Effects of Metals on the Reductive Dechlorination
of Chlorophenols

Jack Jones
U.S. Environmental Protection Agency
Athens, GA

and

In Chul Kong
Technology Applications, Inc.
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. Asamodel 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 were examined using unadapted
and chlorophenol(CP)-adapted freshwater sediment slurries (pH
7.0) in the presence and absence of added metal salts (CuCL,
CdCl, K,Cr,0,). The time required for unadapted control sedi-

55

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 a20 ppm concentration of
the chloride salts of Cu(II) and Cd(ID) 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 100 ppm (or greater) ‘Cr(VI), as
K,Cr,0,, caused total inhibition of dechlorination of 2-and3-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. and 200 ppm Cd(I]) concentrations, respectively. Addition
of CuCl, 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(II), 0 to 30 ppm for Cd(II), and
0 to 28 ppm for Cr(V1) over the range of metal salts added (20 to
200 ppm).

The effects of metal salts (20 to 200 ppm of Cu(II), Cd(I1)) 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


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
Tesearchers 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 adryerandaCO, 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).

119

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 CO,, CH,, disappearance of parent compound by chro-
matographic assay, COD, and pH. 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,

5 U.S. GOVERNMENT PRINTING OFFICE:1992-648.003/60054


Section Two
Bioremediation Field Initiative

The Bioremediation Field Initiative is one of the major research components of EPA’s Bioremediation Research
Program. The initiative was undertaken in 1990 to expand the nation’s field experience in bioremediation techniques.
Objectives of the initiative are to more fully document the performance of full-scale applications of bioremediation;
to provide technical assistance to site managers; and to provide regular information on treatability studies, design, and
operation of bioremediation projects, including cost and performance data. The initiative is currently tracking
bioremediation activities at more than 130 Superfund sites, RCRA corrective action facilities, and Underground
Storage Tank sites nationwide.

Full-scale performance evaluation is the subject of several recent projects. At a Montana Superfund site, researchers
studied the bioremediation of contaminated ground water and soil both in situ and in aboveground facilities,

EPA researchers reported results of the first 6 months of a 2-year field study of bioremediation using bioventing
begun in 1991 at two sites in differing climatic zones: near Fairbanks, Alaska, and near Salt Lake City, Utah.
Bioventing is the process of supplying oxygen in situ to soil microbes to stimulate biodegradative activity with the
goal of destroying toxic compounds in the ground.

A treatability study and design assessment examined the use of nitrate for in situ bioremediation of an aquifer
contaminated with petroleum hydrocarbons, Results showed that remediation is controlled by the rate of application
of nitrates and microbial reaction rates, Researchers recently studied a method for delivering methane and oxygen
nutrients directly to a microbial Population for in situ bioremediation of an aquifer contaminated with trichloroeth-
ylene. Another recent project involved a full-scale site characterization and treatability study of trichloroethylene in
@ ground-water plume and the potential for enhanced anaerobic rather than aerobic biological transformation,

A demonstration project at a confined treatment facility (CTF) is simulating field conditions on the rate and extent
of polychlorinated biphenyl (PCB) biotransformation under varied biotreatment scenarios. In another Project,
Tesearchers are characterizing the nutritional, physiological, and environmental factors that enhance the growth of
desired microbial populations for dechlorinating hazardous wastes. EPA is also investigating the effects of toxic heavy
metals on the anaerobic biotransformation of organic compounds.

21




3.0 : : 7 —
2-Methyiphenol
2-Ethylphenol 4
2-Propylphenol 4

oxic capacity/anoxic capacity

©, mg

Figure 2. Ratio of oxic and anoxic GAC adsorptive capacity for three ortho-substituted phenols.

115


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

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, andJ.H.
Sammons. 1989. Mobilization of aviation gasoline from
aresidual source. Research Journal WPCF, Vol. 61, pp.
1684-1690.

venting by a field-scale system has been reported to be roughly 2.  Kampbell, D.H. and J.T. Wilson. 1991. Bioventing to
$50 per ton (U.S. EPA Report 540/A5-89/003, 1989), while treat fuel spills from underground storage tanks. Jour.
incineration costs are more than 10 times this amount. A cost Haz. Materials 28, pp. 75-80.
estimate of about $15 per yd’ sandy soil for bioventing treatment
ataJP-4 jet fuel contaminated site has been reported by Vogel (4). 3. Twenter,F.R.,T.R. Cummings, and N.G. Grannemann.
1985. Ground-water contamination in East Bay Town-
Total costs incurred for the pilot-scale bioventing systems over ship, Michigan. U.S. Geological Survey Water-Re-
the 6-month period for construction, operation, field monitoring, sources Investigations Report 85-4064, Lansing, Michi-
and sample analyses were near $147,000. Expansion of the gan.
injection well grid from 3,000 ft? to 60,000 ft? for a field-scale
system would increase construction costs about $69,400. Less 4, Vogel, C. 1991. Enhanced in situ biodegradation of
extensive field monitoring and sample analyses could be insti- petroleum hydrocarbons through soil venting. Techdata
gated to offset the increase in construction cost. Maintaining total RDV 91-7, AirForce E&S Center, Tyndall AFB, Florida.
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 yd° of contaminated
subsurface.
Table 4,  Ground-Water BTEX
Sample Wall depth Benzene (mg/L) Total BTEX (mg/L)
January 1991
MWw2N 17 5 1,018
MW2N at <6 440
mwes 18 <5 9
Mwes 20 43 1,028
Total <43 $2,495
April 1991
wen 17 <5 149
MW2N 21 174 492
mwes 18 <6 39
MWw2S 20 7 444
Total $181 $1,124
October 1991
mw2N 17 <6 51
MW2N 2t 5 159
Mw2S 18 <5 19
MWw2S 20 6 31
Total so $260

64


Fundamental Studies on the Treatment of VOCs in a Biofilter

R. Govind, V. Utgikar, Y. Shan, W. Zhao
Department of Chemical Engineering
University of Cincinnati
Cincinnati, OH

and

DF. Bishop, S.I. Safferman, G.D. Sayles
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, OH

The Superfund Amendments and Reauthorization 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, trichloroethylene
(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
atthe 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 biofilter packed with ceramic
pellets (celite, Manville Corporation, California). Complete deg-
tadation of the five compounds was obtained in the experiments.
The TCE concentration was reduced by about 35 percent.

ri)

The third study on the five compounds describes the perfor-
mance of a biofilter 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 biofilter. 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 biofilter 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 biofilter 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.


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 otherrecalcitrantcreosote Constituents as sole sources
of carbon and energy for growth (3, 5, 6). In addition, an axenic
Culture of Pseudomonas 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-

- Bioreactor 2
Bioreactor | (2a, b, and c)
CSTR BATCH
Ground-water
feed Ground-water giuent
equilization/ LA > re
or feed tank Tank
Membrane
Concentrate indigenous microbes strains
or strains CRE1-13 EPAS505,
and SR3
. Effluent
Discharge —€ A polishing [~ he
A = Sampling points

Figure 1. Ground-water bioremediation system.

1






82

Table 2. Phenanthrene In The Libby Soil Treatment Facility (mg/kg)

Sampling Date 9/18/91



Development of Small-Scale Evaluation Techniques for
Fungal Treatment of Soils

Steven J. 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 concentations.
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


References
1. RE. 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 Jn Situ Bioreclamation, R.E. Hinchee
and R. F, Olfenbuttel, eds. Butterworth-Heinemann,
Boston, pp. 541-545.


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 nottoapply 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 AC1100
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 AC1100 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 AC1100 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.

60

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,
and A.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., DJ. Schlemm, R.P. Lyon Ill, 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 organicacidsandalcohols. Applied Environ.
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.


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 isespecially 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. Air Force at Eielson Air Force
Base (AFB) near Fairbanks, Alaska. The site has JP-4 jet fuel-
contaminated unsaturated soil where a spill occurred in associa-
tion witha fuel distribution network. The contractor operating the

31

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

AtEielson 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 O,, and 10 to 18 percent co,,
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 O, and 0.03 percent CO,,.


Table 1. Soil Gas Hydrocarbons (mg/L)
North Plot South Plot
Date One Meter Three Meters One Three meters
10/22/90 1,940 2,060 4,220 2,960
10/29/90 860 910 1,960 1,210
11/12/90 430 490 760 550
oot 180 200 200 260
04/04/91" 400 990 600 1,580
04/18/91 310 820 480 1,160
05/02/91 78 52 66 38
08/08/91 32 40 20 20
10/07/91 16 16 14 14
* System started after 3-month shutdown for winter.
Table 2. Fuel Carbon Mass in North Plot (mg/tt?)
09/25/90
Rept Rep2 x
Avove water table 43,640 54,020 48,830
Below water table 221,630 232,800 227,220
10/16/91
Rep 1 Rap 2 Rep 3 x
Avove water table 253 172 480 302
Below water table 147,980 234,390 150,360 177,580
Table 3. Fuel Carbon Mass in South Plot and South Control (mgft?)
South Plot
Rep1 Rep2 x
Avove water table 37 27 32
Below water table 105,335 62,502 83,920
South Control
Rep 1 Rep2 x
Avove water table 12,090 4,510 8,300
Below water table 54,616 103,910 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 toremediate
the unsaturated test area subsurface contamination of aviation

gasoline. Surface emissions, once acclimation was attained 1
monthafter system startup, did notexceed 104g TPH/L and were
less than 1 yg 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 the October 1991 water-table level. Final benzene concentra-
tions were less than 5 pg/L compared to initial levels up to 133
Hg/L in the underlying ground water near the water table. Large


Pre-Screening |...
No potential
Potential fungal strains
--| Sensitivity |... -! Soil Bending:
No potential
Potential fungal strains
Growth
Strains with best growth potential
Fate Soil Pan
Data Evaluation
Figure 1. Tiered treatability 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
compoundsare 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

100

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 paramountas the protocol is notintended 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 torecognize 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.




