-
0.7-
0.6
live
killed control
chemical control
T~
10
T
T~
20
T
T"
30
40
live
killed control
chemical control
150
live
killed control
chemical control
live
killed control
chemical control
HexaCDD formation
respectively.
^^^
dehalogenation occurs. Moreover, the low extraction efficien-
cies obtained for all PCDD/F after extended incubation in a
relatively high organic carbon inoculum (1.5 percent) may indi-
cate that the (higher) chlorinated congeners are strongly sorbed
and rendered biologically unavailable. To test this hypothesis,
some microcosms were spiked with 2 mgAg PCDD/F, but data
are not available at the time of this report.
References
1 Brown, J.F., Jr., D.L. Bedard, M.J. Brennan, J.C.
Carnahan, H. Feng, and R.E. Wagner. 1987. Polychlo-
rinated biphenyl dechlorination in aquatic sediments.
Science 236:709-712.
2. Quensen, J.F. III, S.A. Boyd , and J.M. Tiedje. 1990.
Dechlorination of four commercial polychlorinated bi-
phenyl mixtures (Aroclorsf by anaerobic microorgan-
isms from sediments. Appl. Environ. Microbiol.
56:2360-2369.
3. Alder, A.C., M. Haggblom and L.Y. Young. 1990.
Dechlorination of PCBs in sediments under sulfate-
reducing and methanogenic conditions. Abstr. Annu.
Meet. Am. Soc. Microbiol., Anaheim, California, Q
47:296.
4. Nies, L. and T.M. Vogel. 1990. Effect of organic
substrates on dechlorination of Aroclor® 1242 in anaero-
bic sediments. Appl. Environ. Microbiol. 56:2612-2617.
5. Adriaens, P. and D. Grbic-Galic. 1991. Evidence for
reductive dehalogenation of highly chlorinated dioxins
and dibenzofurans. Abstr. Dioxin '91 Conference,
Chapel Hill, North Carolina.
6. Alford-Stevens, A.L., J.W. Eichelberger, T.A. Bellar,
and W.L. Budde. 1986. Determination of chlorinated
diben/o-p-dioxins and dibenzofurans in environmental
samples using high resolution mass spectrometry. EPA
Report: Physical and Chemical Methods Branch.
80
-------
200
.8 100 •
c
o
O
• Added
0 Active
E3 Autoclaved
H7CDD
H6CDD
H6CDDi
Figure 2.
PCDD/F added to and recovered from final Hudson River
extracts after a 16-month incubation period.
Figure 3. PCB homolog profile of an Aroclor ° 1242 standard (A)
and Hudson River sediment-inoculated microcosms at
time 0 (B) and after 16 months (C).
81
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-------
Bacterial Degradation of Trichloroethylene
In a Gas-Phase Bioreactor
Malcolm S. Shields and Rhonda Schaubhut
University of West Florida
Gulf Breeze, FL
and
Michael Reagin
U.S. Environmental Protection Agency
Technical Resources Inc.
Gulf Breeze, FL
Introduction
The application of the constitutive trichloroethylene (TCE)-
degrading bacterium, Pseudomonas cepacia G4 5223 Phel in
bioreactors designed for the removal of TCE from contaminated
air or ground water was investigated. To exploit the unique
metabolic capabilities of this bacterium, two bioreactor designs
were used; both employed a fixed biofilm associated with the
surface of an inert support material. Various materials were
assessed for their abilities to support a TCE-degrading popula-
tion of Phel. One reactor received air-entrained TCE along with
a continuous inflow of nutrient medium. Strain G4 (the non-
constitutive parent of G4 5223 Phel) was unable to affect TCE
concentration in such a vapor phase bioreactor (100 percent re-
feed), whereas Phel (the constitutive TCE degrader) completely
removed the 80 urn TCE present. Tests of strain Phel in a
continuous flow vapor phase reactor resulted in an average of
92.1 percent removal of TCE at an average input concentration
of 12.6 urn (0 percent re-feed) over a 72-hour test.
Results and Discussion
Ground water contamination by TCE and related chlorinated
compounds (dichloroethylenes and vinyl chloride) is a subject of
overwhelming concern. Currently, treatment of TCE-contami-
nated soil and water relies primarily on pump-and-treat systems
in which TCE is distilled away from the water under vacuum or
is air stripped and transferred onto an adsorbent such as charcoal
In either event, the result is simply the transfer of the pollutant to
another location. The capacity to destroy the contaminant at the
site via bioremediation could provide significant environmental
and economic benefits.
All TCE-degrading, non-recombinant bacterial systems char-
acterized to date, require chemical induction. Because of this co-
metabolic requirement, the !bioremediation potential of these
bacteria is limited to situations in which a constant input of
mducer can be maintained. The use of such organisms in a
bioreactor requires a sensitive balancing of the inducer chemical
growth substrate (if different from the inducer), TCE, and oxy-
gen. The simplifications inherent to a bioreactor colonized by a
constitutive TCE degrader are obvious. P. cepacia strain G4
5223 Phel is an excellent candidate for use in bioremediation
systems in view of its ability to mineralize TCE constitutively
(2,3,4).
Because P. cepacia G4 5223 Phel does not use TCE as a
growth substrate, it was necessary to consider bioreactor designs
that would allow continual nutritional supplementation of the
biofilm. Consequently, six different physical support materials
were tested for the ability to maintain an active population of
Phel. Several colonizable surfaces were found to serve in this
role. One bioreactor achieved colonization of the support mate-
rial to > 1.4 x 108 bacteria gram-'.
The toxicityofTCEandits breakdown products in thebioreactor
were also analyzed. Tests with Phel indicated a toxicity due to
the metabolism of TCE at concentrations above 500 urn (similar
to the metabolite toxicities reported for isoprene oxidizing bac-
teria (1)). &
Due to the inactivity of the wild-type strain toward TCE in the
absence of an aromatic inducer, p. cepacia G4 was an ideal
negative control for the bioreactor studies. Air-entrained TCE
was introduced to columns containing an axenic biofilm of either
Pseudomonas cepacia G4 5223 Phel or Pseudomonas cepacia
04 with 100 percent re-feed, and no supplementation of oxygen
or nutrients (Figure 1). Gas chromatographic analysis of the
column atmospheres indicated TCE breakthrough occurred at
approximately 0.3 hours in both columns. While the column
containing G4 attained a concentration of approximately 50 urn
(approximately 6,000 ppb) and maintained this concentration
throughout the 20.3-hour test period, the G4 5223 Phel-contain-
mg column never achieved a return TCE concentration above 13
Jim (approximately 1,700 ppb) at the initial breakthrough and
declined to undetectable levels by 20.3 hours. In addition an
83
-------
Recirculating Vapor Phase Biofilter
Unidentified GC Peak
o
70
60-
50-
40-
30-
20-
10-
0
uM TCE G4
uM TCE Phe1
20000-
10000-
Q.
C
0-
-10000
1 2
Hours
1 2
Hours
Figure 1. TCE concentration in the atmospheres of two independent relating gas-phase biofilters inoculated with Pseudomonas cepacia 34
or G4 5223 Phe1.
unidentified transient compound, which we interpreted to be a
metabolite, was produced only in the G4 5223 Phe 1 column. It is
represented in Figure 1 as the area under a cryptic peak occurring
at a retention time of 5.6 minutes.
The second bioreactor tested was a continuous flow, gas-phase
biofilter that allows constant addition of nutrients and air-en-
trained TCE to the biofilm. Nutrients were pumped to the top of
the column where they flowed down through the inert support.
Spent medium was pumped from the bottom of the column to a
waste bottle, where it was sampled for residual TCE. Air and TCE
vapor were pumped into the column at the bottom and allowed to
Continuous Flow Fixed Film Bioreactor
LU
exit at the top. Input and output ports were sampled to assess TCE
concentrations before and after exposure to the biofilm. Input and
output concentrations over a 72-hour trial are shown in Figure 2.
Air-entrained TCE was fed at 3.25 mL/min, through a column
packed with 231 cm3 of crushed oyster shell colonized by P.
cepacia G4 5223 Phel. Continuous input of medium into the
column was maintained at 0.37 mL/min. The average TCE input
and output concentrations over this time were 12.6 and 1.0 ^m,
respectively, yielding an overall TCE removal of 92.1 percent
with 0 percent re-feed. The overall efficiency during the test
period was found to correspond to 0.51 mg TCE/hour/kg support
material (at 2.5 g/cm3).
Clearly, P. cepacia G4 5223 Phel growing in a thin aqueous
biofilm is capable of degrading air-entrained TCE. These results
indicate that the potential exists for coupling bioremediation with
methods that generate TCE-air mixtures at contaminated sites.
Such a coupling of physical and biological processes may greatly
enhance current efforts to reduce the amount of TCE in contami-
nated aquifers.
References
1 Ewers J., D. Freier-Schroder, and H.J. Knackmuss.
1990. Selection of trichloroethene(TCE) degrading bac-
teria that resist inactivation by TCE. Arch. Microbiol.
154:410-413.
2 Shields, M.S., and M.J. Reagin. Construction of a
Pseudomonas cepacia strain constitutive for the degra-
dation of trichloroethylene and its evaluation for field
and bioreactor conditions. Abstract, 91 st General Meet-
ing of the American Society for Microbiology, Dallas,
Texas.
Figure 2. TCE degradation within a continuous feed vapor-phase
bioreactor.
84
-------
3. Shields, M.S., S.O. Montgomery, S.M. Cuskey PJ
Chapman, and P.H. Pritchard. 1991. Mutants of
Pseudomonas cepacia strain G4 defective in catabo-
lism of aromatic compounds and trichloroethylene
Appl. Environ. Microbiol. 57:1935-1941.
4. Shields, M.S., S.O. Montgomery, PJ. Chapman S M
Cuskey, and P.H. Pritchard. 1989. Novel pathway of
toluene catabolism in the trichloroethylene-degrading
bacterium G4. Appl. Environ. Microbiol. 55:1624-1629.
85
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-------
Anaerobic Bidegradation of 5-Chlorovanillate as a Model Substrate for the
Bioremediation of Paper-Milling Waste
Barbara R. Sharak Genthner, Gary Mundfrom
Technical Resources, Inc.
