EPA/BOO/
H-94/075
A
United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/R-94/075
June 1994
Symposium on
Bioremediation of
Hazardous Wastes:
Research,
Development, and
Field Evaluations
Abstracts
ANA Hotel San Francisco
San Francisco, CA
June 28-30, 1994
PROTECTION
AGENCY
DALLAS, TEXAS
II
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EPA/6fW94/075
June 1994
Symposium on Bioremediation of Hazardous Wastes:
Research, Development, and Field Evaluations
Abstracts
ANA Hotel San Francisco
San Francisco, CA
June 28-30, 1994
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Printed on Recycled Paper
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Disclaimer
The projects described in this document have been funded wholly or in part by the U.S.
Environmental Protection Agency (EPA), and the abstracts have been reviewed in accordance
with EPA's peer and administrative review policies and approved for presentation and
publication. Mention of trade names or commercial products does not constitute endorsement
or recommendation for use.
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Contents
Bioremediation Field Initiative
Intrinsic Bioremediation of TCE in Ground Water at an NPL Site in St. Joseph,
Michigan
John T. Wilson, James W. Weaver, Don H. Kampbell, U.S. EPA Ada, OK 3
Enhanced Reductive Dechlorination of Chlorinated Ethenes
Zachary C. Hasfon, Pramod K. Sharma, James N.P. Blade, Perry L McCarty,
Stanford University, Stanford, CA 11
Bioventing of Jet Fuel Spills I: Bioventing in a Cold Climate With Soil Warming at
Eielsen AFB, Alaska
Gregory D. Say/es, Richard C. Brenner, U.S. EPA, Cincinnati, OH; Robert E.
Hinchee, Andrea Leeson, Battelle Memorial Institute, Columbus, OH; Catherine M.
Vogel, U.S. Air Force, Armstrong Laboratories, Jyndall AFB, FL; Ross N. Miller,
U.S. Air Force, Center for Environmental Excellence, Brooks AFB, JX 15
Bioventing of Jet Fuel Spills II: Bioventing in a Deep Vadose Zone at Hill AFB, Utah
Gregory D. Say/es, Richard C. Brenner, U.S. EPA, Cincinnati, OH; Robert E.
Hinchee, Battelle Memorial Institute, Columbus, OH; Robert Elliott, Hill AFB, UT . . . , 22
In Situ Bioremediation of a Pipeline Spill Using Nitrate as the Electron Acceptor
Stephen R. Hutchins, John T. Wilson, Don H. Kampbell, U.S. EPA, Ada, OK 29 ^
Performance Evaluation of Full-Scale In Situ and Ex Situ Bioremediation of Creosote
Wastes in Ground Water and Soils
Ronald C. Sims, Judy L Sims, Darwin L. Sorensen, David K. Stevens, Utah State
University, Logan, UT; Scott G. Muling, Bert E. Bledsoe, John E. Matthews, U.S.
EPA, Ada, OK; Daniel Pope, Dynamac Corporation, Ada, OK 35
Bioventing Soils Contaminated With Wood Preservatives
Paul 7. McCauley, Richard C. Brenner, Fran V. Kremer, U.S. EPA, Cincinnati, OH;
Bruce C. Alleman, Battelle Memorial Institute, Columbus, OH; Douglas C.
Beckwith, Minnesota Pollution Control Agency, St. Paul, MN 40
Field Evaluation of Fungal Treatment Technology
John A Glaser, U.S. EPA Cincinnati, OH; Richard T. Lamar, Diane M. Dietrich,
Mark W. Davis, Jason A. Chappelle, Laura M. Main, U.S. Department of
Agriculture, Madison, Wl 46 v
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Performance Evaluation
Integratfng Health Risk Assessment Data for Bioremediation
Larry D. C/axton, S. Elizabeth George, U.S. EPA, Research Triangle Park, NC 57
Construction of Noncolonizing E. Coli and P. Aeruginosa
Paul S. Cohen, University of Rhode Island, Kingston, Rl 59
Field Research
Field-Scale Study of In Situ Bioremediation of TCE-Contaminated Ground Water and
Planned Bioaugmentation
Perry L McCorty, Gary Hopkins, Stanford University, Stanford, CA 65
Geochemistry and Microbial Ecology of Reductive Dechlorination of PCE and TCE in
Subsurface Material
Guy W. Sewell, Candida C. West, Hugh Russell, U.S. EPA, Ada, OK; Susan A.
Gibson, William G. Lyon, ManTech Environmental Research Services Corp.,
Ada, OK 69
Application of Laser-Induced Fluorescence Implemented Through a Cone
Penetrometer To Map the Distribution of an Oil Spill in the Subsurface
Don H. Kampbell, Fred M. Pfeffer, John T. Wilson, U.S. EPA, Ada, OK; Bruce J.
Nielsen, Armstrong Laboratory, Tyndall AFB, FL 76
Effectiveness and Safety of Strategies for Oil Spill Bioremediation: Potential and
Limitations
Joe Eugene Lepo, University of West Florida, Pensacola, FL; C. Richard Cripe,
P.H. Pritchard, U.S. EPA, Gulf Breeze, FL 80
Pilot-Scale Research
Pilot-Scale Evaluation of Alternative Biofilter Attachment Media for Treatment of
VOCs
Francis L. Smith, George A. Sorial, Makram T. Sufdan, Prati'm Biswas, University of
Cincinnati, Cincinnati, OH; Richard C. Brenner, U.S. EPA, Cincinnati, OH 89
Biological Treatment of Contaminated Soils and Sediments Using Redox Control:
Advanced Land Treatment Techniques
Margaref J. Kupferle, In S. Kim, Guanrong You, Tiehong Huang, Maoxiu Wang,
University of Cincinnati, Cincinnati, OH; Gregory D. Sayles, U.S. EPA, Cincinnati,
OH; Douglas S. Upton, Levine-Fricke Consulting Engineers, Emeryville, CA 98
Research Leading to the Bioremediation of Oil-Contaminated Beaches
Albert D. Venosa, John R. Haines, U.S. EPA, Cincinnati, OH; Makram J. Suidan,
Brian A. Wrenn, Kevin L Strohmeier, B. Loye Eberhart, Edith L. Holder, Xiaolan
Wang, University of Cincinnati, Cincinnati, OH 103
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Engineering Optimization of Slurry Bioreactors for Treating Hazardous Wastes
John A Closer, Paul T. McCau/ey, U.S. EPA Cincinnati, OH; Ma/id A Dosani,
Jennifer S. P/aff, E. Radha Krishnan, /.T. Environmental Programs, Inc.,
Cincinnati, OH 109
Development and Evaluation of Composting Techniques for Treatment of Soils
Contaminated With Hazardous Wastes
Car/ L Potter, John A Closer, U.S. EPA Cincinnati, OH; Ma/id A Dosani, Srinivas
Krishnan, Timothy Deets, E. Radha Krishnan, I.T. Environmental Programs, Inc.,
Cincinnati, OH 116
Remediation of Contaminated Soils From Wood-Preserving Sites Using Combined
Treatment Technologies
An id P. Khodadoust, Gregory J. Wilson, Malcram T. Suidan, University of
Cincinnati, Cincinnati, OH; Richard C. Brenner, U.S. EPA Cincinnati, OH 120
Process Research
V
Metabolic and Ecological Factors Affecting the Bioremediation of PAH- and
Creosote-Contaminated Soil and Water
P.H. Pritchard, U.S. EPA Gulf Breeze, FL; Jian-Er Lin, Technology Resources, Inc.,
Gulf Breeze, FL; James G. Mueller, Suzanne Lanfz, SBP Technologies, Inc.,
Gulf Breeze, FL 129
Metabolic Pathways Involved in the Biodegradation of PAHs
Peter J. Chapman, Richard Eaton, U.S. EPA Gulf Breeze, FL; Sergey A Se/ifonov,
University of Minnesota, St. Paul, MM; Magda Grifoll, University of Barcelona,
Spain : 139
Environmental Factors Affecting Creosote Degradation by Sphingomonas
paucimobilis Strain EPA505
James G. Mueller, Suzanne E. Lanfz, SBP Technologies, Inc., Gulf Breeze, FL;
P.H. Pritchard, U.S. EPA Gulf Breeze, FL 143
Molecular Genetic Approaches to the Study of the Biodegradation of Polycyclic
Aromatic Chemicals
Richard W. Eaton, Peter J. Chapman, U.S. EPA Gulf Breeze, FL; James D.
Nifterauer, Technical Resources, Inc., Gulf Breeze, FL, and University of Arkansas
for Medical Sciences, Little Rock, AK 150
Comparison of Sulfur and Nitrogen Heterocyclic Compound Transport in Creosote-
Contaminated Aquifer Material
Ean M. Warren, E. Michael Godsy, U.S. Geological Survey, Menlo Park, CA 153
Modeling Steady-State Methanogenic Degradation of Phenols in Ground Water at
Pensacola, Florida
Barbara A Belcins, E. Michael Godsy, Donald F. Goer/ifz, U.S. Geological Survey,
Menlo Park, CA 158
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Anaerobic Biodegradation of 5-Chlorovanillate as a Model Substrate for the
Bioremediation of Paper-Milling Waste
B.R. Sharak Genfhner, 8.O. Blattmann, Avanti, Corp., Gulf Breeze, FL; P.H.
Pritchard, U.S. EPA, Gulf Breeze, FL 164
Characterization of a 4-Bromophenol Dehalogenating Enrichment Culture:
Conversion of Pentachlorophenol to Phenol by Sediment Augmentation
Xiaoming Zhang, National Research Council, National Academy of Sciences,
Washington, DC; W. Jack Jones, John E. Rogers, U.S. EPA, Athens, GA 171
Stimulating the Microbial Dechlorination of PCBs: Overcoming Limiting Factors
John F. Quensen, III, Stephen A Boyd, James M. T/'ed/e, Michigan State
University, East Lansing, Ml; John E. Rogers, U.S. EPA, Athens, GA 1 75
Potential Surfactant Effects on the Microbial Degradation of Organic Contaminants
Stephen A Boyd, John F. Quensen, III, Mahmoud Mousa, Jae Woo Park,
Michigan State University, East Lansing, Ml; Shaobai Sun, William Inskeep,
Montana State University, Bozeman, MJ 180
Enhanced Dechlorination of PCBs in Contaminated Sediments by Addition of Single
Congeners of Chloro- and Bromobiphenyls
W. Jack Jones, John E. Rogers, U.S. EPA, Athens, GA; Rebecca L. Adams,
Technology Applications, Inc., Athens, GA 1 84
Effect of Heavy Metal Availability and Toxicity on Anaerobic Transformations of
Aromatic Hydrocarbons
John H. Pardue, Ronald D. DeLaune, William H. Patrick, Jr., Louisiana State
University, Baton Rouge, LA 1 89
Biodegradation of Petroleum Hydrocarbons in Wetlands Microcosms
Rochelle Araujo, Marirosa Molina, U.S. EPA, Athens, GA; Dave Bachoon,
University of Georgia, Athens, GA; Lawrence D. LaPlante, Technology
Applications, Inc., Athens, GA 1 94
Biodegradation of Petroleum Hydrocarbons in Wetlands: Constraints on Natural and
Engineered Remediation
John H. Pardue, Andrew Jackson, Ronald D. DeLaune, Louisiana State University,
Baton Rouge, LA 201
Anaerobic Biotransformation of Munitions Wastes
Deborah J. Roberts, Farrukh Ahmad, University of Houston, Houston, TX; Don L.
Crawford, Ronald L Crawford, University of Idaho, Moscow, ID 206
Covalent Binding of Aromatic Amines to Natural Organic Matter: Study of Reaction
Mechanisms and Development of Remediation Schemes
Eric J. Weber, Dalizza Colon, U.S. EPA, Athens, GA; Michael S. Elovitz, National
Research Council, Athens, GA 213
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Kinetics of Anaerobic Biodegradation of Munitions Wastes
J/'ayang Cheng, Mafcram T. Suidan, University of Cincinnati, Cincinnati, OH;
Albert D. Venosa, U.S. EPA Cincinnati, OH 218
Biodegradation of Chlorinated Solvents
Sergey A Se/ifonov, Lisa N. Newman, Michael E. She/ton, Lawrence P. Wackeff,
L/nivers/fy of Minnesota, St. Paul, MN 223
Characterization of Bacteria in a TCE Degrading Biofilter
Alec W. Breen, A/ex Rooney, Todd Ward, John C. Loper, Ralcesh Govind,
University of Cincinnati, Cincinnati, OH; John R. Haines, U.S. EPA
Cincinnati, OH 229
Bioremediation of TCE: Risk Analysis for Inoculation Strategies
Richard A Snyder, Malcolm S. Shields, University of West Florida, Pensaco/a, FL;
P.H. Pritchard, U.S. EPA Gu/f Breeze, FL 234
Studies on the Aerobic/Anaerobic Degradation of Recalcitrant Volatile Chlorinated
Chemicals in a Hydrogel Encapsulated Biomass Biofilter
Ralcesh Govind, P.S.R.V. Prasad, University of Cincinnati, Cincinnati, OH; Do/loff
F. Bishop, U.S. EPA Cincinnati, OH 238
Poster Session
Pilot-Scale Evaluation of Nutrient Delivery for Oil-Contaminated Beaches
Michael Boufadel, Malcram T. Suidan, University of Cincinnati, Cincinnati, OH;
Albert D. Venosa, U.S. EPA Cincinnati, OH 245
Metabolites of Oil Biodegradation and Their Toxicity
Peter J. Chapman, Steven S. Foss, Douglas P. Middaugh, William S. Fisher, U.S.
EPA Gulf Breeze, FL; Michael E. Shelton, University of Minnesota, St. Paul, MN;
Simon Akkerman, University of West Florida, Pensacola, FL 246
The Use of In Situ Carbon Dioxide Measurement To Determine Bioremediation
Success
Richard P.J. Swannell, AEA Technology, Oxon, United Kingdom; Francois X.
Merlin, CEDRE, Plouzane, Brest, France 248
Toxicant Generation and Removal During Crude Oil Degradation
Linda E. Rudd, Jerome J. Perry, North Carolina State University, Raleigh, NC;
Larry D. C/axfon, Virginia S. Houk, Ron W. Williams, U.S. EPA, Research
Triangle Park, NC 248
Intrinsic Bioremediation of JP-4 Jet Fuel Contamination at George AFB, California
John T. Wilson, Michael L. Coolc, Don H. Kampbell, U.S. EPA Ada, OK 254
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Field Treatment of BTEX in Vadose Soils Using Vacuum Extraction or Air Stripping
and Biofilters
Rakesh Govind, University of Cincinnati, Cincinnati, OH; E. Rac/ha Krishnan,
Gerard Henderson, International Technology Corporation, Cincinnati, OH; Dolloff
F. Bishop, U.S. EPA, Cincinnati, OH 255
TCE Remediation Using a Plasmid Specifying Constitutive TCE Degradation:
Alteration of Bacterial Strain Designs Based on Field Evaluations
Malcolm S. Shields, Allison Blake, Michael Reagin, Tracy Moody, Kenneth
Overstreet, Robert Campbell, University of West Florida, Pensacola, FL; Stephen
C. Francesconi, P.H. Pritchard, U.S. EPA, Gulf Breeze, FL 258
Dechlorination With a Biofilm-Electrode Reactor
John W. Norton, MaJcram T. Suidan, University of Cincinnati, Cincinnati, OH;
Albert D. Venose, U.S. EPA, Cincinnati, OH 259
Degradation of a Mixture of High Molecular-Weight Polycyclic Aromatic
Hydrocarbons by a Mycobacferium Species
/. Kelley, A. Selby, Carl E. Cemiglia, U.S. Food and Drug Administration,
Jefferson, AR 262
Potentiation of 2,6-Dinitrotoluene Bioactivation by Atrazine in Fischer 344 Rats
S. Elizabeth George, Robert W. Chadwick, Michael J. /Cohan, Joycelyn C. Allison,
Larry D. C/axfon, U.S. EPA, Research Triangle Park, NC; Sarah H. Warren, Ron W.
Williams, Integrated Laboratory Systems, Research Triangle Park, NC 263
Effects of Lactobacillus Reuteri on Intestinal Colonization of Bioremediation Agents
Mitra Fiuzat, Walter J. Dobrogosz, North Carolina State University, Raleigh, NC; /
S. Elizabefh George, U.S. EPA Research Triangle Park, NC 264 V
Bioavailability Factors Affecting the Aerobic Biodegradation of Hydrophobic
Chemicals •
Pamela J. Morn's, University of Florida/U.S. EPA Gulf Breeze, FL; Suresh C. Rao,
University of Florida, Gainesville, FL; Semen Akkerman, University of West
Florida/U.S. EPA, Gulf Breeze, FL; Michael E. She/ton, University of
Minnesofa/U.S. EPA Gulf Breeze, FL; Peter J. Chapman, P.H. Pritchard, U.S. EPA,
Gulf Breeze, FL 268
Use of Sulfur Oxidizing Bacteria To Remove Nitrate From Ground Water
Michael S. Davidson, Thomas Cormack, Harry Ridgway, Grisel Rodriguez, Orange
County Water District, Fountain Valley, CA 270
Engineering Evaluation and Optimization of Biopiles for Treatment of Soils
Contaminated With Hazardous Waste
Carl L Potter, John A G/aser, U.S. EPA Cincinnati, OH 272
Factors Affecting Delivery of Nutrients and Moisture for Enhanced In Situ
Bioremediation in the Unsaturated Zone
James G. Ufaer, Ronghui Liang, University of Cincinnati, Cincinnati, OH; Paul T.
McCauley, U.S. EPA Cincinnati, OH 273
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The Bioremediation in the Field Search System (BFSS)
Fran V. Kremer, U.S. EPA Cincinnati, OH; Linda B. Diamond, Susan P.E.
Richmond, Jeff B. Box, /van B. Rudnicki, Eastern Research Group, Inc.,
Lexington, MA 274
Hazardous Substance Research Centers
In Situ Attenuation of Chlorinated Aliphatics in Glacial Alluvial Deposits
Michael J. Barcelona, Mark A Henry, Walter J. Weber, Jr., University of Michigan,
Ann Arbor, Ml (Great Lakes and Mid-Atlantic Hazardous Substance Research
Centerj 279
In Situ Bioremediation of Chlorinated Solvent Ground-Water Contamination: Scaling
up From a Field Experiment to a Full-Scale Demonstration
Perry L. McCarty, Gary D. Hopkins, Mark N. Goto, Stanford University, Stanford,
CA (Western Regional Hazardous Substance Research Center) 281
Bioavailability and Transformation of Highly Chlorinated Dibenzo-p-Dioxins and
Dibenzofurans in Anaerobic Soils and Sediments
Peter Adriaens, Quingzhai Fu, University of Michigan, Ann Arbor, Ml (Great Lakes
and Mid-Atlantic Hazardous Substance Research Center) 283
Localization of Tetrachloromethane Transformation Activity in Shewane/la Putrefaciens
MR-1
Erik A Petrovslcis, Peter Adriaens, Timothy M. Vogel, University of Michigan, Ann
Arbor, Ml (Great Lakes and Mid-Atlantic Hazardous Substance Research Center) . . . 284
Formation and Transformation of Pesticide Degradation Products Under Various
Electron Acceptor Conditions
Paige J. Novalc, Gene F. Parkin, Craig L Just, University of Iowa, Iowa City, IA
(Great Plains and Rocky Mountain Hazardous Substance Research Center) 285
Bioremediation of Aromatic Hydrocarbons at Seal Beach, California: Laboratory and
Field Investigations
Harold A Ball, Gary D. Hopkins, Eva Orwin, Martin Reinhard, Western Region
Hazardous Substance Research Center, Stanford, CA (Western Region Hazardous
Substance Research Center) 291
Pneumatic Fracturing To Enhance In Situ Bioremediation
John R. Schuring, Thomas A Bo/and, New Jersey Institute of Technology, Newark,
NJ; David S. Kosson, Shankar Venkatraman, Rutgers University, Piscataway, NJ
(Northeast Hazardous Substance Research Center) 294
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Bioremediation Field Initiative
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Intrinsic Bioremediation of TCE in Ground Water at an NPL Site
in St. Joseph, Michigan
John T. Wilson, James W. Weaver, and Don H. Kampbell
Robert S. Kerr Environmental Research Laboratory, U.S. Environmental Protection Agency,
Ada, OK
Introduction
The ground water at the St. Joseph, Michigan, National Priority List (NPL) site is contaminated
with chlorinated aliphatic compounds (CACs) at concentrations in the range of 10 mg/L to 100
mg/L. The chemicals are thought to have entered the shallow sandy aquifer either through
waste lagoons, which were used from 1968 to 1976, or through disposal of trichloroethylene
(TCE) into dry wells at the site (1). The contamination was determined to be divided into eastern
and western plumes, as the suspected sources were situated over a ground-water divide. Both
plumes were found to contain TCE, cis- and trans-1,2-dichloroethylene (c-1,2-DCE and t-1,2-
DCE), 1,1-dichloroethylene (1,1-DCE), and vinyl chloride (VC).
Previous investigation of the site indicated that natural anaerobic degradation of the TCE was
occurring because of the presence of transformation products and significant levels of ethene
and methane (2,3). The purpose of this presentation is to provide the results of later sampling
of the western plume near Lake Michigan, to estimate the contaminant mass flux, and to
estimate apparent degradation constants. The estimates are based on visualization of the data
representing each measured concentration by a zone of influence based on the sample spacing.
The presentation of the data is free from artifacts of interpolation, and extrapolation of the data
beyond the measurement locations is controlled.
Data Summary
In 1991 three transects (1,2, and 3 on Figure 1) were completed nearthe source of the western
plume (2). The three transects consisted of 1 7 borings with a slotted auger. In 1992 two
additional transects (4 and 5 on Figure 1) were completed consisting of 9 additional slotted
auger borings. In each boring, water samples were taken on roughly 1.5 m (5 ft) depth
intervals. Onsite gas chromatography was performed to determine the width of the plume and
to find the point of highest concentration. Three of the transects (2, 4, and 5) were roughly
perpendicular to the contaminant plume. Of the remaining transects, transect 1 crosses the
plume at an angle and transect 3 lies along the length of the plume. The perpendicular
transects form logical units for study of TCE biotransformation.
The site data from the transects are visualized as sets of blocks centered around the
measurement point. The blocks are defined so that the influence of a particular measured
concentration extends halfway to the next measurement location both horizontally and vertically.
Thus, the presentation of the data is simple and direct. The visualization of the data is
performed on a Silicon Graphics Indigo workstation using a two-dimensional version of the fully
1994 Symposium on Bioremediation of Hazardous Wastes
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three-dimensional field-data analysis program called SITE-3D, which is under development at
the Robert S. Kerr Environmental Research Laboratory.
The mass of each chemical per unit thickness and the advective mass flux of each chemical are
calculated by summing over the blocks. By following this procedure, the measured chemical
concentrations are not extrapolated into the clay layer under the site. Neither are they
extrapolated beyond a short distance from the measurement locations (5 ft vertically and 50 ft
to TOO ft horizontally). Other interpolation schemes, such as inverse distance weighting or
kriging, could also be used to estimate the concentration field and perform the mass estimates.
Figures 2 and 3 show the distributions of VC and TCE at transect 5 using a logarithmic, black-
and-white "color" scale. Notably, the maximum VC concentration at transect A was 1,660^g/l
and at transect 5 was 205 /Mg/L . The maximum TCE concentration at transect 4 was 8,720
/ig/L and at transect 5 was 163 /*g/L . As noted previously for other portions of the site (2,4),
the contamination is found near the bottom of the aquifer. The highest concentrations of VC
and TCE do not appear to be co-located. In Table 1, mass estimates are presented for the
perpendicular transects ordered from furthest upgradient (transect 2) to furthest downgradient
(transect 5). The data in Table 1 represent the mass in a volume of aquifer that has an area
equal to the cross-sectional area of the transect and is 1.0 m thick in the direction of ground-
water flow.
Advective Mass Flux Estimates
Results from the calibrated MODFLOW model of Tiedeman and Gorelick (4) were used to
estimate the ground-water flow velocity at each transect. The estimate is an upper bound
because the modeled vertical component of flow was neglected in the present analysis. The
head drop from one location to the next was assumed to generate horizontal flow only.
Tiedeman and Gorelick (4) also represented the aquifer by single values of hydraulic conductivity
and porosity. They gave, however, 95-percent confidence limits for the hydraulic conductivity.
Well yields estimated for each sample location indicate declining hydraulic conductivity toward
the west, i.e., towards Lake Michigan and transects 4 and 5. Thus, using the single parameter
values from the MODFLOW simulations may overestimate the flux of water into the lake.
As would be expected, the advective mass fluxes decline toward the downgradient edge of the
plume. There, the concentrations are lower due to either transient flow or degradation of the
TCE. Notably the mass fluxes using the average hydraulic conductivity result in a total flux of
13 kg/y of TCE, c-1,2-DCE, t-1,2-DCE, 1,1 -DCE, and VC at transect 5. This value contrasts
with the total flux of these CACs of 310 kg/y at transect 2, near the source of contamination.
Thus, there is a 24.4-fold decrease in mass flux of CACs across the site. Given the 95 percent
confidence limits on the hydraulic conductivity determined by Tiedeman and Gorelick (4), the
total range of mass flux of these five chemicals is from 205 kg/y to 420 kg/y at transect 2 and
from 8.4 kg/y to 1 7 kg/y at transect 5. The range of fluxes at transect 5 is an upper bound on,
and the best estimate of, the flux into Lake Michigan.
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Apparent Degradation Constants
The mass per unit thickness of TCE at transects 2, 4, and 5 was used to estimate apparent first-
order degradation constants. The constants are estimated by applying the first order rate
equation
In
ci
X At
to the site data, where GJ is the average concentration in the transect j, cj+1 is the average
concentration in the downgradient transect j+1, At is the advective travel time for TCE to move
between the transects, and A is the apparent degradation constant. The mass per unit thickness
data for TCE and the cross sectional area were used to determine the average concentrations
Cj and C|+1 in the up- and downgradient transects. The porosity, bulk density, fraction organic
carbon, organic carbon partition coefficient (5), ground-water gradient, and distance between
the transects were used to determine the advective travel times. The values used in Equation
1 are given in Table 3. From these quantities, the apparent degradation constant for TCE was
determined to be -0.0076/week from transect 2 to 4 and -0.024/week from transect 4 to 5.
References
1. Engineering Science, Inc. 1990. Remedial investigation and feasibility study, St.
Joseph, Michigan, phase I technical memorandum. Liverpool, NY.
2. Kitanidis, P.K., L. Semprini, D.H. Kampbell, and J.T. Wilson. 1993. Natural anaerobic
bioremediation of TCE at the St. Joseph, Michigan, Superfund site. Symposium on
Bioremediation of Hazardous Wastes: Research, Development, and Field Evaluations.
EPA/600/R-93/054. pp. 57-60.
3. McCarty, P.L., and J.T. Wilson. 1992. Natural anaerobic treatment of a TCE plume
at the St. Joseph, Michigan, NPL site. In: U.S. EPA. Bioremediation of hazardous
wastes (abstracts). EPA/600/R-92/126. pp. 47-50.
4. Tiedeman, C., and S. Gorelick. 1993. Analysis of uncertainty in optimal groundwater
contaminant capture design. Water Resour. Res. 29(7):2139-2153.
5. U.S. EPA. 1990. Subsurface remediation guidance table 3. EPA/540/2-90/011 b.
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Table 1. Mass per Unit Thickness (kg/m) at St. Joseph, Michigan
Chemical
VC
1,1 -DCE
t-1,2-DCE
c-l,2-DCE
TCE
Methane
Ethene
Ethane
TOC
Chloride
Sulfate
NO3-
Nitrogen
NH4-
Nitrogen
TKN-
Nitrogen
Transect
2
1.523
0.2377
0.566
12.32
10.67
5.855
0.6847
no data
no data
129.9
37.05
2.904
1.835
2.987
1
1 .8969
0.0816
0.5059
5.1127
5.5804
5.4826
0.8925
no data
no data
148.8
34.376
2.471
2.5609
3.8357
4
0.4868
0.01451
0.03628
i .890
1.397
4.620
0.1747
0.2085
12.63
213.1
95.78
4.421
0.4562
0.6353
5
0.0481 1
0.001047
0.007041
0.2832
0.02821
1.373
0.004901
0.001689
8.314
156.2
66.19
8.247
0.2256
0.3646
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Table 2. Mass Flux (kg/y) at St. Joseph, Michigan
Chemical
VC
1,1 -DCE
t-1 ,2-DCE
c-1,2-DCE
TCE
Methane
Ethene
Ethane
TOC
Chloride
Sulfate
NCy
Nitrogen
NH4-
Nitrogen
TKN-
Nitrogen
Transect
2
18.81
2.934
6.995
152.1
131.7
72.29
8.453
no data
no data
1604
457.4
35.85
22.66
36.88
1
36.03
1.551
9.609
97.11
106.0
104.1
16.95
no data
no data
2826
652.9
46.93
48.64
72.85
4
10.69
0.3185
0.7963
41.48
30.67
101.4
3.836
4.577
277.2
4678
2102
97.05
10.01
13.95
5
1.676
0.03648
0.2453
9.868
0.9829
47.86
0.1708
0.05885
289.7
5444
2306
287.4
7.861
12.70
-------
Table 3. Chemical and Hydraulic Values Used in Estimating Apparent Degradation Rates
Transect
2
4
5
Area with
non-zero
TCE
concentra-
tion
(m2)
1592
2774
1943
Mass per
unit
thickness
from
SITE-3D
(kg/m)
10.67
1.397
0.0282
Average TCE
concentra-
tion in the
transect
(kg/m3)
CjOnd Cj+1 in
Equation 1
6.70e-3
5.04e-4
1 ,44e-5
Distance
between
transects
(m)
260
160
Gradient
estimated
from
Tiedeman
and
Gorelick
(1993)
7.3e-3
1.1 e-2
"Re-
tarded
seepage
velocity
for TCE
(m/d)
0.11
0.156
Estimated
travel time
between
transects
(weeks)
At in
Equation 1
340
•
145
"Constants used in seepage velocity calculation
Hydraulic conductivity: 7.5 m/d
Retardation factor for TCE: 1.78 = 1 +
Porosity, 9: 0.30
Bulk density pb: 1.86 g/cm3
r^: 126mL/g
L: 0.001
-------
St Joseph, Michigan
NPL Site
Figure 1. St. Joseph, Michigan, NPL site plan.
St. Joseph, Michigan
Vinyl Chloride
transect: 5
mass: 0.4811 E-01 Kg/m
tsi 152
10 feet ,
•pprox. N
t54 tS3 tSS
100 feet
Concentration
ug/L
250000.1
25000.
2500. |
j
250.0
25.00
2.500
0.2500
0.0250
Figure 2. VC distribution at transect 5.
-------
St. Joseph, Michigan
Trlchloroethene
transect: 5
mass: 0.2821 E-01 Kg/m
151 152
154 153 tSS
Ground »urt»ct
10 feet
approx. N
100 feet
Concentration
ugA.
250000.
25000.
2500.
250.0
25.00
2.500
0.2500
0.0250
Figure 3. TCE distribution at transect 5.