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 pg/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
Tate was equivalent to that observed in the clean-aquifer micro-
cosms, the actual nitrate demand was at least four times as high

4

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). Itis 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-m? area that is to
receive ground-water recharge containing 10 mg/L NO,-N at an
application rate of 680 m’/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
tate. 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 Treatment — Uncontaminated Contaminated

BTEX Nitrate only = 0.248
Nitrate plus BTEX 1.65 0.275

Nitrate-N Nitrate only 2.80 2.92
Nitrate plus BTEX 2.80 3.06



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 Phe1. One reactor received air-entrained TCE along with
a continuous inflow of nutrient medium. Strain G4 (the non-
constitutive parent of G4 5223 Phe1) was unable to affect TCE
concentration in such a vapor phase bioreactor (100 percent re-
feed), whereas Phe (the constitutive TCE degrader) completely
removed the 80 um 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 um (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.
Ineither 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
inducer can be maintained. The use of such organisms in a
bioreactor requires sensitive balancing of the inducer ‘chemical,

83

growth substrate (if different from the inducer), TCE, and oxy-
gen. The simplifications inherent to a bioreactor colonized bya
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
(2,34).

Because P. cepacia G4 5223 Phel does not use TCE asa
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 10® bacteria gram".

The toxicity of TCE andits breakdown productsin the bioreactor
were also analyzed. Tests with Phel indicated a toxicity due to
the metabolism of TCE at concentrations above 500 um (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
G4 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 um
(approximately 6,000 ppb) and maintained this concentration
throughout the 20.3-hour test period, the G4 5223 Phel-contain-
ing column never achieved areturn TCE concentration above 13
um (approximately 1,700 ppb) at the initial breakthrough, and
declined to undetectable levels by 20.3 hours. In addition, an




Table of Contents

Page

Executive Summary
Introduction

Section One: Site Characterization

Site Characterization at Pipeline Spill at Park City, Kansas:
Estimating Hydraulic and Geochemical Constraints on Bioremediation ....

Field Demonstration of Innovative Bioremediation Strategies for
PCP and Creosote ....

Case Study on Site Characterization at a TCE Plume: St. Joseph,
Michigan, NPL Site

Approaches to the Characterization of Trichloroethylene Degradation by
Pseudomonas cepacia G4 .....

Section Two: Bioremediation Field Initiative ....

Progress in the Field Applications of Bioremediatiot

Evaluation of Full-Scale /n Situ and Ex Situ Bioremediation
of Creosote Wastes .....

Optimizing Bioventing in Shallow Vadose Zones and Cold Climates:
Eielson AFB Bioremediation of a JP-4 Spill

Optimizing Bioventing in Deep Vadose Zones and Moderate Climates:
Hill AFB Bioremediation of a JP-4 Spill......

Use of Nitrate to Bioremediate a Pipeline Spill at Park City, Kans:
Projecting from a Treatability Study to Full-Scale Remediation...

Design and Treatability Study of /n Situ Bioremediation of Chlorinated
Aliphatics by Methanotrophs at St. Joseph, Michigan

Natural Anaerobic Treatment of a TCE Plume:
St. Joseph, Michigan, NPL Site .....

Proposed PCB Biodegradation Study at the Sheboygan River Confined
Treatment Facility ...



@ _Ground-water monitoring well

CW = Soil vapor cluster well

A-A’ = Cross section trace

(1.5) = TPH in ground water (mg/L)(9/91)

50 feet

Figure 1. Plan view of the joint U.S, EPA and U.S. Air Force bioventing activities at Hill AFB, near Salt Lake City, Utah. CW are 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

38

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


Section Four
Process Research

Process research involves isolating and identifying microorganisms that carry out biodegradation processes and
developing new biosystems for treatmentof environmental pollutants. Although the majority of conference presentations
dealt with laboratory-scale work, itis 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 ata 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
fora 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


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 overa2-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.

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
chrysosporium: 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., MJ. Larsen, and T.K. Kirk, 1990. Sensi-

tivity to and degradation of pentachlorophenol by
Phanerochaete spp. App. Environ. Microbiol. 56:3519-

3526.
Table 1. Experimental Design for Fungal Treatment Effectiveness
Treatment Fungus/Control Inoculum Loading Plots
1 P. chrysosporium 5% B
2 “ 10% CHKV,X
3 P. sordida 10% E
4 P. chrysosporium/ 5% F
T. hirsuta 5%
5 Inoculum carrier control 10%
6 No treatment control -
7 T. hirsuta 10%
8 P. chrysosporium 13% I,
9 “ 10% (Day O) J,Q,U,Z
3% (Day 14)
10 Wood chip control - 1MO,W
1000
[ No treatment control
800 }
600

PCP (mg/kg)

400

200

Carrier control

Phanerochaete sordida

20

Figure 2. Pentachlorophenol depletion by Phanerochaete sordida.

30
Time (days)

78


United States Office of Research and EPA/600/R-92/126

Environmental Protection Development August 1992
Agency Washington, DC 20460

SEPA Bioremediation of
Hazardous Wastes

»> 2D


Table 1. Potential Variations in Experimental Parameters

Parameter Potential Variations

Compound DDT, other pesticides, PCBs, PCP

Co-substrate Pure compounds: acetate, methanol, ethanol
Complex source: primary and anaerobic sludge

ORP Methanogenic, sulfate-reducing, nitrate-reducing

(anaerobic)

Organism Source Enriched from contaminated materials, waste
treatment facilities, other research groups

Temperature Ambient (20°C), mesophilic (35°C), thermo-
philic (70°C)

Soil Highfow organic content, sand, silt, clay

References

104

1.

Abramowicz, D.A. 1990. Aerobic and anaerobic
biodegradation of PCBs: A review. Crit. Rev.
Biotechnol. 10:3, pp. 241-251.

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.

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
RF. Olfenbuttel, eds. Butterworth-Heinemann, Bos-
ton, pp. 428-436.


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.


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 agents 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 laury] alcohol. Surfactant concen-
trations ranged from 3 x 10* 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

105

percent), enhanced the dechlorination rate of HCB compared to
unamended sediments. The enhanced activity is characterized by
adramatic increase in 1,3,5 trichlorobenzene (TCB) after 55 1060
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 to3 g/L). Withina 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


Proposed
corer > 1?
injection
system Contaminant
Pilot plume
injection
system Waste pit area
Infiltration trench
Bioreactor
housing
Rock
fiter
LTu LTu
Regional
round-water
Scale (feet) NY flow
0 300 600
©} Monitoring well C1 Injection well
© Extraction well

Figure 1. _ In situ bioremediation of upper aquifer (trom 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 achemically 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
cone 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.

26

Land Treatment Units

Total estimated contaminated soil volume for treatment is
45,000 yd? (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 yd? of
contaminated soil, and a2-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 sampling 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 #1 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
#1, and lift#1 was placed in cell #2. A third sampling eventin late


3.

Shields, M.S., S.O. Montgomery, $.M. Cuskey, P.J.
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,

Shields, M.S., S.0. Montgomery, P.J. 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


A B
1.0
09
08
07
—s— live —1— live
07 —e— killed control 06 —e— killed control
—— chemical control —m— chemical control
05
o6-—J T_T TT T_T rr rt
° 10 20 30 40 0 10 20 30 40
150
14 aI D
1.0 125
09
3 100
08 —t— live §
—e— killed contro! 2
07 —m— chemical control] 2 live
5
8 50 —e— Killed control
06 8 —a— chemical control
os 25 HexaCDD formation
04 tt 0
0 10 20 30 40 0 10 20 30 40
Time (weeks)

Figure 1. Fate of H,CDD (A), H,CDDi (B), and H,CDD (C) in Hudson River sediments, and product formation in H, CDD-spiked microcosms (D).
The values for the C/Co plots represent averages of thrae and two replicates for active microcosms and killed or chemical controls,

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 mg/kg PCDD/F, but data
are not available at the time of this report.

References
1. Brown, JF., Jr., D.L. Bedard, MJ. Brennan, J.C.
Carnahan, H. Feng, and RE. Wagner. 1987. Polychlo-
rinated biphenyl! dechlorination in aquatic sediments.
Science 236:709-712.

2. — Quensen, J.F. Ill, S.A. Boyd , and J.M. Tiedje. 1990.
Dechlorination of four commercial polychlorinated bi-
phenyl mixtures (Aroclors)® by anaerobic microorgan-
isms from sediments. Appl. Environ. Microbiol.
56:2360-2369.

80

3. Alder, A.C., M. Hiiggblom 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® 1242inanaero-
bicsediments. Appl. Environ. Microbiol. 56:2612-2617.

5,  Adriaens, P. and D. Grbié-Galié. 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. Bichelberger, T.A. Bellar,
and W.L. Budde. 1986. Determination of chlorinated
dibenzo-p-dioxins and dibenzofurans in environmental
samples using high resolution mass spectrometry. EPA
Report: Physical and Chemical Methods Branch.