. Gulf Breeze, FL
and
Richard Devereux
U.S. Environmental Protection Agency
Environmental Research Laboratory
Gulf Breeze, FL
5-chlorovanillate (5CV) was chosen as a model compound for
studying the biodegradation of paper-milling effluents because it
is readily available and it contains the methoxyl-, chloro- and
carboxyl side groups that are present on the aromatic chlorinated
compounds in paper-milling effluent. Using this approach an
anaerobic enrichment was developed that degraded 5CV 3-
chlorocatechol (3CC) and catechol were detected as intermedi-
ates of degradation (Figure 1). If 5mM bromoethanesulfonic acid
(BBS A) was included in the enrichment, 5-chloroprotocatechuate
(5CP), protocatechuate, and vanillate were also detected (Figure
2). From the temporal sequence and amounts of intermediates
formed, we concluded that the major pathway of 5CV degrada-
tion was stepwise demethoxylation to 5CP, decarboxylation to
3CC, and dechlorination to catechol, which was subsequently
completely degraded (Figure 3). The addition of BES A appeared
to inhibit decarboxylation and demethoxylation, allowing us to
detect 5CP, as a major intermediate, and vanillate and
500
w—V 5-chlorovanillate
A—A 3-chlorocatechol
0—0 catechol
10
30 40
Time (days)
Figure 1. Degradation of 5-chlorovanillatQ in the first transfer of an
Eleven Mile Creek primary anaerobic 5C V enrichment.
250-
c
o
O
• • 5-chlorovanillic acid
A A 3-chlorocatechol
0—0 catechol
d d 5-chloroprotocatechuic acid
V—V protocatechuic acid
O—O vanillic acid
5 10
Time (days)
15
Figure 2. Degradation of 5-chlorovanillate in the first transfer of an
B
-------
C°°H
CH3OH
CO2 + CH4
ci T ocH3 ci T OH
OH OH Cl
5-chlorovanilhate 5-chloroprotocatechuate 3-chlorocatechol
ci-
COOH
OCH,
CH3OH
OH
vanilliate
catechol
COOH
OH
protocatechuate
Figure 3. Proposed pathway for the complete degradation of 5-chlorovanillic acid.
lated in defined pyruvate medium and designated strain 3162pyr.
Physiological studies placed strain 3162pyr in the desulfoviridin-
negative, desulfovibrio-like group of rod-shaped sulfate reduc-
ers that include Desulfovibrio strain Norway 4, and the newly
formed genus Desulfomicrobium. Placement of strain 3162pyr
as a new species in this group was confirmed by 16S rRNA
analysis. 16S rRNA analysis of the co-culture is currently under-
way to identify the curved rod.
Strain 3162pyr degraded benzoate, 3CB, and 3-bromobenzoate
in the presence of pyruvate, but did not degrade 5CV. Benzoate
has not been detected as an intermediate in the degradation ot
3CB or 3-bromobenzoate. These data suggest that the curved rod
is responsible for dechlorination of 5CV. Studies are underway
to confirm the role of the two species in 5CV transformation.
The original 5CV enrichment, its derivatives, and the co-
culture will be tested for their ability to detoxify paper-milling
waste obtained from a local paper-milling company. Untreated
and treated waste samples will be assayed in a test for toxicity/
tetratogenicity with embryonic inland silversides, Memdia
beryllina. Comparisons can be made to determine the efficacy of
the various consortia in detoxification.
88
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Methanogenic Degradation of Heterocyclic Aromatic
Compounds by Aquifer-Derived Microcosms
E. Michael Godsy, Donald F. Goerlitz
U.S. Geological Survey
Menlo Park, CA
and
Dunja Grbic Galic
Stanford University
Stanford, CA
Introduction
The ultimate fate of heterocyclic aromatic compounds in
subsurface environments is controlled by various transport and
transformation processes. Potentially the most important but
currently one of the least understood processes affecting ground-
water quality is biotransformation of these pollutants by indig-
enous microorganisms under anoxic conditions. Heterocyclic
compounds are frequently encountered in the environment be-
cause they are major constituents of both fossil and synthetic
fuels and of many pesticide mixtures (e.g., creosote). Heterocy-
clic aromatic hydrocarbons that contain nitrogen or sulfur in their
ring structure have been shown to degrade under methanogenic,
denitrifying, and sulfate-reducing conditions.
The mechanism of the initial methanogenic oxidation of
benzothiophene was shown to be hydroxylation of the heterocy-
clic ring, followed by reduction and cleavage (1). After cleavage,
the substituent side chains and the remaining homocyclic ring
were subjected to various reactions, including oxidation, decar-
boxylation, desulfurylation, and O-methylation. These reactions
were followed by reduction and cleavage of the homocyclic ring,
6-oxidation, and mineralization to CH4, CO2, and H2S. The major
degradation pathway intersected segments of both the phenol and
benzoic acid pathways. A minor pathway was observed starting
with the oxidation of the homocyclic ring with subsequent ring
reduction, ring cleavage, degradation of the remaining S-hetero-
cyclic ring, and mineralization (1).
There have been several reports of the persistence of interme-
diate compounds during laboratory anaerobic degradation stud-
ies of heterocyclic aromatic compounds. Oxindole has been
shown to be an intermediate of the anaerobic biodegradation of
indole (2). 2(lH)-quinolinone and l(2H)-isoquinolinone were
observed as intermediates of quinoline and isoquinolinone, re-
spectively (3). Godsy et al. (3) also observed that during
downgradient travel in an aquifer and in laboratory microcosms,
quinoline and isoquinoline were very rapidly oxidized to 2(1H)-
quinolinone and l(2H)-isoquinolinone, respectively. The latter
two compounds persisted for some time before they were miner-
alized to CH4 and CO2.
In this study, we present evidence that the major factor affect-
ing the fate of creosote-derived heterocyclic compounds in
methanogenic environments is microbial conversion to CH4 and
CO2. The pathways and thermodynamics of microbial conver-
sion by a complex, aquifer-derived, methanogenic consortium is
determined for a number of these compounds.
Materials and Methods
The sample site is located in Pensacola, Florida, adjacent to an
abandoned wood-preserving plant (3). The wood-preserving
process consisted of steam pressure treatment of pine poles with
creosote and/or PCP (pentachlorophenol). For more than 80
years, large but unknown quantities of wastewater, consisting of
extracted moisture from the poles, cellular debris, creosote, PCP,
and diesel fuel from the treatment processes, were discharged to
nearby surface impoundments. The impoundments were unlined
and in direct hydraulic contact with the sand-and-gravel aquifer.
Contamination of the ground water resulted from the accretion of
wastes from these impoundments.
Microcosms used for the study were prepared in 4 L glass
sample bottles and contained approximately 3 kg of aquifer
material anaerobically collected from the approximate centroid
of the active methanogenic zone (3). The compounds of interest
were added to 2.5 L of mineral salts solution at concentrations
similar to those found in the aquifer: 10 to 40 mg/L. Amorphous
FeS was used as the reducing agent (4) to ensure methanogenic
conditions. The microcosms were prepared, stored, and sampled
in an anaerobic glove box containing an O-free Ar atmosphere
at 22°C.
Degradation pathways for the various compounds were deter-
mined by computer-aided gas chromatography-mass spectrom-
etry (GC/MS/DS) analysis of intermediate compounds that ap-
peared in the growth liquor during biodegradation (3). Concen-
89
-------
Oxidation of the
Homocyclic
Ring
Oxidation of the
Heterocyclic
Ring
8-quinolinone
I
I
I
I
I
y
H3C
H3C"
2,3-dimethylpyridine
8-quinolinol
2-quinolinol
2(1 H)-quinolinone
Hydrolytic Cleavage of
the Oxidized Ring
aniline
benzoic acid
phenol
Ring Cleavage
Acetogenesis
acetic acid, formic acid
\
I
y
Methanogenesis
a, CH4, NH3 *
* Proposed intermediate-not found in microcosms but found in ground-water samples.
G) Ammonia not determined.
Figure 1. Methanogenic degradation pathway for quinoline in aquifer-derived microcosms.
trations of CH4 and CO2 in the head space and dissolved in the
mineral salts were determined by gas chromatography (GC) (5).
Results
Compounds detected in the growth liquor just before the onset
and during methanogenesis of quinoline are shown in Figure 1.
Compounds that were found in the organic-free controls are not
shown and presumably do not arise from the degradation of
quinoline. Autoclaved controls did not produce any detectable
amounts of CH4 and CO2.
During oxidation of the nitrogen-containing heterocyclic com-
pounds indole, quinoli ne, and isoquinolinone to oxindole, 2(1H)-
quinolinone, and l(2H)-isoquinolinone, respectively, the parent
compounds were initially stoichiometrically oxidized. The oxi-
dation did not support growth. The oxidized intermediates per-
sisted for some time after the initial oxidation of the parent
compounds before they were degraded to CH4 and CO2.
The mass balances on the complete degradation of the above
compounds yielded 87.6 percent of theoretical gas production for
oxindole, 91.7 percent for 2(lH)-quinolinone, and 88.5 percent
for l(2H)-isoquinolinone. The conversion values obtained for
the above compounds are well within the expected range.
Discussion and Conclusions
It is evident, based on the report by Pereira et al. (6) of the
incorporation of water into the homocyclic ring, that the first step
in biodegradation consisted of oxidation, followed by the cleav-
age of the N-heterocyclic ring. After cleavage of this ring, the
substituent side chains and remaining homocyclic rings were
subjected to various reactions, including oxidation, decarboxyla-
90
-------
Table 1. Free Energy Changes During Oxidation of the Tested Heterocyclic Compounds
Compound
' kJ.mol-'(8,9)
Indole to Oxindole
C8H7N + H20 ->C8H7NO + 2 H* + 2e~
Quinoline to 2(1H)-quinolinone
C9 H7 N + H2O-*C9H7 N) + 2 H * + 2e-
Isoquinoline to 1 (2H)-isoquinolinone
C9H7N + H2O->C9H7NO + 2 H * + 2 e-
Oxindole to CH4 and CO2
C8H7NO + 6.5 H2O + H* -> 3.75 CO2 + NH + 4.25 CH4
2(1H)-quinolinone to CH4 and CO.,
C9H7NO + 7.5 H20 + H* -» 4.25 CO2 + NH + 4.75 CH4
1(2H)-isoquinolinone to CH4 and CO2
C9H7NO + 7.5 H2O + H* -» 4.25 CO2 + NH + 4.75 CH4
168
84
56
-362
-298
-303
Note: AG °'(aq) from (8) or estimated using the method of Jobak (9).
tion, and deamination, with ultimate conversion to benzoic acid
and phenol. The conversion to benzoic acid and phenol most
likely follows the previously described pathways (2,7), consist-
ing of the reduction of the homocyclic ring, cleavage of this ring,
6-oxidation, and mineralization to CH4, CO2, and NH3. A minor
pathway first reported by Godsy and Grbic-Galic (1) for
benzothiophene was also observed for quinoline, presumably for
isoquinoline, but not for indole. The pathway consists of oxida-
tion of the homocyclic ring with subsequent ring reduction, ring
cleavage, degradation of the remaining N-heterocyclic ring, and
mineralization. Nitrogen-containing aliphatic acids, however,
were not detected in the microcosms to confirm that 2,3-
dimethylpyridine was converted to CH4 and CO2. Nitrogen-
containing aliphatic acids were detected in ground-water samples.