10
-------
Enhanced Reductive Dechlorination of Chlorinated Ethenes
Zachary C. Hasten, Pramod K. Sharma, James N.P. Black, and Perry L McCarty
Western Region Hazardous Substance Research Center, Stanford University, Stanford, CA
Introduction
Reductive dehalogenation of trichloroethylene (TCE) to cis-1,2-dichloroethylene (c-l,2-DCE),
trans-1,2-dichloroethylene (t-l,2-DCE), vinyl chloride (VC), and ethene was found to be
occurring at a site in St. Joseph, Michigan, by indigenous microbial populations under
anaerobic conditions (1). This has raised two possibilities for further study: 1) that the natural
anaerobic processes at the site may be sufficient to bring about site remediation alone or 2) that
the natural process will be incomplete without some enhancement. Further site characterization
is now under way by the EPA Robert S. Kerr Environmental Research Laboratory to determine
the extent of natural onsite transformation. This study aims to determine whether enhancement
of the anaerobic process might be beneficial, what microorganisms are responsible for the
natural transformation, and what is an effective primary substrate to add to the ground water
for enhancing the remediation in s/fu. For comparison, aquifer material from a site in Victoria,
Texas, is also being evaluated. This site is contaminated by tetrachloroethylene
(perchloroethylene, orPCE) and is being actively bioremediated by the addition of benzoate and
sulfate (2).
Methods
Aquifer material for this study was obtained aseptically in the absence of oxygen from both St.
Joseph and Victoria sites. The potential of the St. Joseph aquifer material for TCE
transformation and the effect of adding different primary substrates were studied using 25 ml
test tubes as small laboratory columns (3). The fluid within the test tubes was exchanged after
incubation periods ranging from 1 to 4 months with filter-sterilized site ground water that was
amended with a primary substrate and TCE. Control columns received TCE-amended,
filter-sterilized ground water without an added primary substrate. Between fluid exchanges, the
openings were sealed, and the columns were incubated without fluid exchange in a room
temperature anaerobic glovebox containing 1 percent to 10 percent hydrogen. Each primary
substrate was fed to yield 100 mg/L chemical oxygen demand (COD) to provide similar
reducing equivalents for each column. Each column was fed only one substrate from the time
the column was prepared.
In addition, microcosms consisting of 125 ml bottles containing aquifer material and site
ground water were used to simulate in s/fu conditions with the Victoria aquifer material. Only
110 ml of saturated aquifer material was used in the bottles to allow for sampling of the liquid
from the remaining 15 ml, and to provide for bed fluidization during mixing. These microcosms
were incubated without headspace.
Enrichments were developed by the addition of Victoria aquifer-material to a basal medium (4).
This enrichment was subsequently transferred to aquifer-material-free media. The effect of
different metabolic inhibitors was studied using an inoculum from a benzoate enrichment culture
1994 Symposium on Bioremediatlon of Hazardous Wastes 11
-------
into 1 60 ml bottles filled with 120 ml of defined media amended with PCE, benzoate, yeast
extract, and the respective inhibitor.
Results
The possibility of enhancing biodegradation by the addition of various primary substrates was
studied using columns of St. Joseph material. Table 1 shows the resulting concentrations of TCE
dechlorination products after a typical 6-week incubation period. Following this exchange, the
ethanol-fed column was switched to benzoate and immediately performed similar to the column
that had been fed benzoate from the start.
Of the primary substrates tested, benzoate addition consistently stimulated the most complete
dechlorination. Similar results were obtained with the microcosms containing Victoria aquifer
material (data not shown). No significant lag time before the onset of dechlorination was
observed with either material.
In the St. Joseph unfed column control, partial dechlorination of TCE to c-l,2-DCE was
observed over several exchanges spanning several months. This may have been associated with
oxidation of natural organics within the aquifer material or of hydrogen that diffused into the
column from the glovebox gases. Victoria microcosms also showed some dechlorination of PCE
to TCE in the unfed controls.
For column studies with St. Joseph material, site ground water was used that included 0.49 mM
nitrate and 0.50 mM sulfate. During incubation in the substrate-amended columns, nitrate and
sulfate were consumed completely, and varying amounts of methane were produced. Nitrate
also disappeared in the unfed control, but no sulfate was consumed or methane produced.
Dechlorination accounted for less than 2 percent of the substrate utilized; nitrate reduction,
sulfate reduction, and methanogenesis accounted for the rest.
After several exchanges, the primary substrate-fed columns became clogged. Small entrapped
bubbles were visible in the columns as well as a noticeable amount of black precipitate.
Considering the amount of primary substrate added to the columns, up to about a fifth of the
pore volume could have been filled by methane formation. The extent of the clogging caused
by iron sulfide precipitate or biomass is unknown, but after a few months, during which the
columns sat unfed, the entrapped bubbles visibly decreased and the columns became
unclogged. Bubbles also formed in the Victoria microcosms, but they were allowed to come to
the surface during daily shaking and were removed during analysis.
PCE was not dechlorinated within 2 months in microcosms containing a defined mineral media
amended with only benzoate, while the addition of benzoate and 0.05 percent yeast extract
stimulated dechlorination of all the PCE completely to ethene (data not shown). The addition
of benzoate and sulfate stimulated partial dechlorination, as did the addition of yeast extract
alone.
Studies of the effects of various metabolic inhibitors were conducted to better understand the
role of sulfate-reducing and methanogenic bacteria. Table 2 lists duplicate live bottles from a
3-month incubation with 0.416 mM benzoate, 0.01 percent yeast extract, and various
amendments, including 2 mM sulfate, 0.5 mM bromoethanesulfonic acid (BESA), and 0.5 mM
12
-------
molybdate, where applicable. No dechlorination was observed in uninoculated or sterile
controls. t-l,2-DCE and 1,1-dichloroethylene were not observed in the enrichment cultures.
Summary and Conclusions
Studies with aquifer material from both contaminated sites have shown that all primary substrates
tested were capable of stimulating dechlorination of some PCE orTCE to ethene, with benzoate
consistently stimulating the most complete degradation. High sulfate concentrations appear to
inhibit dechlorination, although no dechlorination was observed in microcosms incubated
without some sulfate or yeast extract. The addition of molybdate reversed sulfate inhibition, but
here dechlorination stopped at c-1,2-DCE. These data show that the anaerobic dechlorination
of PCE or TCE to ethene can be enhanced by the appropriate addition of a primary substrate
and yeast extract or sulfate.
References
1. McCarty, P.L., and J.T. Wilson. 1 992. Natural anaerobic treatment of a TCE plume,
St. Joseph, Michigan, NPL site. In: U.S. EPA. Bioremediation of hazardous wastes
(abstracts). EPA/600/R-92/126. Cincinnati, OH. pp. 47-50.
2. Beeman, R.E. 1994. In situ biodegradation of ground-water contaminants. U.S. Patent
No. 5,277,815.
3. Siegrist, H., and P.L. McCarty. 1987. Column methodologies for determining sorption
and biotransformation potential of chlorinated aliphatic compounds in aquifers. J.
Contam. Hydrol. 2:31-50.
4. Tanner, R.S., and R.S. Wolfe. 1988. Nutritional requirements of A/lefhanomicrob/um
mob//e. Appl. Environ. Microbiol. 54:625-628.
13
-------
Table 1. Concentration of TCE Dechlorination Products After 6 Weeks of Incubation in
St. Joseph Aquifer Material Columns*
Added
Substrate
None
Benzoate
Lactate
Sucrose
Ethanol
Methanol
Acetate
Compoui
TCE
20.5
0
0.5
2.4
3.6
9.3
10.4
ids remai
cDCE
4.8
0
4.1
5.4
3.7
6.3
3.4
ning after s
1,1-DCE
0
0
0
0.7
0.6
0.7
0.9
x weeks
VC
0
11.6
13.5
16.1
14.1
7.9
7.1
of incuba
Ethene
0
14.4
5.8
5.1
2.4
1.6
1.9
tion (nM)
Sum
25.3
26.0
23.9
29.7
24.4
25.8
23.7
*t-l ,2-DCE was also present in some columns in trace amounts.
Table 2. Effects of Inhibitors on Dechlorination*
Amendments
Benzoate and Yeast
Extract
Benzoate, Yeast Extract,
and BESA
Benzoate, Yeast Extract,
and Molybdate
Benzoate, Yeast Extract,
and Sulfate
Benzoate, Yeast Extract,
Sulfate, and Molybdate
Benzoate, Yeast Extract,
Sulfate, and BESA
(xmoles
PCE
0.00
0.00
0.00
0.00
0.05
0.01
1.06
1.07
0.00
0.00
0.97
0.96
remainin
TCE
0.00
0.00
0.00
0.00
0.06
0.01
0.30
0.31
0.00
0.00
0.33
0.35
g in dupli
cDCE
0.00
0.00
0.00
0.00
1.52
1.65
0.15
0.13
1.78
1.65
0.24
0.24
cate bott
VC
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
les after in
Ethene
1.70
1.76
1.63
1.62
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
cubation
Sum
1.70
1.76
1.63
1.62
1.63
1.67
1.51
1.50
1.78
1.65
1.54
1.55
*Values for PCE and its dechlorination products from duplicate cultures incubated for 3 months
at room temperature.
-------
Bioventing of Jet Fuel Spills I:
Bioventing in a Cold Climate With Soil Warming at Eielson AFB, Alaska
Gregory D. Sayles and Richard C. Brenner
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Robert E. Hinchee and Andrea Leeson
Battelle Memorial Institute, Columbus Division, Columbus, OH
Catherine M. Vogel
U.S. Air Force, Armstrong Laboratories, Tyndall Air Force Base, FL
Ross N. Miller
U.S. Air Force, Center for Environmental Excellence, Brooks Air Force Base, TX
Introduction
Bioventing is a process that supplies oxygen in situ to oxygen deprived soil microbes by forcing
air through unsaturated contaminated soil at low flow rates (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.
Previous work (2) has demonstrated that biodegradation rates associated with bioventing are
temperature dependent. Briefly, the goal of the current study is to demonstrate bioventing in a
cold climate and to evaluate several low-intensity soil warming methods for the ability to
maintain greater than average soil temperatures and rates of biodegradation.
The EPA Risk Reduction Engineering Laboratory, with resources from EPA's Bioremediation Field
Initiative, began a 3-year field study of ;h s/fu 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 has occurred in association with a fuel
distribution network. The contractor operating the project is Battelle Memorial Institute,
Columbus, Ohio. This report summarizes the first 2Vi years of operation.
Methodology
Site history, characterization, installation, and monitoring were summarized previously (3,4,5).
Figure 1 shows a plan view of the project.
Briefly, four 50 ft x 50 ft test plots have been established, all receiving relatively uniform injection
of air. The four test plots are being used to evaluate three soil warming methods:
1994 Symposium on Bioremediation of Hazardous Wastes 15
-------
• Passive warming: Enhanced solar warming in late spring, summer, and early
fall using a clear plastic covering over the plot; and passive heat retention the
remainder of the year by applying insulation to 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.
• Buried heat tape warming: Warming by heat tape buried at a depth of 3 ft and
distributed throughout the plot 5 ft apart. The tape heats at a rate of 6 W/ft,
giving a total heat load to the plot of roughly 1 W/ft2.
The contaminated control consists of contaminated soil vented with injected air with no artificial
method of heating.
The passively heated, actively heated, and control test plots were installed in the summer of
1 991, and the heat tape plot was installed in September 1992. Air injection/withdrawal wells
and soil gas and temperature monitoring points are distributed throughout the site. (See Figure
1.) Heating of the actively heated plot was discontinued in July 1993 to compare heated and
unheated biodegradation rates at the same location.
Periodically, in situ respirometry tests (6) are conducted to measure in situ oxygen uptake rates
by the microorganisms. These tests allow estimation of the biodegradation rate as a function
of time and, therefore, as a function of ambient temperature and soil warming technique. The
rate of oxygen use can be converted into the rate of petroleum use by assuming a stoichiometry
of biodegradation (4). Quarterly comprehensive and monthly abbreviated in situ respiration tests
were conducted.
Final soil hydrocarbon analyses will be conducted in the summer of 1 994 and compared with
initial soil analyses to document actual hydrocarbon loss due to bioventing.
Results
Evaluation of Soil Warming Methods
Figure 2 displays the average temperature of each plot and at an uncontaminated background
location as a function of time during the study. By applying warm water to the plot, the
temperature of the actively heated plot was maintained in the range of 10°C to 25°C, compared
with the contaminated (unheated) control plot where the minimum winter temperature is roughly
0°C. When heating of the actively heated plot was terminated in July 1993, its temperature
followed the temperature of the unheated control plot closely, as expected.
The ability to control temperature in the passively heated plot was not as successful. The
temperature of the passively heated plot roughly mimicked the contaminated control plot
temperature except during the summer of 1992, when the passively heated plot was roughly 5°C
warmer than the control plot. The insulation applied during the winter has been marginally
16
-------
successful at best, providing 1°C to 2°C temperature elevation in the passively heated plot
relative to the control plot.
Heating by buried heat tape in the surface heated plot has been successful at maintaining
temperatures between 10°C and 22°C year-round. Temperatures achieved in this plot in the
summer were much higher than those maintained in the winter because, although the heat input
was constant, the ambient temperature was much higher in the summer.
Rate of Biodegradation
The rate of jet fuel biodegradation, estimated by in s/fu respirometry tests, as a function of time
for each plot is shown in Figure 3. The influence of temperature on the rate is clear: the actively
warmed and surface warmed plots maintained rates two to three times greater than the
unheated control plot year-round. The small difference in temperature between the passively
warmed and the control plots (see Figure 2) is reflected in the small difference in respective rates
measured in these plots.
Researchers commonly believed that bioremediation systems should be shutdown for the winter
in any cold climate because microbial activity is thought to approach zero at these low
temperatures. The rate was nonzero (roughly 0.5 mg/kg/day), however, in the unheated control
plot in the middle of winter in Alaska, when the average temperature of the plot was roughly
0°C (see Figure 2).
After July 1 993, when heating of the actively warmed plot was discontinued, the rate observed
in this plot was not significantly different than the rate measured from the unheated control plot,
consistent with the similar temperatures of these two plots.
Conclusions
Application of warm water and heat generated by electrical resistance has been successful at
maintaining summer-like temperatures in the soil year-round. The enhanced temperatures in
the plots provided elevated rates of biodegradation. The passively warmed plot has performed
only marginally better than no heating (the contaminated control) with respect to temperature
and rate.
At the conclusion of this study, a cost-benefit analysis will be conducted to compare the
performance of the heating methods in terms of rate enhancement versus cost of heating.
References
1. Hoeppel, R.E., R.E. Hinchee, and M.F. Arthur. 1991. Bioventing soils contaminated
with petroleum hydrocarbons. J. Indust. Microbiol. 8:141-146.
17
-------
2. Miller, R.N., R.E. Hinchee, and CM. Vogel. 1991. A field-scale investigation of
petroleum hydrocarbon biodegradation in the vadose zone enhanced by soil venting at
Tyndall AFB, Florida. In: Hinchee, R.E., and R.F. Olfenbuttel, eds. In situ
bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 283-302.
3. Sayles, G.D., R.C. Brenner, R.E. Hinchee, C.M. Vogel, and R.N. Miller. 1992.
Optimizing bioventing in shallow vadose zones and cold climates: Eielson AFB
bioremediation of a JP-4 spill. In: U.S. EPA. Symposium on bioremediation of
hazardous wastes (abstracts). EPA/600/R-92/126. Washington, DC (May), pp. 31-35.
4. Leeson, A., R.E. Hinchee, J. Kittel, G. Sayles, C.M. Vogel, and R.N. Miller. 1 993.
Optimizing bioventing in shallow vadose zones and cold climates. Hydrological Sci.
38(4):283-295.
5. Ong, S.K., A. Leeson, R.E. Hinchee, J. Kittel, C.M. Vogel, G.D. Sayles, and R.N. Miller.
1994. Cold climate applications of bioventing. In: Hinchee, R.E., et al., eds.
Hydrocarbon bioremediation. CRC Press, pp. 444-453.
6. Ong, S.K., R.E. Hinchee, R. Hoeppel, and R. Schultz. 1991. In situ respirometry for
determining aerobic degradation rates. In: Hinchee, R.E., and R.F. Olfenbuttel, eds.
In situ bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 541-545.
18
-------
\
N
-WO1 T.
e *-i o ••« o ••» o ••
o M
Tuhviy
O • Oroundwiter monHortng (Mil
• -Mr Injection/withdrawal mil
t • Thr**-lw«l Mil g» prob*
T - Ttw««-l«v*l thtrmocoupl* prab*
O -Alrln|*cllan/MNhdrawilw*U
(Cumnlly nal In UM)
Figure 1. Plan view of the EPA/U.S. Air Force bioventing system at Eielson AFB near Fairbanks,
Alaska. "S" represents three-level soil gas monitoring points, T' represents three-level
temperature probes, and" " and "•" represent inactive and active air injection wells,
respectively. Instrumentation in the lower left is the uncontaminated background
location.
19
-------
Acuve Warming
Passive Warming
Contaminated Control
Surface Warming
Uncontaminaled Background
1991
1992
1993
Figure 2. Average temperature of each plot and at an uncontaminated background location
at the Eielson AFB bioventing site as a function of time during the study.
20
-------
Active Warming
Passive Warming
Contaminated Control
Surface Warming
August November || January
October || January
1991
1992
1993
Figure 3. Average rate of jet fuel biodegradation of each plot at the Eielson AFB bioventing
site, as measured by in situ respirometry, as a function of time during the study.
21
-------
Bioventing of Jet Fuel Spills II:
Bioventing in a Deep Yadose Zone at Hill AFB, Utah
Gregory D. Sayles and Richard C. Brenner
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Robert E. Hinchee
Battelle Memorial Institute, Columbus Division, Columbus, OH
Robert Elliott
Hill Air Force Base, UT
Introduction
Bioventing is a process that supplies oxygen in situ to oxygen deprived soil microbes by forcing
air through unsaturated contaminated soil at low flow rates (1). Unlike soil venting or soil
vacuum extraction technologies, bioventing attempts to stimulate biodegradative activity while
minimizing stripping of volatile organics, thus destroying the toxic compounds in the ground.
Bioventing technology is especially valuable for treating contaminated soils in areas where
structures and utilities cannot be disturbed because bioventing equipment (air
injection/withdrawal wells, air blowers, and soil gas monitoring wells) is relatively noninvasive.
The EPA Risk Reduction Engineering Laboratory, with resources from EPA's Bioremediation Field
Initiative, began a 3-year field study of in situ bioventing in the summer of 1 991 in collaboration
with the U.S. Air Force at Hill AFB near Salt Lake City, Utah. The site has JP-4 jet fuel
contaminated unsaturated soil, where a spill occurred in association with overfilling of an
underground storage tank. The contractor operating the project is Battelle Laboratories,
Columbus, Ohio. This report summarizes the first 2V£ years of the study.
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
of the organic contaminant.
Methodology and Results
See previous reports (2,3) for additional details.
Site Description/Installation
The site is contaminated with JP-4 from depths of approximately 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 disturbing these structures. Apian view
of the installation is shown in Figure 1. The single air injection well installed in December 1 990,
continuously screened from 30 ft to 95 ft below grade, is indicated. "CW" wells are soil gas
22 1994 Symposium on Bioremediation of Hazardous Wastes
-------
"cluster wells," where independent soil gas samples can be taken at 10-ft intervals from 10 ft to
90 ft deep; CW1 through CW3 installed in April 1991, CW4 through CW9 were installed in
September 1991. A cross section of the site along path M' in Figure 1 is shown in Figure 2.
The injection well and the soil gas monitoring wells are indicated. Initial soil total petroleum
hydrocarbon (TPH) concentrations measured from the locations indicated are given also.
Air Injection
To determine the influence of air injection rate on biodegradation and volatilization rates,
various air injection rates have been used during this study:
• August 1991 to October 1992 and December 1992 to April 1993, 67 frVmin
• October to December 1992 and April to June 1993, 40 frVmin
• July to November 1993, 117 frVmin
• November 1 993 to present, 20 frYmin
Soil Gas Composition
Monthly soil gas measurements during venting are conducted. Soil gas O2/ CO2, and total
hydrocarbons are measured at each depth in all wells, providing a three-dimensional map of
soil gas composition in the vadose zone.
fn Situ Respiration Tests
For each flow rate used, an in situ respirometry test (4) is conducted to evaluate the in situ
biodegradation rate. Rates are measured at each soil gas monitoring location. Table 1 shows
rates at three original well locations averaged over depth versus time over a 2-year period.
These wells are close enough to the injection well that changes in the air injection flow rate did
not significantly change oxygen levels at these locations (data not shown). Lower rates with time
suggest that bioventing is removing petroleum hydrocarbons from the site at a significant rate.
Operational Paradigm for Bioventing in Deep Vadose Zones
Bioventing of this system appears to degrade jet fuel by two mechanisms: 1) providing oxygen
for bioremediation of jet fuel contaminated soils near the injection well (Figure 2) and 2)
transporting oxygen and volatilized jet fuel components into the surrounding, relatively
uncontaminated soils (Figure 2), where the organic vapors are biodegraded. Otherstudies have
demonstrated in situ hydrocarbon vapor biodegradation (5-8). Evidence also exists here to
support this operational paradigm. Based on soil gas measurements averaged from August and
September 1993 from all depths in all monitoring wells, Figure 3 shows CO2 produced versus
O2 consumed as the air stream passes from the injection well to the monitoring point. The
approximately linear relationship indicates that oxygen is being converted stoichiometrically to
carbon dioxide at all locations, contaminated or not. Thus, hydrocarbon vapors are degraded
as they are transported through the uncontaminated soils.
23
-------
Based on data taken in April and September 1 991, a preliminary best-fit linear model for the
rate of oxygen uptake versus soil gas TPH and soil TPH was developed:
Rate(%02/hr) = 2.5 x TO'5 C^^pprnv) + 5.7 x 1Q-4 C^mg/kg) (1)
where C^g,,.™ and C^m^ are soil gas TPH and soil TPH concentrations, respectively. Clearly,
the soil gas hydrocarbon vapors contribute significantly to the total oxygen demand. Thus, jet
fuel vapor degradation is a significant mechanism for total jet fuel removal at Hill AFB. The rate
function Rate(Csoi| gas jpH/Qoii TPH) 's plotted in Figure 4. This model will be reassessed as
additional soil gas data is reviewed.
Soil Sampling
Final soil hydrocarbon analyses will be conducted in the summer of 1 993 and compared with
initial soil analyses to document actual hydrocarbon loss due to bioventing.
References
1. Hoeppel, R.E., R.E. Hinchee, and M.F. Arthur. 1991. Bioventing soils contaminated
with petroleum hydrocarbons. J. Indust. Microbiol. 8:141-146.
2. Sayles, G.D., R.C. Brenner, R.E. Hinchee, CM. Vogel, and R.N. Miller. 1992.
Optimizing bioventing in deep vadose zones and moderate climates: Hill AFB
bioremediation of a JP-4 spill. In: U.S. EPA. Symposium on bioremediation of
hazardous wastes (abstracts). EPA/600/R-92/126. Washington, DC (May).
3. Sayles, G.D., R.E. Hinchee, R.C. Brenner, and R. Elliott. 1993. Documenting
bioventing of jet fuel to great depths: A field study at Hill Air Force Base, Utah. In:
U.S. EPA. Symposium on bioremediation of hazardous wastes: Research, development,
and field evaluations (abstracts). EPA/600/R-93/054. Washington, DC (May).
4. Ong, S.K., R.E. Hinchee, R. Hoeppel, and R. Schultz. 1991. In situ respirometry for
determining aerobic degradation rates. In: Hinchee, R.E., and R.F. Olfenbuttel, eds.
In situ bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 541-545.
5. Ostendorf, D.W., and D.H. Kampbell. 1990. Bioremediated soil venting of light
hydrocarbons. Haz. Waste Haz. Mat. 7:319-334.
6. Kampbell, D.H., and J.T. Wilson. Bioventing to treat fuel spills from underground
storage tanks. J. Haz. Mat. 28:75-80.
7. Miller, R.N., R.E. Hinchee, and C.M. Vogel. 1991. A field-scale investigation of
petroleum hydrocarbon biodegradation in the vadose zone enhanced by soil venting at
Tyndall AFB, Florida. In: Hinchee, R.E., and R.F. Olfenbuttel, eds. In situ
bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 283-302.
24
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8. Kampbell, D.H., J.T. Wilson, and CJ. Griffin. 1992. Performance of bioventing at
Traverse City, Michigan. In: U.S. EPA. Symposium on bioremediation of hazardous
wastes. EPA/600/R-92/126. Washington, DC (May), pp. 61-64.
Table 1. Rates of Biodegradation, Averaged Over the Depth and
Measured by In Situ Respirometry, at the Three
.Original Soil Gas Monitoring Wells
Rate (mg/kg/day)
Well
CW1
CW2
CW3
September 1 991
1.1
0.26
0.54
September 1 992
0.59
0.13
0.26
October 1 993
0.31
0.16
0.12
25
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A O
CW9
• WW - Ground Water Monitoring Well
O CW - Sofl \fepor Cluster Wefl
(1.5) - TPH In Ground Water (mg/L) (9/91)
A-A1 » Cross Section Ttece
Figure 1. Plan view of the joint EPA and U.S. Air Force bioventing activities at Hill AFB, near
Salt Lake City, Utah. IW is the air injection well, and CW are cluster soil gas
monitoring wells.
26
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1SW
Figure 2. Cross-section view of the bioventing installation at Hill AFB. Cross section follows the
path AA' in Figure 1. Initial soil TPH concentrations measured at various depths at
the wells are indicated.
27
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i
10
i
15
i
20
25
O2 Consumed (%
Figure 3. CO2 produced versus O2 consumed as the air stream passes from the injection well
to each soil gas monitoring point. Data indicate biological activity at all soil gas
monitoring well locations.
6
d
12000 S*r*
10000 .
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In Situ Bioremediation of a Pipeline Spill Using Nitrate as the Electron Acceptor
Stephen R. Hutchins, John T. Wilson, and Don H. Kampbell
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory,
Ada, OK
Introduction
In the late 1970s, leakage of refined petroleum products from an underground pipeline
contaminated approximately 24,000 square meters of a shallow water-table aquifer in Park City,
Kansas. Aerobic in situ bioremediation was initiated but was unsuccessful due to plugging of
the injection wells or sediments adjacent to the well screen by gas and iron precipitates.
Nitrate was selected as an alternative electron acceptor that might avoid some of the problems
with plugging.
Approach
Ground water from the aquifer was amended with sodium nitrate and ammonium chloride and
returned to the area of the hydrocarbon spill through a series of infiltration wells that were
installed in a grid. The wells were spaced 6.1 m apart. The study area contained 157
infiltration wells, spaced over 5,800 m2, which received 3,000 m3 of water in a tracer test,
followed by 39,400 m3 of water containing 4,136 kg of sodium nitrate (an average of 1 7 mg/L
nitrate nitrogen). The circulated water also contained 50 to 60 mg/L sulfate.
Figure 1 plots the cumulative flow of ground water to the infiltration wells against time. Flow
was unhindered for the first 150 days of operation, then the system plugged over the next 100
days.
A total of 7.3 m of recharge was applied to the spill, of which 6.8 m contained nitrate.
Procedure To Distinguish Flushing From Biodegradation of BTEX
The site was cored, and vertically stacked continuous cores from the same borehole were
analyzed to determine the total mass of BTEX compounds in the aquifer. To estimate the mass
of BTEX compounds in ground water in contact with the hydrocarbon spill, monitoring wells were
installed in the boreholes used to acquire the cores. The screened interval on the monitoring
well was equivalent to the depth interval containing NAPL hydrocarbons.
The following procedure was used to determine the total mass of BTEX compounds in the aquifer
under a unit surface area. The concentration of BTEX compounds in individual core samples
(gm/kg) were multiplied by the vertical interval that each core represented (M), then multiplied
by the bulk density of sandy aquifer material (1,800 kg/m3). The masses in the depth intervals
1994 Symposium on Bioremediation of Hazardous Wastes 29
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represented by the cores were then summed to determine the total mass of each BTEX
compound under each square meter (Table 1).
The concentration of BTEX compounds in water under each square meter was determined by
multiplying a square meter by the length of the well screen to determine the volume sampled,
then by 0.3 to estimate the volume of ground water, then by the concentration of BTEX
compounds in ground water sampled from the well (Table 1). The volume of aquifer sampled
by the well to estimate mass in ground water and the volume summed to estimate total mass
were equivalent.
The ratio of mass in water to total mass determines the fraction of total mass that can be flushed
away each time water in the sampled volume is exchanged by the infiltrating ground water
(Table 1).
The volume of water in the sample volume was considered equivalent to a pore volume in a
column experiment; the infiltration of ground water was expressed in pore volumes. The mass
of each BTEX compound remaining after one pore volume of flushing should equal the initial
mass, multiplied by 1.0 minus the ratio of mass in water to total mass. The mass of each BTEX
compound remaining after any number of pore volumes of flushing should equal the initial total
mass, multiplied by 1.0 minus the ratio of water/total, raised to an exponent equal to the
number of pore volumes flushed through the spill.
FinalMass - InitialMass (1.0 - Water/Total)1'0"'™""""
This approach was used to predict the reduction in contaminant concentration due to flushing
and to separate the effects of flushing from biodegradation. Over 90 percent of benzene was
removed from ground water during the demonstration. However, flushing accounted for most,
if not all, of this removal (Figure 2). Over 95 percent of toluene and ethylbenzene was
removed, and biodegradation accounted for most of the removal (see Figure 3 for toluene
removal). Removal of xylenes varied from 68 percent to 76 percent; most of the removal was
accounted for by biodegradation (Figure 4).
Estimate of Treatment Effectiveness
If the concentration of BTEX compounds in ground water and in the NAPL are in equilibrium,
Raoult's Law can be used to put an upper boundary on the total mass of contaminant removed
by in situ bioremediation. Concentrations of individual BTEX compounds were compared before
and after remediation to determine fractional removal in ground water. The fractional removals
in ground water were multiplied by the initial total mass of each BTEX compound to estimate
total mass removals.
The amount of BTEX degraded during denitrification is equivalent to the amount of nitrate-
nitrogen applied. Apparently, considerably more BTEX was removed than could be explained
by the quantity of nitrate supplied (Table 2). In fact, there was more removal than could be
accounted for by either denitrification or flushing. Sulfate in well 60A was less than 1.0 mg/L
prior to the start of infiltration; during infiltration concentrations ranged from 57 mg/L to 93
mg/L. During the course of the demonstration, concentrations of sulfate in monitoring well 60G
30
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in the study area were near 10 mg/L, when concentrations of sulfate were in the range of 50
mg/Lto 60 mg/L in the infiltrated water. Removal of 40 mg sulfate per liter by sulfate reduction
could have accounted for as much as 230 gm/m2 of total BTEX removal. If this is the case,
naturally occurring sulfate in the infiltrated ground water was more important as an electron
acceptor than the nitrate that was intentionally added. Concentrations of methane in the
infiltrated water ranged from 4.8 mg/L to 6.3 mg/L while concentrations in well 60A ranged
from 2.8 mg/L to 3.7 mg/L. Methanogenesis cannot explain the missing mass of BTEX
compounds.
The assumption of chemical equilibrium may also be in error, and much of the BTEX may not
have been in contact with the ground water. In this case the total BTEX removed would be
overestimated, and the nitrate demand that was exerted would represent that portion of the
hydrocarbons that exchanged readily with the ground water.
Table 1. Concentration of BTEX Compounds in Ground Water and in the Aquifer at Site 60A,
the Most Contaminated Site in the Study Area*
Compound
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
Mass in Water (gm/m2)
2.01
2.57
1.02
0.958
1.26
0.776
Total Mass (gm/m2)
17.6
102
72
68
161
78.3
Water/Total
0.114
0.0252
0.0142
0.0141
0.00783
0.00991
*Subsurface concentrations are expressed as the total mass in the vertical interval under a
square meter of land surface area.