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-
bonisroughly 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)
0 (feet) 9 3000 6000
o— 1 1
x 5—
Se |_ =
Sit] |_ ys
T|- 2
Water
table | |— 25—
Cla — 30
z SE] 35 —
bee ‘Sand id Wee
fsés7 Sand and gravel Al go —

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


5. Freedman, D.L. and J.M. Gossett. 1989. Biological 7. Major, D.W., W.W. Hodgins, and B.J. Butler. 1991.
reductive dechlorination of tetrachloroethylene and Field and laboratory evidence of in situ biotransforma-
trichloroethylene to ethylene under methanogenic con- tion of tetrachloroethene to ethene and ethane at a
ditions. Appl. and Environ. Microbiol. 55(9):2144- chemical transfer facility in North Toronto. In On-Site
2151. Bioreclamation, R.E. Hinchee and R.F. Olfenbuttel,

eds. Butterworth-Heinemann, Stoneham, p. 147-171.

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 8. Keck Consulting Services, Inc. Hydrogeological Study
culture inthe absence of methanogenesis. Appl. Environ. Report. Section 10, Lincoln Township, Berrien County,
Microbiol. 57(8):2287-2292. Michigan, June 1986.
Table2. 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 @ TCE 1,1DCE eDCE tDCE ve Ethene CH, Total
1-2-70° mg/l 4.03 0.09 4.14 0.81 319 6.62 4.61
Equiv. COD
Decrease, mg/l 0.00 0.01 0.66 0.13 0.41 11.32 18.44 30.97
% of Equiv. COD 0.0 0.0 at 0.4 1.3 36.6 59.5
1-3-75° mg/l 12.80 0.27 16.90 0.67 5640 2.25 6.62
Equiv. COD
Decrease, mg 0.00 6.04 2.70 0.17 28.80 3.84 26.00 61.5
% of Equiv. COD 0.0 0.1 44 0.2 46.8 6.2 42.2
21-75" mg/l 0.44 0.09 13.40 0.21 1.46 3.15 7.43
Equiv. COD
Decrease, mg/ 0.00 0.01 2.14 0.03 0.74 5.39 29.72 38.03
% of Equiv. COD 0.0 0.0 5.6 Ot 1.9 14.2 78.1
2-6-65" mg/t 0.51 4.70 0.03 2.66 4.27 11.72
Equiv. COD
Decrease, mg! 0.00 0.00 0.75 0.00 1.36 5.98 46.88 54.97
% of Equiv. COD 0.0 0.0 1.4 0.0 25 10.9 85.3
3-2-80" mg/h 2.53 0.01 0.90 0.03 0.27 4.87 11.59
Equiv. COD
Decrease, mg/l 0.00 0.00 0.14 0.00 0.14 8.32 46.36 54.96
% of Equiv. COD 0.0 0.0 0.3 0.0 0.3 15.1 84.4

@ First value is transect number, second value is borehole number, and third value is depth of sample below ground surface in feet.

50


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.

Stull, D.R., E.F. Westrum, Jr., and G.C. Sinke. 1969.
The chemical thermodynamics of organic compounds.
John Wiley and Sons, New York.

Reid, R.C., J.M. Prausnitz, and B.E. Poling. 1985. The
properties of gases and liquids. McGraw-Hill Book Co.,
New York.

92


Evaluation of Enhanced In Situ Aerobic Biodegradation of CIS- and
Trans-1-Trichloroethylene and CIS- and Trans-1,2-Dichloroethylene by

Phenol-U

ing 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 7 —
4 tDCE
e-DCE
08 4 ZY
TCA
3g _ TCE
3
= 065
3
8
8 4
e
i=}
§ 044
3
& J —
024
0.0 =

Phenol

Toluene

Ammonia

Figure 1, Relative fractions of CACs degraded in the ground-water microcosms following stimulations with different primary growth substrates.

71


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). Airis injected from 2 to 6 ft below grade at arate of 2.5
ft?/min/well, spaced 30 ftapart. 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 ftbelow

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 warmed 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 ofa portion of theactively warmed plotis shown in Figure
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

Passive
Control N
° ° °
||
° [@)
Oo
mo 8 @ Ground-water monitoring well
a rn
© Air injection’withdrawal well
Active
aA &  Three-level soil gas probe
I Three-level thermocouple probe
Site trailer
Scale
a
Active 0 25! 50"
warming
system

Figure 1. Plan view of the joint U.S, EPA and U.S. Air Force bioventing activities at Eielson AFB, near Fairbanks, Ak.

32


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-

R(m)

25 75

© PN g2aen a

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


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 mera position
of the parent phenol molecule induced the least susceptibility to
this polymerization reaction.

113

Table 1. Organic Compounds Chosen for This Study
2-Methylphenol Catechol
3-Methylphenol Resorcinol
4-Methylphenol Hydroquinone
2-Ethylphenol 2,4-Dichlorophenol
3-Ethylphenol 2,4,6-Trichlorophenol
4-Ethylphenol Pentachlorophenol
2-Propylphenol Benzene
3-Propyiphenol Toluene
4-Propylphenol Nitrobenzene
2-Nitrophenol Chlorobenzene
3-Nitrophenol 2,4-Dinitrotoluene
4-Nitrophenol

Indole
2-Chlorophenol Quinoline
3-Chlorophenol Methylquinoline
4-Chlorophenol

Aniline
2-Hydroxybenzoie Acid Pyridine
3-Hydroxybenzoic Acid 2-Naphthol
4-Hydroxybenzoic Acid

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 for three ortho-substituted phenols
(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


Table 1.

Half Reactions Indicating Electron Equivalents of Change and Associated. Equivalent COD Decrease Associated with Change

Equiv. COD

decrease
Transform Mol. g COD/g
Product wt Half reaction Product*
Methane 16 CO, + 8H + 8@ > CH, + 2H,0 4
DCE 97 CHCI=CCI, + Ht + 26 > CHCl=CHCI + CI- 0.16
ve 62.5 CHCI=CCI,+ 2H* + de > CH,=CHCI + 2CI- 0.51
Ethene 28 CHCI=CCl, + 3H* + 62 > CH,=CH, + 3CI- 1.77

* From [ O, + 4H" + 4e"~> 2H,0 ], the half-teaction for oxygen, one electron eq
COD Decrease = (8rvMol. Wi), where n is the number of electrons in the halt.

References

1.

McCarty, P.L. et al. 1991. In situ methanotropohic
bioremediation for contaminated ground water at St.
Joseph, Michigan. In On-Site Bioreclamation 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-

livalent of COD equals one-fourth mole of molecular oxygen or 8 grams, thus Equiv.
-reaction.

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-
99.

——
100 Meters

10 ug/L. Contours

Figure 2. Chlorinated aliphatic compound 10 mg/L contours at St. Joseph, Michigan.

49


formation processes have occurred in the aquifer, and not only
are responsible for the presence of c-DCE and VC, butalso 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
amethanotrophic 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 thatthe 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 tothe design of asystem 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 downgradientas a final polishing step.
This scenario is currently being investigated.

References
1, McCarty, P.L. et al. 1991. Jn 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.

2. 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.

3.  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.

4, 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.

5. Wilsonetal. 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 Development Pro-
gram.

6. McCarty, P.L. andJ.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.

7. 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.




Recirculating Vapor Phase Biofilter

TCE uM

—e— UMTCE GS
—f— uMTCE Phe1

Hours

Figure 1.
or G4 5223 Phet.

unidentified transient compound, which we interpreted to be a
metabolite, was produced only in the G4 5223 Phe1 column. Itis
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 forresidual TCE. Airand TCE
vapor were pumped into the column at the bottom and allowed to

Continuous Flow Fixed Film Bioreactor

—E-  Input
—e— Output

TCE uM

Hours

Figure 2. TCE degradation within a continuous feed vapor-phase
bioreactor.

Unidentified GC Peak
20000:
—e— Phet
a os
i 100004
$
3
é
£
=
8 Of
-10000: T T
0 1 2 3
Hours

TCE concentration in the atmospheres of two independent recirculating gas-phase biofiters inoculated with Pseudomonas cepacia G4

exit at the top. Input and output ports were sampled to assess TCE
concentrations before and after exposure to the biofilm. Inputand
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 cm* of crushed oyster shell colonized by P.
cepacia G4 5223 Phe1. 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 pm,
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/cm’).

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 MJ. Reagin. Construction of a
Pseudomonas cepacia strain constitutive for the degra-
dation of trichloroethylene and its evaluation for field
and bioreactor conditions. Abstract, 91st General Meet-
ing of the American Society for Microbiology, Dallas,
Texas.


4790 rs - 4790
tiger cwr Abas)
4770 4770
4750 4750
a
Qo
=
2 4730 4730
:
3
2 ano 4710
4690 4690
10.
4670 =| 0.3 mg/l 4670
[=| sand with gravel and clay Screened interval CW — = _ Soil gas luster well
Silty sand _Zz_ 890 = TPH in soil (mg/kg){9/91)
= as Perched water (approx. surface) 15mg. TPHin ground water
Figure 2. Cross-sectional view at Hill AFB 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
toevaluate 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.

39

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.

S.K. Ong, R.E. Hinchee, R. Hoeppel, and R. Schultz,
1991. Jn situ respirometry for determining aerobic
degradation rates. In /n Situ Bioreclamation; R.E.
Hinchee and R.F, Olfenbuttel, eds. Butterworth-
Heinemann, Boston, pp 541-545.