The corresponding compounds during benzothiophene degrada-
tion were detected, confirming that thiophene-2-ol was degraded
to CH4 and CO2 (1). 2,3-dimethylpyridine was removed from the
contaminated ground water faster than dilution or dispersion
could account for during downgradient movement in the aquifer
at the study site. This was presumably due to biodegradation.
The oxidation of the parent compounds was found to be a
process that requires energy (Table 1); therefore, these reactions
must be coupled to the reduction of an unidentified compound(s).
The first step in the anaerobic degradation of aromatic com-
pounds is the reduction of the aromatic ring (7). The electrons
from the oxidation of the nitrogen heterocycles are quite possibly
used for this purpose. It is not apparent why the parent com-
pounds are rapidly oxidized, then remain for long periods of time
(days to months) before they are converted to CH4 and CO2 (3),
if not for the purpose of obtaining electrons for the reduction of
other aromatic compounds (e.g., phenol and benzoate). It is not
clear at this time whether the organism(s) that are responsible for
the oxidation of the parent compound are involved in the
methanogenic degradation of the oxidized products.
References
1. Godsy, E.M. and D. Grbic-Galic. 1989. Biodegradation
pathways for benzothiophene in methanogenic micro-
cosms. In: U.S. Geological Survey Toxic Substances
Hydrology Program—Proceedings of the Technical
Meeting, Mallard, G.E. and S.E. Ragone, eds. Phoenix,
Arizona, September 26-30,1988. U.S. Geological Sur-
vey Water Resources Investigations Report 88-4220,
pp. 559-564.
2. Berry, D.F., E.L. Madsen, and J.M. Bollag. 1987. Con-
version of indole to oxindole under methanogenic con-
ditions. Appl. Environ. Microbiol. 53:180-182.
3. Godsy, E.M., D.F. Goerlitz, and D. Grbic-Galic. 1992.
Methanogenic biodegradation of creosote contaminants
in natural and simulated ground-water ecosystems.
Ground Wat. 30:232-242.
4. Brock, T.D. and K. O'Dea. 1977. Amorphous ferrous
sulfide as a reducing agent for culture of anaerobes.
Appl. Environ. Microbiol. 33:254-256.
5. Godsy, E.M. 1980. Isolation of Methanobacterium
bryantii from a deep aquifer by using a novel broth-
antibiotic disk method. Appl. Environ. Microbiol.
39:1074-1075.
6. Pereira, W.E., C.E. Rostad, D.M. Updegraff, and J.L.
Bennett. 1988. Microbial transformations of azaarenes
in creosote-contaminated soil and ground water: Labo-
ratory and field studies. International Conference on
Water and Wastewater Microbiology. Newport Beach,
California, pp. 5-1 to 5-7.
91
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7. Young, L.Y. 1984. Anaerobic degradation of aromatic
compounds. In Microbial degradation of organic com-
pounds. D.T. Gibson, ed. Marcel Dekker, New York,
New York, pp. 487-523.
8. Stull, D.R., E.F. Westrum, Jr., and G.C. Sinke. 1969.
The chemical thermodynamics of organic compounds.
John Wiley and Sons, New York.
9. Reid, R.C., J.M. Prausnitz, and B.E. Poling. 1985. The
properties of gases and liquids. McGraw-Hill Book Co.,
New York.
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Chemical Interactions and pH Profiles in Microbial Biofilms
Joseph R.V. Flora, Makram T. Suidan, and Pratim Biswas
University of Cincinnati
Cincinnati, OH
and
Gregory D. Sayles
U.S. Environmental Protection Agency
Cincinnati, OH
Biofilm processes are important to water quality control and
are applied in numerous areas such as in ground-water treatment,
in conventional and hazardous-wastewater treatment, and in the
prediction of fate and effect of pollutants in natural stream
systems. Substrate utilization in biofilms has been modeled
traditionally by coupling Pick's law with Monod reaction kinet-
ics. An inherent assumption in these models is that the pH
remains constant within the biofilm. However, biological reac-
tions can effect changes in the local pH. Nitrification and the
utilization of chlorinated organics produces acid equivalents and
causes a decrease in pH, while denitrification consumes acid
equivalents and causes an increase in pH. Production and utiliza-
tion of carbon dioxide also alters chemical equilibrium. The rates
of substrate utilization and growth of microorganisms are pH
dependent and may vary significantly from the bulk solution to
the attachment wall of biofilms. This is specially important for
methanogens, which can tolerate only a very narrow pH range,
and for nitrifying bacteria, which cannot tolerate low pH. Fur-
thermore, diffusional resistance to mass transfer creates pH
gradients within the biofilm and alters the speciation of com-
pounds into forms unavailable for microbial consumption (e.g.,
an increase in pH transforms carbon dioxide into bicarbonate and
inhibits autotrophic organisms).
A detailed understanding of the chemical interactions within
the biofilm is essential for the optimization of reactor operation.
An approach was developed to analyze the effect of pH on
substrate transport. This approach incorporates ionic mass trans-
port effects accurately and accounts for corrections for the
activity, electrophoresis, relaxation, andhydration of ions. Mod-
els will be developed for general autotrophic and heterotrophic
biofilms, including nitrifying, denitrifying, and dechlorinating
systems. Preliminary results for an autotrophic biofilm are de-
scribed.
Inorganic carbon in the form of dissolved CO2 is a major
nutrient required by autotrophs (such as algae) and may easily
become limiting in systems used for wastewater treatment. Using
Monod kinetics to describe the rate of substrate utilization of
inorganic carbon and assuming the rate of CO production
through endogenous respiration to be first order with respect to
the biomass density, a steady-state mass balance on inorganic
carbon yields,
dJco2 +dJHctf, +dJCo§- bX,. kX,[COgl
dx dx dx Y "KS + |
(1)
where J is the molar flux of the species; x is the direction
perpendicular to the biofilm surface; b is the respiration coeffi-
cient; Xf is the biomass density; Y is the yield of microorganisms
per mass of substrate; k is the maximum utilization rate; Ks is the
Monod half-saturation constant for CO2; and [CO2] refers'to the
molar concentration of CO2. The characteristics of the system
dictate the appropriate expressions for the flux to be used. In this
study, three scenarios utilizing various expressions for the flux
are compared. Case 1 refers to a dilute system where Pick's law
is used for neutral solutes and the Nernst-Planck equation is used
for ionic solutes; case 2 incorporates activity corrections using
Davies' equation; and case 3 includes activity corrections with
additional corrections for electrophoretic effects, relaxation, and
hydration using the Stefan-Max well equations. Equation (1) with
relationships for chemical equilibrium and electroneutrality was
solved assuming HCO3", CO/-, H+, OH", and Na+ as the ions
present within the system.
The profiles of CO2, HCO3", CO32', and pH for a biofilm depth
of 1,000 ]im with a bulk pH of 7.0 and an alkalinity of 1 meq/L
for case 1 are shown in Figure 1. As dissolved CO2 is consumed,
the pH increases dramatically within the first 100 um of the
biofilm. Consequently, the dissolved CO2 concentration de-
creases rapidly in the first 50um due to the coupled effects of pH
on the speciation of inorganic carbon and the net substrate
utilization of CO2. HCO3" remains relatively constant during the
93
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1
1.5
1.0
0.5
0.0
11
10
200 400 600 800
Distance into the Biofilm, ^M
1000
Figure 1.
Concentration profiles of the inorganic carbon species and
pH for Lf= 1,000 \im, bulk pH = 7.0, bulk alkalinity =
1 meq/L (case 1).
first 50 (am of the biofilm and subsequently decreases throughout
the biofilm as the rise in pH transforms the dominant inorganic
carbon species to CO32\ Finally, the CO32 concentration in-
creases due to the shift in chemical equilibrium and eventually
decreases because CO2 is still utilized even in the deeper portions
of the biofilm. Thus, the utilization of a substrate whose specia-
tion is dependent on pH may be severely limited in deep biofilms.
Figure 2 shows the variation of the total flux as a function of the
overall biofilm depth for a bulk alkalinity of 1 meq/L for various
bulk pH values for case 1. The increase in flux is linear with the
biofilm depth at small values of Lf while the flux approaches a
constant value at large Lf. The flux is limited by the rate of
reaction within the biofilm when Lf is small and is limited by the
rate of mass transfer for large values of Lf. A decrease in pH,
while retaining the same alkalinity, provides a higher flux when
the biofilm is deep because a greater fraction of the inorganic
carbon in the bulk is CO2. Thus, simple strategies such as altering
the bulk conditions by changing the pH or increasing the buffer
intensity of the system will optimize utilization of the biofilm
depth.
Other results of the study show that incorporating activity
corrections in the analysis (case 2) results in a decrease in the
calculated fluxes for deep biofilms. The activity of each species
is reduced in systems with an ionic strength less than 1 m and
results in an overall decrease in the driving force for the flux. This
decrease is proportional to the bulk alkalinity for the system
investigated and a larger difference develops between the fluxes
at high and low alkalinity values. When the biofilm depth is
small, a decrease in the driving force is insignificant because the
flux is limited by the rate of reaction within the biofilm. The
calculated fluxes for cases 2 and 3 are identical and corrections
for electrophoretic effects, relaxation, and hydration are negli-
gible.