31
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Table 2. Use of Raoult's Law To Estimate the Total Mass of Contaminants Removed by Nitrate-
Based Bioremediation at 60A, the Most Contaminated Site in the Study Area
Compound
Benzene
Toluene
o-Xylene
m-Xylene
p-Xylene
Ethyl-
benzene
Concentration in Well 60A
HA)
Initial
2010
2570
776
1260
958
1020
Final
174
77.9
209
297
304
26.5
Fraction
Removed
From
Water
(percent)
0.913
0.970
0.732
0.764
0.683
0.974
Initial
Concentration
in Core
Material
(gm/m2)
17.6
102
78.3
161
68
72
Total BTEX removed
Maximum attributed to nitrate as electron acceptor
Maximum attributed to flushing
Balance, attributed to sulfate as electron acceptor
Mass
Removed
(gm/m2)
16.1
98.9
57.2
123
46.4
70.2
411.2
118
131
163
32
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45000 -
40000
35000 •
J5 30000 -f
I 25000
.a 20000 t
.0
r3, 15000 +
0
Cumulative Flow
50 100 150
Days After Addition of Nitrate
200
250
Figure 1. Cumulative flow of ground water amended with nitrate to the study area (m3).
Benzene Depletion
0
10 15
Pore Volumes
Predicted
Measured
20
25
Figure 2. Comparison of benzene depletion to that expected from flushing alone.
Concentrations in
33
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3000 T
i- 2500 -r
o
Toluene Depletion
10 15
Pore Volumes
20
Predicted
Measured
25
Figure 3. Comparison of toluene depletion to that predicted from flushing alone.
Concentrations in/
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Performance Evaluation of Full-Scale In Situ and Ex Situ Bioremediation of
Creosote Wastes in Ground Water and Soils
Ronald C. Sims, Judy L. Sims, Darwin L. Sorensen, and David K. Stevens
Utah State University, Logan, UT
Scott G. Huling, Bert E. Bledsoe, and John E. Matthews
U.S. Environmental Protection Agency, Ada, OK
Daniel Pope
Dynamac Corporation, Ada, OK
The Champion International Superfund Site in Libby, Montana, was nominated by the Robert S.
Kerr Environmental Research Laboratory as a candidate site for performance evaluation as part
of the EPA-sponsored Bioremediation Field Initiative. Two forms of wood preservative were used
at the site: creosote, containing polycyclic aromatic hydrocarbons (PAHs), and loose
pentachlorophenol (PCP). PAHs are currently the primary components of concern at the site.
The performance evaluation project is directed by Dr. Ronald Sims of Utah State University.
The bioremediation performance evaluation consisted of three phases: 1) summarize previous
and current remediation activities, 2) identify site characterization and treatment parameters
critical to the evaluation of bioremediation performance for each of the bioremediation
treatment units, and 3) evaluate bioremediation performance based on this information.
Three biological treatment processes are addressed in the bioremediation performance
evaluation: 1) surface soil bioremediation in a prepared-bed, lined land treatment unit (LTU);
2) treatment of extracted ground water from the upper aquifer in an aboveground fixed-film
bioreactor; and 3) in si'fu bioremediation of the upper aquifer at the site. A description of the
site with accompanying figures appears in the abstract book from the 1993 EPA-sponsored
Symposium on Bioremediation of Hazardous Wastes (1).
Biological Treatment Processes
The LTU has been used for bioremediation of contaminated soil taken from three primary
sources, including tank farm, butt dip, and waste pit areas. Contaminated soil was excavated
and moved to one central location, the waste pit. Soil pretreated in the waste pit area is further
treated in one of two prepared-bed, lined land treatment cells (LTCs). Total estimated
contaminated soil volume for treatment is 45,000 yd3 (uncompacted). Contaminated soil
cleanup goals (dry-weight basis) are: 1) 88 mg/kg total (sum of 10) carcinogenic 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 2,3,7,8-dioxin equivalent.
The LTU comprises two adjacent 1 -acre cells. Components of the soil bioremediation system
for each LTC include the treatment zone, liner system, and leachate collection system. Each cell
is lined with low-permeability materials to minimize leachate infiltration from the unit.
Contaminated soil is applied and treated in lifts (approximately 9-in. thick) in the designated
1994 Symposium on Bioremediation of Hazardous Wastes 35
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LTC. When reduction of contaminant concentrations in all lifts placed in the LTD has reached
the cleanup goals specified in the Record of Decision (ROD), 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.
Degradation rates, volume of soil to be treated, initial contaminant concentration, degradation
period, and LTC size determine the time required for remediation of a given lift. Based on an
estimated 45-day time frame for remediation of each applied lift as determined by Champion
International, an estimated 45,000 yd3 of contaminated soil, and a 2-acre total LTD surface
area, the projected time to complete soil remediation is 8 to 10 years.
The upper aquifer aboveground treatment unit provides biological treatment of extracted ground
water for removal of PAHs and PCP prior to reinjection via an infiltration trench. The biological
treatment consists of two fixed-film reactors operated in series. The first reactor is heated and
has been used for roughing purposes, while the second has been used for polishing and
reoxygenation of the effluent prior to reinjection. The system was commissioned in February
1990.
Extracted ground-water treatment system components include equalization and biotreatment.
Equalization system components include four ground-water extraction wells and an equalization
tank, which consists of a cylindrical horizontal flow tank with a nominal hydraulic residence time
of 6 hours at a flow rate of 10 gpm. The bioreactor treatment system components include
nutrient amendment, influent pumping, bioreactor vessels, aeration, heating, and effluent
pumping. The components of the aboveground treatment system for extracted ground waterare
shown in the 1993 Symposium abstract book (1).
The pilot upper aquifer area in situ bioremediation system involves the addition of oxygen and
inorganic nutrients to stimulate the growth of microbes. The initial source of oxygen was a
hydrogen peroxide injection system that was designed to maintain a concentration of
approximately 100 mg/L of hydrogen peroxide. Injection flow rate was approximately 100 gpm
into three injection clusters. Inorganic nutrients in the form of potassium tripolyphosphate and
ammonium chloride are continuously added to achieve concentrations in the injection water of
2.4 mg/L nitrogen and 1 mg/L phosphorus.
The ROD calls for cleanup levels in the upper aquifer of 40 parts per trillion (ppt) total
carcinogenic PAHs, 400 ppt for total noncarcinogenic PAHs, 1.05 mg/L for PCP, 5 /ig/L for
benzene, 50/ig/Lfor arsenic, and a human health threat no greater than 10'5 for ground-water
concentrations of other organic and inorganic compounds.
Performance Evaluation Activities
Performance of the soil bioremediation system in the LTCs involved evaluating the reduction in
concentration of PAHs and PCP with time and with depth within the LTD. The primary purpose
of the LTD soil sampling program in this project was to determine the statistical significance and
extent of contaminated soil treatment at this site. A quantitative expression of 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 the
biodegradation potential of similar type wastes under similar conditions at other sites.
36
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In most soils and disturbed soil materials, physical and chemical properties are not distributed
homogeneously throughout the volume of the soil material. The variability of these properties
may range from 1 percent to greater than 100 percent of the mean value within relatively small
areas. Chemical properties, including contaminants, often have the highest variability. A first
approximation of the total variance in monitoring data can be defined by the following equation:
v, = vs/k + vyk*n
where k is the number of samples, n the number of analyses per sample, k*n the total number
of analyses, V, the total variance, Va the analytical variance, and V$ the sample variance. In
general, sampling efforts to minimize V, result in the most precision. Analytical procedures
frequently achieve precision levels (V^k*^ of 1 percent to 10 percent, while soil sampling
variation (VJ may be greater than 35 percent. Sampling designs that reduce the magnitude of
Vs should be employed where possible. Therefore, the sampling procedures used in this
evaluation were designed to minimize V, and to provide representative information about the
transformation of PAHs and PCP within the LTCs.
The LTD was sampled in May, June, July, and September 1991, and in September 1992.
Field-scale investigations concerning PAH and PCP concentrations were supported by laboratory
mass-balance investigations of radiolabeled compounds for determination of mineralization as
well as humification potential for target contaminants.
Performance evaluation of the upper aquifer aboveground fixed-film treatment system involved
evaluating the bioreactor system. Treatment evaluation focused on characterizing performance
regarding system capability to remove PAHs and PCP from the ground water, and on optimizing
operation within the bioreadors. The aboveground treatment system was sampled during 1 991
and 1 992 for chemical, physical, and biological parameters. In addition, a pilot-scale reactor
was constructed and operated to evaluate abiotic reactions of chemicals present in the water
phase within the bioreactors. The information generated from the sampling and monitoring of
the full-scale reactor and from the operation of the pilot-scale reactor was combined with data
provided by Champion International to provide an in-depth evaluation of performance.
Performance evaluation of the in situ bioremediation system focused on characterization of the
water phase, the solid phase (aquifer materials), and oil associated with the aquifer solid
material. The aquifer was sampled during 1 991 and 1992. An evaluation of the water phase
included measurements of dissolved oxygen (DO) concentrations, the inorganic chemicals iron
and manganese to evaluate potential abiotic demand for injected hydrogen peroxide, and the
concentrations of PAHs and PCP. An evaluation of the aquifer solid phase has included PAHs
and PCP concentrations in treated and background areas at the site. Laboratory mass balance
experiments using radiolabeled target compounds were used in conjunction with field-scale
measurements to provide additional information concerning biotic reactions (mineralization) and
potential abiotic reactions (poisoned controls).
Summary of Results
Analyses of over 300 soil samples were performed from which greater than 5,000 individual
chemical concentrations were determined forthe 16 priority pollutant PAH compounds using gas
chromatography/mass spectrometry (GC/MS) and for pentachlorophenol (PCP) using a gas
37
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chromatography/electron capture detector (GC/ECD). Results of chemical analyses indicated
that target remediation levels for target chemicals were achieved using mean values at each
depth evaluated in each LTC, with only two exceptions where mean concentrations were only
slightly higher than the target remediation levels. As a result of obtaining vertical samples at
each sampling event, downward migration of target chemicals through the LTD was not
observed. Soil within the LTU was detoxified to control uncontaminated soil levels. Toxicity
information was based upon results of using both the Microtox™ assay to measure water extract
toxicity and the Ames Salmonella typhimurium mammalian microsome mutagenicity assay (Ames
assay) to measure mutagenicity of soil solvent extracts. Detoxification to nontoxic levels was
evident in all samples evaluated for both Microtox™ and Ames assays.
Results of the laboratory evaluation of soil microbial metabolic potential demonstrated that PCP
and phenanthrene, the two chemicals evaluated using radiolabeled carbon, could be
metabolized to carbon dioxide by indigenous microorganisms present in the contaminated soil
matrix present at the site at temperature and moisture conditions representative of the site. In
addition, significant volatilization of PCP or phenanthrene is unlikely based upon the laboratory
evaluation. The information obtained in the laboratory evaluation corroborated the interpretation
of apparent decrease in target chemical concentrations in field samples within the LTU and in
the in situ aquifer samples at the Libby site as due to biological processes rather than
physical/chemical processes.
Results of the aboveground fixed-film bioreactor indicated that removal of PCP and PAHs from
extracted ground water was strongly influenced by hydraulic retention time (HRT). The system
removed greater than 80 percent of PCP and 90 percent of PAHs at a flow rate of 10 gallons
per minute (gpm), with an HRT of 30 hours. At a flow rate of 10 gpm, the effluent
concentrations of PCP and total PAHs were 0.3 mg/L to 0.9 mg/L and less than detection
(30 ftg/L), respectively. When the flow rate was increased to 15 gpm, with an HRT of 20 hours,
removal of both PCP and PAHs decreased significantly. At the 15-gpm flow rate, effluent
concentrations of PCP and total PAHs were 6 mg/L to 12 mg/L and 0.6 mg/L to 6 mg/L,
respectively. Additional limitations of DO and nutrients are addressed in the final report.
Results of the in situ treatment evaluation indicated that, with respect to the ground-water phase,
total PAHs and PCP were present at lower concentrations in wells considered to be under the
influence of the treatment injection system consisting of nutrients and hydrogen peroxide, while
total PAHs and PCP were present at higher concentrations in wells considered to be outside of
the influence of the injection system. An evaluation of the water phase in monitoring wells
demonstrated the presence of reduced inorganic compounds, including iron and manganese,
with concentrations inversely related to DO concentrations. These chemicals may exert a
demand on the oxygen supplied by the hydrogen peroxide and reduce the oxygen available for
microbial utilization.
With respect to the nonaqueous phase liquid (NAPL) phase, both total PAHs and PCP were
found in the highest concentrations in the NAPL, greater than 10,000 mg/L and 1,000 mg/L,
respectively, than in any other phase sampled at the Champion International Site. These results
indicate that there is potential contamination of the upper aquifer remaining in the form of a
nonaqueous phase that represents significant potential contamination of the ground water by
transfer of contaminants from the NAPL phase to the ground-water phase.
Total PAH and total petroleum hydrocarbons (TPH) were present within the aquifer
sediment/NAPL samples at concentrations of 5 mg/kg to 687 mg/kg and 70 mg/kg to 2,525
38
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mg/lcg, respectively. The heterogeneous distribution of total PAH, PCP, and TPH contaminants
was consistent among three boreholes evaluated from the water table to the deepest sampling
point. Target chemicals associated with sediment/NAPL interfaces may be more difficult to
bioremediate in situ than chemicals in the aqueous phase due to limitations of mass transport
of oxygen and nutrients from the water phase to the NAPL phase that contain target chemicals.
Chemical mass balance evaluations conducted using radiolabeled target chemicals in the
laboratory demonstrated that aquifer materials from the site contained indigenous
microorganisms that had the ability to mineralize phenanthrene. Up to 70 percent of the
radiolabeled carbon became incorporated into the aquifer matrix and was nonsolvent
extractable. Nosignificant phenanthrene mineralization or incorporation of radiolabeled carbon
was observed in poisoned controls. PCP mineralization, however, was insignificant (less than 2
percent), with results similar for nonpoisoned and poisoned samples.
The three biological treatment processes evaluated at the Libby, Montana, site represent a
treatment train approach to site decontamination, where each of the treatment processes are
biological. The soil phase is treated in the LTD system, and any leachate produced can be
treated in the aboveground bioreactor before it is returned to the LTU as part of soil moisture
content control and treatment of low levels of PAHs and PCP in the effluent. The in situ
treatment system addresses the oil and solid phases in the subsurface. At the Libby site,
therefore, a different biological process was chosen for remediation of each contaminated phase
(soil, oil, and water).
Performance Evaluation Reports
While the extended abstract presented in this report has been abridged concerning site
characterization and treatment results, separate reports have been prepared for EPA that address
each of the three biological treatment systems at the site in detail: 1) soil bioremediation in the
prepared-bed LTU, 2) aboveground fixed film system for extracted ground water, and 3) in situ
treatment. Information generated from full-scale characterization and monitoring, pilot-scale
studies, and laboratory treatability studies was combined with information provided by Champion
International to provide an integrated evaluation of bioremediation performance at the Libby,
Montana, site. The information obtained can be used to evaluate and select rational
approaches for characterization, implementation, limitations, and monitoring of bioremediation
at other sites.
References
1. Sims, R.C., J.E. Matthews, S.G. Huling, B.E. Bledsoe, M.E. Randolph, and D. Pope.
1 993. Evaluation of full-scale in sifu and exsifu bioremediation of creosote wastes in
soils and ground water. In: U.S. EPA. Symposium on bioremediation of hazardous
wastes: Research, development, and field evaluations (abstracts). EPA/600/R-93/054.
Washington, DC (May).
39
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Bioventing Soils Contaminated With Wood Preservatives
Paul T. McCauley, Richard C. Brenner, and Fran V. Kremer
U.S. Environmental Protection Agency, Office of Research and Development,
Cincinnati, OH
Bruce C. Alleman
Battelle Memorial Institute, Columbus, OH
Douglas C. Beckwith
Minnesota Pollution Control Agency, St. Paul, MN
Introduction
The Reilly Tar and Chemical Corporation operated a coal tar distillation and wood-preserving
plant, known as the Republic Creosote Company, in St. Louis Park, Minnesota, from 1 91 7 to
1 972. During this period, wastewater discharges as well as drips, spills, and dumping from the
wood-preserving processes resulted in creosote and coal tar contamination of about 80 acres
of this site and the underlying ground water. In 1972, the City of St. Louis Park purchased the
site from the Reilly Tar and Chemical Corporation for land use. All onsite buildings were
dismantled and removed, and the soil was graded and covered with 3 ft of topsoil for
beautification and odor control.
In 1978, the Minnesota Department of Health began analysis of ground water from municipal
wells in St. Louis Park and neighboring communities for carcinogenic and noncarcinogenic
polycyclic aromatic hydrocarbons (PAHs). The discovery of significant concentrations of
regulated PAHs in six St. Louis Park wells resulted in their shutdown during the period of 1978
to 1981. St. Louis Park is currently maintaining gradient control of the contaminated ground-
water plume by pumping and treating. With the exception of a tar plug in one well, little PAH
source contamination has been removed. Without source control of the PAHs, pumping and
treating of contaminated ground water may be required for several hundred years.
Background
Bioventing is a proven technology for in situ remediation of various types of hydrocarbon
contaminants. The technology has been successfully used to remediate sites contaminated with
gasoline (1), aviation fuels (JP-4 and JP-5) (2,3), and diesel fuel (4). A biological treatment
process, bioventing uses low-rate atmospheric air (or oxygen enriched air up to 100-percent
oxygen) injection to treat contaminated unsaturated soil in situ. The air flow provides a
continuous oxygen source that enhances the growth of aerobic microorganisms naturally present
in soil, with minimal volatilization to the atmosphere of any volatile organic compounds that may
be present in the soil. The size of the treatment area is defined by the number of wells installed,
the size of the air blower used, and site characteristics such as soil porosity. The current research
evaluates the potential of bioventing to remediate soils contaminated with PAHs.
40 1994 Symposium on Bioremediation of Hazardous Wastes
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Methods
Site Description
A 50 ft x 50 ft control and a 50 ft x 50 ft bioventing treatment plot were established on the site
during the original soil gas survey (Figure 1). The first 3 ft of soil at the test plots is
unconta mi noted topsoil applied after the cessation of industrial use (Figure 2). A dense, 3-in.
to 6-in., hard-packed layer separates the topsoil from the porous sandy layer, which extends to
below the water table (8 ft to 10 ft below the ground surface). Most of the PAH contamination
was found in the sandy layer.
PAH Sampling
Composite soil samples (120 soil borings per plot) were taken for PAH analysis and prepared
by homogenizing the soil obtained from the 4 ft to 8 ft depth of each boring. The resultant
boreholes were filled immediately with bentonite. The PAH soil analyses were recorded as zero-
time PAH concentrations.
Venting Well
A single-vent bioventing system was installed at the center of the treatment area (Figure 2). The
vent (injection) well was screened from 7 ft to 11 ft below the surface and packed with sand.
The vent well was then sealed with bentonite from the 5 ft depth to the surface.
Soil Gas Sampling Well
Twelve soil gas probes were installed along diagonals drawn from the comer of the square
treatment area (Figure 2), and four were installed in the comers of the no-treatment control
area. The soil gas probes were constructed so that the soil gas withdrawal points and
thermocouples were located at 4, 6, and 8 ft below the ground surface.
Respirometry
Initial O2 and CO2 measurements were obtained using stainless steel gas probes withdrawing
air from measured intervals below the ground surface to Gas Teck® O2 and CO2 meters. The
gas measurements were expressed as percentages of total soil gas. Gas samples for the zero-
time sampling in November were extracted using the newly installed soil gas sampling wells.
Initial sampling indicated that, due at least in part to the highly pervious soil at the Reilly site,
injected air was migrating from the test plot 125 ft to 180 ft into the unaerated control plot.
A 10-ft deep bentonite slurry wall was constructed across the near wall of the control plot. The
slurry wall and reduced air injection pressures and flow rates effectively prevented further
unwanted aeration of the control plot.
41
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Shutdown Respiration Tests
Shutdown respiration tests are being conducted for 2 weeks at quarterly intervals. Soil gases
are brought to atmospheric O2 and CO2 levels in the test plots by pumping ambient air into the
ground. When ambient O2 and CO2 levels are achieved and documented, the air flow into the
ground is stopped. Soil gases levels are taken over measured intervals until an O2 utilization
rate is defined. The air flow was set at 10 frVmin, which translated at this site to a pressure of
3.5 in. of H2O.
Results
In the summer of 1992, a field team from the Risk Reduction Engineering Laboratory (RREL),
Biosystems Branch, conducted a soil gas survey at the Reilly site and determined that soil gases
were below the estimated 5-percent oxygen threshold required for aerobic metabolism (5).
Under a cooperative project involving the Bioremediation Field Initiative, the Superfund
Innovative Technology Evaluation (SITE) Demonstration Program, and RREL's Biosystems
Program, a pilot-scale bioventing field demonstration for PAH bioremediation was initiated at
the Reilly site in November 1992.
Soil PAH analysis demonstrated significant contamination in both plots. The treatment plot
demonstrates an order-of-magnitude decrease in PAH concentration on the eastern side of the
plot. The control plot is contaminated to about the same degree as the western half of the
treatment plot.
Quarterly shutdown respiration tests have shown respiration rates ranging from below detection
(Figure 3) to 0.484 percent O2 per hour (Figure 4). The highest respiration rates were found
in the western half of the treatment area, where PAH contamination was also shown to be the
heaviest. Current average measured respiration rates are consistent with a 14-percent reduction
in PAH contamination per year.
Summary and Conclusion
A 3-year evaluation program was initiated in November 1992 with the zero-time sampling. In
situ respiration tests are being performed four times each year to determine oxygen utilization
and CO2 evolution rates. These data can be converted to estimated biodegradation rates to
estimate the disappearance of PAHs (6). Because of the strong partitioning of PAHs to soil,
long-term bioventing is expected to be necessary to fully remediate the site. The target PAH
removal rate for this 3-year project is 30 percent. Successful achievement of this rate would
project total cleanup in 10 to 15 years.
References
1. Ostendorf, D.W., and D.H. Kampbell. 1990. Bioremediated soil venting of light
hydrocarbons. Haz. Waste Haz. Mat. 7:319-334.
42
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2. Sayles, G.D., R.C. Brenner, R.E. Hinchee, A. Leeson, C.M. Vogel, R. Elliot, and R.N.
Miller. 1994. Bioventing of jet fuel spills I: Bioventing in a cold climate with soil
warming at Eielson AFB, Alaska. Presented at the U.S. EPA Symposium on
Bioremediation of Hazardous Wastes: Research, Development, and Field Evaluations,
San Francisco, CA (June).
3. Sayles, G.D., R.C. Brenner, R.E. Hinchee, and R. Elliott. 1994. Bioventing of jet spills
II: Bioventing in a deep vadose zone at Hill AFB, Utah. Presented at the U.S. EPA
Symposium on Bioremediation of Hazardous Wastes: Research, Development, and Field
Evaluations, San Francisco, CA (June).
4. Kampbell, D.H., and J.T. Wilson. 1991. Bioventing to treat fuel spills from
underground storage tanks. J. Haz. Mat. 28:75-80.
5. Ong, S.K., R.E. Hinchee, R. Hoeppel, and R. Schultz. 1991. In situ respirometry for
determining aerobic degradation rates. In: Hinchee, R.E., and R.F. Olfenbuttel, eds.
In situ bioreclamations, applications, and investigations for hydrocarbons and
contaminated site remediation. Boston, MA: Butterworth-Heinemann. pp. 541-545.
6. Hinchee, R.E., and S.K. Ong. 1992. A rapid in situ respiration test for measuring
aerobic biodegradation rates of hydrocarbons in soils. J. Air Waste Mgmt. Assoc.
42(10):!,305-1,312.
43
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Louisiana Avenue
Treatment Plot
Vent
o
Water
Level
Well
Control Plot
Junction
Box
O
Water
Level
Well
o
Fence
Not to Scale
Figure 1. Placement of injection and soil gas sampling wells in the control and treatment plots.
44
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Bioventing Injection and Soil Gas Sampling Wells
Air Sampling Wells 1 1 | Air Injection Well
Soli Surface •
1 1
-11 Feet
[3 Top Soil
OH Hard Packed Layer
(I7!] Coarse Sandy Layer
[_] Water Table
| Bentonlte
H Sand and Gravel
Figure 2. Air injection and soil gas sampling wells installed in the treatment plot.
RESPIRATION CURVE (MP- K)
SHUTDOWN TES~ll
10
ode
0 40 80 120 160 200 240
Time, hours
* 4 FOOT • 6 FOOT » 8 FOOT
-6HO
15
12
6 f
SHUTDOWN TEST
20!
£ 10 -
0 40 80 120 160 200 240
Time, hours
Figure 3. Solid symbols represent O2. Figure 4. Solid symbols represent O2.
Hollow symbols represent CO2. Hollow symbols represent CO2.
45
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Field Evaluation of Fungal Treatment Technology
John A. Glaser
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Richard T. Lamar, Diane M. Dietrich, Mark W. Davis, Jason A. Chappelle,
and Laura M. Main
U.S. Department of Agriculture, Forest Products Laboratory, Madison, Wl
Introduction
Bioaugmentation of soil contaminated with pentachlorophenol (PCP) using selected strains of
lignin-degrading fungi has been shown to result in extensive and rapid decrease in the PCP
concentrations for two soils under field treatment conditions (1,2). In different soils studied
under laboratory conditions, the same behavior was observed and extensively evaluated by
means of determining the pollutant mass balance in the soils (3,4). Initial products of fungal
biotransformation were identified. PCP concentrations in excess of 1,000 mg/kg were 80
percent to 90 percent biotransformed in soil by selected fungi in 56 days (Figure 1).
A two-phase project, consisting of a treatability study in 1991 and a demonstration study in
1992, was conducted at an abandoned wood treating site in Brookhaven, Mississippi, to
evaluate fungal treatment effectiveness under field conditions. The study site, located 60 miles
south of Jackson, was identified as a removal action site for EPA Region 4. While the wood
treating facility was in operation, two process liquid lagoons were drained and excavated. The
sludge was mounded above the ground surface in a Resource Conservation and Recovery Act
(RCRA) hazardous waste treatment unit. The excavated material provided the contaminated soil
for both phases of the project. The demonstration phase was undertaken as a Superfund
Innovative Technology Evaluation (SITE) Program Demonstration Project.
The fungal treatment processes reported herein were conducted at Brookhaven because the site
characteristics were suitable for conducting field investigations, not because the investigators
desired to promote fungal treatment as one of the treatment options for the site.
Methodology
The demonstration study was designed to evaluate the ability of a single fungal strain
(Phanaerochaete sordida) to degrade PCP in soil. The soil pile was sampled and analyzed for
PCP and creosote components (i.e., polycyclic aromatic hydrocarbons [PAHs]) prior to
developing the test site. Analysis of the laboratory results identified sections of the pile with PCP
concentrations of less than 700 mg/kg. These sections were used to supply the contaminated
soil for both phases of the study.
A test location was constructed on an uncontaminated portion of the wood treating site. The
base for the test plots was formed by using uncontaminated soil to provide a 1 -percent to 2-
percent slope to promote better drainage. Soil beds (Figure 2) were constructed of galvanized
46 1994 Symposium on Bioremediation of Hazardous Wastes
-------
sheet metal. For the demonstration study, the P. sordida treatment plot measured 30.5 m x
30.5 m and the treatment and inoculum control plots measured 7.6 m x 15.25 m. Plot
dimensions were determined in conjunction with SITE program personnel. A concrete pad was
constructed to assist tiller entry into the different plots and to decontaminate the tiller as it was
moved from plot to plot.
Within each plot, the base soil was graded for a V-shaped indentation in the central portion of
the plot to permit leachate collection. A leachate collection system was installed to direct the
liquid discharge from all test plots to a central location for testing and treatment. After
installation of the leachate system, 25 cm (10 in.) of clean sand was layered into each test plot
followed by a 25-cm (10-in.) lift of contaminated soil.
The treatment plot received 10 percent by weight of an infested inoculum containing P. sordida.
The no-treatment control received no amendments. The inoculum control plot consisted of
contaminated soil amended with noninfested inoculum carrier. All plots were tilled on the same
schedule, weather permitting. The fungal inoculum was developed jointly with the L.F. Lambert
Spawn Co. of Coatesville, PA. The prepared inoculum and inoculum carrier were shipped to the
site by refrigerated transport.
The contaminated soil was sized through a 2.5-cm (1 -in.) mesh screen using a Read Screen All
shaker screen having a capacity 8.4 m3/hr (10 ydVhr). The soil was deposited in separate piles
on a polyethylene tarp. Further homogenization was accomplished by mixing different portions
of screened soil. The soil was then mixed with the 10 percent by weight fungal inoculum in a
Reel Auggie Model 2375 Mixer and applied to the treatment plots using a front end loader.
After inoculation with fungi, each plot was irrigated and tilled with a garden rototiller. Soil
moisture was monitored on a daily basis throughout the study and maintained at a minimum
of 20 percent. Ambient and soil plot temperatures were recorded daily throughout the study. Plot
tilling was scheduled on a weekly basis 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 after 1, 2, 4, 8, 12, and 20 weeks of
operation (Figure 3).
Results
The demonstration study was conducted over a 5-month period between June and November
1992. The greatest removal of PCP (Table 1) was achieved in the plot inoculated with P.
sordida. Over the course of the study, this treatment regime produced 69-percent
transformation of PCP from the contaminated soil initially having a pH of 3.8. Significant
precipitation occurred throughout the study, leading to unexpected excursions from the
prescribed treatment protocol specified by the Risk Reduction Engineering Laboratory (RREL)
Food Products Laboratory (FPL) developers. Lack of tilling clearly compromised the ability to
evaluate the fungal treatment technology.
Information collected by both the SITE program and the RREL/FPL effort demonstrated that
fungal activity in the treatment plot was significantly lower than expected at the beginning of the
study. Fungal activity in the inoculum control increased significantly during the study, which is
most likely attributable to infestation with a wild-type fungal species.
47
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Summary demonstration removal data for the soil contaminants is presented in Table 1 for the
treatment using P. sord/'c/a. Concentration decreases of the three- and four-ring PAHs were
consistently greater following fungal treatment. Larger ring PAHs persisted in both the treatment
and control plots.
Summary and Conclusions
Treatment of PCP by fungal application had a significantly greater effect when compared with
controls. Loss of fungal activity was detected in both the fluorescein diacetate and ergosterol
analyses (Figures 4 through 8). The specified RREL/FPL treatment protocol could not be
followed in the required time frame due to excessive precipitation during the testing period. The
missing component of the protocol was the specified tilling of the treatment beds. The treatment
data clearly show that the inoculum control was infested with a wild-type fungal species, which
contributed to the biotransformation of the targeted pollutants in that plot.
Treatment by the selected fungal species was observed for PCP concentrations in excess of
1,000 mg/kg, which is greater than any reported concentrations treated using bacterial inocula
(Figure 9). Despite the remarkable differences in soil composition and characteristics for the
Wisconsin and Mississippi sites, consistent biotransformations of 80 percent to 90 percent were
observed for PCP. One notable soil feature that apparently does not affect fungal treatment is
soil pH, which, for the Wisconsin and Mississippi sites, was 3.5 and 9.2, respectively.
References
1. Lamar, R.T., and D. Dietrich. 1990. In situ depletion of pentachlorophenol from
contaminated soil by Phanerochaefe spp. Appl. Environ. Microbiol. 56:3,093-3,100.