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

This project is a component of research by the University of
Cincinnati/U.S. Environmental Protection Agency to develop
aerobic biofilters capable of the biodegradation of volatile or-
ganic compounds (VOCs). Standard microbial methods and
DNA probe technologies will be used to characterize the micro-
organisms that grow and operate as biodegradation agents in the
biofilters. This information will contribute to the successful
optimization of reactor design and its operational criteria in the
biodegradation of VOCs in aerobic and anaerobic configura-
tions. The specific objectives of this project are to (1) provide a
tapid index of the microbial diversity of biofilter samples based
upon DNA amplification fingerprinting (DAF) techniques, (2)
identify the predominant organisms in biofilter samples, and (3)
isolate, identify, culture, and characterize specific microorgan-
isms or microbial subpopulations recognized as having an impor-
tant active function in the desired biodegradative processes.

109

The profile pattems generated by the DAF method will be
useful in providing an index of biodiversity that would be
Tepresentative of the particular sample taken from the biofilter,
Approaches to the second and third objectives will include the
use of PCP-amplified DNA for Ppathway-specific gene probe
sequences available from the literature, and for 16S rDNA
sequence-specific probes for particular microorganisms identi-
fied from the biofilters.

Identification of the microbial strains and mixtures that are
characteristic of efficient bioreactor operation is a long-term
Tesearch aspect of this project.


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.

118

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 asa 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, and Robie J. Vestal for
their past and current consultation on this research.


United States

Environmental Protection Agency

Center for Environmental Research Information
Cincinnati, OH 45268

Official Business
Penalty for Private Use
$300

EPA/600/R-92/126

Please make all necessary changes on the label below,
detach or copy, and return to the address in the upper
left-hand corner.

If you do not wish to receive these reports CHECK HERE CL];
detach, or copy this cover, and return to the address in the
upper feft-hand corner

BULK RATE
POSTAGE & FEES PAID
EPA
PERMIT No, G-35



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




\ North Plot
\ DG 4" PVC ball valve.
+ DA
\ +
\ <-—#__# DE
\ eo _2
\ | °
‘ |
\ HIE
N eee
\ | ‘T —O
\ 1 ——
\ |
Vapor withdrawal line =| \
a Outer boundaries of project area
Vapor reinjection line ly
\ por reinjection Ii '
began \ Air injection line ———————} 2" Brass ball valve
© Monitoring well - MW Hl
© Air injection wall - Al \ My
© Vapor extraction well - VW \ Reduced to
4 Vapor reinjection welll -VRI \ 2" PVC line {
> Pressure and gas probe - Cl in this area —=——
© Moisture and temperature probe - MTP \ tle 10 0 10 2
--- Fence N Feet
+ Existing soil cluster wells \ Blower 1" =20'
\ Building
Figure 1. Bioventing flow line plan.

System Monitoring

Blower flow rates, water table levels, monitoring well nutri-
ents + aromatic hydrocarbons (BTEX) + dissolved oxygen,
vertical profile oily residue, subsoil moisture + temperature, soil
gas carbon dioxide + oxygen + combustible gas, and hydrocar-
bon surface emissions were assayed on a regularly scheduled
basis.

Time Requirement

Venting of the subsurface by both pilot-scale systems was done
at a rate equivalent to a calculated residence time of 24 hours.
Gasoline vapor throughout the unsaturated subsurface increased
toamaximum of near 6,000 mg/L total petroleum hydrocarbons
in3 days. A gradual decline to near 600 mg/L occurred 3 weeks
later, after which the decline was much less. The bioventing
systems were shut down in January for the winter because the
upper subsurface was frozen. When the soil had thawed, the
venting was restarted in April (Table 1). After 5 months of
venting, the soil gas hydrocarbon concentration of 50 mg/L was
attained. This level was selected as the clean-up requirement for
two reasons. Uncontaminated subsurface meter readings were
usually in the 10 to 40 mg/L range. Very low concentrations of

62

less than 50 mg/L may not be sufficient to stimulate enzyme
production needed to degrade the gasoline hydrocarbons.

Extent of Treatment

Total petroleum hydrocarbon concentration in the subsurface
above the water in the bioventing area ranged up to 2,060 mg/kg
in September 1990, up to 38 mg/kg in October 1991. A compari-
son of vertical profile fuel carbon mass in the north plot for
September 1990 and October 1991 is shown in Table 2. Unsatur-
ated subsurface reduction in oily-phase residue was more than 99
percent between the two sampling periods. Oily-phase residue in
the saturated subsurface below the water table was reduced 22
percent during the same time period. Some natural losses likely
occurred from volatilization, dissolution, and biodegradation.

South plot data (Table 3) compared to the control showed a
large reduction in unsaturated subsurface oily-phase residue, but
only a slight reduction below the water table.

From January 1991 to October 1991, total BTEX concentra-
tions in monitoring well water decreased tenfold (Table 4). All
benzene levels in October 1991 were less than 5 ppb. Several


Lake
Michigan |

100 Meters

Chemical Oxygen Demand
(mg/l)

@ Monitoring well
— _Transect

Railroad
Red Arrow
Highway
Transects
1 feo e
10
3
Ne YL
Le
e
400
e
150 le
e
@
e 200

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
2is 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 toethene 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 notconsidered, 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
tepresents 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 CACsto
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.


Bioremediation of Soils and Sediments Contaminated with Aromatic Amines

EJ. Weber and J.E. Rogers
U.S. Environmental Protection Agency
Environmental Research Laboratory
Athens, GA

and

DLL. 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 tosoil 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

97

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 '*C-aniline suggested that revers-
ible cation exchange processes do not contribute significantly to
aniline sorption in the natural sediment-water system, and that




Pvul

Xmni Mtul
Seal J Noel
Wz 0 ,
mT
Ppal sul. % Hindill
Sra ‘
f Bgill
Ap
‘ Pstl
pucLv1
4,42 kb
AIwNI

Aftill

Figure 2.
derived from pL V59 are unshaded.

Restriction map of pUCLV1 showing positions of the EcoRI gene (r), the temperature sensitive EcoRI methylase (m-ts), the
B-lactamase gene (Ap), and the origin of. replication (ori). Portions of the vector

derived from pUC19 are shaded. Portions of the vector

Aval Hincll

Sat|———

Figure 3.
P. cepacia DNA are shaded.

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 of
approximately 50 and 150 kb. In an effort to enhance the fitness
‘of Phel for use in bioremediation, a third plasmid, pRO101, was

18

(EcoRI!)

Restriction map of pCDOS showing positions of the chloramphenicol acetyltransferase gene (Cm), remnants of the EcoRI gene (r') and
the EcoRI methylase (m '), and the catechol-2,3-dioxygenase gene of P, cepacia (C230). Portions
Portions of the plasmid derived from pLV59 are unshaded.

of the plasmid derived from

introduced via conjugation. Plasmid pRO101, 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 pRO101 had lost the large
resident plasmid (pPhe1L). Those organisms, designated Phel
cure, lost the ability to degrade TCE and to transform trifluoro-
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


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 Testand Evaluation Facility. 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, CO,, H,O, 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

101

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.


oe

Table 4. Total Carcinogenic PAHs In The Libby Soil Treatment Facility (mg/kg)

_|

Sampling Date 9/18/91

eet South =
Sample -
Depth : 1132 |
28

Lift 4 352 194 187 131 241 282 247
Lift 3 53 134 101 83 51 78 719 228
Lift 2 47 81 84 93 59 66 60 206
Lift 1 38 74 86 53 35 19 105 23
17 18 19 20 21 22 23 24
Lift 4 165 352 250 317 201 192 241 110
Lift 3 165 217 42 160 115 187 88 172
Lift 2 94 94 117 90 114 129 15 174
Lift 1 35 232 43 102 22 116 104 17
16 15 14 13 12 11 10 9
Lift 4 223 254 150 130 223 172 185 244
Lift 3 108 370 106 159 193 149 229 151
Lift 2 357 86 51 124 72 220 mi
Lift 1 116 92 71 131 125 113 68 114
1 2 3 4 5 6 7 8
Lift 4 44 187 128 151 17 111 211 134
Lift 3 81 120 136 63 19 131 150 113
Lift 2 143 154 274 59 88 172 113 11
Lift 1 23 116 188 119 151 192 212 68



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


(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 and9).

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.


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


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, andt-DCE
at concentrations of 40 g/L, 20 ug/L, and 20 pg/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 Jn 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).

72

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 pg/L, 40 pg/L, and 40 pg/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
aconcentration 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, 2m
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 andc-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
inconcentration. 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




Oxidation of the Sy Oxidation of the
Homocyclic Heterocyclic
¢ : . ‘
Ring Px 2a». Ring
a N NS
,
@ oa quinoline a
HOH - @ _
g& wv
I= S= Or
sa sm
N N N~~on n7So
° OH H
8-quinolinone 8-quinolinol 2-quinolinol 2(1H)-quinolinone
1
'
1
'
'
Hydrolytic Cleavage of v
the Oxidized Ring C
NH2
aniline
benzoic acid
phenol
-
oa
Ring Cleavage Pe
‘Acetogenesis oa

sa ord
acetic acid, formic acid 4“

v

Methanogenesis

(002, CH4, NHg*

*

@ Ammonia not determined.

Proposed intermediate-not found in microcosms but found in ground-water samples.

Figure 1. Methanogenic degradation pathway for quinoline in aquifer-derived microcosms.

trations of CH, and CO, 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 CH, and CO,.

During oxidation of the nitrogen-containing heterocycliccom-
pounds indole, quinoline, and isoquinolinoneto oxindole, 2(1H)-
quinolinone, and 1(2H)-isoquinolinone, respectively, the parent
compounds were initially stoichiometrically oxidized. The oxi-
dation did not support growth. The oxidized intermediates per-

90

sisted for some time after the initial oxidation of the parent
compounds before they were degraded to CH, and CO,.