1
.92
o
E
0.25
0.20
0.15
0.10
0.05
0.0
bulk pH=6
bulk pH=7
bulk pH=8
100 200 300 400 500
Figure 2. Total flux of inorganic carbon into the biofilm for various L,
and bulk pH, and for a bulk alkalinity = 1 meq/L (case 1).
94
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Decontamination of PCB-Contaminated Sediments Through
the Use of Bioremediation Technologies
Daniel A. Wubah
University of Georgia
Athens, GA
and
W. Jack Jones and John E. Rogers
U.S. Environmental Protection Agency
Athens, GA
Polychlorinatedbiphenyl (PCB) congeners are relatively inert,
stable to oxidation and heat, and have high viscosities, high
refractive indices, low vapor pressures, high hydrophobicities,
and excellent dielectric characteristics. As a result of these
properties, PCBs were used in a variety of industrial applications,
including organic diluents, plasticizers, hydraulic fluids, heat
exchange and transfer fluids, solvent extenders, and microscope
immersion oils (1). Some of these properties also contributed to
their accumulation in soils, various elements of the food chain,
and sediments of streams, lakes and estuaries (2). During the past
two decades, concern has increased with respect to the eco-
toxicological and human implications of the accumulation of
PCBs (3,4). In 1971, PCBs were restricted to closed systems in
the United States. Eight years later, a total ban on their manufac-
ture and use was ordered by the Congress. Residual contamina-
tion of PCBs continues to be a problem, however, particularly
along the banks of the Great Lakes where PCBs were used.
Pollution control agencies have launched a concerted effort to
reduce the levels of accumulated PCBs in the environment
because of their persistence, toxicity, and tendency to
bioconcentrate. Polychlorinated biphenyls can be transformed
into less harmful compounds by both chemical and physical
methods, e.g., ultraviolet radiation, microwave treatment,
supercritical water processes, and incineration (5). However,
microbial degradation appears to offer the most cost-efficient
method of reducing levels of PCBs in contaminated sediments.
Consequently, the coupling of anaerobic dechlorination to aero-
bic metabolism has been suggested as a possible method to
reduce the levels of accumulated PCBs in the environment.
Results from bench-scale studies suggest that PCBs can be
biodegraded by both aerobic and anaerobic microbial processes
(4,6). Aerobic bacteria usually degrade congeners with one to
five chlorine atoms, but not the higher chlorine-substituted
congeners. However, the rate and extent of degradation are
influenced by the position of the chlorine atoms on the biphenyl
ring (7, 8). For example, congeners with open 2,3- or 3,4-
positions are degraded faster than congeners with chlorine atoms
at these positions. These sites are necessary for the attachment of
a dioxygenase enzyme to effect the fission of the biphenyl ring.
Several bacterial strains that can biodegrade PCBs aerobically
have been isolated from enrichments of soils and sediments from
contaminated sites. These strains vary in their abilities to degrade
PCBs; some degrade only mono-, di-, and trichlorobiphenyls,
and others degrade tetra- and pentachlorobiphenyls (7). Anaero-
bic reductive dechlorination of PCBs was first proposed by
Brown and coworkers (9). Later reports from laboratory studies
provided unequivocal evidence that reductive dechlorination
could occur in anaerobic sediments (10). Microorganisms that
are capable of anaerobic dechlorination often degrade the highly
chlorinated PCB congeners, but not the less chlorinated conge-
ners and the biphenyl nucleus (4). There is also a preferential
removal of chlorine from the para and meta positions. The
products of anaerobic dechlorination, therefore, are mainly ortho-
substituted and have fewer chlorine atoms on the biphenyl ring.
Such congeners are less toxic and may serve as substrates for
aerobic dechlorinating bacteria (10).
The objective of this study is to determine the dechlorinating
capacity of PCB-contaminated sediments from selected Great
Lakes sites. The dechlorination of indigenous PCB congeners in
sediment from the Saginaw River was determined under aerobic
conditions. The initial chromatographic profile of Saginaw sedi-
ment indicated a preponderance of congeners with one to four
chlorine atoms, which suggests that the original Aroclor mixture
had been transformed. An additional aerobic incubation for 6
months resulted in a further net loss in the concentration of total
congeners reported to be amenable to aerobic degradation. These
congeners include4-chlorobiphenyl, 2;2'4-trichlorobiphenyl (tri-
CB), 2,4',5-tri-CB, 2,4,4'-tri-CB, 3,4,4'-tri-CB, 2,3',5',6-
tetrachlorobiphenyl (tet-CB), and 2,2',3,4'-tet-CB.
95
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Sediments collected from the Ashtabula River and the
Sheboygan Harbor and bay area were incubated under anaerobic
conditions. The Ashtabula sediment was spiked with 2,3,3',4,4'-
pentachlorobiphenyl (penta-CB), 2,3,3',4,4',5-hexachlo-
robiphenyl (hexa-CB), and 2,2',3,4,5,6,6'-heptachlorobiphenyl
(hepta-CB). Each congener was added individually and in com-
bination with the other two congeners. Reductive dechlorination
was observed after lag periods of 5,3, and 4 months for the penta-
CB, hexa-CB, and hepta-CB, respectively. The Sheboygan sedi-
ment was spiked with three concentrations of 2,2',3,3 ,4,5,6,6'-
octachlorobiphenyl (octa-CB). After a lag phase of 5 months, a
decrease in the higher chlorinated congeners, including the added
octa-CB, and a concurrent increase in the lower chlorinated
congeners were observed in all the samples amended with the
octa-CB, but not in the unspiked and autoclaved sterile controls.
These preliminary data suggest the presence of actively dechlo-
rinating microorganisms in the sediments from these three Great
Lakes sites. Further research will include studies to determine the
factors that will enhance the dechlorinating activity of these
microorganisms.
References
1. Hutzinger, O., S. Safe, and V. Zitko. 1974. The Chem-
istry of PCBs. CRC Press, Cleveland, Ohio.
2. Safe, S., L. Safe and M. Mullin. 1987. Polychlorinated
biphenyls: environmental occurrence and analysis. In
PCBs: Mammalian and Environmental Toxicology.
Springer-Verlag, Heidelberg.
3. Kalmaz, E.V. and G.D. Kalmaz. 1979. Rev. Ecol.
Model. 6:223.
4. Abramowitz,D.A. 1990. Crit. Rev. Biotechnol. 10:241.
5. Johnston,L.E. 1985.Environ. HealthPers. 60:339-346.
6. Unterman, R. 1990. Aerobic biodegradation of PCBs.
In Biological Remediation of Contaminated Sediments,
with Special Emphasis on the GreatLakes, Jafvert, C.T.
and E. Rogers, eds. U.S. Environmental Protection
Agency, Athens, Georgia. Publication No. EPA/600/9-
91/001.
7. Bedard, D. L. and M. L. Haberl. 1990. Microb. Ecol.
20:87.
8. Furukawa, K., N. Tomizuka, and A. Kamibayashi.
1979. Appl. Environ. Microbiol. 38:301.
9. Brown, J.F., R.E. Wagner, D.L. Bedard, M.J. Brennan,
J.C. Carnahan, RJ. May, and TJ. Tofflemire. 1984.
Northeast Environ. Sci. 3:167.
10. Quensen, J.F., J.M. Tiedje and S.A. Boyd. 1988. Sci-
ence 242: 752.
96
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Bioremediation of Soils and Sediments Contaminated with Aromatic Amines
E.J. Weber and I.E. Rogers
U.S. Environmental Protection Agency
Environmental Research Laboratory
Athens, GA
and
D.L. Spidle
Technology Applications Inc.
c/o U.S. Environmental Protection Agency
Athens, GA
and
K.A. Thorn
USGS
Arvada, CO
Aromatic amines comprise an important class of environmen-
tal contaminants. Concern over their environmental fate arises
from the toxic effects that certain aromatic amines exhibit toward
microbial populations and reports that they can be toxic and/or
carcinogenic to animals. Aromatic amines can enter the environ-
ment from the degradation of textile dyes, munitions, and numer-
ous herbicides, including the phenylureas, phenylcarbamates,
and acylanilides. Because these chemicals are synthesized from
aromatic amines, loss of aromatic amines to the environment may
also result from production processes or improper treatment of
industrial waste streams. The high probability that contamination
of soils, sediments, and ground-water aquifers with aromatic
amines will occur necessitates the development of in situ biore-
mediation techniques for their treatment.
A number of studies suggest that aromatic amines become
covalently bound to the organic fraction of soils and sediments
through oxidative and/or nucleophilic coupling reactions (1,2). It
generally is accepted that, once bound, the bound residue is less
bioavailable and less mobile than the parent compound. Thus,
procedures that enhance the irreversible binding of aromatic
amines to soil constituents could potentially serve as aremediation
technique.
Model studies suggest that oxidative enzymes derived from
soil microorganisms play a significant role in catalyzing the
formation of bound residues (3). Stimulation of these naturally
occurring enzymes could provide an effective in situ method for
the treatment of soils, sediments, and ground-water aquifers
contaminated with aromatic amines. For example, Berry and
Boyd (4) were able to enhance covalent binding of the potent
carcinogen, 3,3'-dichlorobenzidine(DCB),in soil by the addition
of highly reactive substrates (i.e., ferulic acid and hydrogen
peroxide). They concluded that by providing the indigenous
peroxidase enzymes with highly reactive substrates, the number
of reactive sites for covalent binding was increased, which led to
the enhanced incorporation of DCB.
The goals of our research are to gain a better understanding of
the role that extracellular enzymes play in the covalent binding of
aromatic amines to natural organic matter (NOM) and to develop
effective methods for remediation of contaminated soils and
sediments by the stimulation of indigenous oxidative coupling
enzymes.