2. Lamar, R.T., J.W. Evans, and J.A. Glaser. 1993. Solid-phase treatment of a
pentachlorophenol contaminated soil using lignin-degrading fungi. Environ. Sci.
Technol. 27:2,566-2,571.
3. Lamar, R.T., J.A. Glaser, and T.K. Kirk. 1990. Fate of pentachlorophenol (PCP) in
sterile soils inoculated with white-rot basidiomycete Phanerocriaefe chrysosporium:
mineralization, volatilization, and depletion of PCP. Soil Biol. Biochem. 22:433-4-40.
4. Davis, M.W., J.A. Glaser, J.W. Evans, and R.T. Lamar. 1993. Field evaluation of the
lignin-degrading fungus Pfianerocfiaefe sordida to treat creosote-contaminated soil.
Environ. Sci. Technol. 27:2,572-2,576.
5. U.S. EPA. 1994. Technology evaluation report: Bioremediation of PCP-and creosote-
contaminated soil using USDA-FPL/USEPA-RREL's fungal treatment technology, Vol. 1.
Final draft.
6. Lamar, R.T., M.W. Davis, D.M. Dietrich, and J.A. Glaser. 1994. Treatment of a
pentachlorophenol- and creosote-contaminated soil using the lignin-degrading fungus
Phanerochaete sordida. Submitted paper.
48
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Table 1. Summary Results for Demonstration Study (5,6)
Percentage Removal
Analyte
No-Treatment
Control
Inoculum
Control
Treatment
(P. sordida)
PCP
(RRELI/FPL data)
2-Ring PAHs
3-Ring PAHs
4-Ring PAHs
5-Ring PAHs
Total PAHs
13
19
70
83
46
14
65
71
30
48
72
67
25
66
69
69
46
64
58
27
59
49
-------
% TREATME
15
PCP(ug/g)
576
IW00D CHIP CONTROL
471
1017
Figure 1. Treatability study performance.
50
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Laachate Collection
InoculumMTrcatm*
ontrol mControl
100 it:
100ft.
25 ». 25 It.
Figure 2. Brookhaven demonstration treatment plot perspective.
Figure 3. Sampling plan layout.
51
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Treatment Control
Inoculum Control
Treatment
Dilution Soil
500 400 300 200 100 » 10 100 1000
DWeek 1 ED Week 20
Figure 4. Total fungal biomass (mg/kg) by fluorescein diacetate staining.
Treatment Control
Inoculum Control
Treatment
Dilution Soil
20 15 10 5 0 5 10 15 20
DWeek 1 BWeek 20
Figure 5. Active fungal biomass (mg/kg) by fluorescein diacetate staining.
52
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Treatment Control
Inoculum Control
Treatment
Dilution Soil
656
844
658
|
!?
a.
I
9i
42.6
120
94.8
162
1000 800600 400 200 0 200 400 600 800 1000
DWeek 1
UWeek 20
Figure 6. Total bacterial biomass (mg/kg) by fluorescein diacetate staining.
Treatment Control
Inoculum Control
Treatment
Dilution Soil
14
Q
133
80,1
89.3
120 100
"
•-.'•.,
29
26
3
iH
80 60 40 20 0 20 40
60
80 100 120 140
DWeek 1 DWeek 20
Figure 7. Active bacterial biomass (mg/kg) by fluorescein diacetate staining.
53
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Cone (mg/kg)
Found Expected
Inoculum 241
Raw Soil 0.2
Inoculated 4
Soil
24
Figure 8. Ergosterol evaluation.
1,400
1248
Time (Weeks)
Figure 9. PCP concentration depletion.
12
— No Treat
+ Inoculum
-••Treatment
20
54
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Performance Evaluation
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Integrating Health Risk Assessment Data for Bioremediation
Larry D. Claxton and S. Elizabeth George
U.S. Environmental Protection Agency, Health Effects Research Laboratory,
Research Triangle Park, NC
Introduction
Scientific literature clearly indicates that our environment contains individual substances,
combinations of substances, and complex mixtures that are hazardous to human health.
Additionally, some environmental microorganisms historically considered nonpathogens have
been shown to cause disease when humans are exposed under "nontypical" conditions. To
protect public health, those involved in remediation efforts must understand the potential for
adverse health effects from environmental contaminants and microorganisms before, during, and
after any type of remediation. When bioassay information coupled with chemical
characterization indicates a measurable loss of toxicity and testing of applied microorganisms
(if any) shows no adverse effects, one can have increased confidence that the remediation effort
will have its intended effect.
Because any human exposure to toxicants in bioremediation sites is most likely to be of the
chronic, low-concentration type, the toxicological endpoint of greatest concern typically is
carcinogenesis. Some investigators report an increased frequency of cancers in counties
surrounding hazardous waste sites. One study reported that age-adjusted gastrointestinal (Gl)
cancer mortality rates were higher than national rates in 20 of 21 of New Jersey's counties. The
environmental variables most frequently associated with Gl cancer mortality rates were
population density, degree of urbanization, and presence of chemical toxic waste disposal sites
(1). In a study of 339 U.S. counties (containing 593 waste sites) where contaminated ground
drinking water is the sole source water supply, the association between excess deaths due to
cancers of the lungs, bladder, stomach, large intestine, and rectum and the presence of a
hazardous waste site (HWS) was significant when compared with all non-HWS counties (2).
Although studies such as these do not prove causality between cancer incidence and release of
hazardous substances from waste sites, they do raise serious questions that should be examined
through more precise research.
There are numerous reasons why large gaps exist in curability to assess the health significance
of environmental exposures to chemicals in our environment. Exposure cannot be readily
quantified by measuring body burdens of contaminants, because rapid metabolism of toxic
agents prevents measurable accumulation. Due to complexities of toxin uptake, toxicologists
do not fully understand the relationships between environmental exposure and body burden (i.e.,
the amount of a toxin reaching and interacting with biological targets). Even more problematic
are the possible antagonistic and synergistic interactions that can possibly nullify predictions
based on the toxicity of individual compounds.
Bioremediation involves increasing the numbers of pollutant-degrading microorganisms to a
level at which they can have a significant effect in a timely fashion. This increase in the
microbial population also increases the likelihood of human exposure to these microorganisms.
Because environmental organisms do have some potential to cause adverse health effects,
1994 Symposium on Bioremediation of Hazardous Wastes 57
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researchers must develop methods to screen bioremedlotion microorganisms for the ability to
induce adverse effects.
The Health Effects Research Program
To address the adverse health effects questions associated with bioremediation, the EPA's Health
Effects Research Laboratory (HERL) has developed an integrated program that addresses key
issues. In collaboration with other EPA laboratories, HERL examines 1) the toxicity of known
HWS contaminants, their natural breakdown products, and their bioremediation products; 2) the
development of methods to screen microorganisms for potential adverse health effects; 3) the
potential for adverse effects when chemical/chemical and chemical/microorganism interactions
occur; and 4) the development of methods to better extrapolate toxicological bioassay results
to the understanding of potential human toxicity. The program is carried out using known HWS
pollutants, samples from microcosm studies that model the biodegradation within waste sites,
and actual waste site samples. The HERL program attempts to coordinate its own efforts with
those of the other cooperating EPA laboratories and academic researchers funded through
cooperative agreements.
HERL projects can be grouped into four categories: 1) the infectivity and pathogenicity of
environmentally released microorganisms, 2) the toxicity of metabolites of environmental
toxicants, 3) the toxicity of products of bioremediation, and 4) development of microbial
constructs that decrease the likelihood of adverse human health effects.
This talk will give a brief overview of the specific research ongoing within the HERL program,
how the research is interrelated, and how the information coming from this program could affect
developing risk assessment methods.
References
1. Najem, G., I. Thind, M. Lavenhar, and D. Louria. 1983. Gastrointestinal cancer
mortality in New Jersey counties and the relationship with environmental variables. Int.
J. Epidemiol. 12:276-289.
2. Griffith, J., R. Duncan, W. Riggan, and A. Pellom. 1989. Cancer mortality in U.S.
counties with hazardous waste sites and ground water pollution. Arch. Environ. Health
44:69-74.
58
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Construction of Noncolonizing E. Coli and P. Aeruginosa
Paul S. Cohen
Department of Biochemistry, Microbiology, and Molecular Genetics,
University of Rhode Island, Kingston, Rl
Introduction
The wall of the mammalian large intestine consists of an epithelium containing brush border
epithelial cells and specialized goblet cells, which secrete a relatively thick (up to 400 ^um),
viscous mucus covering (1). The mucus layer contains mucin, a 2-MDa gel-forming
glycoprotein, and a large number of smaller glycoproteins, proteins, glycolipids, and lipids (2-4).
For many years, we have been interested in how Escherich/'a col/ and Salmonella typhimurium
colonize the large intestines of mice and have come to the conclusion that growth in the mucus
layer is essential (5). Moreover, when E. col/and S. ryph/murium are grown in intestinal mucus
in vitro, they synthesize surface proteins that are not synthesized during growth in normal
laboratory media (6). These results led us to envision two approaches for obtaining strains of
E. co// that are perfectly healthy when grown in normal laboratory media but are unable to
colonize the large intestines of mice. The first approach is to identify and mutate E. coli genes
that are necessary for growth or survival in mucus and determine whether such mutants are
unable to colonize. The second strategy is to identify major nutrients, for growth of E. coli in
mucus, isolate mutants unable to utilize these nutrients, and determine whether such mutants are
unable to colonize. Here we report that we have been successful in both approaches with E.
coli and have now obtained strains that are unable to colonize but that are completely healthy
in the laboratory. These strains should be as effective as their parents for gene cloning yet more
effective for containment of recombinant DMA.
Technological exploitation of modern genetic techniques now holds great promise for use of
members of the genus Pseudomonas for environmental purposes (e.g., as agricultural
biopesticides [7], as detoxifiers of chemicals [8], and in prevention of ice nucleation on plants
[9]). For obvious reasons, the strains to be released into the environment must be strong,
competitive organisms. Unfortunately, strong, competitive pseudomonads can be oppor-
tunistically pathogenic (10,11). Human exposure to these microorganisms may occur in the
agricultural or industrial setting during production or application. Because a high concentration
of these microorganisms may be found in the air and water, exposure and subsequent disease
may occur through inhalation and ingestion. Clearly, strong, competitive Pseudomonas strains
should be constructed that are unable to colonize the lung tissue or the intestines of humans and
animals to minimize the possibility of opportunistic infections resulting in debilitating disease.
Here we report our initial attempts at obtaining such strains using the approaches outlined
above for E. coli.
1994 Spposium on Bioremediation of Hazardous Wastes 59
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Background
E Co//
E. coli F-18 was isolated from the feces of a healthy human in 1977 and is an excellent
colonizer of the streptomycin-treated mouse large intestine. Its serotype is rough:Kl :H5. E. co//
F-l 8Col", a poor colonizing derivative of E. co// F-l 8, contains all the E. co// F-l 8 plasmids,
and its serotype is also rough:Kl :H5. These strains were used in experiments designed to
determine why E. coli F-l 8Col" is a poor colonizer and to identify major nutrients required for
successful E. coli colonization of the mouse large intestine.
Pseudomonas Aeruginosa
P. aeruginosa AC869 is an environmental strain that has been engineered to utilize 3,5-
dichlorobenzoate as the sole source of carbon and energy (11) but which has been found to
be pathogenic for mice when administered intranasally (11). This strain was used in experiments
to determine changes associated with growth in mouse lung and cecal mucus preparations in
vitro.
Results
E Co/;
E. co// F-18 DMA was randomly cloned into E. coli F-18 Col" using the plasmid pRLB2. The
entire bank was fed to three streptomycin-treated mice, and all three mice selected the same
clone which contained a 6.5 kb insert. This insert increased the colonizing ability of E. coli F-
1 8Col" approximately 1-million-fold. After subcloning and sequencing, we identified the gene
responsible for the observed increased colonizing ability: /euX, which encodes a leucine tRNA
specific for the rare leucine codon DUG. An E. coli K-12 derivative, E. co//XAcsupP, contains
a defective /euX gene. This strain was found to be unable to colonize the large intestines of
streptomycin-treated mice; i.e., mice fed 1010 colony forming units (CFU) were essentially free
of the strain by Day 11 postfeeding. In contrast, streptomycin-treated mice fed 1010 CFU of E.
coli XAc supP containing the cloned /euX gene colonized indefinitely at 1O7 CFU per gram of
feces. Here, then, is an E. coli K-12 strain that is perfectly healthy when grown in normal
laboratory media but is unable to colonize the mouse intestine.
Glucuronate, a major carbohydrate in mouse cecal mucus, i.e., 0.6 percent by dry weight (12),
is metabolized in E. coli via the Ashwell pathway (13). Mutants unable to grow using
glucuronate as the sole source of carbon were isolated after mini-TnlO mutagenesis. One of
the mutants was unable to metabolize glucuronate, gluconate, and galacturonate, suggesting
that it was lacking 2-keto-3-deoxy-6-phosphogluconic aldolase (EC 4.2.1.14), an enzyme
encoded by the ec/a gene (14). The mutant ec/a gene was transduced into wild-type E. coli K-
12, and the E. coli F-18 eda" strain and the E. coli K-12 ec/a" strain were each fed to
streptomycin-treated mice (I010 CFU per mouse). Both strains were essentially eliminated from
the mouse intestine by Day 9 postfeeding. When the eda' mutants were complemented with the
previously cloned eda+ gene, both strains colonized indefinitely at between 107 and 108 CFU
per gram of feces. We are presently constructing E. coli F-18 and E. coli K-12 supP" ec/a"
60
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double mutants to determine whether such mutants are even more rapidly eliminated from the
mouse large intestine.
P. Aeruginosa
Rabbit antisera were raised against P. aerug/nosa AC869 grown in Luria broth, mouse lung
mucus, and mouse cecal mucus. P. aerug/nosa AC869 grown in these media were subjected
to SDS-PAGE and immunoblotting using the three different rabbit antisera as probes.
Surprisingly, the major change in P. aerug/nosa AC869 observed when grown in either mouse
lung mucus or cecal mucus was a huge increase in O-side chain containing lipopolysaccharide
(IPS). In support of this view, P. aerug/nosa AC869 grown in Luria broth was found to be
untypeable with respect to IPS, whereas the same strain grown in either mouse lung mucus or
cecal mucus was typed as O6. [IPS serotyping was kindly performed at the Statens Seruminstitut
in Copenhagen, Denmark.] This finding was of great interest, since P. aerug/nosa strains without
O-side chain on their LPS are known to be serum sensitive; i.e., they are killed by normal human
serum (15). We are therefore presently attempting to isolate mutants of P. aerug/nosa AC869
that do not make O-side chain when grown in either mouse lung mucus or cecal mucus. It is
hoped that such mutants will be perfectly healthy when grown in laboratory media, will remain
capable of metabolizing 3,5-dichlorobenzoate, yet will be nonpathogenic when inoculated
intranasally into mice.
Summary and Conclusions
The genes /euX and eda have been shown to be critical for E. coli colonization of the
streptomycin-treated mouse large intestine. These findings have allowed us to obtain E. coli K-
12 strains that grow well in normal laboratory media but are unable to colonize the
streptomycin-treated mouse large intestine. Moreover, these strains are easily transformed with
pBR322-based plasmids containing chromosomal DMA inserts. Developing healthy E. co//K-12
strains for recombinant DMA work that will not colonize the human intestine now appears
possible.
We have shown that P. aerug/nosa AC869 synthesizes more O-side chain (O6) when grown in
either mouse lung mucus or cecal mucus than in Luria broth. Since P. aerug/nosa strains that
lack O-side chain are serum sensitive, its seems likely that such mutants of P. aerug/nosa AC869
will be less pathogenic in the lungs of mice. Experiments designed to test this hypothesis are
currently in progress.
References
1. Neutra, M.R., and J.F. Forstner. 1987. Gastrointestinal mucus: Synthesis, secretion,
and function. In: Johnson, L.R., ed. Physiology of the gastrointestinal tract, 2nd ed.
New York, NY: Ravan Press, p. 975.
2. Kim, Y.S., A. Morita, S. Miura, and B. Siddiqui. Structure of glycoconjugates of
intestinal mucosal membranes. Role of bacterial adherence. In: Boedecker, E.G., ed.
Attachment of organisms to the gut mucosa, Vol. II. Boca Raton, FL: CRC Press, Inc.
p. 99.
61
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3. Allen, A. 1981. Structure and function of gastrointestinal mucus. In: Johnson, L.R.,
ed. Physiology of the gastrointestinal tract. New York, NY: Ravan Press, p. 617.
4. Slomiany, A., S. Yano, B.L. Slomiany, and G.B.J. Glass. 1978. Lipid composition of
the gastric mucus barrier in the rat. J. Biol. Chem. 253:3,785.
5. Cohen, P.S., B.A. McCormick, D.P. Franklin, R.L. Burghoff, and D.C. Laux. 1991. The
role of large intestine mucus in colonization of the mouse large intestine by Escher/ch/a
co// F-l 8 and Salmonella typhimurium. In: Wadstrom, T., A.M. Svennerholm, H. Wolf-
Watz, and P. Klemm, eds. Molecular pathogenesis of gastrointestinal infections. New
York, NY: Plenum Press, p. 29.
6. McCormick, B.A., D.C. Laux, and P.S. Cohen. Unpublished results.
7. Obukowicz, M.G., F.J. Perlak, K. Kusano-Kretzmer, EJ. Mayer, S.L. Bolten, and L.S.
Watrud. 1986. Tn5-mediated integration of the delta-endotoxin gene from Bacillus
thuringiensis into the chromosome of root-colonizing pseudomonads. J. Bacterial.
168:982.
8. Leahy, J.G., and R.R. Colwell. 1990. Microbial degradation of hydrocarbons in the
environment. Microbiol. Rev. 54:305.
9. Lindow, S.E. 1985. Ecology of Pseudomonas syringae relevant to field use of Ice'
deletion mutants constructed ;n vitro for plant frost control. In: Halvorson, H.O., D.
Pramer, and M. Rogul, eds. Engineered organisms in the environment: Scientific issues.
Washington, DC: American Society for Microbiology, p. 23.
'/ 10. George, S.E., MJ. Kohan, D.A. Whitehouse, J.P. Creason, and L.D. Claxton. 1990.
Influence of antibiotics on intestinal trad survival and translocation of environmental
Pseudomonas species. Appl. Environ. Microbiol. 56:1,559.
r" 11. George, S.E., MJ. Kohan, DA Whitehouse, J.P. Creason, C.Y. Kawanishi, R.L.
Sherwood, and L.D. Claxton. 1991. Distribution, clearance, and mortality of
environmental pseudomonads in mice upon intranasal exposure. Appl. Environ.
Microbiol. 57:2,420.
12. Krivan, H.C., and P.S. Cohen. Unpublished results.
13. Ashwell, G. 1962. Enzymes of glucuronic and galacturonic acid metabolism in
bacteria. Methods Enzymol. 5:190.
14. Folk, P., H.L. Komberg, and E. McEvoy-Bowe. 1971. Isolation and properties of
Escherichia coll mutants defective in 2-keto 3-deoxy 6-phosphogluconate aldolase
adivity. FEBS Lett. 19:225.
15. Dasgupta, T., T.R. de Kievit, H. Masoud, E. Altman, J.C. Richards, I. Sadovskaya, D.P.
Speert, and J.S. Lam. 1994. Characterization of lipopolysaccharide-deficient mutants
of Pseudomonas aeruginosa derived from serotypes O3, O5, and O6. Infed. Immun.
62:809.
62
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Field Research
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Field-Scale Study of In Situ Bioremediation of TCE-Contaminated Ground Water
and Planned Bioaugmentation
Perry L. McCarty and Gary Hopkins
Western Region Hazardous Substance Research Center, Stanford University, Stanford, CA
Introduction
Trichloroethylene (TCE) and other lesser halogenated ethenes are biodegradable through
aerobic co-metabolism. Here, microorganisms that possess oxygenases for initiating the
oxidation of either aliphatic or aromatic hydrocarbons or ammonia fortuitously can oxidize the
chlorinated alkenes to unstable epoxides. The epoxides degrade further to inorganic end
products through a combination of chemical and biological transformations. To carry out in situ
biodegradation of such chlorinated ethenes in ground water, the appropriate aliphatic or
aromatic hydrocarbon or ammonia must be added to the ground water as a substrate both to
grow a sufficient population of the desired organisms and to supply the energy required for
maintaining activity of the oxygenase. Field studies to evaluate the potential of aerobic co-
metabolism of TCE and other chlorinated alkenes have been conducted at the Moffett Naval
Air Station in Mountain View, California, since 1985 (1 -3). Methane, phenol, and toluene have
now been added to ground water at this site to determine their effectiveness as primary
substrates for chlorinated ethene degradation.
The above studies have shown the effectiveness of microorganisms indigenous to the subsurface
environment at Moffett Field for degrading chlorinated alkenes. One potential problem in
attempting to translate the results at the Moffett Field site to other field sites is that the same
primary substrates may not stimulate the growth of microorganisms with similar effectiveness.
Many different microorganisms can grow on the primary substrates found effective for TCE co-
metabolism, but their effectiveness for this purpose can vary widely. To better ensure a high
degree of effectiveness, an ability to apply bioaugmentation successfully with organisms known
to be capable of high rates of biotransformation is highly desirable. In addition, phenol and
toluene, substrates found to be highly effective as primary substrates, are also hazardous
chemicals. Use of microorganisms that can use less hazardous chemicals as primary substrates
while maintaining a high degree of effectiveness is desirable. Efforts to evaluate
bioaugmentation at the Moffett Field site are now under way.
A summary of the results from the Moffett Field test site using indigenous organisms is described
below, as are plans for in situ bioaugmentation.
Moffett Field Test Results
Over the past several years, methane and phenol have been evaluated for their effectiveness
in stimulating aerobic co-metabolic degradation of a range of chlorinated alkenes. During this
past year, toluene was evaluated as well. The results of these studies are summarized in Table
1. The concentrations of the primary substrates added were based upon their oxygen
consuming potential, which was about 20 mg/L. Thus, the added dissolved oxygen
1994 Spposium on Bioremediation of Hazardous Wastes 65
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concentration, achieved by adding pure oxygen, was maintained somewhat above this, or from
26 mg/L to 30 mg/L The results indicate that TCE was much more effectively transformed with
phenol and toluene than with methane. In addition, both phenol and toluene were much more
effective at degrading cis-1,2-dichloroethylene (c-1,2-DCE) than methane, while methane was
better at degrading trans-1,2-dichloroethylene (t-1,2-DCE). All primary substrates were highly
effective at vinyl chloride (VC) oxidation. The one problem compound here was
1,1 -dichloroethylene (1,1 -DCE), which was only evaluated with phenol. Here, only 54 percent
degradation was achieved, and the presence of this compound was found to be very detrimental
to TCE degradation, apparently because of the toxicity of the degradation intermediates.
Laboratory studies with methane indicated a similar effect.
One concern with the addition of either phenol or toluene as primary substrates for TCE co-
metabolism is the concentration remaining after biodegradation. The Moffett Field studies
indicated that within 1 day of travel time from the point of injection, both compounds were
reduced by biodegradation from the mg/L range to below 1 /*g/L. Here, sufficient oxygen was
present for effective oxidation. The EPA maximum contaminant level (MCL) and maximum
contaminant level goal (MCLG) for toluene in drinking water is 1,000 ^g/L, and the taste and
odor threshold is in the range of 20 /
-------
The laboratory studies will be conducted during the first ongoing year of this study. Field
implementation is planned for the second year of study. The different institutions involved in this
study will share in the evaluation of the effectiveness of bioaugmentation. The Moffett Field site
offers a good opportunity in general for a comparative evaluation of different approaches to In
situ biodegradation of chlorinated aliphatic compounds, and offers promise for evaluating
bioaugmentation as well.
Acknowledgments
The studies reported here were supported by EPA through the Robert S. Kerr and Gulf Breeze
Environmental Research Laboratories, the Biosystems Program, and the Western Region
Hazardous Substance Research Center, and by the U.S. Department of Energy. These agencies
have not reviewed this publication, and no official endorsements by them should be inferred.
References
1. Hopkins, G.D., L. Semprini, and P.L. McCarty. 1993. Microcosm and In situ field
studies of enhanced biotransformation of trichloroethylene by phenol-utilizing
microorganisms. Appl. Environ. Microbiol. 59(7):2,277-2,285.
2. Hopkins, G.D., J. Munakata, L. Semprini, and P.L. McCarty. 1993. Trichloroethylene
concentration effects on pilot field-scale in situ ground-water bioremediation by
phenol-oxidizing microorganisms. Environ. Sci. Technol. 27(12):2,542-2,547.
3. Semprini, L., P.V. Roberts, G.D. Hopkins, and P.L. McCarty. 1990. Afield evaluation
of in situ biodegradation of chlorinated ethenes: Part 2. Results of biostimulotion and
biotransformation experiments. Ground Water 28:715-727.
67
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Table 1. Summary of the Effectiveness of Different Primary Substrates for In Situ
Co-metabolic Biodegradation of Chlorinated Ethenes at the Moffett Field
Test Site
Primary Substrates
Methane Phenol Toluene
Primary substrate concentrations (mg/L) 6.6 12.5 9
Dissolved oxygen concentrations (mg/L) 26 30 28
Percent Percent Percent
Target Compounds removal removal removal
VC 95 >98 NE
1,1-DCE NE 54 NE
t-l,2-DCE 92 73 75
c-l,2-DCE 42 92 >98
TCE lj> 94 93
NE = Not evaluated
68
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Geochemistry and Microbial Ecology of Reductive Dechlorination of PCE and TCE in
Subsurface Material
Guy W. Sewell and Candida C. West
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory,
Ada, OK
Susan A. Gibson and William G. Lyon
ManTech Environmental Research Services Corp., Robert S. Kerr Environmental Research
Laboratory, Ada, OK
Hugh Russell
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory,
Ada, OK
Introduction
Chloroethenes are among the most common organic contaminants of ground water. In the
subsurface and other anaerobic environments, they can be transformed through a biologically
mediated, step-wise, reductive removal of chloride ions, known as reductive dechlorination.
Potentially this process can lead to nonchlorinated products that are environmentally acceptable.
Unfortunately, more mobile and toxic daughter products are intermediates. If the process
"stalls," as it often seems to in the subsurface, before reaching nonchlorinated end products, the
reductive dechlorination process may increase potential risks to human and environmental
health. Thus, the reductive dechlorination process can exacerbate or attenuate the problems
created by the release of chloroethenes such as trichloroethylene (TCE) or tetrachloroethylene
(PCE) to the subsurface and ground-water environments. In these studies, we have attempted
to identify the environmental parameters that control the onset and extent of the dechlorination
activity.
Three areas of investigation have been the focus of efforts by Robert S. Kerr Environmental
Research Laboratory researchers on the reductive dechlorination of chloroethenes. The first is
the effects of alternate electron acceptors, commonly found in the subsurface, on the reductive
dechlorination process. The second is to develop a conceptual understanding of microbial
populations and interactions that carry out the process. The third is directed toward identifying
organic compounds that can serve as sources of reducing equivalents for the dechlorination
process under native conditions or as a component of an active biotreatment application.
Results and Discussion
Saturated sandy subsurface sediments from near the municipal landfill in Norman, Oklahoma,
were collected and used as the test material in these studies. The subsurface environment from
which the material was collected is impacted by landfill leachate and classified as methanogenic.
This material has been previously shown to contain microbial populations capable of reductively
1994 Symposium on Bioremediation of Hazardous Wastes 69
-------
dechlorinating PCE (1). Figure 1 demonstrates the microorganisms' capacity for complete
dechlorination of PCE in long-term batch enrichments.
Alternate Electron Acceptor Studies
Under anaerobic conditions, the oxidation of organic compounds is linked to the reduction of
electron acceptors other than oxygen. In the subsurface may be present many different electron
acceptors, such as nitrate, ferric iron, sulfate, carbonate, or organic contaminants, such as
chloroethenes. If multiple acceptors are present in physiologically acceptable concentrations,
then the predominant terminal oxidation process is linked to the acceptor that will yield the most
energy. As this acceptor becomes limiting, the acceptor with the next highest energy yield is
utilized, and so on, until the acceptor with the lowest energy yield is utilized, which is usually
carbonate (methanogenesis). Previous research suggests that in the subsurface, reductive
dechlorination may be only a minor fate (less than 10 percent) for the reducing equivalents
generated during the anaerobic oxidation reactions (2). Whetherthis noncompetitiveness is due
to the physiological limitation of the organisms involved, the low potential energy of reactions
coupled to reductive dechlorination, or as-yet-unrecognized environmental parameters is
unclear.
Laboratory microcosm studies indicated that nitrate was extremely inhibitory to the reductive
dechlorination process (Figure 2). In the presence of nitrate, oxidizable organic carbon is
quickly utilized by microorganisms in the test material. Whether this was the only mechanism
of inhibition was unclear. Sulfate appeared to be partially inhibitory underthe conditions tested.
Again, competition for electron donor appeared to be the mechanism of inhibition. In
experiments with different initial concentrations of sulfate, significant dechlorination activity
appeared only after sulfate concentration fell below 400 fiM (Figure 3).
Microbial Process Studies
Formation of a conceptual model is the first step in the development of valid mathematical
descriptions of in situ reductive dechlorination processes. In an effort to define the metabolic
processes involved in these reactions and to enhance our understanding of the ecology of the
reductive dechlorination process, we have studied the effects of metabolic inhibitors (2-
bromoethanesulfonic acid [BESA], molybdate, and vancomycin) on butyrate, ethanol, and
formate driven reductive dechlorination of PCE in aquifer microcosms. Molybdate (5 mM) and
BESA (1 mM and 10 mM) are used as specific inhibitors of sulfate-reduction and
methanogenesis, respectively. Vancomycin (100 ppm) is used as a general eubacterial inhibitor.
Molybdate appears to be an effective inhibitor of reductive dechlorination underthe conditions
tested. BESA completely inhibited dechlorination in microcosms at 10 mM, but only partially
inhibited activity at 1 mM (Table 1). The results of experiments, such as those shown in Table
1, suggest that the dechlorinating organisms access the same pool of reducing equivalents as
the terminal oxidizing organisms.
Electron Donor Studies
We have shown in the laboratory that the availability of a suitable electron donor is essential for
dehalogenation of PCE and TCE to occur at appreciable rates in oligotrophic subsurface
environments (3,4). We and other groups have identified a wide variety of organic electron
donors that can drive biodehalogenation of chloroethenes (2-9). Conceptually, any organic
substance capable of being catabolized under anaerobic conditions should be able to support
70
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or"drive" reductive dechlorination. At some sites, however, chloroethene plumes are undergoing
dechlorination where significant amounts of anthropogenic material is not detected. Physical
interactions of chloroethenes with indigenous organic matter in soil, sediment, and aquifer solids
are important processes controlling the fate and transport of contaminants in the subsurface (10-
12). In many instances, organic carbon concentrations of aquifer solids are assumed to be
negligibly low, and in soils they are assumed to decrease exponentially with surface depth. We
have tested a working hypothesis that under certain conditions, the release of chlorinated solvent
could mobilize soil organic material, which could then serve as an anaerobically metabolizable
carbon source that will drive the dechlorination of chloroethenes.
Organic carbon was extracted from a spodic soil high in humic and fulvic acid concentrations,
collected from the vadose zone of the Sleeping Bear site in Michigan. Distilled water and
distilled water saturated with TCE were used as extractants. The presence of TCE was observed
to improve the extractability of organic compounds (although the specific identity of these
compounds is unknown at this time, as is the mechanism of extraction). Experiments were
conducted in which microcosms were spiked with the soil carbon extracts in a range of
concentrations. The extracted organic material served as the primary carbon/energy source for
subsurface microorganisms in the microcosms. The microcosms were monitored over time to
determine the ability of the extractable organic carbon to support the dechlorination of PCE.