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(1H)-quinolinone, and 88.5 percent
for 1(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-


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
trichloroethylene 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.


12

6.2 mg/L

Normalized conc. (C/Co)

High Phenol
12.4 mg/L

4 tT edce injection
stopped
0 T T T
to) 200 400

Time (hours)

800 1000 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
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-
ethylene by abacterial isolate. Appl. Environ. Microbiol.
52:383-384,

Nelson, M.J.K., S.O. Montgomery, W.R. Mahaffey,
and P.H. Pritchard. 1987. Biodegradation of trichloro-
ethylene and involvementof an aromatic biodegradative
pathway. Appl. Environ. Microbiol. 53:949-954,

Nelson, M.J.K.,S.O. Montgomery, and P.H. Pritchard.
1988. Trichloroethylene metabolism by microorgan-
isms that degrade aromatic compounds. Appl. Environ.
Microbiol. 54: 604-606.

73

Wackett, L.P. and D.T. Gibson. 1988. Degradation of
trichloroethylene by toluene dioxygenase in whole-cell
studies with Pseudomonas putida F1. Appl. Environ.
Microbiol. 54:1703-1708.

Fox, B.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.


Anaerobic Degradation of Highly Chlorinated Dibenzo-p-Dioxins
and Dibenzofurans

P. Adriaens and D. Grbié-Galié
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 g/kg of the following polychlo-
rinated 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 CDD;
1,2,3,4.6,7,9-HCDD; 1,2,4,6,8-PentaCDF: and 1,234,678.
H,CDF. Previously, a hexachlorinated congener was identified
as a metabolite in H.CDD 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 aconcentration or
bioavailability threshold below which noreductive dehalogenation
occurs or at rates not measurable within the time of this project.

Background

Reductive dehalogenation of polychlorinated biphenyls (PCBs)
with ahigh 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 PCB congeners (3,4). The PCBs
presumably serve as an alternative “electron sink” to CO, for
methanogenic bacteria. Because of the structural similarity 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 Teported that after 2
months, approximately 9 g/kg of hexa-chlorinated dioxin was
formed in Hudson River sediment-inoculated microcosms spiked
with HeptaCDD (5). The concentration of the intermediate
decreased to 6 g/kg after 9 months (Figure 1, D), but no lesser
chlorinated congeners resulting from further dehalogenation
were detected (5). Ina parallel series of experiments, partitioning
the PCDD/F in microcosms inoculated with aquifer material was
shown to influence the interpretation of long-term incubation
Tesults, 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.

79

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® 12A2. 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 tecovery of PCDD/
Fin 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 Tecovery 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
trichlorobiphenyls 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, alt ough 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) and was 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


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 Pseudomonas 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 utilizin, ig 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

1


Anaerobic Transformation and Degradation of Chlorobenzoates and
Chlorophenols Under Four Reducing Conditions

MM. 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
substituent and 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 jm) 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,
oriron- 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. CH,OCI +4.5 HO —> 3.25 CH, + 2.75 CO, +H*+Cr
2. C,H,OCI + 3.25 SO? +4H,O —> 6 HCO, +3.25H,S +0.5H"+Cr
3. C,H,OCI + 5.2NO, + 4.2H* —> 6CO, +26N,+46H,0+Cr
4. C.H,OCI + 27 Fe(II) + 17H,O —> 6 HCO, + 27 Fe(Il) + 33 H*
Chlorobenzoates:
5. CH,O,Cl+5H,O —> 3.5 CH, +3.5 CO, +H’ +Cr
6. C,H,O,Cl + 3.5802 +5H,O —> 7HCO, +35 HS +H'+Cr
7. C/H,O,Cl + 5.6NO, +4.6 Ht —> 7CO, +2.8N, +48 HO + Cr
8. CH,O,Cl+ 29 Fe(II) + 19 H,O —> 7HCO, +29 Fe(II) +36 H*

Degradation of 2-,3-, and 4-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 CO, (Equation 2). Under denitrifying condi-
tions, 3- and 4-chlorobenzoate were degraded coupled to sto-

107


Fungal Treatment 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, Asystematic 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-
tachlorophenol 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 (RCB) 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 RCB 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 witha garden 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 , ROB Study
1
, fo fe G iH ti
1 z yz 2} !
1 iR Q J ' '
1 gf id ! Rxs
1
1
' '
Ss |r K]t |
1
1 2 z
vu IN |M '
1
\ sl id 1 {
1 roc css '
1 1
1 fw x 'fE Cc
' 4 '
(i 1
1 _of ' 2
7 Sm mm my
'
tir D
1
1
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eee eee Olle
Figure 1. Treatment plot arrangement.


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 same bioprocess. 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.

103

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-line pH, oxidation-reduction potential (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.


COOH COOH

CH,OH

4+

cl OCH;

ca
OH

§-chlorovanilliate

»

5-chloroprotocatechuate

COOH

vanilliate

COz OH ct OH
s . Of — — cO,+ OY,
OH OH OH
OH cl

3-chlorocatechol

catechol

COOH

Os, OF
OH

oon, Ts0H Y

OH

protocatechuate

Figure 3, Proposed pathway for the complete degradation of &-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 RNA 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 SCV. Benzoate
has not been detected as an intermediate in the degradation of

88

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 SCV 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, Menidia
beryllina. Comparisons can be made to determine the efficacy of
the various consortia in detoxification.


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 af 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

56

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 chlorophenolsin anoxic
sediments.


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 M.L. Habewrl. 1990. Microb. Ecol.
20:87-102.

52

Furukawa, K., N. Tomizuka, and Kamibayashi. 1979.
Appl. Environ. Microbiol. 38:301-310.

Brown, J.F.,R.E. Wagner, D.L. Bedard, M.J. Brennan,
J.C. Carnahan, R.J. May, and TJ. Tofflemire. 1984.
Northeast Environ. Sci. 3:167-179.

Quensen, J.F., J.M. Tiedje, and S.A. Boyd. 1988. Sci-
ence 242:752-754.

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.


12.

tachlorophenol in ground water: chemical and biologi-
cal assessment. Appl. Environ. Microbiol. 57:1277-
1285.

Resnick, S.M. and P.J. Chapman. 1990, Isolation and
characterization of a pentachlorophenol (PCP)-degrad-
ing gram-negative bacterium. Abstr. 90th Ann. Meeting
Am. Soc. Microbiol., Q70, p. 300.

14


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
Wester Region Hazardous Substance Research Center
Stanford University
Stanford, CA

Introduction

The St. Joseph, Michigan, National Priority List industrial site
isarelatively 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 pg/L, 160 pg/L,
and 70 g/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, to2.7 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, suchas oxygen and
methane needed for the development of the methanotrophic
bacterial population. The mixing method being investigated uses
arecirculation 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,

_

Methane

Vadose zone

Recirculation unit —
Ground water

A

N= 2285
x
2
2

Figure 1. Recirculation system for in situ remediation at the St.
Joseph site.


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 10 m
below the ground surface in a fine sand aquifer that is about 20
m deep.

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

47

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 the early 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 70 percentof 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




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.

Olsthoorn, T.N. 1982. The clogging of techarge wells,
Assembled within the research program of the VEWIN
association, Rijswijk, the Netherlands. August 1982,
pp. 15-17.

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.


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 Treatment Facility (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 usider aerobic conditions (1,2).
Thus, an aerobic treatment phase contributes to more complete

51

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 chlorinated PCBs. Furthermore, we
have recent evidence of ortho-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
aminimum of 1 yearand perhaps for 2 years to provide sufficient
time for potentially significant in situ 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.


Table of Contents (continued)

Characterization of Microorganisms, Microbial Consortia, and Microbial
Processes for the Reductive Dechlorination of Hazardous Wastes ...

Effects of Metals on the Reductive Dechlorination of Chlorophenols
Section Three: Performance Evaluation

Development of Comparative Genotoxicity Risk Methods for Evaluating
Alternative Hazardous Waste Control Technologies...

Performance of Bioventing at Traverse City, Michigan

Bioventing of a Gasoline Spill at Traverse City, Michig:
Practical Engineering Considerations .....

Clearance and Pulmonary Inflammatory Response after Intranasal
Exposure of C3h/HeJ Mice to Biotechnology Agents...

Section Four: Process Research...
Evaluation of Enhanced In Situ Aerobic Biodegradation of
cis- and trans-1-Trichloroethylene and cis- and trans-1,2-Dichloroethylene
by Phenol-Utilizing Bacteria
Fundamental Studies on the Treatment of VOCs in a Biofilter......

Fungal Treatment of Pentachlorophenol

Anaerobic Degradation of Highly Chlorinated Dibenzo-p-Dioxins and
Dibenzofurans

Bacterial Degradation of Trichloroethylene in a Gas-Phase Bioreactor

Anaerobic Biodegradation of 5-Chlorovanillate as a Model Substrate
for the Bioremediation of Paper-Milling Waste .....

Methanogenic Degradation of Heterocyclic Aromatic Compounds by
Aquifer-Derived Microcosm ....

Chemical Interactions and pH Profiles in Microbial Biofilms .....

Decontamination of PCB-Contaminated Sediments through the Use of
Bioremediation Technologies

2189

2095





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

23

tomore 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.


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 to63 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 ata
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-
tus, 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,
twicea 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 Threshold Limit 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 Lfour.

Water quality samples analyzed for nutrients and BTEX are
obtained monthly by sampling at two depths in each of the 14-
point monitoring wells.