Initially, our research efforts have focused on elucidating the
mechanisms by which aromatic amines (i.e., aniline and substi-
tuted anilines) bind irreversibly to sediments, soils, and dissolved
organic matter. Accordingly, the sorption kinetics of a series of
2- and 4-substituted anilines were measured in a sediment-water
system. In general, the rate and extent of sorption were found to
increase with the pKa of the aniline. Decreasing pH enhanced the
sorption of the 4-substituted anilines in a silt-clay system. Se-
quential extraction studies with 14C-aniline suggested that revers-
ible cation exchange processes do not contribute significantly to
aniline sorption in the natural sediment-water system, and that
97
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irreversible covalent binding to the organic matter of the sedi-
ment matrix dominates the sorption process. In other studies,
treatment of a sediment-water system with 2,4-DNP or hydroxy-
lamine, chemicals known to form stable adducts with carbonyl
groups, 24 hours prior to the addition of the 4-methoxyaniline
effectively blocked the sorption of the aniline, suggesting the
importance of covalent binding through nucleophilic addition to
carbonyl groups in the sediment matrix.
To provide direct spectroscopic evidence for the type of
covalent linkages formed between aromatic amines and natural
organic matter, 15N NMR was used to analyze fulvic acid that had
been treated with l5N-aniline. INEPT and ACOUSTIC 15N-NMR
spectra exhibited resonances for imine, anilide, anilino-quinone,
and anilino-hydroquinone nitrogens, providing further evidence
that covalent binding of aniline occurs through nucleophilic
addition to carbonyl moieties found in organic matter.
Experiments are currently being designed to increase the
binding capacity of sediment-water systems by enhancing the
activity of the indigenous enzymes through the addition of
reactive substrates such as naturally occurring phenols and
hydrogen peroxide. Future studies will focus on assaying soil,
sediment, and aquifer materials for peroxidase activity and
measuring the capacity of these systems to covalently bind
representative aromatic amines. In addition, we will use tech-
niques for isolating extracellular soil enzymes to transfer activity
from soils with high activity to soils with low activity.
References
1. Paris, G.E. 1980. Covalent binding of aromatic amines
to humates. 1. Reactions with carbonyls and quinones.
Environ. Sci. Technol. 14:1099-1106.
2. Jafvert, C.T. and E.J. Weber. 1990. Sorption of ioniz-
able organic compounds to sediments and soils. Unpub-
lished report.
3. Bollag, J. and W.B. Bollag. 1990. A model for enzy-
matic binding of pollutants in the soil. Intern. J. Environ.
Anal. Chem. 39:147-157.
4. Berry, D.F. and S.A. Boyd. 1985. Decontamination of
soil through enhanced formation of bound residues.
Environ. Sci. Technol. 19:1132-1133.
98
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Development of Small-Scale Evaluation Techniques for
Fungal Treatment of Soils
Steven I. Safferman
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
Frederic Baud-Grasset
Visiting Scientist, U. S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
Sandrine Baud-Grasset
International Technology Corporation
U.S. Environmental Protection Agency Test & Evaluation Facility
Cincinnati, OH
and
Richard T. Lamar
U.S. Department of Agriculture
Forest Products Laboratory
Madison, WI
Standardized bioremediation protocols that mimic full-scale
treatments will allow for consistent and economical evaluations
and comparisons of remediation options under controlled condi-
tions. These protocols are essential for developing technologies,
such as lignin-degrading fungi (LDF) processes, where the data
base of experience is limited. They not only help decrease the
chance of failure but can also be a research tool for further
development. In this research, a draft protocol to assess fungal
treatment of contaminated soils has been developed and is
currently being tested and modified. A tiered approach is being
utilized (Figure 1), starting with simple visual laboratory analy-
ses and progressing to bench-scale soil pans. The component
experiments and preliminary results from the first test soil are
described in the following paragraphs.
The protocol is initiated by examining data concerning site-
specific conditions including the characteristics of the soil and
concentrations of the pollutants. In addition, the feasibility of a
full-scale land treatment (i.e., adequate space and temperature) at
the site must be assessed, as land farming is currently the only
method available to apply LDF technologies. If the LDF tech-
nologies appear to be feasible, soil samples will be characterized
for the target concentrations, nutrient contents, and other impor-
tant characteristics, and a simple sensitivity test will be initiated.
In the sensitivity test, fungal strains that have been identified in
the pre-screening stage are plated on a medium that has been
amended with the target pollutants at varying concentrations.
Since the conditions for growth are optimal for the fungus during
this experiment, the lack of growth can be attributed to an
inhibiting effect of the pollutants at the test concentration.
Fungal strains that are not relatively resistant to the target
pollutants will be tested in the growth tier of the protocol. This
stage allows for the selection of .fungal species, inoculum load-
ings, nutrient loadings and compositions, and soil blending (to
reduce pollutant concentration) that appear to have the most
potential to remediate the soil. Fungal-inoculated and non-inocu-
lated site soils are incubated for 4 to 6 weeks. The endpoint
analyses depend on the experience with the target compounds
and the soil's characteristics. If extensive LDF data are already
99
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Literature
^
Discontinue
-^
\ Soil Blending
Strains with best growth potential
Figure 1. Tiered testability approach.
available for the target pollutant(s), a visual growth evaluation
might suffice. For a pollutant that has been studied to a lesser
extent, disappearance of the parent compound, analysis of sus-
pected intermediates, and simple toxicity assays are necessary.
The results from the growth studies will determine the number
of fungal strains and other bioremediation variables that will
need to be examined during the fate and soil pan stages of the
protocol. The fate study seeks to determine how the target
compounds are removed, i.e., by bioremediation or abiotic mecha-
nisms. Radio-labeled compounds are spiked into clean reference
soils and incubated in reactors similar to those used in the growth
study. All possible routes of target compound disappearance are
monitored during a4- to 6-week incubation period, thus allowing
for mass balances.
The soil pan study attempts to mimic field conditions. Twelve-
in. square, 18-in. deep stainless steel pans with gravel under
drains and leachate collection systems are currently being uti-
lized. The pans are housed in glove boxes to protect researchers
and also to control air flow through the box. Moisture and
temperature are controlled to reflect field conditions. Common
hand-held garden tools are used periodically to mix and aerate the
soil. The performance of fungal treatments is reported as target
compound disappearance and soil detoxification relative to the
various control pans. The controls are treated identically to
fungal treatments, except no inoculum is added. No attempt to
eliminate indigenous microorganisms is needed as the advan-
tages gained by the fungal treatment are sought. Controls that
contain wood chips and/or sawdust/bran fungal growth matrix
(used as part of the fungal treatment) are utilized. Each treatment
is duplicated in two independent pans. The soil in the pans can be
subsampled without a significant change in soil quantity so that
the kinetics of degradation can be determined. Leachate samples
are also easily obtained for analyses.
The use of toxicity assays to determine if the treatments are
reducing the toxicity of the soil and to predict appropriate times
to conduct expensive chemical analyses is also being considered
for inclusion in this protocol. Assays that are currently being
studied include plant germination, root elongation, earthworm
lethality and reproduction, earthworm strand break, plant genetic
damage using Tradescantia paludosa, and the Microtox system.
When interpreting toxicity data, care must be exercised as the
toxicity data from an individual assay only apply to the organism
tested and inference to another is not valid. In addition, the effect
of long-term toxicity cannot be assessed from short-term assays.
The evaluation of the technology is, therefore, based on the fate
studies, soil pan experiments, and toxicity assays. Caution in the
interpretation at these early stages of the development of LDF
technologies is paramount as the protocol is not intended to verify
the success of the LDF treatment in the field to detoxify the soil
or to provide full-scale design parameters. Rather, it is intended
to provide an economical means to determine if more detailed
fate and field-scale treatability studies are warranted.
The original draft protocol was tested on creosote-contami-
nated soil from a Superfund site. The protocol described above
reflects modifications that were deemed necessary from experi-
ence with this soil. Preliminary results found clear distinctions
between the control and fungal treated pans, pollutant removal
rates for different inoculum loadings, and different removal rates
for lower and higher molecular weight compounds.
Additional soils will be tested using the revised protocol before
a finalized guidance document is prepared. The protocol, how-
ever, should not be limited to fungal treatments, but could be
modified to apply to virtually any land treatment application. The
possibility of simplifying the protocol should also not be dis-
counted. A comparison of results from this protocol with field
data would be valuable in assessing its true predictive powers and
is planned as part of this program.
Acknowledgements
The authors would like to recognize and express their apprecia-
tion to the following researchers within the listed collaborating
organizations. International Technology Corporation: Radha
Krishnan, Roy Haught, Gary Lubbers, and Jennifer Platt;
Microbics Corporation: Anthony Bulichand and John Cusick;
Technical Applications, Inc.: Yong Benyaghoub and Joe Bifulco;
U.S. EPA Environmental Monitoring and System Laboratory:
Bernard Daniel, Lina Chang, Susan Cormier, Denise Gordon,
John Meier, T. Reddy, and N. Kate Smith; U.S. EPA Risk
Reduction Engineering Laboratory: Leo Fichter, Alfred Kornel,
Vincent Salotto, and Lawrence Winslow; Western Illinois Uni-
versity: Te-Hsiu Ma.
100
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Development of Aerobic Biofilter
Design Criteria for Treating VOCs
George A. Sorial, Makram T. Suidan, Francis L. Smith, and Paul J. Smith
University of Cincinnati
Cincinnati, OH
and
D. Strunk, P. Biswas, and Richard C. Brenner
U.S. Environmental Protection Agency
Cincinnati, OH
The objective of this project is to reduce aerobic biotreatment
by biofiltration of air contaminated with hazardous volatile
organic compounds (VOCs). The VOCs to be studied are tolu-
ene, ethylbenzene, methylenechloride, trichloroethylene, and
chlorobenzene.
Previous investigations have demonstrated that such VOCs
can be rapidly and efficiently biodegraded in biofilters. Reactor
plugging due to excess biomass was identified as a complicating
problem. This research will evaluate different biological attach-
ment media, as well as operational strategy alternatives for the
purpose of optimizing reactor performance and design.
The experimental apparatus will be started in April 1992 at the
EPA Test and Evaluation Facili ty. This apparatus consists of four
separate biofilters, complete with all necessary support equip-
ment Each biofilter system is provided with independent tem-
perature and humidity control, and is insulated to prevent tem-
perature fluctuations. The simplified process description is as
follows. Standard 100 psi utility air is purified by separating out
particulates, CO2, H2O, and VOCs, producing essentially bone
dry, pure air at ambient conditions (4 to 40°C). After let-down to
10 psi, the air is chilled to 4°C and then reheated to 15°C to ensure
the constant temperature in the air, regardless of ambient condi-
tions. From this point, air to each biofilter is independently
metered on mass flow control. The air flows to a co-current,
recirculated, temperature-controlled, packed, column for
humidification. A bypass permits independent control of relative
humidity. The humidified air flows to the VOC feed, where liquid
VOCs are injected and evaporated. A bypass permits adjustment
of the flow conditions at the feed injection point. The completed
feed air flows to the biofilter, which is configured for operation
in either a co- or counter-current mode. A temperature sensor in
the feed air line feeds back to control the humidifier circulation
heater. Buffer and nutrient solutions are sprayed as needed onto
the top of the biofilters.