Figure 4 shows the results of the microcosm experiments, which indicate the loss of PCE over
time for both types of extracts when present in sufficient concentrations. The dechlorination of
PCE in the active experimental treatments correlated with the production of TCE and
dichloroethylene (DCE) daughter products (data not shown), indicating that the extracts provide
a metabolizable electron donor capable of supporting microbial consortia responsible for
reductive dechlorination of PCE.
Summary and Conclusions
In situ reductive dechlorination holds significant potential for use in natural (passive) and active
in situ remediation methods. For reductive biodehalogenation to gain acceptance as a viable
alternative to conventional physical and biological treatment methods, however, it must be
predictable and well understood. Information and operational experience are needed
concerning the environmental parameters, microbial interactions, and metabolic responses that
control the initiation, rate, and extent of these degradation processes in the subsurface. An
understanding of the controlling mechanisms and the incorporation of such mechanisms into
predictive models and operational designs should allow more accurate assessment of the
applicability and implementation of anaerobic remediation of chloroethenes at chloroethene-
contaminated sites.
References
1. Suflita, J.M., S.A. Gibson, and R.E. Beeman. 1988. Anaerobic biotransformation of
pollutant chemicals in aquifers. J. Indust. Microbiol. 3:179-194.
2. Sewell, G.W., and S.A. Gibson. 1991. Stimulation of the reductive dechlorination of
tetrachloroethane in anaerobic aquifer microcosms by the addition of toluene. Environ.
Sci. Technol. 25:982-984.
71
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3. Gibson, S.A., and G.W. Sewell. 1 992. Stimulation of reductive dechlorination of
tetrachloroethane (PCE) in anaerobic aquifer microcosms by addition of short-chain
organic acids or alcohols. Appl. Environ. Microbiol. 58(4):1,392-1,393.
4. Gibson, S.A., D.S. Robinson, H.H. Russell, and G.W. Sewell. 1 994. Effects of addition
of different concentrations of mixed fatty acids on dechlorination of tetrachloroethane
in aquifer microcosms. Environ. Toxicol. Chem. 13(3).-453-460.
5. Freedman, D.L., and J.M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions.
Appl. Environ. Microbiol. 55:2,144-2,151.
6. Scholz-Muramatsu, H., R. Szewzyk, U. Szewzyk, and S. Gaiser. 1990.
Tetrachloroethylene as electron acceptor for the anaerobic degradation of benzoate.
FEMS Microbiol. Lett. 66:81-86.
7. DiStefano, T.D., J.M. Gossett, and S.H. Zinder. 1991. Reductive dechlorination of
tetrachloroethane to ethene by an anaerobic enrichment culture in the absence of
methanogenesis. Appl. Environ. Microbiol. 57:2,287-2,292.
8. Barrio-Lage, G.A., F.Z. Parsons, R.S. Nassar, and P.A. Lorenzo. 1987.
Biotransformation of trichloroethene in a variety of subsurface materials. Environ.
Toxicol. Chem. 6:571-578.
9. Fathepure, B.Z., and S.A. Boyd. 1988. Dependence of tetrachloroethylene
dechlorination on methanogenic substrate consumption by Mefhanosarc/na sp. strain
DCM. Appl. Environ. Microbiol. 54:2,976-2,980.
10. Karickhoff, S.W. 1981. Semi-empirical estimation of sorption of hydrophobic pollutants
on natural sediments and soils. Chemosphere 10:833-846.
11. Schwarzenbach, R.P., and J. Westall. 1 981. Transport of nonpolar organic compounds
from surface water to ground water: Laboratory sorption studies. Environ. Sci. Technol.
15:1,360-1,366.
12. Dzombach, D.A., and R.G. Luthy. 1984. Estimating adsorption of polycyclic aromatic
hydrocarbons on soils. Soil Sci. 137:292-308.
72
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Table 1. Effects of Various Inhibitors on Reductive Dechlorination Activity in Norman Landfill
Sediments
Donor
Treatment
BESA(lOmM)
BESA(1 mM)
Mo (5 mM)
Mo/SCV
(5/10 mM)
SO4= (10 mM)
Vancomycin
hydrochloride
(100 ppm)
Formate Ethanol
RDC DC RDC DC
0 +/- 0 -
0 +/- 0 -
0 - 00
0 00
0 +/- - +
0 +/- - +/-
Butyrate
RDC DC
0
-
0 0
0 0
0 +
-
RDC = Reductive dechlorination activity relative to positive control
DC = Electron donor catabolism relative to positive control
Mo = Molybdate (Na2MoO4»2H2O)
0 = No activity
+/- = No significant change relative to positive control
= Decreased activity relative to positive control
+ = Increased activity relative to positive control
n = Five each treatment
73
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o
1
4-1
0)
u
c
_o
*3
"o
PCE
vinyl chloride
ethene
1350
1400
1450
1500
Time (days)
Figure 1. Production of ethene and vinyl chloride from repeated PCE spikes overtime in long-
term Norman Landfill sediment enrichments. TCE and DCE intermediates not shown.
_E.
c
o
u
8
o
£
1
O
5
PCE (methanogenic)
PCE (10mM Sulfate)
PCE (10mM Nitrate)
TCE (methanogenic)
80
100
Figure 2. Effects of nitrate and sulfate on the dechlorination of PCE versus time in Norman
Landfill microcosms. Values are an average of five replicants. DCE intermediates not
shown.
74
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No Added Sulfate
_ 0.5 mM Added Sulfate
40
0 10 20 30 40 50 60
Time (days)
CO 0 10 20 30 40 50 60
Time (days)
0 mM Added Sulfate
3
to
10 20 30 40 50 60
Time (days)
^ 5.0 mM Added Sulfate
E
.g
+3
S
**
0)
o
o
O
0)
••-»
CO
M—
3
CO
0 10 20 30 40 50 60
Time (days)
Figure 3. Effects of different initial sulfate concentrations on the onset of reductive
dechlorination activity. -Cl is carbon-chloride bonds reduced and is equal to [TCE]
+ 2[DCE]. Values are an average of five replicants.
a.
c
o
1
o
c
o
O
III
No Extract
Abiotic
100 ml TCE/water Extract
50 ml TCE/water Extract
10 ml TCE/water Extract
100 ml water Extract
50 ml water Extract
10ml water Extract
100 150
Time (days)
200
250
Figure 4. Effects of water and water/TCE extracts on reductive dechlorination of PCE in
Norman Landfill microcosms. TCE and DCE intermediates not shown.
75
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Application of Laser-Induced Fluorescence Implemented Through a Cone
Penetrometer To Map the Distribution of on Oil Spill in the Subsurface
Don H. Kampbell, Fred M. Pfeffer, and John T. Wilson
U.S. Environmental Protection Agency, Ada, OK
Bruce J. Nielsen
Armstrong Laboratory, Tyndall Air Force Base, FL
Introduction
Field monitoring at spill sites usually involves collection and analysis of ground water, soil gas,
and/or core material. Applications for soil gas are limited to volatile contaminants in the vadose
zone. Ground-water assays are useful but detect only contaminants associated with the aqueous
phase. Total contamination of the subsurface, especially by petroleum hydrocarbons, is best
measured by vertical profile core sampling and analyses. Our field site characterization studies
of fuel spills involve vertical profile core sampling for direct analysis of combustible gas and
solvent extractions for total petroleum hydrocarbons (TPH) by infrared spectrometry or for
aromatic hydrocarbons by gas chromatography and mass spectrometry.
Objective
The objective of the study was to demonstrate the usefulness of a laser-induced fluorescence
cone penetrometer (LIF-CPT) as an inexpensive and rapid alternative to taking core samples for
defining the three-dimensional boundaries of an immiscible oily phase. Data are for use in the
Bioplume model to determine the amenability of the site to intrinsic bioremediation.
Operative Components
Dakota Technologies, Inc., and Applied Research Associates, Inc., under contract with the U.S.
Air Force (Armstrong Laboratory's Environics Directorate), have developed a LIF-CPT tool for
mapping the distribution of petroleum hydrocarbons as nonaqueous phase liquids (NAPLs).
Principal individuals from the two organizations involved in development and application of the
specific LIF-CPT probe used in this study are Wesley L. Bratton, Randy St. Germain, Martin L.
Gildea, Greg D. Gillispie, and James O. Shinn. Basic operating components are an optical
system to deliver tuneable laser radiation into an optical fiber for transfer downward through a
cone penetrometer to a sensor tip equipped with a sapphire window. The subsurface material
next to the window fluoresces upon exposure to laser radiation. This fluorescence radiation is
transmitted back to the surface, where intensity, fluorescent lifetime, and wavelength are
measured.
The LIF-CPT was calibrated for condensed ring aromatic hydrocarbons (specifically, the
naphthalene class), which are common constituents of petroleum products. Acquired data were
stored on a floppy disk for later processing. Data plots were also displayed on a monitor screen
76 1994 Symposium on Bioremediation of Hazardous Wastes
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for direct interpretation as the probe moved downward. The LIF-CPT was also used for
continuous profiling of soil stratigraphy and collection of soil gas, ground-water, and core
samples.
Field Site
The field study site was used extensively as a firefighting training area from 1 950 to the mid-
1 980s. Fire training pits were flooded with water, and waste jet fuel mixed with oil and solvents
was floated on the water and ignited. The burning oil was extinguished. Any unburned oil
infiltrated after these exercises. Pits were constructed in about 70 ft of sand above a confining
layer of clay. The lithology is unconsolidated and uniform glacial outwash sand. The water
table is about 30 ft below the ground surface. The ground-water seepage velocity is about 0.4
ft/day.
Less than 3 hours were required to acquire LIF data, recover the tools, decontaminate, and
move to the next site. Using the LIF-CPT to collect cores for analyses took 12 hr. Samples
could not be collected more than 3 ft below the water table. A conventional hollow stem auger
would have required 24 hr to acquire the same samples. The LIF-CPT can detect petroleum
hydrocarbons in material below the water table where material cannot be recovered as cores.
Results
Vertical profile LIF-CPT probe responses were obtained at nine locations within the study area.
Figure 1 shows probe responses in a longitudinal transect through the fire training area parallel
to the direction of ground-water flow. Strip chart displays for each location depict relative
fluorescence measurements. Location 84D was within the fire pit. Beginning at 15 ft below
the land surface, a LIF-CPT response positive for NAPL was obtained. This response extended
another 30 ft downward to a position 5 ft below the water table. A core taken at the water
table contained 125,000 mg TPH/kg soil. From combined LIF-CPT and TPH information, an
estimated 85 percent of the oily phase is present above the watertable. Remediation by vadose
zone venting may be able to remove a major fraction of the subsurface contaminated mass.
Test locations 84L and 84F were 100 ft apart and 700 ft downgradient from the fire pit (Figure
1). NAPL was present in the capillary fringe at both locations. Core material collected at the
watertable depth at location 84F contained 2,050 mg TPH/kg soil. Location 84K, located 100
ft downgradient from 84F, did not have a positive response to LIF-CPT probing. Therefore, the
leading edge of the oily-phase plume was concluded to be less than 100 ft beyond 84F.
Figure 2 is a display of the TPH and LIF-CPT results for location 84D and shows a direct
relationship with the two parameters. Other information will be presented to show that results
obtained for specific fuel aromatic hydrocarbons also show a direct relationship with TPH and
LIF-CPT results.
77
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Discussion
The LIF-CPT probe used as an onsite rapid assay tool successfully mapped in three dimensions
the oily-phase plume studied. Applications of the LIF-CPT technology will be investigated at
other field spill sites. We are continuing system development to apply the LIF-CPT method to
characterization studies at sites with different classes of hydrocarbons present.
84L-LIF "$- * 84F-LIF 84K-UF A
Leading edge
of oily phase
Figure 1. LIF response versus elevation at sampling locations.
78
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Figure 2. LIF and TPH versus depth at location 84D.
79
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Effectiveness ond Safety of Strategies for Oil Spill Bioremediation:
Potential and Limitations
Joe Eugene Lepo
Center for Environmental Diagnostics and Bioremediation, University of West Florida,
Pensacola, FL
C. Richard Cripe and P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Background
A variety of commercial agents are available for use in oil spill bioremediation. Selection of
appropriate bioremediation agents or bioremediation strategies for use in the field, however, has
been complicated by the lack of standard tests for assessing agent effectiveness and
environmental safety. Acknowledging this problem, EPA began an effort to develop protocols
for assessing effectiveness and safety of putative commercial bioremediation agents (CBAs)
based on a tiered approach (1,2).
Protocol validation for open-water and beach spill scenarios has progressed using selected CBAs
and positive control regimes. CBAs were characterized by vendors as m/crob/a/, nutrient,
enzyme, dispersant, and other. Tier I involves the gathering of pertinent information from
vendors on potentially hazardous components in the agents, putative mechanism(s) of action,
and methods and rates of application. Tier II monitors oil biodegradation in a closed,
shake-flask test system in which the oil is physically agitated. Tier III oil spill simulation tests are
designed to model field conditions thought to significantly affect CBA effectiveness in open water
or on sandy beaches; effluents can be monitored for washed out petroleum hydrocarbons or
monitored for toxicity. Tier IV testing will be an actual field evaluation of the protocol test
systems, conducted on a controlled release of oil or a "spill of opportunity."
Because of the nature of bioremediation, nutrients are common components of CBAs; however,
most forms of inorganic nitrogen exhibit some toxicity to aquatic organisms. The concern for
product toxicity is addressed at the Tier III level with two 7-day chronic estimator tests associated
with effluent toxicity evaluations that use a crustacean (Mysidopsis bahia, mysids) and a fish
(Menidia beryllina, inland silversides) (3). The mysid test has three endpoints—survival, growth,
and fecundity—while the fish test focuses on survival, growth, and development. In addition to
evaluation of toxicity of CBA alone, CBA toxicity is also assessed in the presence of a sublethal
water soluble fraction of oil to examine potentially detrimental interactions.
This report focuses on results of protocol development for CBA effectiveness and environmental
safety using the Tier III open-water and sandy beach test systems.
TIER III Test Systems
The Tier III open-water test system provides an intact, undisturbed oil-on-water slick in a
flow-through design. A constant influx of seawater below the oil slick removes CBA microbes
80 1994 Symposium on Bioremediation of Hazardous Wastes
-------
and nutrients that do not remain associated with the oil slick, as would be expected at a field
site. Test duration is 7 days. Effluent is split: one stream for oil residue analysis and the other
for toxicity testing. The slick is analyzed at the end of the test. If a significant amount of oil is
mobilized from the slick surface to the water column below (e.g., from biosurfactant production),
a subsequent test assesses the biodegradability of the transported oil.
The Tier III oiled beach test system provides a sandy beach substratum, colonized for 1 week
by microflora indigenous to seawater. The systems model tidal influx and egress. The surface
is oiled and 2 days later a CBA or other bioremediation strategy is applied. Beach test systems
run for 28 days, after which the oil residues can be extracted for analysis. Effluents are collected
for analytical or toxicological examination.
Forthe purpose of the Tier III protocol, generic environmental parameters were selected for both
the open-water and the beach test systems. The oil was applied to a 0.5-mm thickness,
turbulence was standardized, and temperature was set to 20°C. The oil was artificially
weathered (4) to simulate the loss of volatiles expected following a spill and to minimize changes
in composition due to loss of volatile components. Gulf of Mexico seawater (30 parts per
thousand salinity) provided a source of hydrocarbon-degrading microorganisms capable of
responding to increased nutrients in the presence of crude oil as a carbon source.
Two treatments, in three replicates each, are used: 1) a control with oil alone and 2) a
treatment with both oil and CBA. Criteria for evaluating the effectiveness of bioremediation in
the Tier III open-water test systems are based on statistically significant (p 5 0.05) reductions
in the weight of oil and in the amount of selected gas chromatography/mass spectrometry
(GC/MS) analytes remaining in the test vessels and test-system effluent relative to the control
vessels and effluents.
Supplemental research (in progress) will examine the effects of environmental parameters (e.g.,
salinity, temperature, water turbulence, increased treatment time or increased CBA application
rates) on the effectiveness of the CBAs to provide more site-specific information.
Results
Validation of Open-Water Test System Using Positive Controls and CBAs
To establish baseline performance for the Tier III open-water test systems, we used
positive-control treatments that were surrogates for either nutrient CBAs or microbial CBAs.
Three conditions were tested: 1) Gulf of Mexico seawater control, 2) seawater amended with
nutrients (to test for the ability of nutrients to enhance the degradation capability of the natural
degraders), and 3) nutrient-amended seawater supplemented with a daily inoculation of
hydrocarbon-degrading bacteria as a test of competent, high levels of microbial biomass.
The effectiveness of the positive control in the open-water test system is presented in Table 1 as
a percent of the oil remaining relative to controls to which neither nutrients nor microbes were
supplied. Values represent an average of three replicate test chambers. The number of the
GC/MS endpoints out of a total of 70 analytes that were significantly reduced relative to the
control for each agent is also tabulated. Nutrients alone failed to stimulate biodegradation by
the microbial population indigenous to Gulf of Mexico seawater. Several analyte endpoints,
81
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however, were significantly different as the result of action by the hydrocarbon-degrading
bacteria in the presence of nutrients.
Table 1 also reports the results of six CBAs selected as representatives of each CBA type. Each
was applied to the oil slick in the test systems according to the instructions supplied by the
vendor. Of the six, only the nutrient CBA gave a promising result, effecting a change in 1 8 of
the GC/MS analytes and a statistically significant reduction (although only 1 percent) in total oil
residue weight. In contrast, the nutrient-amended seawater treatment of the positive control
experiment effected a statistically significant change in only one of the GC/MS analytes.
Only in the positive control experiment in which nutrients were supplied continuously and oil
degrading bacteria were applied daily did we find effects on a relatively large number of
endpoints as well as substantial reduction in the total weight of the oil recovered.
Validation of Oiled Beach Test System Using CBAs
Table 2 shows the percent of oil and oil components remaining in the test systems after 28 days
of exposure to four CBAs in Tier III beach test systems. The control treatments, in which
seawater flushed the systems in the same tidal regimes as in the CBA chambers, lost substantial
amounts of the lower molecular weight polycyclic aromatic hydrocarbons. Positive control
experiments and experiments in which we attempted to run sterile control treatments have
suggested that the disappearance is biologically mediated, although whether the compounds
have been washed intact from the test systems or catabolized is still being investigated.
Environmental Safely of CBAs
An important ecotoxicological consideration for CBAs is the possible production of toxic
metabolites. This is addressed at the Tier III level with a mysid 7-day chronic estimator test on
the effluent from the open-water and beach test systems. A key assumption is that the test
system designs are conservative with respect to dilution; thus, if toxicity is not observed under
these mixing scenarios, it is unlikely to occur in a field application. Increased toxicity (compared
with the toxicity of effluent from control systems containing only oil) exceeding that of the product
alone (from Tier II testing) would suggest the need for further studies that focus on potentially
toxic metabolites. Table 3 indicates that the open-water effluent from most CBAs demonstrated
low or no toxicity. Safety has not yet been evaluated using the beach test system.
One application of toxicology came as a result of adapting a 10-day amphipod (Leptocheirus
plumulosus) (5) sediment toxicity test to evaluate potential toxic metabolites associated with the
sand of the beach test system after the 28-day CBA efficacy test. We observed that oiled
sediment, whether subjected to bioremediation or not, was toxic. Although this phenomenon
prevented accurate assessment of potential toxic metabolites in the sediment, it led to research
to determine whether toxicity testing could be used as an efficacy endpoint, focusing on the
potential of a CBA to render an oiled sediment suitable for amphipod recolonization. The
results of preliminary studies will be discussed.
82
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Conclusions
We have completed validation of the open-water and sandy-beach testing systems. Thus far,
the CBAs examined during our protocol development work have shown little toxicity and should
pose little environmental threat to the organisms tested when applied according to the vendor's
suggested regime. Some CBAs effected significant changes in one or more targeted
hydrocarbons relative to the control; however, it should be emphasized that the sum of all
GC/MS analytes is less than 6 percent of the total oil. Moreover, no substantial decreases in
oil residue weights were associated with treatment by CBAs.
By daily addition of microbial biomass and nutrients to the open-water system, however, we were
able to demonstrate the greatest biodegradation of oil components within the 7-day period,
including a significant weight loss; i.e., there were significant decreases in 30 of the 70 GC/MS
analytes. Thus, we conclude that the test system itself was capable of giving a measurable
response, although its accurate modeling of actual site-specific field conditions remains to be
evaluated. These results may also indicate that the recommended application rates of CBAs are
insufficient to produce substantial changes in oil biodegradation. Daily or more frequent
additions may be untenable in some open-water field situations (e.g., large-area spills);
however, spills of a more confined nature may be reasonably treated with higher or more
frequent applications.
There are substantial barriers to effective performance of oil-spill CBAs, among them dilution
rates, nutrient and biomass limitations, and a limited time in which a CBA can remain in contact
with the oil spill. Efficacy indices from analytical chemistry, coupled with assessments of toxicity
for CBAs, should provide useful information to an on-scene coordinator. These limitations will
be discussed in the light of our experience with the Tier III effectiveness protocol.
Acknowledgments
Validation of the effectiveness protocol for Tier III open-water and beach test systems as well as
the ecotoxicology for Tier II and Tier III was performed through a cooperative agreement
(CR-81 8991 -01) between the University of West Florida Center for Environmental Diagnostics
and the EPA Environmental Research Laboratory at Gulf Breeze. The following people
contributed ideas and technical assistance during the development of this project: Wanda Boyd,
Mike Bundrick, Peter Chapman, Jim Clark, Carol Daniels, Barbara Frederick, Tim Gibson,
Wallace Gilliam, Jeff Kavanaugh, Joanne Konstantopolis, Tony Mellone, Len Mueller, Neve
Norton, Jim Patrick, Bob Queries, Mike Shelton, Scott Spear, Phil Turner, Ling Wan, George
Ryan, Vicki Whiting, Diane Yates, and Shiying Zhang.
References
1. Lepo, J.E. 1993. Evaluation of Tier III bioremediation agent screening protocol for
open water using commercial agents: Preliminary report. EPA/600/X-93/001.
University of West Florida/U.S. Environmental Protection Agency, Gulf Breeze
Environmental Research Laboratory, Gulf Breeze, FL.
83
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2. National Environmental Technology Application Corporation (NETAC). 1993. Oilspills
bioremediation products testing protocol methods manual. Pittsburgh, PA: University
of Pittsburgh Applied Research Center (August).
3. U.S. EPA. 1988. Short-term methods for estimating the chronic toxicity of effluents and
receiving waters to marine and estuarine organisms. EPA/600/4-87/028. Washington,
DC.
4. International Organization for Standardization. 1989. Crude petroleum oil:
Determination of distillation characteristics using 15 theoretical plates columns. Draft
international standard ISO/DIS 8708.
5. Schlekat, C.E., B.L. McGee, and E. Reinharz. 1992. Testing sediment toxicity in
Chesapeake Bay with the amphipod Leptocheirus p/umu/osus: An evaluation. Environ.
Toxicol. Chem. 11:225-236.
Table 1. Percentage of Analyte Remaining Relative to Controls After 7 Days of Treatment With
Bioremediation Agents or Positive Control Regimes
3CBA OR POSITIVE CONTROL TREATMENT
ANALYTE IM/M _E _N IM/M M/D JD +IM +IM/M
97 102 *92 92 103 105 94 **34
C30 101 100 99 96 102 100 99 **57
PHYTANE 103 103 99 101 102 101 102 95
PRiSTANE 103 99 103 104 101 99 104 98
FLUORENE 102 106 96 105 102 107 99 95
CHRYSENE 103 117 95 107 **90 114 96 95
PHENANTRENE 102 102 99 102 99 103 102 *97
/V-ALKAIMES 98 105 **92 96 102 106 96 **40
AROMATICS 102 105 98 102 102 103 103 97
TOTAL OIL 99 101 *99 103 99 102 102 *93
bENDPOINTS 5 1 18 6 1 0 1 30
treatment type: E = enzyme, N = nutrient, D = dispersant, M = microbial, +N =
nutrient positive control, +N/M = nutrient positive control + microbes
bNumber of endpoints showing a statistically significant change at 0.05 or less.
* p < 0.05; ** p < 0.01
84
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Table 2. Tier III Effectiveness Results of Beach System Tests With CBAs
PERCENT REMAINING HYDROCARBON ANALYTE
CBA Type
C18
Phytane
C18/Phyt
Fluorene
Dibenzothioph
Phenanthrene
Chrysene
Gravimetric
Nutrient
**33
**86
**39
32
52
53
106
*a92
Control
90
93
97
23
51
48
106
94
Nutrient/
Microbial
**20
**53
**37
29
52
51
104
**89
Nutrient/
Microbial
**25
85
29
43
68
67
100
*91
Control
89
89
100
39
68
68
99
96
Dispersant
89
86
100
50
81
84
100
94
amean of 2 replicates; all others were means of 3 replicates
* p < 0.05; ** p < 0.01
Table 3. Tier III Results of 7-Day Chronic Estimator Tests With Mysidopsis bah/a
Max. Effluent
CBAb Cone. (%)
E 63
N 55
N/M 66
D 10
7-DAY LC50
(95% C.I.I
>63 survival
growth
fecundity
>55 survival
growth
fecundity
>66 survival
growth
fecundity
3.7 survival
(3 - 4.6) growth
fecundity
Comparison to Oil Control3
NOEC LOEC
63
63
63
55
55
55
66
66
66
3
NE
3
NE
NE
NE
NE
NE
NE
NE
NE
NE
10
NE
__c
Comparisons were made between the effluent from control systems that contained oil
alone with those from systems containing oil and the CBA
CBA types as defined in the note to Table 1.
cFecundity data at these effluent concentrations greater than 3% are disregarded because
no females were found alive.
NOEC = no observed effect concentration; LOEC = lowest observed effect concentration;
NE = no effect
85
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Pump #1
(800 ml/day)
Air
Pump #2
(20,000 ml/day)
Microcosm &
Synchronous Motor
^
fc-
Acidified
Effluent
Bottle
C~^i
A
Stir Plate
Effluent
for
Toxicity
Testing
Pump #3
400 ml/day
Figure 1. Tier III simulated open-water oil spills test system.
Acidified
Effluent
for
Analysts
View of Bottom of
Inner Beaker
EPA/RU93036-00
Figure 2. Tier III simulated oiled beach test system.
86
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Pilot-Scale Research
-------
Pilot-Scale Evaluation of Alternative Biofilter Attachment Media for
Treatment off VOCs
Francis L Smith, George A. Serial, Makram T. Suidan, and Pratim Biswas
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Richard C. Brenner
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
Since enactment of the 1990 amendments to the Clean Air Act, the control and removal of
volatile organic compounds (VOCs) from contaminated air streams has become a major public
concern (1). Consequently, considerable interest has evolved in developing more economical
technologies for cleaning contaminated air streams, especially dilute air streams. Biofiltration
has emerged as a practical air pollution control (APC) technology for VOC removal. In fact,
biofiltration can be a cost-effective alternative to the more traditional technologies, such as
carbon adsorption and incineration, for removal of low levels of VOCs in large air streams (2).
Such cost effectiveness is the consequence of a combination of low energy requirements and
microbial oxidation of the VOCs at ambient conditions.
Preliminary investigations (3) were performed on three media: 1) a proprietary compost mixture;
2) a synthetic, monolithic, straight-channeled (channelized) medium; and 3) a synthetic,
randomly packed, pelletized medium. These media were selected to offer a wide range of
microbial environments and attachment surfaces and different air/water contacting geometries.
The results of this preliminary work demonstrated that 95+ percent VOC removal efficiency
could be sustained by all three media at a toluene loading of 0.725 kg COD/m3-d, but at
different empty bed residence times (EBRTs). For the pelletized medium, this performance could
be achieved at an EBRT of 1 min, for the channelized medium at 4 min, and for the compost
medium at 8 min. Both synthetic media developed headless over time, with the pelletized
medium showing a pressure drop in excess of several feet of water after sustained, continuous
operation. These results left open the question of which medium could provide the optimum
combination of high VOC elimination efficiency at high loading with minimum pressure drop.
This paper discusses the continuing research being performed for development of biofiltration
as an efficient, reliable, and cost-effective VOC APC technology. The objectives of the recent
research were to conclude the evaluation of the three media and to develop workable strategies
for the removal and control of excess biomass from the (ultimately) selected pelletized
attachment medium.
Experimental Apparatus
The biofilter apparatus used in this study consists of three independent, parallel biofilter trains,
each containing 4 ft of attachment medium: biofilters A, B, and C. A detailed schematic and
1994 Symposium on Bioremediation of Hazardous Wastes 89
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equipment description is given elsewhere (-4). Biofilter A was filled with a proprietary compost
mixture, B with a Corning Celcor* channelized medium, and C with a Manville Celite* pelletized
medium. Biofilters A and B are square and have an inner side length of 5.75 in.; biofilter C
is round, with an inside diameter of 5.75 in. The air supplied to each biofilter is highly purified
for complete removal of oil, water, CO2, VOCs, and particulates. After purification, the air flow
for each biofilter is split off, injected with VOCs, humidified, and fed to the biofilters. The air
feed is mass flow controlled, and the VOCs are metered by syringe pumps. The flow direction
of the air and nutrient inside each biofilter is downward. Each biofilter is insulated and
independently temperature controlled.
Buffered nutrient solutions are fed to biofilters B and C. A detailed description of the nutrient
composition is given elsewhere (4). Each of these biofilters independently receives a nutrient
solution containing all the necessary macro- and micronutrients, with a sodium bicarbonate
buffer. The nutrients required in biofilter A were included as part of the original compost
mixture.
Results
Biofilter A
This biofilter run on the compost medium was made to evaluate the effects of temperature and
then loading on toluene removal efficiency. Figures 1 a and 1 b summarize the biofilter
performance. The biofilter was started up and, after some operational difficulties, stabilized by
Day 10 at 52°F, 50 ppmv toluene, 2 min EBRT, and a removal efficiency of about 58 percent.
On Day 1 7, the temperature was raised to 60°F, resulting in a rise in efficiency to about 75
percent, which decreased after Day 24 into the 60s, and after Day 32 into the 50s. On Day
41, the temperature was increased to 70°F, resulting in a gradual increase in efficiency to about
75 percent by Day 47. On Day 53, the temperature was increased to 80°F, resulting in an
increase in efficiency into the low 80s. On Day 61, the temperature was increased to 90°F,
resulting in a further increase in efficiency to the mid-90s (Figure la). After Day 77, the feed
was increased slowly to about 95 ppmv toluene, resulting in a drop in efficiency to about 88
percent. Further increases in the feed concentration to a maximum of 1 80 ppmv toluene on
Day 139 resulted in a further decline in efficiency to about 58 percent (Figure 1 b). The run was
terminated on Day 215.