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 toxicological 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 representrisk of
unknown magnitude” (1),

NCPs, therefore, limit the ability todo 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 pathwaysand to testeach

59

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 nota 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 persistentin 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-Tismetabolized to 2,4,5-trichlorophenol (2,4,5-TCP),2,5-
dichlorohydroquinone (DCHQ), and 5-chloro-2-h ydroxy-
hydroquinone (CHHQ) before the arene ting 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 AC1100, has the ability to degrade 2,4,5-
T. Dr. Sangodkar indicates that AC1100 metabolizes 2,4,5-T
through the intermediate CHHQ and that catechol isaby-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 per plate 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-
Tand 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




Ground surface
Plastic sheeting

Styrofoam Insulation

Ground-water
‘monitoring
well (MW-3)

Plywood Themeocavbles Air injection/withdrawal well
Blue viny! na Three-level
cover Ts soil gas probe
[16 7 B B

1; Brown sand
and gravel

+f Bae At (s4) A [° (s5) A

C (86)

(Sarg a

ILLE

Ba
Gray sandy gravel

aa

Figure 2. Cross-sectional view of a portion of the actively warmed test plot at Eielson AFB showing the various wells, the temperature monitoring
davices, the heated water delivery system and the ground coverings. Vertical axis is depth from the surface in feet.

compared with the initial soil analysis to document actual hydro-
carbon loss due to bioventing.

Results

Progress to date includes installation of the pilot-scale equip-
ment and initial soil sampling for total hydrocarbons during July
and August 1991, continuous soil temperature monitoring since
August 25, 1991, weekly soil gas monitoring, and in situ respira-
tion tests conducted periodically since early October 1991. As of
early January 1992, all test plots were being adequately aerated.
The active, passive, and control plots, and the background
location showed average soil gas oxygen levels of 12,20, 15, and
20 percent, respectively. Soil gas oxygen levels should be main-
tained above at least 5 percent oxygen to avoid oxygen limitation
of microbial activity.

Figure 3 shows average soil temperatures for each plot and the
background location as a function of time through early January
1992, Clearly, the active-warming strategy is functioning suc-
cessfully: as ambient temperatures fell during the fall to an
average winter value of about -20°C, the actively warmed plot
remained above 10°C (except for a short period near 8°C), while
the temperature in the other plots and the background dropped
steadily toward 0°C.

The passively warmed plot is not maintaining a significantly
higher temperature than the control plot. However, this plot did
not benefit from a summer of enhanced solar heating using the
plastic mulch because installation and startup occurred late in the
summer. Insulation, not mulch, was laid down during the instal-
lation in anticipation of the upcoming winter. The success of the
passive heating strategy will be evaluated during the winter of
1993 after a full season of solar heating.

33

The temperatures in Figure 3 are reflected in the measured
oxygen uptake rates and biodegradation rates from an in situ
respirometry test conducted in early December 1991. The results
shown in Table 1 reveal the biodegradation rate in the actively
warmed plot is at least twice the rate in either of the other two
plots. In addition, the biodegradation rate in the actively warmed
Plot of 2.9 mg/kg/day is consistent with biodegradation rates
observed in other bioventing studies at various locations ( 1). The
conversion from oxygen uptake rates to biodegradation rates is
accomplished by assuming that the biodegradation reaction
follows the stoichiometry in Equation 1:

CH,

Gist)

+ 9.50, = 6CO, +7H,O a)

The biodegradation rate associated with the control plot was
greater than the rate associated with the passively warmed plot.
This difference is unexplainable at this time. Another year of data
will be needed to determine if this difference is a consistent trend
or random variation.

Reduced water flow rate from the buried soaker hoses due to
the accumulation of silt around the hoses caused inefficient
performance of the actively warmed plot in its early operation.
The low water flow rate resulted in the steady decline of the
temperature between Day 30 and 56 (Figure 3). Our goal is to
maintain the soil in the actively warmed plot between 10° and
20°C in the winter. To remedy this situation, weekly high-
pressure pulses are employed to clear the hoses. The pulsing
began on Day 56 (noted in Figure 3) and resulted in increased soil
temperature. Discovering the water flow rate that maximizes
heating and maintains adequate soil oxygen levels will be an
important outcome of this field study.


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 Kuhnetal. (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

53

dechlorination. Additions of individual substrates, including
formate, butyrate, propionate, acetate, anda 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 contro! 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 followin ig the second


Table 1. Relevant Characteristics of P. cepacia G4 and Tn5-derivative Strains.

Designation TCE* TEMP Stability © Phenotype Putative Tn Location
P. cepacia G4 He + NA NA NA

P. cepacia G4 5220 +i - stable C230 cdo

P. cepacia G4 5223 - - yields TCE © Tom tom regulatory

P. cepacia G4 5227 - - stable Tom tom

P. cepacia G4 5223 Phet +0 +o stable Tom+c unknown

Ability to degrade trichloroethylene
Ability to produce hydroxytrtluoroheptadienoic acid from trllucromethylphenol.
Relative stability of phenotype after several passages on a nonselective medium,
Inducible activity.

Constitute activity,

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 Phe1), which does not seem to carry the Tn5 insertion
on the large plasmid.

In this system, Tom activity carried
be detected by the conversion of’ TFMP totrifluoromethylcatechol
(TFMC). TEMC is,
pCDOS,

number plasmid (pCDOS). Plasmid pCDOS (Figure 3) is an over-
producer of C230, which carries the Cm‘
replication from pLV59. The high level of: expression of the C230
gene is fortuitous in that it will facilitate the detection of Tom
activity carried on the primary cloning vector.

gene and origin of

on the cloning vector can

in turn, a substrate for the C230 expressed by
Producing the bright yellow-ring cleavage product

TFHA. Thus, the combination of plasmids pUCLV1 and pCDO5

While Phe1 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 toalter expression of the gene,toplacethegene —_ removal

allow for a rapid and sensitive detection of recombinant Tom
activity. While P. 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

may provide even more powerful ways to address the

ina well-defined genetic background and the potential to produce —_ problem in the future.
purified enzyme(s) for specialized field applications. To facili-
tate screening for Tom activity in potential clones, a novel host- References

vector system has been constructed. The vector, puUCLV1 (Fig- Ll
ure 2), carries the EcoRI endonuclease and methylase genes from
pLV59 (1), and the B-lactamase gene and replication origin of
pUC19 (5). The EcoRI 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 EcoRI. EcoRI expression can be
inactivated by cloning into one of three unique restriction sites
within the EcoRI 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, see Figure 1) onalowcopy-

19

O’Connor, C.D, and G.O. Humphreys, 1982, Gene
20:219.

Shields, M.S. and MJ. 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, P.J. 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.






200 7 60
I Added 50
@ Active
Autoclaved 40

30

A Aroclor 1242 standard

20
10

Concentration (ug/kg)

1 2 3 4 5 6 7
B "Weathered" 0 months.

8

H7CDD H6CDD H6CDDi

40

Figure 2, PCDD/F added to and recovered from final Hudson River
extracts after a 16-month incubation period.

% area

30
20
10

Cc 16 months.

1 2 3 4 5 6 7
Homologs

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






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 for phosphorus 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 testresults 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
isutilized asa substrate by the microbial populations. As fluores-
cein is released, the solution changes color as a function of the

117

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




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

Dataon 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 forasmall 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
Teduction 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
os 5 aE ae
So : :
Sample |
Depth BT
Lift 3 229 435 169
Lift 2 30 155 33 12 12 144 19 9
Lift 1 16 9 29 33 30 147 19 1
Lift 3 257 217 42 «167 303 325 761 275
Litt 2 4 15 21 0 58 40 15
Lift 1 31 28 38 17 29 380 94 9
Lift 3 189 321367 204 440 336 1243 1121
Lift 2 0 9 21 17 ll it 35 2
Lift 1 42 17 28 64 27 34 52
Lift 3 275 891 798 710 212 55 667 663
Lift 2 226 26 177 18 24 12.5 67 29
Lift 1 156 13 165 176 136 24 18 56
Sa ee |



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, and Rodenticide Act; the Resource Conservation and Recovery Act; and
the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA, known as Superfund) all
tequire 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
toreleases of hazardous materials that endanger public health or the environment, Also, the Superfund Amendments and
Reauthorization Act of 1980 (SARA) expands EPA’s authority in research and development, training, health assess-
ments, community right-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 itis 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 ofa product or a process being more toxic than the original
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
Initiative.

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
technology transfer.


Table 2, | Summary Core Data for Locations 60B and 60H, Cell #2

Parameter Units 60B 60H
Depth interval m BLS* 2.4-8.5 2.7-8.9
Contaminated zone(s) m BLS 2.46.9 6.0-7.2
Uncontaminated zone(s) mBLS 6.9-8.5 2.7-6.0
7.28.9
Most contaminated interval** =m BLS 6.4-6.6 6.7-6.8
Mean BTEX concentration ng/kg dry wt 42.6 24.3
Mean TPH concentration ng/kg dry wt 2,360 686
Peak BTEX concentration mg/kg dry wt 353 412
Peak TPH concentration mg/kg dry wt = 10,500 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

42

microbial rate of nitrate removal, the penetration zones would be
5.8 to6.1 m for the Core 60B location and 6.3 to6.6m 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, $.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.

Hutchins, S.R. 1991. Optimizing BTEX biodegrada-
tion under denitrifying conditions. Environ. Toxicol.
Chem. 10:1437-1448.


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 tesiduals/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






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 ina 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

15

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.




Table of Contents (continued)

Bioremediation of Soils and Sediments Contaminated with
Aromatic Amines

Development of Small-Scale Evaluation Techniques for
Fungal Treatment of Soils....... a

Development of Aerobic Biofilter Design Criteria for Treating VOCs..