Initially, each biofilter will be operated in different modes.
Biofilters "A" and "B" have square cross sections of a 5.74 in.
inner wall, and contain 48 in. of Corning Celcor* channelized
media. One will be operated in a co-current mode and the other
in a counter-current mode. Biofilters "C" and "D" have circular
cross sections of a 5.74 in. ID. Both will be operated in a co-
current mode. Biofilter "C" contains 48 in. of Manville Celite
BioStars* pelletized attachment media. Biofilter "D" contains
Clairtech Bioton* peat mixture.
101
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Sequential Anaerobic-Aerobic Treatment of
Contaminated Soils and Sediments
Gregory D. Sayles, Grace M. Lopez, and Dolloff F. Bishop
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
In S. Kim, Guanrong You, and Margaret J. Kupferle
Department of Civil and Environmental Engineering
University of Cincinnati
Cincinnati, OH
and
Douglas S. Lipton
Levine-Fricke Consulting Engineers
Emeryville, CA
Introduction
Many highly chlorinated aromatics and aliphatics can be
destroyed microbiologically most rapidly by sequential anaero-
bic-aerobic treatment. In general, the biochemical pathway
providing the highest rate for the initial steps of microbial
destruction of the highly chlorinated organics is anaerobic reduc-
tive dechlorination. Once partially dechlorinated, the resulting
compounds typically degrade faster under aerobic, oxidizing
conditions.
For example, polychlorinated biphenyls (PCBs) are candidate
contaminants for this sequential treatment (1,2,3). It is known
that the meta and para chlorines are removed by anaerobic
reductive dechlorination; however, the ortho chlorines are only
very slowly removed by the samebioprocess. Aerobic organisms
remove the ortho chlorine and complete the mineralization of the
compound relatively quickly. Thus, sequential anaerobic-aero-
bic treatment should provide relatively rapid destruction of
PCBs. This process should also be applicable to other highly
chlorinated aromatics, such as pesticides, and the wood preserva-
tive PCP.
The objective of this project (begun in spring 1992) is to
conduct fundamental and applied research in order to develop
sequential anaerobic-aerobic landfarming and composting tech-
nologies capable of biologically treating soils or sediments
contaminated with highly chlorinated aromatic compounds and
other low-solubility compounds susceptible to sequential treat-
ment. Because no experimental results are available yet, only the
proposed methodology is discussed below.
Methodology
The project will commence with bench-scale studies con-
ducted with aqueous slurries of soils or sediments to minimize
mass transfer limitations. Samples of soils from contaminated
sites will be screened for adapted microorganisms and from other
sources such as other research groups. Each test will be initiated
by spiking the soil or sediment with (1) contaminant of interest,
(2) co-substrate, (3) adapted microorganisms, and (4) minimal
salts media. The bench-scale studies will be conducted with the
following run-in parallel: (1) serum-bottle reactors dedicated to
soil and aqueous sampling, and (2) custom-designed flask reac-
tors dedicated to on-linepH,oxidation-reductionpotential(ORP),
and gas-production measurements. Table 1 lists the system
parameters that will be varied to provide a fundamental under-
standing and optimization of the sequential process.
Success of the anaerobic and aerobic phases of the treatment
will be determined by observing the disappearance of the parent
contaminant and following the appearance and disappearance of
daughter compounds. Rates of degradation of parent and daugh-
ter compounds as a function of time during the degradation
process will indicate optimal time for switching from anaerobic
to aerobic conditions. Degradation rates will also be correlated
with ORP.
Later, soil columns or pans simulating in situ or landfarming
techniques and lab-scale composting reactors will be employed
to demonstrate the sequential treatment technology under field-
like conditions.
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Table 1. Potential Variations in Experimental Parameters
Parameter
Compound
Co-substrate
ORP
(anaerobic)
Organism Source
Temperature
Soil
Potential Variations
DDT, other pesticides, PCBs, PCP
Pure compounds: acetate, methanol, ethanol
Complex source: primary and anaerobic sludge
Methanogenic, sulfate-reducing, nitrate-reducing
Enriched from contaminated materials, waste
treatment facilities, other research groups
Ambient (2CPC), mesophilic (3&C), thermo-
philic (70° C;
High/low organic content, sand, silt, clay
References
1. Abramowicz, D.A. 1990. Aerobic and anaerobic
biodegradation of PCBs: A review. Crit. Rev.
Biotechnol. 10:3, pp. 241-251.
2. Bedard, D.L. 1990. Bacterial transformation of poly-
chlorinated biphenyls. In Biotechnology and Biodegra-
dation, D. Kamely, A. Chakrabarty, G.S. Omenn, eds.
Advances in Applied Biotechnology Series Vol. 4.
Portfolio Pub., Co., The Woodlands, Texas, pp. 369-
388.
3. Anid, P.J., L. Nies, and T.M. Vogel. 1991. Sequential
anaerobic-aerobic biodegradation of PCBs in the river
model. In On-Site Bioreclamation, R.E. Hinchee and
R.F. Olfenbuttel, eds. Butterworth-Heinemann, Bos-
ton, pp. 428-436.
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Influence of Low Levels of Nonionic Surfactants
On the Anaerobic Dechlorination of Hexachlorobenzene
Patricia L. Van Hoof
University of Georgia
Athens, GA
and
John E. Rogers
U.S. Environmental Protection Agency
Athens, GA
Surfactants solubilize pollutants into micellar solution in the
presence of soil or sediment, effectively desorbing them from
these solids (1). This phenomenon has been suggested as a
potential tool to enhance the treatment of contaminated sedi-
ments and soils (2,3,4). The rationale is that surfactant micelles
or emulsions will solubilize precipitated or sorbed compounds,
making them more readily available for a variety of treatment
processes: biological remediation, pump-and-treat, and soil-
washing operations.
The potential of surfactants to enhance the bioavailability of
highly sorbed compounds to microbial organisms has recently
been investigated (5,6,7,8). Generally, surfactant concentrations
greater than that required for micelle formation, i.e., the critical
micelle concentration (cmc), are examined. With few excep-
tions, microbial degradation rates are slowed or completely
inhibited at these high surfactant levels. Similarly, investigations
of the effectiveness of antibacterial agents on microorganisms
demonstrate that, at high concentrations of nonionic surfactants
(0.01 to 0.1 percent w/v), the activity of the agent is suppressed.
Two possible explanations for these results are that the surfactant
is solubilizing components of the microbial membrane, or the
solute or agent is less available to the organisms. Interestingly, at
low surfactant concentrations (0.001 to .01 percent w/v), i.e.,
those less than the cmc, the activity of the agents is enhanced. To
further explore the effects of various levels of nonionic surfac-
tants on the capacity of microorganisms to degrade hydrophobic
organic compounds, the dechlorination of hexachlorobenzene
(HCB) in anaerobic sediment suspensions was examined.
Anaerobic pond sediments (5 to 7 percent w/v) were amended
with one of three nonionic surfactants: Tween 80 and 85,
polyethoxylated sorbitan monooleate, and trioleate, respectively,
and Brij 30, a polyethoxylated lauryl alcohol. Surfactant concen-
trations ranged from 3 x 10"4 percent (w/v), well below the cmc,
to 5 percent (w/v), well above the cmc. The addition of Tween 80
at concentrations less than the cmc, and just above it (<0.1
percent), enhanced the dechlorination rate of HCB compared to
unamended sediments. The enhanced activity is characterized by
a dramatic increase in 1,3,5 trichlorobenzene (TCB) after 55 to 60
days. Complete dechlorination of HCB occurs between 75 to 100
days. Generally, a similar response was observed for all surfac-
tant concentrations less than 0.1 percent. By comparison,
unamended suspensions were fairly inactive, producing only
small amounts of TCB. Sediment suspensions previously accli-
mated to HCB also demonstrated enhanced activity with low
levels of Tween 80. The more hydrophobic Tween 85 sorbed to
a greater extent, and lowered the aqueous concentration of HCB
by an order of magnitude. This reduced level of aqueous HCB
may explain why enhanced dechlorination activity in these
suspensions occurred more sporadically relative to those receiv-
ing Tween 80 amendments.
At concentrations well above the cmc for Tween 80, the
aqueous concentration of HCB was enhanced one to two orders
of magnitude over that of controls, which were generally at half-
aqueous saturation (2 to 3 |^g/L). Within a week, however, Tween
80 was degraded enough that the aqueous HCB concentration
had decreased to control levels. The HCB dechlorination activity
was initially similar to that of the controls, but after 40 days it
ceased.
Additions of Brij 30 at concentrations less than the cmc (<0.03
percent) did not alter the dechlorination activity of HCB in fresh
sediments. At concentrations greater than the cmc, aqueous
concentrations of HCB remained elevated, indicating that the
surfactant was not being degraded. However, additions of Brij 30
at these high concentrations totally inhibited HCB dechlorination
activity in both fresh and acclimated sediments.
Our results demonstrate that, although micellar concentrations
of nonionic surfactants have the capacity to transfer hydrophobic
compounds from solid surfaces to a colloidal pseudo-phase, the
addition of the surfactants suppressed, rather than enhanced, the
105
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microbial degradation of the compound. At low surfactant con-
centrations, without any apparent enhanced solubility of HCB,
dechlorination activity is increased—albeit surfactant-depen-
dent. This may be explained by fluidization of the cell membrane
by surfactant monomers resulting in enhanced mass transfer of
the solute (observed by others). Further investigation of surfac-
tant interactions with microbial membranes is required to assess
whether higher concentrations are responsible for membrane
solubilization or reduction of the more available, truly dissolved
solute. In addition, although surfactant structure was shown to be
an important factor, little is known of structure-activity relation-
ships for these systems.
References
1. Jafvert, C.T. 1991. Environ. Sci. Technol. 25:1039-
1045.