Biofilter B
This biofilter run was made on the synthetic channelized medium to evaluate the effects of
temperature and then nutrient feed rate on removal efficiency. The biomass in the channels of
the medium remaining from the previous run was removed by hydroblasting the eight 6-in. high
medium blocks from top and bottom. The comers of these square blocks were filled with grout
to provide a "round" active block. This last step was taken to match a round block cross section
with the round pattern of the nutrient delivery spray nozzle. Figure 2 shows the biofilter
performance as a function of time. The biofilter was started up at 52°F, 50 ppmv toluene, and
2 min EBRT. By Day 36, the removal efficiency had drifted over a range from about 62 percent
to 80 percent. On Day 36, the nutrient feed rate was increased from 30 L/day to 60 L/day,
while keeping the mass loading of the nutrients constant. The increased nutrient flow rate
effectively doubled the wetting cycle from 20 sec/min to 40 sec/min. An immediate increase
90
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in efficiency to 99 percent was observed, which then quickly dropped and ranged by Day 50
between about 30 percent and 70 percent. On Day 50, the nutrient feed rate was increased
to 90 L/day (increasing the wetting cycle to 60 sec/min), but the efficiency dropped from 69
percent and ranged by Day 67 from about 22 percent to 65 percent. On Day 67, the
temperature was raised from 52°F to 60°F, and the efficiency increased to 66 percent. By Day
75, the efficiency was 87 percent, and this level was maintained to Day 83. After Day 83, the
temperature was raised in 10°F steps to 90°F, but the efficiency did not improve. In fact, for
the rest of the run, at 90°F and 60 I/day the efficiency ranged between about 58 percent and
83 percent. The run was terminated on Day 152.
Biofilter C
The first biofilter run on the synthetic pelletized medium was made to evaluate the effects of
pressure drop and then temperature on toluene removal efficiency. The biofilter was charged
with pellets used in the previous run. These pellets were washed by hand in hot water (150°F)
until the accumulated surface biomass had been removed and the pellets were free flowing.
Figure 3 presents the biofilter performance as a function of time. The biofilter was started up
at 52°F, 50 ppmv toluene, and 2 min EBRT. By Day 21, the removal efficiency was 99 percent,
and by Day 27, it had reached 100 percent and remained at this level until Day 50. From Day
51 to Day 57, the EBRT was gradually reduced to 1 min, causing the efficiency to drop to 84
percent. Subsequently, the toluene removal efficiency rapidly increased to the low 90s and
remained in that range until Day 81. On Day 82, the temperature was raised to 60°F, and the
efficiency steadily rose until complete biodegradation of the toluene was reached on Day 89.
This essentially 100-percent efficiency in toluene removal was maintained through Day 97.
During the period between Day 54 and Day 97, pressure drop across the system increased from
0.2 to 5.5 in. water. From Day 97 to Day 111, the efficiency dropped steadily from 100
percent to 86 percent, while the pressure drop increased from 5.5 to 6.0 in. water. On Day
112, the temperature was increased to 70°F, and the efficiency rebounded by Day 113 to a
peak value of 97 percent, after which it dropped to 85 percent by Day 188. On Day 119, the
temperature was raised to 80°F, and the efficiency rose to about 89 percent by Day 120.
During the period from Day 112 to Day 120, the pressure drop increased from 6 to 1 8 in.
water. By Day 128, the efficiency had steadily dropped from 89 percent to 77 percent as the
pressure drop increased from 18 to 27 in. water. This pattern of a steady loss of efficiency with
a coincident increase in pressure drop suggests the development of short circuiting within the
biofilter medium due to biomass accumulation, which results in a significant reduction in actual
contact time. The run was terminated on Day 128.
The second biofilter run on this medium was conducted to evaluate routine biomass control by
backwashing. The biofilter was charged with a 50:50 mixture of fresh pellets and pellets from
the previous run. The used pellets were thoroughly washed by hand in tepid water (90°F) until
the accumulated surface biomass had been removed and the pellets were free flowing. Figure
4 shows the biofilter performance as a function of time. The filter was started up at 90°F, 50
ppmv toluene, and 2 min EBRT. By Day 4, the removal efficiency was 100 percent. (Note: This
second run, started up with pellets washed in tepid water, contrasts with the slower startup in the
first run, where the pellets were washed with hot water.) On Day 8, the feed was increased to
250 ppmv toluene; the efficiency dropped to 97 percent and ranged between 92 percent and
98 percent until Day 25, when it again reached 99 percent. Subsequently, the efficiency
dropped as low as 86 percent before regaining 99 percent on Day 81, after which the efficiency
was nearly always 99+ percent. Initially, backwashing was performed once a week by using
100 L of fresh water at a rate of 6 gallons per minute (gpm). After Day 28, the frequency was
91
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increased to twice per week, and after Day 38, the volume was increased to 200 L. These
changes were made because measurable pressure drop was observed between backwashings.
On Day 73, the backwash rate was increased to 15 gpm to induce full fluidization. Although
the pressure drop increase was minimal, the efficiency did not improve, suggesting some form
of channelizing within the bed. Therefore, on Day 80, the length of the backwash period was
increased to 1 hr by recirculating the backwash water. After this final adjustment, the toluene
removal efficiency, as mentioned above, achieved and sustained 99+ percent. During this latter
period, the total volume of water used per backwash was optimized to 120 L. Of this volume,
70 L were used for the 1 -hr backwash recycle, while the remaining 50 L were used to flush the
released solids from the reactor. Figure 5 shows the development of biomass with time. After
Day 38, the rate of biomass accumulation declined with the increase in the wash volume. After
Day 73, the accumulation rate became nearly zero with the implementation of full fluidization.
Since then, no change in the backwash procedure has been made, and the accumulation of
biomass within the biofilter has leveled off at about 180 g with the pressure drop between
backwashings typically under 0.2 in. of water.
Conclusions and Future Work
A marked improvement in toluene removal efficiency with increasing temperature was
demonstrated in this study for the compost mixture, the channelized medium, and the pelletized
medium. The direct consequence of this finding is that much less medium would be needed for
a biofilter operating at 90°F than at 52°F, resulting in a proportional reduction in capital cost.
The economic tradeoff with the cost of heating the incoming air should usually favor operation
at these warmer conditions.
The modest performance of the compost mixture with respect to increased loading
complemented our earlier findings with respect to decreasing EBRT (3). Unfortunately, implicit
limitations of the experimental apparatus may have resulted in reduced performance.
Specifically, the manufacturers recommended using a width-to-depth ratio of 1:1, rather than
1:8. They also stated that from their experience the only effective means of controlling bed
moisture content was to weigh the entire biofilter. This was impossible with the heavy stainless
steel unit used here, which was bolted to a support frame. Several moisture measurement and
control strategies were attempted, but it was never possible to be certain that the bed moisture
content was consistently at the reported optimum range, i.e., between about 50 percent and 60
percent (5,6). The sometimes erratic performance may have been influenced by variations in
bed moisture content. The best removal efficiencies achieved by the compost mixture, however,
were better than shown by the channelized media but worse than shown by the pelletized media.
The performance of the channelized medium also confirmed our earlier findings that this
medium is distinctly inferior to the pelletized medium (3). The best performance was achieved
during the use of new medium blocks. After biomass accumulation within the channels and
subsequent removal by hosing, the performance never regained the previous, still modest, levels.
Attempts to adjust nutrient flow as a means of testing the effect of the duration of wetting in the
nutrient application cycle did not overcome the previously demonstrated efficiency limitations.
The more erratic performance of this medium after removal of the biomass suggests that this
medium may be unsuitable for sustained efficiency after periodic cycles of biomass removal.
This erratic performance, due to suspected random uneven plugging of channels by biomass,
coupled with its relatively low overall removal efficiency, difficulty in biomass removal, and
92
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intrinsically high medium cost, suggests that this medium may not be a viable option for this
application.
The pelletized medium exhibited the best and most consistent performance of the three media
tested. It rapidly achieved high removal efficiencies at high toluene loadings. As the first run
demonstrated, however, an excessive accumulation of biomass, shown by a rise in the pressure
drop across the medium, results in a substantial loss in efficiency, followed by a very rapid rise
in pressure drop. This suggested that efficient, sustained performance might be achieved
through early and periodic control of biomass accumulation by backwashing. In the second run,
the implementation of a suitable backwashing strategy for biomass control was achieved by
using full medium fluidization. This strategy permitted sustained operation of the biofilter at high
loadings with efficiencies consistently at 99+ percent. According to mass balance calculations,
the biomass retained within the biofilter stabilized at a nearly constant level.
Future work will concentrate on further optimizing the use of the pelletized medium, with the
objective of minimizing the medium volume required for a selected ARC technology application.
References
1. Lee, B. 1991. Highlights of the Clean Air Act Amendments of 1990. J. Air Waste
Mgmt. Assoc. 41(1):16.
2. Ottengraf, S.P.P. 1986. Exhaust gas purification. In: Rehn, H.J., and G. Reed, eds.
Biotechnology, Vol. 8. Weinham, Germany: VCH Verlagsgesellschaft.
3. Sorial, G.A., F.L Smith, PJ. Smith, M.T. Suidan, P. Biswas, and R.C. Brenner. 1993.
Evaluation of biofilter media for treatment of air streams containing VOCs. Paper No.
AC93-070-002. Proceedings of the Water Environment Federation 66th Annual
Conference and Exposition, pp. 429-439.
4. Sorial, G.A., F.L Smith, PJ. Smith, M.T. Suidan, P. Biswas, and R.C. Brenner. 1993.
Development of aerobic biofilter design criteria for treating VOCs. Paper No. 93-TP-
52A.04. Presented at the 86th Annual Meeting and Exhibition of the Air and Waste
Management Association, Denver, CO (June).
5. Bohn, H.L 1993. Biofiltration: Design principles and pitfalls. Paper No. 93-TP-
52A.01. Presented at the 86th Annual Meeting and Exhibition of the Air and Waste
Management Association, Denver, CO (June).
6. Van Lith, C, S.L. David, and R. March. 1990. Design criteria for biofilters. Trans. Inst.
Chem. Eng. 686:127-132.
93
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100
90
"0
u 80
K
o
a
o
70
60
50
Toluene Loading
«
0.45 kg COD/m day
EBRT = 2 minutes
50
60
70
80
90
Temperature, F
Figure 1 a. Effect of temperature on the performance of the compost biofilter.
100
90
w 80
>
O
a
V
W 70
u
a
o
° 60
50
Temperature = 90 F
EBRT = 2 minutes
0.0
/
0.4
0.6 O.B 1.0 1.2 1.4
Toluene Loading, kg COD/day
1.6 1.8
Figure 1 b. Effect of toluene loading on the performance of the compost biofilter.
100
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94
-------
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80
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i toluene Loading
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o
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60 80 100 120 140 160
Sequential Date, days
Figure 2. Performance of channelized biofilter with respect to toluene removal of an EBRT of
2 min.
95
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100
60 F 70 E /BO F
1 min. E 3RT
2 mm. EBRT
Toluene Loading kg GOD/m day
Pressure Drop
llntmay>_—rrfrr^i----—ffTf*O»-CPr^ I i i i t 1 I i i i . I
40 60 80 100
Sequential Date, days
120
140
Figure 3. Performance of pelletized biofilter with respect to toluene removal at 1 min and 2
min EBRT without backwashing.
100
90
80
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£ 50
13 40
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20
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Efficiency
Toluene Loading kg COD/m day
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25
50 75 100 125
Sequential Date, days
150
20
15
Figure 4. Performance of pelletized biofilter with respect to toluene removal at 2 min EBRT
with backwashing.
96
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700
• VSS Produced (nitrogen balance)
o VSS Lost (backwash + effluent liquid)
o VSS Retained
0
0
50 75 100 125
Sequential Date, days
150
175
Figure 5. Development of pelletized biofilter with time (VSS closure).
97
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Biological Treatment of Contaminated Soils and Sediments Using Redox Control-
Advanced Land Treatment Techniques
Margaret J. Kupferle, In S. Kim, Guanrong You, Tiehong Huang, and Maoxiu Wang
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Gregory D. Sayles
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Douglas S. Lipton
Levine-Fricke Consulting Engineers, Emeryville, CA
Introduction
Soils and sediments contaminated with highly chlorinated aromatic compounds such as
polychlorinated biphenyls (PCBs), pentachlorophenol (PCP), hexachlorobenzene (HCB), and
1,1,1 -trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) are found at many of the Superfund sites
that have been placed on the National Priority List for cleanup. Bioremediation has been
proposed as a means for converting these contaminants into less toxic or nontoxic substances.
The biodegradation rates of many highly chlorinated compounds can be accelerated by
controlling the redox potential (or oxidation-reduction potential, ORP) of the treatment
environment. In general, the biochemical pathway providing the highest rate for the initial steps
of microbial destruction of highly chlorinated organics is anaerobic reductive dechlorination.
Once partially dechlorinated, the resulting compounds typically degrade faster under aerobic,
oxidizing conditions. Efficient and complete degradation of highly chlorinated contaminants is
possible when the two redox conditions are sequentially applied.
Sequential treatment techniques have been proposed as a means of treating aqueous wastes
and slurries containing soils contaminated with highly chlorinated aromatic compounds such as
PCBs, PCP, HCB, and DDT, among others (1,2). For example, the meto and para chlorines of
highly chlorinated PCBs are removed by anaerobic reductive dechlorination; however, the orfho
chlorines are only slowly removed by the same bioprocess. Aerobic organisms remove the ortrio
chlorine and complete the mineralization of the compound relatively quickly. Thus, sequential
anaerobic-aerobic treatment should provide relatively rapid destruction of PCBs (3,4). The
process applied to PCB-contaminated sediments has been studied by other research groups
(1,5) and is currently being demonstrated in the field. Woods et al. (6) suggested that an
anaerobic-aerobic sequential treatment strategy would be an attractive treatment alternative for
highly chlorinated phenols because the anaerobic consortium used in their study was capable
of reductively dechlorinating highly chlorinated phenols to monochlorophenols. The
monochlorophenols were not reductively dechlorinated further; however, they are known to
degrade in aerobic treatment processes. Bench-scale work in our own group has evaluated the
applicability of the technology for the treatment of HCB and DDT contamination. Results for
DDT degradation in the anaerobic phase have been encouraging. DDT, which usually
accumulates as 1,1 -dichloro-2,2-bis(p-chlorophenyl)ethane (DDD) underanaerobic conditions,
has been degraded to less chlorinated intermediates such as 2,2-bis(p-chlorophenyl)ethanol
98 1994 Symposium on Bioremediation of Hazardous Wastes
-------
(DDOH) and dichlorobenzophenone (DBP) using a combination of chemical reducing agents
and surfactants in conjunction with anaerobic culture. Aerobic degradation of these
intermediate products is under investigation.
A practical means of applying sequential redox control in field-scale remediation is needed.
Since land treatment is a well-understood, cost-effective means of conducting aerobic biological
treatment of soils contaminated with compounds such as petroleum and polycyclic aromatic
hydrocarbons (PAHs), we have proposed to extend it to include an anaerobic phase to treat
compounds amenable to reductive dechlorination. In this project, methods are being developed
for operating land treatment reactors under anaerobic as well as aerobic conditions so that a
sequential strategy can be readily applied in the field. Methods of applying multiple cycles of
alternating redox conditions to achieve cleanup are also being investigated. During this project
year, these methods will be tested using PCP-contaminated soil in pilot-scale soil pan reactors.
In subsequent project years, we plan to investigate soils from several types of sites, including
sites contaminated with DDT.
Methodology
Reactor operating strategies that deliver adequate anaerobic and aerobic microbial
environments are currently being developed using uncontaminated soil in a pilot-scale unit with
two pans (reactors). Each pan holds approximately 30 kg of soil. Various methods of
maintaining anaerobic conditions in the soil reactor currently are being evaluated, including
simply flooding the soil bed, adding an easily degradable organic compound(s) to serve as an
oxygen scrubber near the surface, and covering the soil bed with an air-impermeable cover to
inhibit the transport of oxygen. Liquid addition and permeate recycle techniques also are being
evaluated during the anaerobic phase of operation. Methods for returning the soil bed to
aerobic conditions will be investigated when the anaerobic phase is complete. The soil bed will
be drained and, if necessary, a vacuum will be applied below the bed to assist in drainage and
aeration of the soil. Bulking agent addition may be required to improve aeration of the soil.
Hand mixing/tilling methods and sample collection methods will be investigated during both
phases.
A source of contaminated soil has been identified, and background information about the site
and the range of contaminants and contaminant concentrations has been obtained. Soil
samples (courtesy of Wildemere Farms, Inc., Lake City, Florida) from various locations at the
American Wood Products site in Lake City, Florida, representing a range of contamination levels
have been analyzed for chlorinated phenolics. A comparison of PCP concentrations in these
samples found by our group and by an independent laboratory is shown in Table 1.
Trace amounts of less chlorinated intermediates were noted in some of the samples analyzed
in our laboratory, but the concentrations were under the method detection limit (~1 mg/kg).
Dioxins, low-level contaminants in technical grade PCP, were analyzed by the independent
laboratory; the congener with the highest concentration was octochlorinated-dioxin at 1 8 ppt,
and the highest risk congener, 2,3,7,8-tetrachlorodioxin, was nondetectable. For the pilot-scale
work at EPA's Test and Evaluation (T&E) Facility, soil will be obtained from two of the sampling
points at the site that represent high and low levels of contamination. Approximately 600 kg
of soil from each sampling location will be required. The soil will be transported to the T&E
99
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Facility, where it will be shredded, sieved, mixed, sampled, characterized, and placed in the
pilot-scale units.
Six pilot-scale units with four pans each, a total of 24 pans, will be employed in this study. The
experimental design is shown in Table 2. Each treatment will be duplicated in separate reactors.
A "clean" soil spiked with PCP will be tested in addition to the two concentrations obtained from
the site. The use of recycle for moving the liquid through the soil versus the maintenance of
stagnant liquid in the pan will be one of the variables tested. Sterile controls will be run in
parallel with each treatment to monitor for abiotic losses. The simplest approach will be tested
first. The soil will be flooded with site water, if it can be obtained, or with deionized water (close
approximation to rainwater) to create anaerobic conditions.
Specific treatment assignments to specific pans in the six four-pan units have been randomly
assigned. Randomization is necessary because this design will be statistically analyzed as a
three-factor analysis of variance (ANOVA) with replication. The three factors are biological
activity, soil "type," and recycle. The dependent variable that will be used to compare treatments
and evaluate treatment effectiveness will be the molar sum of the chlorinated aromatics (parent
compound + metabolites) removed per kilogram of dry soil at a set time interval (e.g., after 4
months in anaerobic treatment and after 2 months in aerobic treatment). Molar concentrations
will be normalized using the initial concentration in each treatment so that the treatments can
be compared statistically using ANOVA techniques.
To supplement the statistical comparison, the pans will be sampled at 2-week interim time
points, and the samples will be analyzed for the parent contaminant and chlorinated aromatic
metabolites to provide insight into the pattern of removal. Other monitoring will include daily
measurement of pH, ORP, and temperature. Total and volatile solids will be determined each
time a soil sample is collected so concentration can be calculated on a dry soil basis and so
soil moisture can be monitored during the aerobic phase.
Serum bottle experiments using soil from the site will be conducted concurrently with the pilot-
scale reactors. In these experiments, alternative treatment strategies including co-substrate and
nutrient amendments and inoculation of acclimated organisms will be explored as means of
improving treatment rate and extent. Pilot-scale evaluation of alternatives found to be optimal
is planned for FY95.
References
1. Zitomer, D.H., and R.E. Speece. 1993. Sequential environments for enhanced
biotransformation of aqueous contaminants. Environ. Sci. Technol. 27(2):227-244.
2. Armenante, P.M., D. Kafkewilz, G. Lewandowski, and C.M. Kung. 1992. Integrated
anaerobic-aerobic process for biodegradation of chlorinated aromatic compounds.
Environ. Prog. 11 (2):113-122.
3. Abramowicz, D.A. 1990. Aerobic and anaerobic biodegradation of PCBs: A review.
Crit. Rev. Microbiol. 10(3):241-251.
100
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4. Bedard, D.L 1990. Bacteria I transformation of polychlorinated biphenyls. In: Kamely,
D., et al., eds. Biotechnology and biodegradation, Vol. 4. The Woodlands, TX:
Portfolio Publishing Co.
5. Avid, P.J., L. Nies, and T.M. Vogel. 1991. Sequential anaerobic-aerobic
biodegradation of PCBs in the river model. In: Hinchee, R.E., and R.F. Offenbuttel,
eds. Onsite bioreclamation. Boston, MA: Butterworth-Heinemann.
6. Woods, S.L, J.F. Ferguson, and M.M. Benjamin. 1989. Characterization of
chlorophenol and chloromethoxybenzene biodegradation during anaerobic treatment.
Environ. Sci. Technol. 23:62-68.
Table 1. Soil Analysis for PCP
Sample
1
2
3
4
5
6
7
8
9
10
11
12
13
14
PCP in Analyzed Soil Samples*
Mean Concentration
(mg PCP/kg dry soil)
12.2
37.8
103
109
8.66
3.54
136
116
209
133
445
69.2
4.21
1.11
Standard Deviation
(mg PCP/kg dry soil)
0.66
1.8
2
12
4.08
0.19
9
7
15
7
38
4.2
1.00
0.22
Data from
Independent Lab
(mg PCPAg soil**)
16.8
46.4
64.5
59.7
3.29
3.08
115
93.3
178
125
N/A
N/A
N/A
N/A
* Three replicates analyzed per sample
** Dry weight not specifically indicated in report
101
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Table 2. Experimental Design for Soil Pan Reactors
Treatment
Biologically
Active
Biologically
Inactivated
No recycle
Recycle
No recycle
Recycle
Contamination Level
Low
2*
2
2
2
High
2
2
2
2
Spiked
Clean
Soil
2
2
2
2
*Two reactors per treatment
102
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Research Leading to the Bioremediation of Oil-Contaminated Beaches
Albert D. Venosa and John R. Haines
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Makram T. Suidan, Brian A. Wrenn, Kevin L Strohmeier, B. Loye Eberhart,
Edith L. Holder, and Xiaolan Wang
University of Cincinnati, Cincinnati, OH
Introduction
During the summer of 1994, EPA, in cooperation with the Delaware Department of Natural
Resources and Environmental Control (DNREQ, plans to conduct a small-scale field study on
the shoreline along Delaware Bay involving bioremediation of crude oil released in small
quantities on 15 identical plots. The goals of this research project are: 1) to obtain sufficient
statistical evidence to determine if bioremediation with inorganic mineral nutrients and/or
microbial inoculation enhances the removal of crude oil contaminating mixed sand and gravel
beaches, 2) to compute the rate at which such enhancement takes place, and 3) to establish
engineering guidelines on how to bioremediate an oil-contaminated shoreline. Prior to
conducting such a study, two important pieces of information need to be defined: 1) the
minimum nitrogen concentration enabling the degrading populations to metabolize the oil
components at their maximum rate at all times and 2) the frequency at which the nutrients must
be added to maintain such a concentration. The first question is being addressed in the
laboratory/ the second in the field. This paper discusses the design and conduct of laboratory
and field experiments and presents some of the preliminary data answering the two questions
posed.
Two nutrient application strategies were tested, one involving a sprinkler system spraying water
soluble nutrients on the plot, the other incorporating a trench situated above the high tide line
but below the underlying water table (1). In the latter method, tracer is applied through a
manifold at the bottom of the trench just before high tide. The underlying ground water carries
the tracer to the treatment zone as tides ebb and flow over time.
Methodology
Laboratory Experiment
To determine the minimum nitrogen concentration needed for maximum biodegradation over
time, six semicontinuous flow respirometric beach reactors able to mimic tidal flow on a beach
(2) were used. A major advantage of this microcosm is its ability to provide continuous,
real-time monitoring of oxygen uptake and carbon dioxide evolution without the need for
destructive sampling. Each tidal flow reactor measures 75 mm in diameter and 260 mm deep
and holds approximately 2 kg beach material. The columns are fed from a 20-L Teflon reservoir
containing a flexible inner Teflon bag. Influent seawater contained inside the flexible bag is
continuously pumped by a "wave" pump into the top of the reactor through a spray nozzle. The
1994 Symposium on ffioremediation of Hazardous Wastes 103
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seawater finally returns to the 20-L carboy outside the Teflon bag to maintain separation
between influent and effluent. The headspace of the reservoir, the reaeration flasks, and the
reactor column are all connected to maintain constant pressure in the system. Oxygen is
supplied automatically to the microcosm system from a respirometer whenever a deficit is
sensed. The cumulative uptake of oxygen is tracked continuously over time, enabling analysis
of reaction kinetics. An experiment was set up in which six different concentrations of nitrate-N
(ranging from 0 mg/L to 10 mg/L) were supplied to the reactors, and biodegradation of
heptadecane was followed continuously. A mixed culture from the shoreline of Delaware,
previously enriched with heptadecane, was used as the inoculum.
Field Experiment
The field study is located on a sandy and slightly gravelly beach south of Slaughter Beach,
Delaware. Surface morphology consists of a loose upper 25-mm thick layer of smooth gravel
ranging in size from 4.75 mm to 19.1 mm atop coarse sand having a moderately homogenous
particle size distribution. Two plots measuring 5 m x 10 m were set up. Two types of wells were
situated within and outside the vicinity of each plot: piezometers and sampling wells. The
piezometers consisted of black iron rods about 2.5 m long and 3.2 cm inside diameter (ID).
The bottoms were fitted with a specially fritted brass tip that allowed water to enter the well
filtered of fine sand or peat particles characterizing the deeper zone of the beach. The
piezometers were equipped with pressure transducers connected to a data logger mounted to
a wooden post in back of and between the plots. The pressure transducers were used to
measure the water head continuously to provide accurate readings of water levels during the
tidal cycles.
The sampling wells were constructed of stainless steel and were also about 2.5 m long.
Openings of 3.2 mm ID were drilled into the sides of the wells starting at 15 cm from the
bottom tip and extending upward at intervals of 15 cm over a total length of 1.8 m. Stainless
steel tubing of the same diameter was welded to these openings. The tubing extended inside
the wells from the openings to above the tops of the wells, where plastic tygon tubing was
attached for collection of water samples via syringe. The openings in the sides of the wells were
covered with a fine-mesh stainless steel screen to filter out particulate matter that might clog the
tubing. Thus, water samples at each depth interval were totally independent from other water
samples, which enabled measurement of tracer concentrations atone depth without influence
from tracer concentrations at other depths.
For the sprinkler plot, 20 kg of LiNO3 was dissolved in 800 L of fresh water. For the trench
application, 30 kg was dissolved in the 800 L because the trench, being 5 m wider than the plot
width, required more tracer for an equivalent amount to reach the desired area of the plot. Two
types of samples were collected at each sampling event: subsurface sand and water from the
sampling wells. The sand samples were collected with a bulb planter at low tide only, water
samples at both low and high tides. Water samples were analyzed for lithium by atomic
absorption spectrophotometry (3). Sediment samples were extracted and filtered, and the pore
water was measured for lithium by activated alumina (AA).
104
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Results
Laboratory Experiment
Figure 1 summarizes results from two of the six reactors. Space limitations preclude presentation
of all the data. Clearly, the reactor fed 10 mg/L NO3'-N exhibited twice the O2 uptake and
CO2 evolution as the reactor fed 0.5 mg/L. Also, the effluent nitrate levels measured in the
reactor fed 10 mg/L were only slightly lower than the influent nitrate levels, whereas effluent
nitrate in the reactor fed 0.5 mg/L declined to virtually undetectable levels. Thus, 0.5 mg/L
nitrogen appears to limit the biodegradative activity. The next higher concentration used in the
experiment was 2.5 mg/L, which gave approximately the same results as the 10 mg/L level.
Another experiment was designed (results not ready at the time of this writing) to determine more
closely the minimum nitrogen level that still provides maximum biodegradation.
Field Experiment
The plots were situated in the high intertidal zone corresponding to where the spring high tide
would flood the entire plot. The tide experienced, however, was a neap tide, which means that
the high tide did not cover the plot at all during the first few days of the experiment. Figure 2
is a three-dimensional mesh graph summarizing the lithium concentrations measured in the
upper 12 cm to 13 cm of sand in the sprinkler plot from time 0 hr to 37 hr after application of
tracer, corresponding to six tidal cycles. Immediately after application, the lithium concentration
in the sediment pore water ranged from approximately 120 mg/kg to 200 mg/kg sand. Thus,
the distribution of the tracer by the sprinkler was not as even as originally hoped. At the next
low tide (12 hr later), the lithium had declined about 50 percent and was more evenly
distributed over the plot surface. At the next low tide (25 hr after application), lithium
concentrations at the bottom of the plot had declined to almost undetectable levels. The
previous high tide had covered this much of the plot, which explains the low levels of tracer
there. Note that the lithium tracer in the upper two-thirds of the intertidal zone, which had not
been wetted by the high tide, still persisted at slightly lower levels than the previous low tide.
At 37 hr, corresponding to the third full tidal cycle, more of the plot had been covered by the
incoming tide as reflected by the lithium concentrations shown in the figure. At the 48-hr mark,
a storm had occurred, causing the tidal waters to completely submerge the plot. Lithium levels
were undetectable (< 1 mg/kg) in the surface sediment from about 55 hr through the remainder
of the experiment, which lasted 10 days. Lithium concentrations in the surface sediment of the
trench plot were undetectable until after the storm event, when low levels of lithium finally
appeared due to underlying water carrying the tracer to the surface.
Tracer levels measured in well water samples from the ground water below the plot (data not
shown) persisted for the duration of the experiment. The tracer moved up and down with the
tides, which is consistent with observations made by Wise et al. (2) in Alaska.
Conclusions
From the laboratory experiment, the minimum nitrogen concentration needed to stimulate
maximum microbial degradation of hydrocarbons is somewhere between 0.5 mg/L and 2.5
mg/L. From the field experiment, it appears that application of fertilizer should be conducted
every day when the tide covers the entire contaminated zone. When the tide only covers the
105
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lower intertidal zone, nutrient application is not needed, since the nutrients will likely persist for
several days. During this period, the microorganisms will be in constant contact with nitrogen
and phosphorus, which will allow time for biostimulation to proceed. For the trench method to
work, two trenches seem to be needed, one for the spring tide and one for the neap tide.
References
1. American Public Health Association. 1989. Direct air-acetylene flame method 311 IB.
In: Standard methods for the examination of water and wastewater, 17th ed.
Washington, DC.
2. Strohmeier, K.L, M.T. Suidan, A.D. Venosa, and J.R. Haines. 1993. A beach
microcosm for the study of oil biodegradation. Poster presented at the Battelle In Situ
and Onsite Bioreclamotion Conference, San Diego, CA.
3. Wise, W.R., O. Guven, F.J. Molz, and S.C. McCutcheon. 1993. Nutrient retention
time in a high-permeability oil-fouled beach. J. Environ. Eng. In press.
106
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0 HOURS
12 HOURS
p'ot L«i
25 HOURS
37 HOURS
200
P'""-«8th. m
Figure 2. Three-dimensional plot showing behavior of lithium tracer during the first 37 hr
after application.
108
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Engineering Optimization of Slurry Bioreactors for Treating Hazardous Wastes
John A. Glaser and Paul T. McCauley
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Majid A. Dosani, Jennifer S. Platt, and E. Radha Krishnan
I.T. Environmental Programs, Inc., Cincinnati, OH
Introduction
Biological treatment of contaminated soil slurries may offer the optimal treatment conditions for
soil bioremediation at an economically viable cost. Despite this promise, slurry bioreactor
treatment of soils has not achieved the status of a durable, reliable, and cost-effective treatment
option. As part of a general program of engineering assessment and optimal design of slurry
bioreactors, both bench- and pilot-scale reactors have been developed to address the pressing
needs for missing operational data associated with slurry bioreactor use. These reactors are
located at the EPA Testing and Evaluation (T&E) Facility located in Cincinnati, Ohio.
Methodology
Application of slurry bioreactors to the treatment of contaminated soil has been conducted with
a variety of soil types (1). Case studies and cost comparison are available, but the information
associated with these studies is incomplete (2). An EPA best demonstrated available technology
(BOAT) study has investigated the application of slurry reactors to creosote-contaminated soil (3).