Sequential Anaerobic-Aerobic Treatment of Contaminated
Soils and Sediments ....

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...

Section Five: Modeling

Prediction of GAC Adsorptive Capacity with and without
Molecular Oxygen

Development of Tools for Monitoring Biofilm Processes ....

Evaluation of Anaerobic Respirometry to Quantify Intrinsic Anaerobic
Biodegradation Kinetics of Recalcitrant Organic Compounds.....



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.

106

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.

Vigon, B.W. and A.J. Rubin. 1989. Journal W.P.C.F.
61:1233-1244.

Gurin, WF. and G.E. Jones. 1988. Appl. Environ.
Micro. 54:937-944,

Van Hoof. P.L. and C.T. Jafvert. 1991. Presentation at
the annual meeting of the Society of Environmental
Toxicology and Chemistry.

Laha, S. and R.G. Luthy. 1991. Environ. Sci. Technol.
25:1920-1930.

Aronstein, B.N., Y.M. Calvillo, and M. Alexander.
1991. Environ. Sci. Technol. 25:1728-1731.


Bioventing of a Gasoline Spill at Traverse 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 ofhydrocarbon-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 silty 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-ftstudy area has been divided into two equal areas
of 45 x 75 ft to evaluate the effects of different flows and
extraction pattems, A conceptual design is presented in Figure 1.
The northern area has an injection system, while the southem
area has an injection and extraction/reinjection system. The

South Plot North Plot
ft, N-
Pumps. Pump
+
*
¥ *
‘ ee tae * * 5
Extraction well with tip + y
of well screen located . |. . . ~ oe
at a depth of approx. . . . .
16 ft below grade .* D .
. ° e
LU .
. : . .
Reinjection well with tip eT RN 4
of well screen located 4 . hz
at a depth of approx. woe XY W
10 ft below grade :
i + oe 10ft . se
= AN Za .
. 4 >A
: ae = ww = \e~
. - oS. eS
16 ft nS 3 .
yo “.
.
° . wf. . .
° , \
Fresh air injection well with Fresh air injection well with tip
tip of well screen located at of well screen located at a depth
a depth of approx. 16 ft of approx. 15 ft below grade
below grade
Figure 1. Conceptual design.






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 will 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 concem 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.

25

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 Teinjection
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 analysisof the existing bioreactors
indicates the following: (1) removal of PAHs and PCP is strongly
influenced by flow rate through the reactor (10 gallons per minute
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.

Apilot-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
Teactor 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


Concentration, mM

200
Distance into the Biofilm, uM.

400 600 800 1000

Figure 1. Concentration profiles of the inorganic carbon species and

pH for L,= 1,000 um, bulk pH = 7.0, bulk alkalinity =
1 meq/t (case 1).

first 50 um of the biofilm and subsequently decreases throughout
the biofilm as the rise in pH transforms the dominant inorganic
carbon species to CO,*. Finally, the CO,* concentration in-
creases due to the shift in chemical equilibrium and eventually
decreases because CO, 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 L, while the flux approaches a
constant value at large L,. The flux is limited by the rate of
reaction within the biofilm when L, is small and is limited by the
rate of mass tansfer for large values of L,. A decrease in pH,

94

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 CO,,. 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.

0.25

0.20-
= bulk pH=6
2
§ 0.15
3
SB  0.10F bulk pH=7
E
= ost bulk pH=8
z

00 1 A . 1

0 100 «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).


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 (SCV) was chosen asa 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 SCV. 3-
chlorocatechol (3CC) and catechol were detected as intermedi-
ates of degradation (Figure 1), If 5mM bromoethanesulfonic acid
(BESA) was included in the enrichment, 5-chloroprotocatechuate
(SCP), protocatechuate, and vanillate were also detected (Figure
2). From the temporal sequence and amounts of intermediates
formed, we concluded that the major pathway of SCV degrada-
tion was stepwise demethoxylation to SCP, decarboxylation to
3CC, and dechlorination to catechol, which was subsequently
completely degraded (Figure 3). The addition of BESA appeared
to inhibit decarboxylation and demethoxylation, allowing us to
detect SCP, as a major intermediate, and vanillate and

500:
O—@ S-chiorovaniliate
a A—A 3-chiorocatecho!
\ O—4 catechol
= 9
: a /
g 250 °
8
2
8 °
/ A
<< \
° T T és T T T
0 10 20 30 40 50 60
Time (days)

Figure 1. Degradation of 5-chlorovanillate in the first transfer of an
Eleven Mile Creek primary anaerobic 5CV. enrichment.

87

500:

@—® s-chiorovanillic acid
A—A 3-chlorocatechol

O—9 catechol

O—1 5-chloroprotocatechuic acid
V—V protocatechuic acid
O—O vanillic acid

Concentration (uM)
g
8

NN
‘A.
ro ——~
10
Time (days)

15 20

Figure 2. Degradation of 5-chlorovanillate in the first transfer of an

Eleven Mile Creek primary anaerobic enrichment with 5mM
BESA added.

protocatechuate as minor products. Enrichments adapted to these
intermediates will be studied using 16S rRNA analysis to identify
bacterial species responsible for a Particular degradative activity
without having to isolate pure cultures. Phylogenetic placement
of the species may also aid in developing enrichment strategies
for obtaining the various bacterial species in pure culture. Once
obtained in pure culture, the role of individual species in the
degradation of SCV will be assessed. A defined mixed culture of
known bacterial species responsible for the complete degrada-
tion of SCV can be devised for further study.

Inaddition to the SCV enrichment described above, an anaero-
bic bacterial co-culture: that dechlorinates 3-chlorobenzoate (CB)
was shown stepwise to dechlorinate SCV to vanillate and
demethoxylate vanillate to protocatechuate. Protocatechuate was
not further degraded. This co-culture contained a gram-negative
small diplobacillus, and curved rod. The diplobacillus was iso-


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 arural 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 isa 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
ft.

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 the problem
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.

Inan 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).

ai

Point bar sand

Figure 1. Geological setting of the site.


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 suchas 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 Fick’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, and hydration 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 CO, 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,

Woog , Woo , Moo§ bX: _ kX+[COd
dx dx dx = Y Ks + [CO] ql)

where J is the molar flux of the species; x is the direction
perpendicular to the biofilm surface; b is the respiration coeffi-
cient; X, is the biomass density; Y is the yield of microorganisms
per mass of substrate; k is the maximum utilization rate; K, is the
Monod half-saturation constant for CO,; and [CO, ] refers to the
molar concentration of CO,. 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 Fick’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-Maxwell equations. Equation (1) with
relationships for chemical equilibrium and electroneutrality was
solved assuming HCO,’, CO,”, H*, OH’, and Na’ as the ions
present within the system.

The profiles of CO,, HCO,’, CO,”, and pH for a biofilm depth
of 1. ,000 pm with a bulk pH of 7. O-and an alkalinity of 1 meq/L
for case 1 are shown in Figure 1. As dissolved CO, is consumed,
the pH increases dramatically within the first 100 um of the
biofilm. Consequently, the dissolved CO, concentration de-
creases rapidly in the first 50pm due to the coupled effects of pH
on the speciation of inorganic carbon and the net substrate
utilization of CO,. HCO, remains relatively constant during the


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, '"N NMR was used to analyze fulvic acid thathad
been treated with 'SN-aniline. INEPT and ACOUSTIC "N-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,

98

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.

Jafvert, C.T. and E.J. Weber. 1990. Sorption of ioniz-
able organic compounds to sediments and soils. Unpub-
lished report.

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.

Berry, D.F. and S.A. Boyd. 1985. Decontamination of
soil through enhanced formation of bound residues.
Environ. Sci. Technol. 19:1132-1133.


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 GAC 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 GF.,
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 T

q, me/g

102

© = 2-Methylphenol
© 2-Ethylphenol

| 2-Propylphenol

101

102

C, mg/L

Figure 1. Anoxic adsorption isotherms for three ortho-substituted phenols.

114


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 pg/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 Teported 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

61

percent, but none greatly hindered the biodegradation process.
‘The rates indicated that ata 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 ina 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.


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

MSS. Taylor and H.G. Brooks
Environmental Health Research and Testing, Inc.
Research Triangle Park, NC

and

ML 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
10° 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.

67

Two clinical isolates, P. aeruginosa, strains PAO! 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,


Table 1. Summary of Toxicological Analyses

Microtox Mysids Ceriodaphnia Embryonic Menidia
Bioreactor Sample EC,, 5 min LC,, 96h LC,, 48h (1% solutions)
Specially selected microorganisms
Feed 0.0185 0.01 0.007 embryotoxic
Biorx 1 2.3645 0.34 0.710 teratogenic
Effluent 1.2956 0.75 0.530 teratogenic
Non-specific microorganisms
Feed 0.1456 0.02 0.079 embryotoxic
Biorx 1 25.057 0.32 0.292 teratogenic
Effluent 54.107 0.16 0.230 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
for4 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.

7 31 ppm 539 ppm 368 ppm
7
100 (| c
7 Be ie
= ae bes
& 2 SS ee 1 input
8 J ] _ Output
3S So :
| /
E
s 4
2
5
2 4-4 .
‘4
g ag
5 4 L
a oa g
x - a
1 0.1 ppm 4.6 ppm — | 5.2 ppm
2-4 | (<1%) (<1%) (<2%)
q a :
JH eo be
T T

Bi-cyclic compounds

3-ring PAHS

4.or more ring PAHS

Figure 2. Reduction of PAH contaminants in creosote-contaminated ground water.