2. Roy, W.R. and R.A. Griffen. 1988. Surfactant- and
chelate-induced decontamination of soil, Report 21;
Environmental Institute for Waste Management Stud-
ies, The University of Alabama.
3. Nash, J.H. 1987. In Field Studies of In Situ Soil Wash-
ing, U.S. Environmental Protection Agency, Cincin-
nati, Ohio. EPA/600/2-87/100.
4. Vigon, B.W. and A.J. Rubin. 1989. Journal W.P.C.F.
61:1233-1244.
5. Gurin, W.F. and G.E. Jones. 1988. Appl. Environ.
Micro. 54:937-944.
6. Van Hoof. P.L. and C.T. Jafvert. 1991. Presentation at
the annual meeting of the Society of Environmental
Toxicology and Chemistry.
7. Laha, S. and R.G. Luthy. 1991. Environ. Sci. Technol.
25:1920-1930.
8. Aronstein, B.N., Y.M. Calvillo, and M. Alexander.
1991. Environ. Sci. Technol. 25:1728-1731.
106
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Anaerobic Transformation and Degradation of Chlorobenzoates and
Chlorophenols Under Four Reducing Conditions
M.M. Haggblom, J. Kazumi, M.D. Rivera, and Lily Y.Young
New York University Medical Center
Inst. of Environmental Medicine and Dept. of Microbiology
New York, NY
and
John E. Rogers
U.S. Environmental Protection Agency
Athens, GA
The anaerobic degradation of halogenated aromatic com-
pounds under methanogenic conditions is well documented.
However, there is little information on the biodegradability of
these compounds under other reducing conditions. We therefore
investigated the anaerobic biodegradation of chlorinated phenols
and benzoic acids coupled to methanogenesis, sulfidogenesis,
denitrification, and iron reduction. Anaerobic enrichment cul-
tures were established on each of the three monochlorophenol
and monochlorobenzoate isomers using sediment samples from
the East River (New York), two locations of the Hudson River
(New York), and two locations of the Nile (Egypt) as inoculum.
The anaerobic biodegradability of the chlorinated phenols and
benzoic acids depended both on the position of the chlorine
substituentand the electron acceptor available. In general, simi-
lar activities were found in all sediments, but with varying rates
of metabolism. Under methanogenic and sulfidogenic condi-
tions, 2-, 3-, and 4-chlorophenol (100 p.m) were completely
removed in 30 to 200 days. Methanogenic conditions generally
allowed for chlorophenol degradation more rapid than under
sulfidogenic conditions. Under denitrifying conditions, signifi-
cant loss was only observed with 2-chlorophenol. Iron-reducing
conditions promoted the degradation of all three mono-
chlorophenols in 90 days. To our knowledge this is the first report
of chlorophenol metabolism under denitrifying or iron-reducing
conditions.
Rapid degradation of 3- and 4-chlorobenzoate under denitrify-
ing conditions was observed in all the enrichment cultures
established with the different sediment samples, with complete
substrate loss observed in 21 days or less. Degradation of 2-
chlorobenzoate only occurred under methanogenic conditions,
with no substrate loss observed under sulfidogenic, denitrifying,
or iron- reducing conditions for over 100 days. Degradation of 4-
chlorobenzoate, on the other hand, only occurred under nitrate-
reducing conditions. 3-chlorobenzoate was the only compound
for which degradation was observed under all four reducing
conditions examined.
Degradation of 2-, 3-, and 4-chlorophenol under methanogenic
and sulfidogenic conditions could be sustained with repeated
refeeding of the respective compound. Degradation of 3- and 4-
chlorobenzoate under denitrifying conditions was also sustained
with repeated refeeding. Although metabolites of chlorophenol
or chlorobenzoate were not detected under sulfate- or nitrate-
reducing conditions, under methanogenic conditions phenol ap-
peared as a transient metabolite of 2-chlorophenol, and benzoate
was detected as a metabolite of 3-chlorobenzoate. Hence, reduc-
tive dechlorination takes place as an initial step under
methanogenic conditions.
The complete mineralization of the chlorophenols and
Chlorobenzoates coupled to methanogenesis, sulfidogenesis,
denitrification, or iron reduction can be described by the stoichio-
metric equations given below.
Chlorophenols:
1. C^OCl + 4.5 H20 — > 3.25 CH4 + 2.75 CO2 + H+ + Cl
2. C^OCl + 3.25 SO/ + 4 H2O — > 6 HCO3 +3.25 H2S + 0.5 H+ + Cl
3. C^OCl + 5.2 NO; + 4.2 H+ — > 6 CO2 + 2.6 N2 + 4.6 H2O + Cl
4. C6H5OC1 + 27 Fe(m)
Chlorobenzoates:
17 H2O — > 6 HCO3 + 27 Fe(H) + 33
5. C^OjCl + 5 H2O — > 3.5 CH4 + 3.5 CO2 + H* + Cl
6. CTHJ01C1 + 3.5 SO/ + 5 H2O — > 7 HCO3 + 3.5 H2S + H+ + Cl
7. C,H502C1 + 5.6 N03 + 4.6 H+ — > 7 CO2 + 2.8 N2 + 4.8 H2O + Cl
8. Cflpfl + 29 Fe(in) + 19 H2O — > 7 HCO, + 29 Fe(H) + 36 H+
Degradation of 2-,3-,and4-chlorophenols in the methanogenic
enrichments was coupled to stoichiometric production of meth-
ane (Equation 1). In the sulfidogenic enrichments,
monochlorophenol degradation was coupled to loss of sulfate
corresponding to that expected for complete oxidation of the
chlorophenol to CO2 (Equation 2). Under denitrifying condi-
tions, 3- and 4-chlorobenzoate were degraded coupled to sto-
107
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ichiometric reduction of nitrate (Equation 7). Analysis on
chlorophenol degradation coupled to iron reduction is in progress.
Our work indicates that chlorinated aromatic compounds can
be degraded under a variety of anaerobic conditions. The anaero-
bic biodegradability of the different monochlorophenol and -
benzoate isomers is dependent on the availability of an electron
acceptor, either carbonate, sulfate, nitrate, or Fe(III). These are
all environmentally significant electron acceptors. For example,
sulfate and iron reduction are important anaerobic processes in
marine environments, while nitrate is found as input into agricul-
tural soils. The availability and the role of these anaerobic
electron acceptors should be considered in understanding the fate
of halogenated aromatic compounds in the environment.
108
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Characterization of Biofilter Microbial Populations
John C. Loper and Alan F. Rope
Department of Molecular Genetics, Biochemistry, and Microbiology
University of Cincinnati
Cincinnati, OH
and
P.R. Sferra
U.S. Environmental Protection Agency
Water and Hazardous Waste Treatment Research Division
Risk Reduction Engineering Laboratory
Cincinnati, OH
„ ^P0"6"1,0^ research by the University of
S. Environmental Protection Agency to develop
aerobic biofilters capable of the biodegradation of volatile or-
DNA nrobeTh ^^^ «*?*** ""^ **
DNAprobetechnologieswillbeusedtocharactenzethemicro-
organisms that grow and operate as biodegradation agents in the
nSf^ri f mf
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Section Five
Modeling
A key to successful application of bioremediation technology is mathematical modeling. Modeling allows scientists
to extrapolate the quantified results of laboratory and pilot research to select the technologies predicted to be most
successful at particular sites.
Two papers presented modeling studies of granular-activated carbon (GAC). The first study further investigated a
previously observed phenomenon, the major effect of molecular oxygen on the adsorptive capacity of GAC for aromatic
organics. The second study is part of continuing research to evaluate quantitative techniques to characterize biofilms
attached to GAC. These biofilms have proven highly effective for treating hazardous waste, but the treatment process
has not been completely measured. The objective of another modeling project is to quantify biodegradation kinetics of
certain organic compounds under anaerobic conditions. The investigators then intend to mix the compounds with sand,
soil, and more complex matrices to measure the resulting products of biodegradation.
111
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Prediction of GAC Adsorptive Capacity With and Without
Molecular Oxygen
Radisav D. Vidic and Makram T. Suidan
University of Cincinnati
Cincinnati, OH
and
Richard C. Brenner
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
Previously reported results by the authors (1,2) revealed that
the presence of molecular oxygen has a major effect on the
adsorptive capacity of GAC for several aromatic compounds
(phenol, o-cresol, and 3-ethylphenol), as well as for natural
organic matter. The adsorptive capacity exhibited in the presence
of molecular oxygen can, in some instances, be up to threefold the
capacity that was attainable under anoxic conditions. Experi-
mental results have demonstrated that this increase in GAC
adsorptive capacity was not due to biodegradation of these
organic compounds in the presence of molecular oxygen (oxic
conditions). It was also discovered that the presence of molecular
oxygen promotes polymerization of these compounds on the
surface of the carbon, which provides an explanation for the
observed phenomenon. Polymerization of phenolic compounds
is achieved through oxidative coupling reactions (3).
The effect of molecular oxygen on the adsorptive capacity of
GAC is now further investigated for several phenolic and other
aromatic organic compounds listed in Table 1. These compounds
were chosen to study the effect of different functional groups
substituted on the parent phenol molecule, as well as to represent
a variety of recalcitrant compounds of special interest in waste-
water treatment.
The effect of the same functional group, substituted on differ-
ent positions of the parent phenol molecule on GAC adsorptive
capacity exhibited under oxic and anoxic conditions, was inves-
tigated for 2-, 3-, and 4-methylphenol. The absence of molecular
oxygen from the test environment yielded very similar GAC
adsorptive capacities for all three compounds. On the other hand,
the substitutional position of the methyl functional group had a
significant impact on the adsorptive capacity of activated carbon
that was exhibited in the presence of molecular oxygen. Oxic
GAC adsorptive capacity for these compounds increased in the
order: 3-methylphenol < 4-methylphenol < 2-methylphenol,
indicating that substituting the methyl group on the meta position
of the parent phenol molecule induced the least susceptibility to
this polymerization reaction.