To systematically evaluate each of the major components of slurry biotreatment, a research
program has been organized along the general divisions of physics, biology, and chemistry.
Each of these divisions is a major contributor to the slurry biotreatment process. The physics of
mixing has been the early focus of the slurry research program. The criteria for optimal mixing
for slurries has not received the required attention. Five different criteria have been advanced
for the chemical processing industry (4-7): 1) maximum uniformity of suspension, 2) complete
off-bottom suspension, 3) complete on-bottom motion of all particles, 4) filleting but no
progressive fillet formation, and 5) height of suspension (cumulative particle size distribution,
percent solids, percent suspension, weight-percent ultimate suspended solids, and percent
ultimate weight-percent settled solids).
For the initial evaluation of the bench-scale reactor (Figure 1), performance was assessed
through the correlation of critical factors contributing to the efficiency of mixing (Figures 2
through 6). Solids composition was investigated for its influence on power consumption and the
rotational speed of the impeller (Figures 2 through 4). Clear optimal ranges for air flow are
evident in the recorded data. The optimal operating conditions are found at the point where the
lowest power is consumed.
A soil from St. Louis Park, Minnesota, was contaminated with creosote constituents and used to
evaluate the performance of bench-scale slurry reactors. The bench-scale bioslurry reactor was
1994 Symposium on Koremediotion of Hazardous Waste 109
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constructed from a 8-L glass conventional resin kettle with a four-port cover fitted with standard
taper joints. The reactor vessel was fabricated to have three sample ports located 5 cm, 10 cm,
and 15 cm vertically from dead center of the reactor bottom. The ports in the reactor cover
permitted introduction of the stirring shaft, influent and effluent gas lines, and a thermocouple
temperature probe into the soil slurry. Operational slurry volume was 6 L or 75 percent of the
total reactor volume.
Ten bench-scale reactors were used to assess the effect of engineering variables on the
degradation of polycyclic aromatic hydrocarbon (PAH) constituents over a 10-week treatment
period.
The experimental design of the treatability study is outlined in Figure 7. Experimental variables
selected for this study were soil loading, rotational speed of the mixing impeller, and dispersant.
Soil solids concentrations of 10 percent and 30 percent (dry weight basis) were tested. Two
mixing speeds were evaluated. A high mixing rate was selected for complete off-bottom
suspension. A low mixing rate was arbitrarily set at 200 rpm lower than the high mixing rate.
Effective high mixing rates were found to be 650 rpm and 900 rpm for the 10-percent and 30-
percent soil solids, respectively. The dispersant (Westvaco, Reax 1OOM) was added to test its
ability to minimize foam production. Foam formation is an operational problem associated with
the application of soil bioslurry technology and is thought to be connected with naturally
occurring organics found in certain soils.
Two separate reactors were operated under abiotic conditions to serve as bioinactive control
reactors. Formaldehyde was used as a biocide in these reactors and maintained at 2-percent
residual concentration.
The following monitoring and operating conditions held constant for the reactors:
• Dissolved oxygen greater than 2 mg/L
• pH range of 6 to 9
• Ambient temperature recorded daily
• Treatment duration of 10 weeks
• Nutrient C:N:P ratio = 100:10:1
• Antifoam as needed to control foam
Results
For purposes of convenience, the individual PAH constituents were grouped into two categories:
two- and three-ring compounds and four- through six-ring constituents. Initial concentration of
total PAHs in the soil prior to treatment were 1,750 ppm in the 10-percent solids loading slurry
and 2,047 ppm in the 30-percent slurry, indicating a degree of heterogeneity in the soil slurry.
The total PAH concentration was reduced to 408 ppm in the 10-percent slurry (runs 1 through
4) and 419 ppm in the 30-percent slurry (runs 5 through 8) after 7 days of treatment. In the
110
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10-percent slurry runs, the concentrations of two- and three-ring PAH compounds decreased
from 709 ppm to 67.4 ppm, and concentrations of four- through six-ring PAHs decreased from
1,041 ppm to 340 ppm; whereas for the 30-percent slurry runs, the concentrations of two- and
three-ring PAH compounds decreased from 798 ppm to 45.1 ppm, and concentrations of four-
through six-ring PAHs in the 30-percent slurry runs decreased from 1,249 ppm to 374 ppm.
Summary and Conclusions
The total PAH concentration was reduced by 85 percent to 90 percent after 70 days of
treatment. The major decrease in PAH concentrations occurred in the first 7 days, where total
PAHs removed ranged from 75 percent to 82 percent. Soil solids concentrations significantly
affected removal rate and the final treatment endpoint (PAH concentration). A maximum
removal for the 30-percent solids loading was achieved after 21 days of treatment. Continued
treatment after 21 days had little effect on further reduction of PAH concentrations. In the 10-
percent solids runs, however, PAH concentrations continued to be reduced between Days 21
and 70. The final concentrations of two- and three-ring and four- through six-ring PAH
categories, as well as total PAHs, for the 10-percent solids runs were half the levels in the 30-
percent solids conditions.
These results show that removal efficiencies are apparently not as sensitive to complete off-
bottom suspension as we had expected. Similarly, removal rates appear to be unaffected by
mixing speed ranges. The dispersant additive did not effectively suppress foam formation or
enhance PAH removal.
This initial study clearly identifies soil solids composition as a major factor controlling treatment
goals. Lower solids compositions and longer treatment duration may favor treatment to lower
PAH concentrations in the soil. Because removal rates observed in this work may be specific
to the soil matrix selected for study, the generalizations arising from this work can be used for
guidance for future applications of soil-slurry bioreactors. Treatability studies are necessary,
however, to determine the most effective operating variables for each waste matrix before em-
barking on any large-scale treatment. Foaming potential of a contaminated soil should be
evaluated prior to treatment to minimize operational problems associated with foam formation
at higher solids concentrations.
References
1. U.S. EPA. 1990. Engineering bulletin: Slurry biodegradation. EPA/540/2-90/016.
Cincinnati, OH.
2. Ross, D. 1990. Slurry-phase bioremediation: Case studies and cost comparisons.
Remediation 1:61-74.
3. U.S. EPA. 1991. Pilot-scale demonstration of a slurry-phase biological reactor for
creosote-contaminated soil. EPA/540/A5-91/009. Cincinnati, OH.
4. Oldshue, J.Y. 1983. Fluid mixing technology. In: Chemical engineering. New York,
NY: McGraw-Hill, pp. 94-124.
Ill
-------
5. Oldshue, J.Y. 1983. Fluid mixing technology and practice. Chem. Eng. pp. 92-108
(June).
6. Oldshue, J.Y. 1 990. A guide to fluid mixing. Rochester, NY: Lightnin.
7. Hicks, R.W., J.R. Morton, and J.G. Fenie. 1976. How to design agitators for desired
process response. Chem. Eng. pp. 102-110 (April).
CLEAN AIR
OR OXYGEN
EFFLUENT GAS
STAINLESS STEEL
GAS TUBE
REACTION KETTLE:
6L WORKING •
VOLUME
MIXER
SLURRY LOADING PORT
TEMPERATURE PROBE
IMPELLER SHAFT
STAINLESS STEEL BAFFLES
SAMPLE PORTS
IMPELLERS
Figure 1. Bench-scale slurry bioreactor.
112
-------
30% Sand/Clay Solids
10,000
1.000:
Q.
cc
•**• Minimum rpm
-'-Power, watts
100
10 15 20
Air Flow (scfh)
45
Figure 2. Complete off-bottom suspension (5 in. between impellers, baffle=design 3).
40% Sand/Clay Solids
•*• Minimum rpm
-i-Power, watts
10
15 20 30
Air Flow (scfh)
45
Figure 3. Complete off-bottom suspension (5 in. between impellers, baffle=design 3).
50% Sand/Clay Solids
1,000
100
10
::::: :n :::::><::::> .". .TTT*
^^
. H '~^ r^*^
Optimal Range
: : : : : 4 :::.:•:::::* ::::::::::
£.i>
2
" I 1
| -+- Rower, watts
O.5
n
10 15 20 30
45
Air Flow (scfh)
Figure 4. Complete off-bottom suspension (5 in. between impellers, baffle=design 3).
113
-------
30% Sand/Clay Solids
100 400 500 600 700
RPM, min
Figure 5. Air flow optimization (5 in. between impellers).
1000
1500
Air Flow
•X-Oschf
+ 10 scfh
*-15 scfh
*20 scfh
*-30 scfh
-»-45 scfh
30% Sand/Clay Solids
•* 0 schf
10 scfh
scfh
-•-20 scfh
scfh
+ 45 scfh
100 400 500 600 700
RPM, min
Figure 6. Airflow optimization (6 in. between impellers).
1000 1500
114
-------
Variable
Run A B
1
2 - +
3 +
4 + +
5
6 - +
7 +
8 + +
9 + +
C D
-
-
-
-
•4-
+
+
+ -
- +
10 + + + +
Variable
Dispersant 0 mg/L 50 mg/L
450/700 rpm 650/900 rpm
Mixing
Speed
Soil
Solids
CH,O
0 mg/L
0 mg/L
50 mg/L
50 mg/L
Figure 7. Experimental design (St. Louis Park soil).
115
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Development and Evaluation of Composting Techniques for Treatment of Soils
Contaminated With Hazardous Wastes
Carl L. Potter and John A. Glaser
U.S. Environmental Protection Agency, Andrew W. Breidenbach Environmental Research
Center, Cincinnati, OH
Majid Dosani, Srinivas Krishnan, Timothy Deets, and E. Radha Krishnan
I.T. Environmental Programs, Inc., Cincinnati, OH
Introduction
Significant progress in optimizing conditions and applying the power of biotechnology to
large-scale com post systems requires a working understanding ofthe processes and mechanisms
involved. Prototype bench-scale units have been designed and tested to evaluate composting
processes using contaminated soils. Identification of suitable co-compost and bulking agents,
appropriate ratios of soil to organic components, and effective aeration strategies and rates
have been selected as major factors requiring investigation.
This research program is designed to develop a thorough engineering analysis and optimization
of composting as a process to treat soil contaminated with hazardous waste. Bench-scale
composters serve as diagnostic tools to estimate the treatment capability of larger systems. Fully
enclosed, insulated reactors permit reliable data collection on mechanisms of metabolism and
the fate of toxic chemicals during soil composting.
We are currently studying the ability of compost microorganisms to biodegrade polycyclic
aromatic hydrocarbons (PAHs) in in-vessel reactors located at the EPA Testing and Evaluation
Facility in Cincinnati, Ohio. Soil contaminated with PAHs was obtained from the Reilly Tar Pit
Superfund site in St. Louis Park, Minnesota, for use in this study.
Background
Composting holds potential to provide low-cost treatment of hazardous waste with minimal
environmental controversy. Commercial compost operations are operated as black-box systems
in that optimization is largely approached through trial and error. Treatment of hazardous waste
cannot be conducted with suboptimal controls to meet the specified endpoints.
Some proponents of compost treatment have claimed significant success in destruction of
hazardous wastes without strong data to support their claims. Disappearance of parent
compounds has been used to claim that microorganisms successfully degraded waste chemicals.
Some toxic chemicals, however, could potentially adsorb to, or react with, humic substances in
the compost and become undetedable by chemical analysis. Such toxicants might later desorb
from humus and migrate to the biosphere. This highlights the need for well-controlled studies
to rigorously document degradation rates and to identify metabolic products of hazardous
116 1994 Symposium on Koremediotion of Hazardous Wastes
-------
chemicals, metabolically active microbial species, and mechanisms of hazardous chemical
transformation in compost systems.
The conventional aerobic compost process passes through four major microbiological phases
identified by temperature: mesophilic (30°C to 45°C), thermophilic (45°C to 75°C), cooling,
and maturation. The greatest microbial diversity has been observed in the mesophilic stage.
Microbes found in the thermophilic stage have been spore forming bacteria (8ac///us sp.) (1) and
thermophilic fungi (2,3). Microbial recolonization during the cooling phase brings the
appearance of mesophilic fungi whose spores withstood the high temperatures of the
thermophilic stage. In the final compost stage, the maturation phase, most digestible organic
matter has been consumed by the microbial population, and the composted material is
considered stable.
Reactor Design
Ten 55-gal, insulated stainless steel composters have been constructed to perform closely
monitored treatability studies. The units stand upright, and air flows up through the compost
mixture. Completely enclosed units permit periodic analysis of volatile organic compounds
(VOCs) and online analysis of oxygen, carbon dioxide, and methane. Cylindrical reactor design
permits mixing of reactor contents by rolling each unit on a drum roller at desired intervals.
Each composter houses four thermocouples connected to a central computer for online
temperature measurements. Thermocouples reside at three equally spaced locations within the
compost mixture, and a fourth thermocouple tracks ambient temperature outside the reaction
vessel. One operational scheme permits temperature control by introduction of ambient air
through a computer-controlled valving system. If the temperature of a unit exceeds a
predetermined value, the computer switches that unit to high air flow to cool the reaction
mixture. After the high-temperature unit cools to a specific temperature, the computer switches
the unit back to low air flow.
Periodic determination of compost moisture content in each reactor unit permits adjustment of
total moisture content in the compost matrix to 40 percent to 50 percent. Moisture condensers
inside compost units promote recycling of moisture. Otherwise, each unit could lose 10 Ib to
15 Ib of water daily.
Current Research
Prototype composter evaluation has proceeded through several different designs. The
performance of each design was evaluated by conducting a treatability experiment using the St.
Louis Park soil. For our design criteria, one particular prototype offered considerable versatility.
This design is currently being converted to stainless steel reactor units.
Current studies focus on defining acceptable operating conditions and process characteristics
to establish suitable parameters for treatment effectiveness. Parameters of interest include
aeration, moisture dynamics, heat production, and physical and chemical properties of the
compost mixture.
117
-------
Aeration studies evaluate porosity (air flow) in the compost system and attempt to discover
relationships between free airspace, forced airflow, and composting rate. Aeration studies also
investigate roles of anaerobic and aerobic metabolism in chemical degradation. Anaerobic
pockets may benefit the process by initiating degradation of recalcitrant compounds, especially
highly chlorinated compounds, via reductive metabolism. After an initial reductive step, aerobic
biodegradation of toxicants may proceed more readily. The research program will attempt to
identify optimal aeration rates and pile mixing frequency for the most effective combination of
anaerobic/aerobic conditions for biodegradation of recalcitrant substrates. These studies will
investigate whether forced anaerobiosis and inoculation with a facultative anaerobe prior to
development of aerobic compost conditions enhances biodegradation of toxic wastes.
Studies on moisture dynamics measure rates of change in moisture content in different regions
of the compost reactor. A compost pile can lose moisture through evaporation and convection.
Moisture dynamics are evaluated in terms of aeration, temperature, and compost composition
(e.g., soil type and co-compost material).
Heat production may be highly variable throughout the compost reactor. We have devised a
method to continually monitor temperature changes (heat production) at various reactor
locations. Bench-top composters are insulated to control heat loss, thereby mimicking a large-
scale compost pile where heat is lost by ventilation and water evaporation more than by
conduction.
Physical properties of the compost mixture include density (g/cm3), pH changes in various
reactor locations, pressure drop across the pile if it is actively aerated, and the fraction of solids,
moisture, and organics. These investigations focus on the potential to enhance biodegradation
by manipulation of physical and biological parameters that influence the process. These studies
will also investigate whether recycling mature compost material into fresh compost enhances
biodegradation of contaminants.
Early microbiological studies will focus on characterizing changes in biological activity during
the four stages of composting. We will also attempt to identify microbial species responsible for
significant biodegradation of PAHs during each compost stage, and look for reappearance of
fungi and other mesophiles (e.g., Acfinomycetes) during the cooling stage.
Future Research
Future investigations will include technical developments necessary to improve composting
applications for degradation of hazardous waste. This will involve increased application of pilot-
scale compost systems in addition to ongoing research in bench-top composters. Emphasis will
be placed on developing techniques for trapping VOCs from pilot-scale systems, determining
mass balance of contaminant degradation in the compost, and identifying microbial species
responsible for biodegradation of contaminants.
Future studies will also attempt to validate extrapolation of results from bench-top to pilot-scale
and field demonstration levels. Maintaining a bench-top system at optimum conditions is
relatively easy compared with a large-scale composter, where optimum conditions will not
prevail at all times. The degree of variance from optimal conditions requires investigation and
approximation in small-scale systems.
118
-------
References
1. Nakasaki, K., M. Sasaki, M. Shoda, and H. Kubota. 1 985. Change in microbial
numbers during thermophilic composting of sewage sludge with reference to CO2
evolution rate. Appl. Environ. Microbiol. 49(1):37-41.
2. Fogarly, A.M., and O.H. Tuovinen. 1991. Microbiological degradation of pesticides
in yard waste composting. Microbiol. Rev. pp. 225-233 (June).
3. Strom, P.P. 1985. Identification of thermophilic bacteria in solid-waste composting.
Appl. Environ. Microbiol. 50(4):906-913.
119
-------
Remediation of Contaminated Soils From Wood-Preserving Sites Using Combined
Treatment Technologies
Amid P. Khodadoust, Gregory J. Wilson, and Makram T. Suidan
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Richard C. Brenner
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
Pentachlorophenol (PCP), a pesticide used as a wood-preserving compound since the 1930s,
has been placed on EPA's National Priority List of pollutants (1). The cleanup of contaminated
soil from PCP manufacturing facilities and wood-preserving sites has been mandated through
the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) (2).
Among technologies employing physical, chemical, and biological processes for the removal of
PCP from contaminated soils, solvent washing followed by biological treatment of the wash fluid
appears to be a viable alternative (3). The selection of the solvent depends on the hydrophobic
nature of the pesticide and the soil wetting capability of the solvent (4,5). Mueller et al. (6)
found that ethanol effectively removed polycyclic aromatic hydrocarbons (PAHs) from wet
contaminated soils. Previously, equal proportions of ethanol and water were found to have the
highest removal efficiencies for aboveground batch extractions of PCP from soil at various
soil:solvent ratios (7). In addition, 50-percent and 75-percent ethanol solutions achieved higher
removal efficiencies at low solvent throughputs in simulated in situ soil flushing experiments.
Chemically synthesized extracts from the soil washing process were treated using an anaerobic,
fluidized-bed granular activated carbon (GAC) bioreactor. The PCP was reduced to an
equimolar concentration of monochlorophenol, which caused inhibition of the biological system.
Reduction of the feed concentration of PCP to 200 mg/L appeared to alleviate reactor inhibition.
Results and Discussion
Solvent Extraction Studies
The effectiveness of the 50-percent ethanol/water mixture was evaluated for the removal of PCP
from soils that had been aged for 3 weeks, 3 months, 6 months, 9 months, and 1 year. The
aging of soil spiked with 100 ppm PCP occurred in the absence of natural weathering, i.e., the
soil was not exposed to ground and atmospheric influences. The 50-percent ethanol/water
solution was used for simulated in situ soil flushing of 20 x 40 and 100 x 140 U.S. mesh soils
and 20 x 40 U.S. mesh soil conditioned at 60°C. The soil washing batch experiments were
conducted on 20 x 40 and 100 x 140 U.S. mesh soils and the clay fraction of the original soil
and on 20 x 40 U.S. mesh soil conditioned at 60°C. The in situ solvent washing (flushing) of
soil was simulated by continuously flushing solvent through a packed bed of soil until the PCP
120 1994 Symposium on Biaremediation of Hazardous Wastes
-------
concentration in the effluent did not decrease. The aboveground soil washing was simulated
by batch extraction tests conducted on PCP-contaminated soil.
The 50-percent ethanol solution, applied as the flushing solvent, consistently produced higher
PCP removal efficiencies at various aging periods from the 100 x 140 U.S. mesh soil than from
the 20 x 40 U.S. mesh soil. The higher PCP recovery from the 100 x 140 U.S. mesh soil was
due to the larger mass transfer area (contact surface) between the solvent and the soil that the
smaller soil particle size provided.
The data in Figure 1 show the results from the batch extraction tests performed on the 100 x
140 U.S. mesh soil. The results indicate that the 50-percent ethanol solution removed more
PCP from the soil than did either the 100-percent ethanol solution or deionized water. Similar
results were obtained for the other soil fractions. This higher recovery of PCP by the 50-percent
ethanol solution was consistent throughout the study. The results also show that PCP recoveries
decreased after 9 months of aging. The PCP removal efficiency for deionized water was lower
than that for the 100-percent ethanol solution after 6 months of aging, indicating that the
solubility of PCP in the hydrophobic solvent was contributing more to the removal of PCP from
the soil than was the superior wetting of soil by water.
In addition to the batch extraction tests with the various ethanol/water mixtures, sonication and
soxhlet extractions with methanol/methylene chloride were carried out on the same soil fractions.
The results shown in Figure 2 indicate that the PCP recoveries from the sonication and soxhlet
extractions of 100 x 140 U.S. mesh soil were not superior to those from the batch extraction
tests performed with 50-percent ethanol solution. Similar results were obtained for the other soil
fractions.
Biological Treatment Studies
Anaerobic, fluidized-bed GAC anaerobic bioreactors were used for the biological treatment of
chemically synthesized extracts (spent solvents) from the soil solvent washing process. The
synthesized spent solvent solution was fed to GAC bioreactors, where the PCP content of the
wash fluid was the biodegradable metabolite and ethanol served as the primary substrate.
The effect of empty bed contact time (EBCT) on the biodegradation of PCP and its degradation
products was examined using the GAC bioreactor (8). Throughout the experiments, the influent
PCP concentration was maintained constant at 100 mg/L by doubling the mass and hydraulic
loadings simultaneously. The EBCTs were based on an effective volume of 7 L (the total volume
of the reactor, 10 L, minus the volume due to a 30-percent carbon expansion) divided by the
total hydraulic flow rate (Table 1).
Effluent concentrations of PCP and its degradation byproducts are shown in Figure 3. Influent
and predicted effluent (with no biological activity) PCP concentrations are also shown. In molar
units, a relationship between influent PCP and the total monochlorophenol concentration in the
effluent indicates nearly complete conversion of the influent PCP to monochlorophenol. PCP
concentration was reduced by at least three orders of magnitude (a greater than 99-percent
transformation) throughout the study. No biological inhibition due to PCP was observed during
any phase, and the EBCT will be further decreased in future work.
121
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Influent chemical oxygen demand (COD) was contributed by PCP, ethanol, and trace salts. As
seen in Figure 4, there was a two-fold increase in the COD loading rate each time the mass
and hydraulic loading rates were doubled (see Table 1). Only 5 percent of the influent COD
persisted in the effluent COD throughout all phases of the study, while 70 percent was
accounted for by the methane produced. The remaining 25 percent of the influent COD was
attributed to biomass production.
Weekly analysis was also performed on the effluent chloride ion concentrations, volatile fatty
acids, and alcohols. The chloride potential is defined as the equimolar amount of chloride from
all potential sources (i.e., all chlorinated phenols in the feed). The delta chloride represents the
difference between the measured effluent chloride concentration and concentration of chloride
in the influent. These analyses confirmed that PCP underwent biological transformation to
monochlorophenols through the removal of four chlorine atoms per molecule of the phenol.
References
1. Cirelli, D. 1978. Patterns of pentachlorophenol usage in the United States of America.
An overview. In: Rao, K.R. Pentachlorophenol. New York, NY: Marcel Dekker, Inc.
pp. 13-18.
2. U.S. EPA. 1989. Superfund Record of Decision (EPA Region 6), United Creosoting
Co., Conroe, Montgomery County, TX (2nd remedial action), report. EPA/ROD/R06-
89/053.
3. U.S. EPA. 1990. Soil washing treatment. Engineering bulletin. EPA/540/2-90/017.
Cincinnati, OH.
4. Voice, T.C., and W.J. Weber, Jr. 1983. Sorption of hydrophobic compounds by
sediments, soils, and suspended solids, Vol. I. Theory and background. Water Res.
17:1,433.
5. Karickhoff, S.W., D.S. Brown, and T.A. Scott. 1979. Sorption of hydrophobic
pollutants on natural sediments. Water Res. 13:241.
6. Mueller, J.G., M.T. Suidan, and J.T. Pfeffer. 1988. Preliminary study of treatment of
contaminated groundwater from the Taylorville gasities site. RR077. Hazardous Waste
Research and Information Center.
7. Khodadoust, A.P., J.A. Wagner, M.T. Suidan, and S.I. Safferman. 1993. Treatment of
PCP-contaminated soils by washing with ethanol/water followed by anaerobic treatment.
In: U.S. EPA. Symposium on bioremediation of hazardous wastes: Research,
development, and field evaluations (abstracts). EPA/600/R-93/054. Washington, DC
(May).
8. Wagner, J.A., A.P. Khodadoust, M.T. Suidan, and R.C. Brenner. 1993. Treatment of
PCP-containing wastewater using anaerobic fluidized-bed GAC bioreactors. Paper No.
AC93-035-003. Proceedings of the Water Environment Federation 66th Annual
Conference and Exposition, pp. 189-200.
122
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Table 1. Operation Summary of Bioreactor
Phase Days of PCP
Operation (g/d)
I
II
m
480-606
607-824
825-999
0.60
1.20
2.40
Ethanol
(g/d)
4.28
8.33
16.66
Flow Rate
(L/d)
6.0
12.0
24.0
EBCT
(hr)
28.01
13.99
7.01
32
OH
-------
g-
I
0>
&
PH
1UU
90
80
70
60
50
40
30
20
10
n
!_
E- ^
~-
~—
"—
E-
E-
E-
—
1 1 1 1
o _JL-^""~"~~'''~^X\
•— ' ^"¥ * ^\,
' >
O
0 Soxhlet
• Sonication
T Soil Washing Batch Test with 50% Ethanol
100 mg PCP/Kg Soil (100 ppm)
i i I i
—
-E
-_
-_
-E
-E
-E
-_
—
0
6 9
Soil Age (Months)
12
15
Figure 2. Sonication and soxhlet extractions of 100 x 140 U.S. mesh soil.
iir
IxlO1
_l
"5
E ixio"
E
e"
% IxIO"'
2
S IxlO-'1
§
o
IxlOJ
<
• Phase I j Phase II Phase HI
: : Effluent PCP (no biological activity)
• • Effluent PCP (actual)
r : O Efnucnt MCPs (actual)
: • Efnucnl DCPs (actual)
! V EfnucntTCPs (actual)
. ; A Effluent Phenol (actual)
•' V * « ^
• Q'T^ Y? • • * ^7
- ^ : ^7V V ^ ^ * * ^ ^
*_,7V v —_ • « •• V^ ^ -fr 7 •$ V
• V *v« V* : S* * • * ^v 77 v "•? V^J* v^"7 vV
• ; • V 7«7 V« 7 •
500 600 700 800 900
Days
i
-
i
Figure 3. PCP and PCP intermediate effluent concentrations.
124
-------
Q
o
o
Influent
0 Gas + Effluent
0 Eflluent
480 500 520 540 560 580 600 620 640 660 680 700 720 740 760 780 800 820 840 860 880 900 920 940
Days
Figure 4. COD balance.
125
-------
Process Research
-------
Metabolic and Ecological Factors Affecting the Bioremediation of PAH- and
Creosote-Contaminated Soil and Water
P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Jian-Er Lin
Technical Resources, Inc., Gulf Breeze, FL
James G. Mueller and Suzanne Lantz
SBP Technologies, Inc., Gulf Breeze, FL
Introduction
Polycyclic aromatic hydrocarbons (PAHs) are a class of potentially hazardous chemicals whose
natural presence in the environment is attributable to a number of petrogenic and phytogenic
sources (1,2). Environments contaminated with large amounts of these chemicals (e.g., creosote
waste, coal tar processing sites) are considered hazardous owing to potential carcinogenic,
mutagenic, and teratogenic effects of specific PAHs (3). Generally, high molecular-weight
(HMW) PAHs, containing four or more fused rings, present the greatest potential hazard to both
the environment and human health (4). There is, consequently, much interest in developing
remedial methods, such as bioremediation, to selectively remove these chemicals from
contaminated environmental materials.
When environmental conditions (e.g., waste load, nutrients, oxygen, pH) are suitable,
biodegradation of low molecular-weight PAHs by indigenous microorganisms readily occurs (5-
7). Under the same conditions, however, biotransformation of HMW PAHs is less likely.
Although bacteria have been isolated in pure culture that grow on HMW PAHs, such as
fluoranthene and pyrene (7-9), strategies for stimulating this activity, as well as the degradation
of other HMW PAHs, in contaminated soils are not readily available. This is due in part to a
poor understanding of the biodegradation ecology of complex mixtures of hydrophobic
chemicals in the environment. How, for example, do microorganisms interact during a
degradation process to promote the degradation of these complex mixtures? Can this
interaction be enhanced through population management of microbial communities or
adjustment of specific environmental conditions? And, have microbial communities in
contaminated soils adapted (genetically and/or physiologically) to utilize hydrophobic PAHs more
effectively? An improvement of our understanding of biodegradation ecology for PAHs and
creosote could, therefore, lead to new and effective strategies for bioremediation of these
contaminants. This paper provides a summary of our research efforts in this area, with specific
attention given to co-metabolic processes, bioavailability, inoculation, and microbial community
adaptation.
1994 Symposium on Bioremediation of Hazardous Wastes 129
-------
Results and Discussion
Co-metabolism
The process of co-metabolism in bioremediation generally refers to the transformation (not
necessarily mineralization) of a hazardous waste chemical(s) as an indirect or fortuitous
consequence of the metabolism of another chemical that a bacterium uses as a source of
carbon and energy (growth substrate). Co-metabolism, an intriguing consequence of broad
enzyme specificity, is one of the important elements in the recent emergence of new
bioremediation strategies. Unfortunately, however, its occurrence in natural microbial
communities is neither well documented nor understood, and the process is difficult to control
in the field. In addition, there are concerns regarding the fate and environmental impact of the
partial oxidation products that are thought to be produced. Successful degradation of HMW
PAHs has been argued to involve extensive co-metabolic reactions (6); that is, enzymes used by
specific bacteria in a microbial community to degrade one type of PAH fortuitously oxidize other
PAHs. Biochemical evidence for this type of reaction is provided in the paper by Chapman
etal.
The importance of co-metabolism in PAH degradation is illustrated by studies in which a
bacterium (Sphingomonas paucimoblis strain EPA505) that used fluoranthene (an HMW PAH
containing four fused rings and a major constituent of most creosote and coal tar wastes) as a
sole source of carbon and energy was found to biotransform many PAHs that were not growth
substrates (10). This included fluorene, pyrene, chrysene, and benzo(a)pyrene. If this bacterium
and other PAH degraders are exposed to the PAH fraction of creosote in a standard shake flask
assay (8) for 10 days and the creosote fraction is monitored by extraction and gas
chromatographic analysis, considerable loss of most of the PAHs occurs even though only a few
of the PAHs are used as growth substrates. A comparison of results from strain EPA505 and
strain N2P5, a bacterium also isolated from creosote-contaminated soil, is shown in Table 1.
Strain N2P5 grew only on two- and three-ring PAHs, such as phenanthrene, and had far less
capacity for this co-metabolic phenotype. A variety of isolates are currently being studied to
more fully characterize this co-metabolic capability. The resulting partially oxidized degradation
products from this co-metabolism have not been specifically identified but are likely to be more
soluble and possibly more biodegradable than the parent compound, perhaps leading to further
degradation or metabolism by other members of a microbial community.