12


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

Polychlorinated biphenyl (PCB) congeners are relatively inert,
stable to oxidation and heat, and have high viscosities, high
tefractive 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

95

influenced by the position of the chlorine atoms on the biphenyl
ting (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 bipheny] 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 include 4-chlorobiphenyl, 2;2’4-trichlorobipheny! (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.


Approaches to the Characterization of Trichloroethylene
Degradation by Pseudomonas Cepacia G4

Charles Somerville and Michael Reagin
Technical Resources, Inc.
Gulf Breeze, FL

and

Robert Gerger
U.S. Environmental Protection A: gency
Gulf Breeze, FL

and

Malcolm Shields
University of West Florida Center for
Environmental Diagnostics and Bioremediation
Gulf Breeze, FL

Introduction in P. cepacia G4 (Table 1). One of the mutants, P. cepacia G4

Inthe presence of aromatic compounds, Pseudomonas: cepacia —_ 5223, lost the ability to degrade TCE, but gave rise to spontane-
G4 expresses a novel pathway for the metabolism of toluene (4). ous “revertants” (€.g., strain Phe1), which express Tom activity
‘The first two steps of the pathway are mediated by toluene ortho- regardless of the presence of aromatics (2,3). While the use of P,
monooxygenase (Tom, Figure 1), an enzyme that also attacks cepacia G4 in the bioremediation of TCE-contaminated Sites is
ttichloroethylene (TCE). To determine the genetic location of the limited by its requirement for aromatic compounds to induce
Tom gene for further study, a series of Tn5 mutants was made Tom activity, P. cepacia G4 5223 Phe1 has no such Tequire-

cH, CH, cH, cH,
OH oH °
a. Enea __ ———— COOH — —a
Tom Tom ‘oH 0280 Lon

cr, cr, oF,
OH °
b. —_ —_—_ cooH Se
Tom 0230
‘OH ‘OH Zon
TFMP TFMC TFHA

Figure 1. The pathway of toluene utilization in P. cepacia. a. Toluene is hydroxylated first to ortho-cresol then to 3-methylcatechol via toluene

7.7.7 trifluoroheptadionoic acid (TFHA), which is not further metabolized by P. cepacia G4. Tom activity is responsible for TCE co-

metabolism in P. cepacia. The appearance of TFHA (bright yellow color) from TFMP indicates the ‘presence of both Tom and C230
activities.

17




Table 1. Free Energy Changes During Oxidation of the Tested Heterocyclic Compounds
Compound GO kd.mol (8,9)
Indole to Oxindole

C,H, N + H,O -»C,H7NO + 2H’ + 26" 168
Quinoline to 2(1H)-quinolinone

C,H,N + H,0-9C,H,N) + 2H* + 26" 84
Isoquinoline to 1(2H)-isoquinolinone

C,H, N +H,O->C,H,NO + 2H*+2e 56
Oxindole to CH, and CO,

C,H, NO + 6.5 H,O + H* + 3.75 CO, + NH + 4,25 CH, -362
2(1H)-quinolinone to CH, and CO,

C,H, NO + 7.5 HO + Ht ~» 4.25 CO, + NH + 4.75 CH, -298
1(2H)-isoquinolinone to CH, and CO,

C,H,NO + 7.5 H,O + H* — 4.25 CO, + NH + 4.75 CH, -303

Note: AG ®,,., 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,
8-oxidation, and mineralization to CH,, CO,, and NH,. A minor
pathway first reported by Godsy and Grbic-Galié (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 ting 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 CH, and CO,. Nitrogen-
containing aliphatic acids were detected in ground-watersamples.
The corresponding compounds during benzothiophene degrada-
tion were detected, confirming that thiophene-2-ol was degraded
toCH, and CO, (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 CH, and CO, (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.

91

References

1.

Godsy, E.M. and D. Grbi¢é-Galié. 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.

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.

Godsy, E.M., D.F. Goerlitz, and D. Grbi¢-Gali¢. 1992.
Methanogenic biodegradation of creosote contaminants
in natural and simulated ground-water ecosystems.
Ground Wat. 30:232-242.

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.

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.

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.


62

Table 3. Pyrene In The Libby Soil Treatment Facility (mg/kg)
Sampling Date 5/6/91

Lift 3 8.9 62.7 95.9 66.1 39.6 80.1 28.4 44.9
Lift 2 11.3 13.3 15.4 ND 198 ND 172 61
Lift 1 23.3 33 81 28.2 143 133.6 30.7 ND



When specially selected bacteria were added to the system, a
one-time inoculum of a 13-member bacterial community with
known degradative abilities was Teconstituted from frozen stock
cultures (strains CRE1 through CREI3) and added to bioreactor
1 ataconcentration of. approximately 1 x 10° cells of each strain/
mL. Daily additions of Pseudomonas paucimobilis stain EPA
505 and Pseudomonas sp. strain SR3 (approximately 1 x 10°
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, reactive-P, total
creosote, and PCP (GC-FID, GC-ECD analyses), and select
biological toxicity assays to monitor bioreactor activity and
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 inthe
system effluent (removal efficiency of >99 percent). Also, the
toxicity and teratogenicity of the bioreactor effluent was signifi-
cantly reduced as determined by MicrotoxTM, embryonic Menidia
beryllina, Mysidopsis bahia, and Ceriodaphnia dubia assays
(Table 1). Notably, the cumulative concentration of 9 HMW
PAHs (each containing 4 or more fused tings) 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
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 demonstratin gan 85
to 90 percent recovery of the chemical tracer.

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 (SBP Technologies, Inc.). Assistance from Scott
Beckman, K.C. Mahesh (SAIC), 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
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,
MJ. Hemmer, G.T. Brooks, and PJ, Chapman. 1991,

13

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: MicrotoxTM and em-
bryonic Menidia. Arch. Environ. Contam, Toxicol.
Submitted.

Mueller, J.G., P.J. 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., P.J. 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 EPA5O5 and identi-
fication of fluoranthene metabolites, Appl. Environ,
Microbiol. Submitted.

Mueller, J.G.,R.J. 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 P.J.
Chapman. 1990. Alternative biological treatment pro-
cesses for remediation of creosote-contaminated mate-
tials: bench-scale teatability studies. EPA/006/9-90/
049. 89 pp.

Mueller, J.G., S.E. Lan, B.O. Blattmann, and P.J.
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 P.J.
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., DP. Middaugh, S.E. Lantz, and PJ.
Chapman. 1991. Biodegradation of creosote and pen-


Section Three
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. Remediation efforts at a contaminated site can produce intermediate compounds that can themselves be
hazardous. Performance evaluation Projects determine the risk of potential health effects and the bioremediation
approaches that best protect public health.

Researchers have determined that nonconventional Pollutants (NCPs) are a potentially 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 determine potential
genotoxicity at remediation sites. Two other: projects evaluated the performance of! bioventing systems to determine their
effectiveness and to monitor possible health risks from surface emissions of hazardous compounds.

A fourth study concemed potential health effects associated with microorganisms genetically engineered for

bioremediation. Investigators exposed laboratory mice to a variety of strains of biotechnology agents and observed the
results on their respiratory systems.

57








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 Grbié Gali¢
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,
8-oxidation, and mineralization toCH,,CO,, and H,S. 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(1H)-quinolinone and 1(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 1(2H)-isoquinolinone, respectively. The latter

89

two compounds persisted for some time before they were miner-
alized to CH, and CO,,.

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 CH, and
CO,. The pathways and thermodynamics of microbial conver-
sion by acomplex, 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-


Table 1. Results of an In Situ Respiration Test in Early December

1991 at Eielson AFB
Oxygen uptake rate Biodegredation rate
Plot (% O,/nr) (mg/kg-day)
Actively warmed 0.15 29
Passively warmed 0.039 0.74
Control 0.071 14
Background 0 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

14

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.

Temperature (°C)

-2

Pressure pulse
on soaker hose

oN 10/1

Figure 3.
August 1991 through early January 1992.

T T T
WWM 12/1 Ww

Time (days)

34

Average temperature of each test plot and ‘background as a function of time at Eielson AFB. Time scale is date beginning in late




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'-heptachlorobipheny!
(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 forthe 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.

96

10.

Kalmaz, E.V. and G.D. Kalmaz. 1979. Rev. Ecol.
Model. 6:223.

Abramowitz, D.A. 1990. Crit. Rev. Biotechnol. 10:241.
Johnston, L.E. 1985. Environ. Health Pers. 60:339-346.

Unterman, R. 1990. Aerobic biodegradation of PCBs.
In Biological Remediation of Contaminated Sediments,
with Special Emphasis on the Great Lakes, Jafvert,C.T.
and E. Rogers, eds. U.S. Environmental Protection
Agency, Athens, Georgia. Publication No. EPA/600/9-
91/001.

Bedard, D. L. and M. L. Haberl. 1990. Microb. Ecol.
20:87.

Furukawa, K., N. Tomizuka, and A. Kamibayashi.
1979, Appl. Environ. Microbiol. 38:301.

Brown, J.F., R.E. Wagner, D.L. Bedard, M.J. Brennan,
J.C. Carnahan, RJ. May, and T.J. Tofflemire. 1984,
Northeast Environ. Sci. 3:167.

Quensen, J.F., J.M. Tiedje and S.A. Boyd. 1988. Sci-
ence 242: 752.


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 tooxygen
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

37

others, bioventing should be available in the very near future as
areliable, 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” wellsare
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 ft'/min depths from
30 to 95 ft.