Table 1. Organic Compounds Chosen for This Study
2-Methylphenol
3-Methylphenol
4-Methylphenol
2-Ethylphenol
3-Ethylphenol
4-Ethylphenol
2-Propylphenol
3-Propylphenol
4-Propylphenol
2-Nitrophenol
3-Nitrophenol
4-Nitrophenol
2-Chlorophenol
3-Chlorophenol
4-Chlorophenol
2-Hydroxybenzoic Acid
3-Hydroxybenzoic Acid
4-Hydroxybenzoic Acid
Catechol
Resorcinol
Hydroquinone
2,4-Dichlorophenol
2,4,6-Trichlorophenol
Pentachlorophenol
Benzene
Toluene
Nitrobenzene
Chlombenzene
2,4-Dinitrotoluene
Indole
Quinoline
Methylquinoline
Aniline
Pyridine
2-Naphthol
Extraction experiments conducted with carbons used in oxic
and anoxic isotherm experiments with these compounds yielded
90 percent recovery of the original compound when the adsorp-
tion occurred under anoxic conditions. On the other hand, the
highest amount of adsorbate was extracted from the GAC used in
the test with 3-methylphenol, which was the compound that
showed the least increase in the GAC adsorptive capacity in the
presence of molecular oxygen.
Anoxic adsorption isotherms forthreeott/zo-substitutedphenols
(methyl, ethyl, and propyl groups) are compared in Figure 1. The
increase in molecular weight of a substituted functional group
induced an increase in the anoxic GAC adsorptive capacity since
113
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the higher molecular weight compounds are usually more hydro-
phobic. However, the relative increase in the adsorptive capacity
in the presence of molecular oxygen decreased as a result of an
increase in the molecular weight of the substituted functional
group (see Figure 2). This behavior is attributed to the stearic
hindrance of the polymerization reaction. Further as the molecu-
lar weight of adsorbate increased, the extraction efficiency ob-
tained for the carbons used in the oxic isotherm tests increased.
This confirms the hypothesis of the diminished polymerization
reaction for the higher molecular weight adsorbates.
Ongoing adsorption isotherm experiments are designed to
evaluate the effects of the solution pH, the G AC type, and the size
of the GAC particle on the extent of the adsorptive capacity
increase that results in the presence of molecular oxygen. Fur-
thermore, attempts will be made to relate the observed phenom-
ena to chemical properties of adsorbates listed in Table 1.
References
1. Vidic, R.D., M.T., Suidan, U.K., Traegner, and G.F.,
Nakhla. 1990. Adsorption isotherms: illusive capacity
and role of oxygen. Water Research 24:10, pp. 1187-
1195.
2. Vidic, R.D. and M.T., Suidan. 1991. Role of dissolved
oxygen on the adsorptive capacity of activated carbon
for synthetic and natural organic matter. Environ. Sci.
Technol. 25:9, pp. 1612-1618.
3. Musso, H., W.I. Taylor, and A.R. Battersby, eds. 1967.
Phenol coupling. In Oxidative Coupling of Phenols.
Marcel Dekker, New York, New York.
103-
102
101
J—I 1—U
2-Methylphenol
2-Ethylphenol
2-Propylphenol
' 1 L
J L.
100
102
103
C, mg/L
Figure 1. Anoxic adsorption isotherms for three ortho-substituted phenols.
114
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3.0
2.5
o
'x
o
'§
1.5
1.0
J 1—i
10o
-I " 1—I r-
2-Methylphenol
2-Ethylphenol
2-Propylphenol
J 1 1—i—i—i—i—i—i T"i-
102
C, mg/L
Figure 2. Ratio ofoxic andanoxic GAC adsorptive capacity for three ortho-substituted phenols.
103
115
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Development of Tools for Monitoring Biofilm Processes
Steven I. Safferman
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH
and
Sandrine Baud-Grasset
International Technology Corporation
U.S. Environmental Protection Agency Test and Evaluation Facility
Cincinnati, OH
The engineering design, process control, and evaluation of
microbial fixed-film processes can be challenging because of the
lack of methods to measure biomass attachment characteristics.
A particularly troublesome biofilm process to monitor, yet one
that is very effective for treating hazardous waste, employs
biological granular-activated carbon (GAC), i.e., biofilms at-
tached to GAC. The adsorptive characteristics of the GAC and its
organic nature preclude using many traditional techniques. The
objective of this continuing research is, therefore, to identify and
evaluate a variety of unique quantitative and semi-quantitative
techniques to characterize biofilm attachment, with emphasis on
biological GAC.
Based on a literature survey and consultation with researchers
in various disciplines, several techniques with potential to achieve
this project's objectives have been identified. These include a
direct measure of biomass and biological activity, semi-quanti-
tative visual assessments of attachment characteristics, and simple
physical characteristics of the media affected by biological
activity. Methods that have been eliminated include chemical
oxygen demand, total organic carbon, washing of the GAC
particle followed by enumeration of the microbial population in
the wash water, volatile solids analyses, and direct microscopic
counting.
All cell walls contain phospholipids and, consequently, the
analysis forphosphorus from an extract can be directly correlated
to the quantity of viable biomass. The application of existing
techniques to GAC required some modifications but are effec-
tive. The variability of the test results and the effect of sample size
are currently being evaluated.
A simple method to measure activity utilizes fluorescein
diacetate (FDA). Fluorescein is released from FDA as the acetate
is utilized as a substrate by the microbial populations. As fluores-
cein is released, the solution changes color as a function of the
microbial activity. The GAC will, however, adsorb fluorescein
and corrections are required to utilize this technique for biologi-
cal GAC. Two important test variables are the size of the sample
and the incubation time, both of which will affect the variability
and sensitivity of the results.
The results from the FDA and lipid measurements are best
interpreted in conjunction with one another. The ratio of FDA to
lipid provides a measure of the activity per unit biomass. Conse-
quently, changes observed in activity per unit biomass can signal
recent changes in the microbial ecology within a reactor.
The three-dimensional image of a specimen's surface revealed
by a scanning electron microscope (SEM) allows direct visual
examination of the attachment characteristics of biomass at-
tached to GAC. SEM examination of several biological GAC
processes has revealed causes of reactor operating difficulties
and provided insight to attachment characteristics. The technique
has been qualitative to date, however, and comparisons of speci-
mens or changes in attachment characteristics with time can be
subjective. A semi-quantitative technique has been developed as
partof this research. A rating system was developed that provides
a framework to evaluate important fluidized-GAC bed reactor
characteristics, including the type and quantity of microorgan-
isms attached to the GAC and the physical characteristics of the
GAC. Verification of the protocol is ongoing and involves the
independent evaluation of GAC samples by several researchers
to determine the extent to which evaluation techniques are
objective.
The media's physical characteristics, i.e., porosity, settling
velocity, and GAC diameter, are all easily measured by simple
laboratory techniques. These parameters are not direct measure-
ments of biological conditions but changes in the physical
properties of the media caused by biological activity. The meth-
ods do not appear to be sensitive to slight changes in biological
117
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conditions; however, they may be useful for designing certain
biological GAC processes.
To develop, evaluate, and compare these biomass monitoring
tools, bench-scale GAC-fluidized beds are being utilized. The
SEM analyses have been found to be effective in monitoring
developing biofilms but the lipid and FDA measurement do not
appear to be sensitive when using reasonable sample sizes during
this developmental stage. However, once a dense biofilm coats
the GAC particles, the lipid and FDA analyses become sensitive
and can discriminate between varied operating conditions. The
SEM is also useful in identifying the general category of micro-
organism present and in trouble-shooting reactor-operational
difficulties.
Future research should focus on assessing the variability and
sensitivity of the techniques described above and determining
their ability to predict the microbial environment within biologi-
cal GAC as well as within other types of media. The use of the
FDA and lipid measurements would also appear to be well suited
for use in mathematical models as a quantitative biomass activity
measure. Replacement of more typical biomass measures with
these measures should be evaluated to determine if improved
predictions are obtained.
Acknowledgement
The authors would like to express their appreciation to Paul L.
Bishop, Kim A. Brackett, Patrick J. Clark, andRobie J. Vestal for
their past and current consultation on this research.
118
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Evaluation of Anaerobic Respirometry to Quantify Intrinsic
Anaerobic Biodegradation Kinetics of Recalcitrant Organic Compounds
Jiayang Cheng, Makram T. Suidan, and Albert D. Veosa
University of Cincinnati
Cincinnati, OH
The objective of the study is to determine the kinetics of
biodegradation of compounds difficult to degrade under strict
anaerobic conditions using a standard respirometer specially
modified for anaerobic operation.
A 12-position N-Con respirometer has been purchased to
conduct the studies. The respirometer, which was originally
designed for use under aerobic conditions, has been modified by
researchers at New Mexico State University to be functional
under anaerobic conditions. Instead of connecting the reaction
vessel to an oxygen cylinder under positive pressure, this vessel
is connected to a vacuum pump under a negative pressure of
approximately 5 psig. As microbes degrade the substrate and
produce gas, the increase in pressure is sensed by the pressure
sensor, which signals the computer to open the exit valve venting
the gas from the vessel headspace. Before release to the sensor,
the product gas is passed through a dryer and a CO2 scrubber. The
total volume of gas exiting the vessel is computed automatically
and recorded on the computer disk for determination of biodeg-
radation kinetic constants. This operating principle can be used
to quantify the biokinetic constants of recalcitrant compounds
such as pentachlorophenol (wood preserving wastes) and 2,4-
dinitrotoluene (ammunition production wastes).
A simple compound such as phenol dissolved in distilled water
will be selected first to test the concept and gain experience with
the operation. Liquor from a conventional anaerobic digester
will serve as the inoculum. Once the operation has been mas-
tered, more recalcitrant compounds such as pentachlorophenol
or 2,4-dinitrotoluene will be investigated in a clean water me-
dium. Sediments and soils contaminated with PCP will be used
as the source of inoculum.
When the anaerobic biokinetic constants have been deter-
mined for the test compounds, a degree of complexity will be
introduced by adding these compounds to a sand matrix and
determining their anaerobic biodegradability while sorbed onto
sand particles. More complexity will be introduced later by using
a loam or clay soil matrix. Finally, mixtures of compounds will
be tested to determine biodegradability in a complex mixture.
Variables to be measured in addition to total gas production
include CO2, CH4, disappearance of parent compound by chro-
matographic assay, COD, andpH. An important control variable
will be temperature. Radio-labeled compounds also will be
studied to determine the fate of the carbon from the parent
compound.
119
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