Other bacteria in nature may behave like these PAH degraders studied in the laboratory, thereby
giving microbial communities the capability of co-metabolism. Few experimental results are
available, however, to show that this is indeed the case. We are conducting experiments to
specifically relate pure culture studies to PAH degradation patterns in natural microbial
communities. At a bioremediation site, where environmental conditions are established to
promote PAH degradation by the indigenous microflora (aeration, inorganic nutrient
amendment, moisture control, etc.), however, co-metabolism may not have its maximum effect
because the PAHs serving as inducers of the enzymatic processes responsible for co-metabolism
are not maintained at sufficient concentrations. As a consequence, it may be reasonable to add
a specific PAH in low concentrations to stimulate microbial communities to co-metabolically
degrade HMW PAHs, thereby more easily bringing PAH concentrations to stipulated cleanup
levels. Clearly, for any long-term bioremediation treatment involving co-metabolism, more
ecological and biochemical research is required.
130
-------
BioavailabilHy
Because of their strongly hydrophobic nature, HMW PAHs usually occur as contaminants in
natural ecosystems and waste treatment systems at mass levels that exceed their water solubility.
In addition, equilibria strongly favor particle-bound chemicals (e.g., sorbed to soils). These
characteristics largely account for the slow biodegradation of HMW PAHs (11). Therefore,
understanding treatment conditions and environmental factors that can be manipulated to
enhance bioavailability and consequently biodegradation is critical to bioremediation
considerations.
It has been suggested that pure cultures of bacteria can use PAH compounds only in the
dissolved state (12-14). Therefore, the dissolution of PAHs may be a prerequisite for initial
oxidation and mineralization. Dissolution rates are usually determined by the solid-liquid contact
surface area and the equilibrium concentration of the PAH compound (11,12,15). Surfactants
can enhance PAH solubilization and dissolution, thus increasing the equilibrium concentration
of the compound in the aqueous phase (16). This should lead to faster degradation rates. It
has been observed, however, that use of surfactants at high concentrations reduced or inhibited
biodegradation (17,18) because of surfactant toxicity to the bacteria used in the study.
On the contrary, Teichm has shown that a variety of nonionic surfactants are nontoxic to a
A/l/cobacter/um sp. that is able to grow on fluoranthene and pyrene and consequently increase
rates of PAH biodegradation (19). Likewise, we have studied the mineralization of 14C-
radiolabeled fluoranthene byS. pauc/mob//s strain EPA505, an organism that grows on this PAH
as a carbon and energy source, and initial rates of mineralization were enhanced by the
presence of the surfactant Triton X-100. An example of this response is shown in Figure 1 (top).
For this experiment, cells were grown in complex medium, washed several times in buffer, and
suspended to a final cell density of 8 x 1010 cells/ml in minimal salts medium containing 20 mg
of unlabeled fluoranthene, approximately 60,000 dpm of HC-fluoranthene, and various
concentrations of Triton X-100. The surfactant concentrations tested were all above the critical
micelle concentration for this surfactant. Initial rates of mineralization were clearly enhanced
by all concentrations of the surfactant. The reduced extent of mineralization at the two highest
surfactant concentrations may have resulted from the sequestering of fluoranthene degradation
intermediates (e.g., leaching from the cells), making them unavailable for mineralization. The
bacterium was clearly able to tolerate high surfactant concentrations, thus emphasizing the
importance of properly selecting PAH-degrading microorganisms that are not inhibited by
surfactants or selecting surfactants that are nontoxic.
Dissolution of a chemical is also a problem in soil slurry systems, where the presence of soil
particles may decrease the aqueous concentration of a PAH compound due to the sorption
effects (20); reduced aqueous concentrations would decrease the rate of biodegradation. The
presence of soil particles in a solution, however, can provide a higher solid-liquid contact
surface area, thereby enhancing the solid-liquid mass transfer. Slow degradation rates, in
essence, are counterbalanced by greater chemical turnover. This is, in fact, true in the case of
fluoranthene degradation. As shown in Figure 1 (bottom), an aqueous suspension of soil
particles (30 mg/mL) and fluoranthene crystals (20 mg) together resulted in greater
mineralization rates than suspensions with only fluoranthene crystals. This would appear to be
the effect of higher solid-liquid contact. Although increases in biomass or in the activity of the
biomass as a result of exposure to soil particles may also explain the effect, this is unlikely since
the biomass (108 cells/mL) and the mineralization rates were initially high. Note that the effect
131
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of soil particles was equivalent to that of adding surfactant, a further indication of increased
dissolution by either material.
In contaminated soils, fluoranthene and other HMW PAHs will likely exist at concentrations far
in excess of their aqueous solubility. Given that undissolved PAHs will not exist as crystals in the
environment, it is important to know if they exist in a form in which soil particles provide higher
solid-liquid contact or in which surfactants can promote greater dissolution, or both.
Research needs to be accelerated in this area because the use of surfactants will almost
assuredly play a significant role in future bioremediation procedures. Also, engineering
strategies for using surfactants or other means of increasing mass transport in the field must be
developed. This should include consideration of how to remove the bioavailability-enhancing
chemical from the field after it has done its job, and how to protect against a negative effect on
contaminant distribution in the field (e.g., seepage into uncontaminated areas).
Bioaug mentation
If we define bioaugmentation as the process of introducing microorganisms of sufficient biomass
into a site in a manner in which it can be documented that the inoculated organism(s) survives
to a point of significantly affecting the fate of a target chemical(s), then very few scientifically
documented examples exist where this process has been successful on a significant scale. Yet,
there are many possible situations in which bioaugmentation of chemically contaminated sites
with microorganisms possessing unique and specialized metabolic capabilities could potentially
be a feasible bioremediation approach. With more careful attention to selection and application
of the inoculants, it is quite reasonable that bioaugmentation could become a major and
effective component of biological cleanup methods.
There are many recognizable limitations to the use of bioaugmentation in bioremediation. Only
a few have been systematically addressed in an experimental sense (20-23). These include the
inability to support the growth and/or activity of the introduced organism because of competition
by the indigenous microflora. Success, however, can be realized by employing specialized
techniques to reduce competition and to maintain a biomass high enough to effect efficient
degradation of the target chemicals. In addition, the contaminated environment will almost
certainly have to be physically modified, perhaps over an extended period, to optimize the
bioaugmentation process. This generally means establishing conditions in which the availability
of oxygen, inorganic nutrients, temperature, degradable substrate, moisture content, etc., are
optimized.
Bioaugmentation using microorganisms with requisite metabolic capacities is one suggested
approach for enhancing biodegradation of these HMW PAHs (6). Although biodegradation of
HMW PAHs by identified microorganisms has been reported, suitable strategies for utilizing these
microorganisms as inocula in the field need to be further developed. We have been
experimenting with the concept of introducing immobilized cells using different encapsulation
procedures (24). For example, polyurethane polymer (PU) has been used to immobilize S.
paudmobilis strain EPA505. The immobilized cells were tested for their ability to mineralize
fluoranthene under these conditions. As shown in Figure 2 (top), there was no significant
difference in fluoranthene mineralization profiles by the PU-immobilized cells of strain EPA505
when compared to nonimmobilized cells. Since the same inoculation size was used in all flasks
during this experiment, the results suggest that the immobilization process does not significantly
132
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affect microbial activity. Cells immobilized in the PU polymer remain active for months when
stored at 4°C.
Active immobilized cells then offer several additional possibilities for further enhancing
biodegradation and environmental control. For example, inclusion of adsorbents in the
immobilization matrix can result in a more rapid uptake of toxic compounds from the
environment, thereby potentially providing greater accessibility of the adsorbed chemical to the
immobilized bacteria. Two issues need to be addressed, however, when using co-immobilized
adsorbents: 1) Is microbial activity affected by co-immobilization with adsorbents? and 2) Is
availability of the adsorbed chemical to the immobilized cells maximal? To study these
questions, diatomaceous earth and powdered activated carbon were co-immobilized with strain
EPA505 in the polyurethane matrix. In Figure 2 (top), it can be seen that the degrading activity
of the cells co-immobilized with the adsorbents was the same as the nonimmobilized cells,
indicating that the degradation of the adsorbed fluoranthene was complete.
Another possibility involves in situ bio re mediation situations, where direct addition of nitrogen
and phosphorous into soil or water may have a negative consequence due to either
enhancement of the activity of undesired indigenous microflora and/or the leaching of the
nutrients into ground water. By co-immobilizing slow-release formulations of nutrients in the
polymer matrix, a major part of the nutrients can be provided to the immobilized cells with
considerably less available for leaching into the environment. In our experiments, slow-release
formulations of nitrogen and phosphorus were co-immobilized with EPA505 in the polyurethane
matrix, then tested in buffer for fluoranthene degradation. As a positive control, the immobilized
cells with external sources of nitrogen and phosphorus (solution of inorganic salts) were also
used. As can be seen in Figure 2 (bottom), co-immobilized nitrogen and phosphorous
supported extensive biodegradation, although the biodegradation rate was slower than with
externally supplied nutrients. Further studies on the effect of release rates of the co-immobilized
nutrients may provide more information for optimizing this approach to bioaugmentation.
Adaptation
We have been characterizing a variety of different fluoranthene-degrading bacteria from around
the world and have shown that the degradation capacity for this PAH is common and distributed
among a variety of bacteria. To date, however, only soils and sediments polluted with PAHs and
creosote have produced fluoranthene degraders. Phenanthrene degraders can be readily
isolated from any type of soil. Thus, the ability to utilize and grow on HMW PAHs represents
an ability to deal with very low available substrate concentrations. What is the source of these
fluoranthene degraders? Are they present in many different environments but only enriched to
the point of detection in polluted soils? Or is there actually gene recruitment occurring in
natural microbial communities that in essence "creates" this metabolic capability? A clearer
understanding of the origins of these organisms has significant implications in bioremediation,
for it may be possible to ultimately adjust environmental or ecological conditions in the field to
accelerate this adaptation process and therefore more readily affect the outcome of a
bioremediation treatment for PAHs.
To this end, we have been characterizing the genetics and physiology of our isolates. As has
been documented for many other catabolic functions for xenobiotic chemicals in bacteria, one
of these organisms harbors the fluoranthene degradative genes on a plasmid. Dr. Tom Lessie,
in our laboratory, has shown that it also contains three mega-plasmids or multiple replicons.
These are quite large, 3,400, 2,300, and 1,200 kilobases in size. The presence of these mega-
133
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plasmids has been reported for other species of Pseudomonas (25), as well as other genera of
bacteria. The physiological and genetic functions of these mega-plasmids are unknown, but
they may be related to the large and broad metabolic capability that these organisms possess
and perhaps even to the ability to degrade fluoranthene. By understanding more about this
genetic makeup, it may eventually be possible to manipulate adaptation in the field in a time
frame that could accelerate or increase the extent of bioremediation.
Summary and Conclusions
The successful bioremediation of PAH-contaminated soils and sediments requires a clear
understanding of the metabolic and ecological factors that can be manipulated to increase the
rate and extent of PAH biodegradation. We provide evidence in this report suggesting that: 1)
co-metabolism may be a potential mechanism for degradation of HMW PAHs, 2) bioava(lability
of PAHs may be improved through the application of surfactants, and 3) the success of
bioaugmentation may be increased by the use of procedures that immobilize PAH-degrading
microorganisms, adsorbents, and/or nutrients. In addition, the knowledge of how microbial
communities become adapted for enhanced PAH biodegradation may play an important role
in developing future strategies for bioremediation.
References
1. Grosser, R.J., D. Warshawsky, and J.R. Vestal. 1991. Indigenous and enhanced
mineralization of pyrene, benzo[a]pyrene, and carbazole in soils. Appl. Environ.
Microbiol. 57:3,462-3,469.
2. National Academy of Science. 1983. Polycyclic aromatic hydrocarbons: Evaluation of
sources and effects. Washington, DC: National Academy Press.
3. Moore, M.N., D.R. Livingstone, and J. Widdows. 1989. Hydrocarbons in marine
mollusks: Biological effects and ecological consequences. In: Varanasi, U., ed.
Metabolism of PAHs in the aquatic environment. Boca Raton, FL: CRC Press, Inc. pp.
291-328.
4. U.S. EPA. 1982. Wood preservative pesticides: Creosote, pentachlorophenol, and the
inorganic arsenical (wood uses). Position Document 213. EPA 540/9-82/004.
Washington, DC.
5. Mueller, J.G., S.E. Lantz, B.O. Blattmann, and P.J. Chapman. 1991. Bench-scale
evaluation of alternative biological treatment processes for the remediation of
pentachlorophenol- and creosote-contaminated materials: Slurry-phase bioremediation.
Environ. Sci. Technol. 25:1,055-1,061.
6. Mueller, J.G., S.E. Lantz, R.J. Colvin, D. Ross, D.P. Middaugh, and P.H. Pritchard.
1993. Strategy using bioreactors and specially selected microorganisms for
bioremediation of ground water contaminated with creosote and pentachlorophenol.
Environ. Sci. Technol. 27:691-698.
134
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7. Cemiglia, C.E. 1993. Biodegradotion of polycyclic aromatic hydrocarbons.
Biodegradation 3:351-368.
8. 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. 55:3,085-3,090.
9. Weissenfels, W.D., M. Beyer, and J. Klein. 1990. Degradation of phenanthrene,
fluorene, and fluoranthrene by pure bacterial cultures. Appl. Microbiol. Biotechnol.
34:528-535.
10. Mueller, J.G., P.J. Chapman, E.O. Blattmann, and P.H. Pritchard. 1990. Isolation and
characterization of a fluoranthene-utilizing strain of Pseuc/omonas paucimoJbf/is. Appl.
Environ. Microbiol. 56:1,079-1,086.
11. Volkerling, F., A.M. Breure, A. Sterkenburg, and J.G. van Andel. 1992. Microbial
degradation of polycyclic aromatic hydrocarbons: Effect of substrate availability on
bacterial growth kinetics. Appl. Microbiol. Biotechnol. 36:548-552.
12. Stucki, G., and M. Alexander. 1987. Role of dissolution rate and solubility in
biodegradation of aromatic compounds. Appl. Environ. Microbiol. 53:292-297.
13. Wodzinski, R.S., and D. Bertolini. 1972. Physical state in which naphthalene and
bibenzyl are utilized by bacteria. Appl. Microbiol. 23:1,077-1,081.
14. Wodzinski, R.S., and J.E. Coyle. 1974. Physical state of phenanthrene for utilization
by bacteria. Appl. Microbiol. 27:1,081-1,084.
15. Thomas, J.M., J.R. Yordy, J.A. Amador, and M. Alexander. 1986. Rates of dissolution
and biodegradation of water-insoluble organic compounds. Appl. Environ. Microbiol.
52:290-296.
16. Edwards, D.A., R.G. Luthy, and Z. Liu. 1991. Solubilization of polycyclic aromatic
hydrocarbons in micellar nonionic surfactant solutions. Environ. Sci. Technol. 25:127-
133.
17. Aronstein, B.N., Y.M. Calvillo, and M. Alexander. 1991. Effect of surfactants at low
concentrations on the desorption and biodegradation of sorbed aromatic compounds
in soil. Environ. Sci. Technol. 25:1,728-1,731.
18. Laha, S., and R.G. Luthy. 1992. Effects of nonionic surfactants on the solubilization
and mineralization of phenanthrene in soil-water systems. Biotechnol. Bioeng.
40:1,367-1,380.
1 9. Tiehm, A. 1994. Degradation of polycyclic aromatic hydrocarbons in the presence of
synthetic surfactants. Appl. Environ. Microbiol. 60:258-263.
20. Guerin, W.F., and S.A. Boyd. 1992. Differential bioavailability of soil-sorbed
naphthalene to two bacterial species. Appl. Environ. Microbial. 58:1,142-1,152.
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21.
22.
23.
24.
25.
Pritchard, P.M. 1992. Use of inoculation in bioremediation. Curr. Opin. Biotechnol.
3:232-243.
Goldstein R.M., L.M. Mallory, and M. Alexander. 1985. Reasons for possible failure
of inoculation to enhance biodegradation. Appl. Environ. Microbiol. 50:977-983.
Comeau, Y., C.W. Greer, and R. Samson. 1993. Role of inoculum preparation and
density on the bioremediation of 2,4-D-contaminated soil by bioaugmentation. Appl.
Microbiol. Biotechnol. 38:681-687.
Lin, J.E., J.G. Mueller, K.J. Peperstaete, and P.M. Pritchard. 1993. Identification of
encapsulation and immobilization techniques for production, storage, and application
of PAH-degrading microorganisms. In: U.S. Naval Research Laboratory report.
Contract No. N00014-90-C-2136 through Geo-Centers, Inc., Newton Upper Falls,
MA.
Hai-Ping, C, and T.G. Lessie. 1994. Multiple replicons comprising the genome
Pseudomonas cepac/a 17616. J. Bacteria. In press.
of
Table 1. Degradation of Creosote PAHs by Selected Bacterial Isolates
compound (mg/L)
naphthalene
thlanaphthene
2-methylnaphthalene
1 -methylnaphthalene
blphenyl
2,6-dimethylnaphthalene
2,3-dlmethy (naphthalene
acenaphthylene
acenaphthene
dlbenzofuran
fluorene
dibenzothlophene
phenanthrene
anthracene
carbazole
2-methylanthracene
anlhraquinone
fluoranlhene
pyrene
benzo(b)(luorene
benz(a)anthracene
chrysene
benzo(b/k)fluoranthene
benzo(a)pyrene
total
unlnoculated (sd)
39.33 (2.47)
1.41 (0.08)
18.88 (0.79)
6.23 (0.18)
3.30 (0.16)
2.85 (0.19)
0.67 (0.04)
0.55 (0.03)
22.46 (1.20)
16.01 (0.84)
19.83 (1.22)
6.85 (0.58)
55.22 (3.00)
2.80 (0.16)
2.94 (0.28)
1.02 (0.55)
5.07 (0.76)
26-53 (2.31)
15.92 (1.40)
2.85 (0.20)
5.94 (2.49)
2.42 (1.12)
1.64 (0.20)
0.60 (0.02)
261.32
EPA505 (sd)
0.04 (0.01)
0.11 (0.03)
0.07 (0.02)
0.04 (0.01)
bdl
0.10 (0.03)
0.06 (0.03)
0.21 (0.07)
bdl
0.12 (0.02)
0.15 (0.05)
0.28 (0.15)
bdl
0.48 (0.10)
0.35 (0.12)
0.21 (0.11)
1.12 (0.21)
bdl
8.39 (0.75)
0.76 (0.08)
5.98 (1.00)
1.77 (0.32)
1.19 (0.21)
0.49 (0.06)
21.92
%reduction
100
92
100
99
100
96
91
62
100
99
99
96
100
83
88
79
78
100
47
73
0
27
27
18
N2P5 (sd)
0.10 (0.06)
0.79 (0.17)
0.17 (0.03)
0.79 (0.12)
0.62 (0.07)
0.50 (0.05)
0.37 (0.09)
0.48 (0.17)
10.06 (1.30)
0.08 (0.01)
0.11 (0.06)
7.42 (0.56)
0.14 (0.01)
1.09 (0.05)
0.43 (0.11)
1.58 (0.06)
4.43 (0.57)
28.46 (4.09)
16.01 (5.90)
2.66 (0.11)
6.02 (0.09)
2.81 (0.09)
2.34 (0.16)
0.94 (0.11)
88.46
%reduction
100
44
99
87
81
82
45
13
55
100
99
0
100
61
85
0
13
0
0
7
0
0
0
0
136
-------
•o
c
9)
c
-------
GO
u
T)
N
15
0)
1
0)
d
IT)
U
O
^
El!
40
30
20
10
O Free cell
• I'll pellet
V I'U pellet with dialo-
moccous enrlh
T PU pnllcl willi octivolccl
carbon
0 *•-•-<
0 100 200 300 400 500
Time (hrs)
GO
50
40
30
20
10
0*M
/
O with external N-I-P source
• with encapsulated slowly
released N-I-P source
i
200
400 GOO
Time (hrs)
000
1000 1200
Figure 2. Mineralization profiles of 14C-fluoranthene by nonimmobilized and polyurethane-
immobilized cells of strain EPA505 with different adsorbents (top) and external and
encapsulated nitrogen and phosphorus sources (bottom) in minimal salts medium.
Inoculum concentration = 8 x 10* cells/ml; fluoranthene concentration = 0.4
mg/mL; Triton X-100 concentration = 0.1 percent; particle concentration = 30
mg/mL
138
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Metabolic Pathways Involved in the Biodegradation of PAHs
Peter J. Chapman
U.S. Environmental Protection Agency, Gulf Breeze, FL
Sergey A. Selifonov
University of Minnesota, St. Paul, MN
Richard Eaton
U.S. Environmental Protection Agency, Gulf Breeze, FL
Magda Grifoll
University of Barcelona, Barcelona, Spain
Introduction
The principal sources of polycyclic aromatic hydrocarbons (PAHs) in the environment are the use
and spillage of fossil fuel-related materials, either petroleum- or coal-derived. Both sources
contain complex mixtures of PAHs but differ in amount and composition. Coal-based materials
such as creosote and coal tar are rich in PAHs, with relatively little alkyl substitution. Petroleum,
on the other hand, generally contains a smaller fraction of PAHs composed of a wide array of
alkyl-substituted homologues. Knowledge of the aerobic biodegradation of PAHs derives largely
from studies of pure bacterial cultures isolated for their ability to utilize for growth single,
unsubstituted aromatic hydrocarbons such as naphthalene, biphenyl, and phenanthrene (1). In
all cases studied, catabolism is initiated by oxygen-adding reactions usually forming cis-
dihydrodiols on arene rings. While biological methods for removal of PAH-containing
environmental contaminants are now seriously considered options for remediation, details of the
processes involved are little understood. For example, little is known of the extent to which
biotransformation (co-metabolism) is involved in the removal of higher molecular weight PAHs
in complex mixtures and the organisms and growth substrates required. Are products of
biotransformation accumulated? What are their environmental effects?
Some recent findings relevant to these questions are summarized below.
Naphthalene Degradation: New Insights
Investigation of reactions of naphthalene degradation catalyzed by enzymes encoded by the
NAH7 plasmid was undertaken using a molecular biological approach involving cloning and
subcloning of pathway genes (2). As a result, a collection of strains of Pseudomonas aerug/nosa
was obtained containing key genetic sequences of the plasmid encoding for the degradative
pathway extending various distances from naphthalene. Such strains were used to accumulate,
under physiological conditions, catabolites of naphthalene otherwise difficult to isolate and
characterize. As a result, /rans-2-hydroxy benzylidene pyruvate was identified as a metabolite
of 2-hydroxy chromene-2-carboxylic acid and a new reaction was recognized as responsible for
formation of salicylaldehyde and pyruvate by means of a novel hydratase-aldolase enzyme.
1994 Symposium on Bioremediation of Hazardous Wastes 139
-------
Degradation of Creosote PAHs
For studies of the bacterial degradation of creosote PAHs, an aromatic hydrocarbon fraction free
of polars, resins, and phenols, with little if any N-heterocyclic material, was obtained by column
chromatography. Enrichments employed this fraction in mineral salts medium to establish
cultures (from creosote-contaminated soils). These were incubated with shaking at 20°C to
24°C in the dark, with transfers biweekly. Amounts of remaining PAHs, determined by GC-FID
after methylene chloride extraction, showed extensive losses of low molecular weight PAHs not
accounted for by abiotic losses. Fluoranthene, pyrene, and PAHs with higher retention times
were recovered essentially unchanged, being associated with insoluble black resinous material
accumulated in cultures. Column chromatography and thin-layer chromatography has shown
this material contains both low molecular weight neutral products and complex polymeric
material. Among the neutral products identified were acenaphthenone, fluorenone, and other
ketones formed from naphtheno-aromatics. Certain of these have previously been shown to
result from the action of bacterial reductive dioxygenases (3).
Naphthalene Dioxygenase Action on Naphtheno-Aromatic Hydrocarbons
With the cloned genes of naphthalene dioxygenase available in a strain of P. aeruginosa (2),
it was possible to investigate the action of a reductive oxygenase on simple naphtheno-aromatic
hydrocarbons and related compounds (4). Induced cells were incubated in buffer with fluorene,
acenaphthene, acenaphthylene, and other hydrocarbons having benzylic functions; products
were extracted for characterization. Fluorenone was identified as a product of fluorene
oxidation, with acenaphthenone formed from acenaphthene and acenaphthylene together with
a c/s-dihydrodiol and acenaphthenequinone in the latter case (Figure 1).
Evidently the first formed secondary alcohols are acted on by broad-specificity cellular
dehydrogenases to give ketonic end products. Apparently anomalous oxidations at benzylic
positions, such as observed here, may be expected in situations where biodegradation of
mixtures of aromatic and naphtheno-aromatic hydrocarbons occurs.
Bacterial Utilization of a Naphtheno-Aromatic: Fluorene
Given that oxidation of benzylic functional groups may be unavoidable when arene dioxygenases
are confronted by naphtheno-aromatics, it was of interest to examine whether such reactions are
involved when bacteria utilize naphtheno-aromatics as growth substrates. Accordingly, the
reactions employed in the utilization of fluorene by a Pseudomonas isolate were investigated.
An earlier study with a different strain (5) suggested that the productive route of catabolism
involved initial aromatic-ring dioxygenation and cleavage and that fluorenone was a dead-end
metabolite. By contrast, the pathway established for the Pseudomonas isolate is initiated by
benzylic oxidation leading to fluorenone formation. Subsequent reactions include formation of
a novel angular diol (6) before opening the central five-membered ring to generate a
dihydroxylated biphenyl carboxylic acid (Figure 2). It would appear that this route (7,8)
represents a significant difference from earlier characterized routes initiated by conversion of
arenes to c/s-dihydrodiols, in that the naphthenic ring is first oxidized and then opened, thereby
accommodating both fluorene and fluorenone.
140
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Organisms possessing this biochemistry are therefore equipped to channel products of
anomalous oxidation by arene dioxygenases into productive catabolic pathways.
References
1. Gibson, D.T., and V. Subramanian. 1984. Microbial degradation of aromatic
hydrocarbons. In: Gibson, D.T., ed. Microbial degradation of organic compounds.
New York, NY, and Basel, Switzerland: Marcel-Dekker, Inc. pp. 181 -250.
2. Eaton, R.W., and P.J. Chapman. 1992. Bacterial metabolism of naphthalene:
Construction and use of recombinant bacteria to study ring cleavage of 1,2-dihydroxy-
naphthalene and subsequent reactions. J. Bacterial. 174:7,542-7,554.
3. Schocken, M.J., and D.T. Gibson. 1984. Bacterial oxidation of the polycyclic aromatic
hydrocarbons, acenaphthene and acenaphthylene. Appl. Environ. Microbiol. 48:10-16.
4. Selifonov, S., M. Grifoll, R.W. Eaton, and PJ. Chapman. 1993. Oxidation of the
naphtheno-aromatic compounds, acenaphthene, acenaphthylene, and fluorene, by
naphthalene oxygenase cloned from plasmid NAH7. Abstr. #Q345. 93rd Annual ASM
Meeting, Atlanta, GA.
5. Grifoll, M., A.M. Solanas, and J.M. Bayona. 1990. Isolation and characterization of
a fluorene-degrading bacterium: Identification of ring oxidation and ring fission
products. Appl. Environ. Microbiol. 58:2,910-2,917.
6. Selifonov, S.A., M. Grifoll, J.E. Gurst, and P.J. Chapman. 1993. Isolation and
characterization of (+)-!,12-dihydroxy-l-hydrofluorenone formed by angular
dioxygenation in the bacterial catabolism of fluorene. Biochem. Biophys. Res. Commun.
193:67-76.
7. Trenz, S.P., K.H. Engesser, P. Fischer, and H-J. Knackmuss. 1994. Degradation of
fluorene by Brev/bacferium sp. strain DPO 1361: A novel c-c bond cleavage
mechanism via 1,10-dihydro-l ,10-dihydroxyfluoren-9-one. J. Bacterial. 1 76:789-795.
8. Grifoll, M., S.A. Selifonov, and PJ. Chapman. 1994. Degradation of fluorene by
Pseudomonas sp. F274: Evidence for a novel degradative pathway. Appl. Environ.
Microbiol. In press.
14)
-------
OH
.0
OH O
Figure 1. Transformation of naphtheno-aromatics by naphthalene dioxygenase.
Phthalic acid
Figure 2. Route of fluorene degradation in Pseudomonas F274.
142
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Environmental Factors Affecting Creosote Degradation by
Sphingomonos paucimobilis Strain EPA505
J.G. Mueller and S.E. Lantz
SBP Technologies, Inc., Gulf Breeze, PL
P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Introduction
The presence of polycyclic aromatic hydrocarbons (PAHs) in soil and ground water is recognized
by EPA as a priority environmental problem. Due to inadequacies intrinsic to the design and
operation of wood treatment facilities of the past, coal tar creosote represents one of the major
anthropogenic sources of excessive PAH concentrations in the environment (1). Coal tar
residues from coal gasification and creosote distillation processes represent another major
source of environmental PAH contamination.
Of the hundreds of locations potentially affected by PAHs from industrial operations, most have
been thoroughly assessed and characterized. In cases where remedial actions to restore soil
and ground water have been prescribed, a variety of treatment alternatives have been evaluated.
Unfortunately, many of the more conventional approaches have proven ineffective and/or
prohibitively expensive. For example, ground-water pump-and-treat approaches have proven
ineffective for PAH-contaminated aquifers (EPA Office of Solid Waste and Emergency Response
[OSWER] memorandum, May 27,1992). For soils, excavation followed by secondary treatment
(e.g., soil washing followed by slurry-phase biotreatment) is of such a scale that costs and
practicability have become prohibitive. In addition, from an end-user's perspective, many
conventional remedial technologies are unacceptable due to regulatory problems and technical
feasibility.
Of the alternative remedial options available for creosote-contaminated soil, bioremediation
may represent a technology of choice. Despite the many potential advantages of
bioremediation, the reported effectiveness of PAH biodegradation in contaminated media has
varied (2). This variability is due to a number of recognized factors, including the presence of
free product as dense nonaqueous phase liquid (DNAPL) and/or light nonaqueous phase liquid
(LNAPL), the heterogenous nature of soil and subsurface matrices, and the use of ineffective
delivery and implementation strategies. From a biological perspective, effective biodegradation
is influenced, in part, by the presence of catabolically competent microflora at a contaminated
site and by certain environmental factors that enhance the activity of this microflora, including
availability and concentration of electron acceptors, inorganic nutrients, and the target
chemical(s). The ability to control and regulate these factors is the foundation for
bioremediation application to PAH/creosote-contaminated soils.
In an effort to enhance the biodegradation of PAHs in the environment, we have recently
focused on several environmental and toxicological factors influencing the ability of
Sphingomonas (Pseudomonas) paua'mobil/s strain EPA505 to mineralize PAHs individually and
in complex mixtures (e.g., creosote). We believe that more effective management of natural
1994 Symposium on Bioreroediotion of Hazardous Wastes 143
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microbial community activities, through control of these factors, may lead to more efficient
bioremediation of soil and water contaminated by PAHs. Additionally, these studies should help
inoculant microorganisms be employed more effectively for site restoration.
Materials and Methods
Evaluation of Temperature and pH Effects
Biometer flasks (3) containing minimal salts medium, radiolabeled fluoranthene or
phenanthrene, and cells of strain EPA505 were used to monitor MCO2 evolution over a range
of pH and temperature. A mixture of unlabeled (10 mg PAH) and 14C-labeled PAH
(approximately 41,000 dpm) was added to 250-mL biometer flasks from acetone stock
solutions, and the solvent was evaporated. To each flask was added 50 ml of Bushnell-Haas,
and the contents were sonicated. The pH of the medium was adjusted with HCI or NaOH. The
buffering capacity of Bushnell-Haas was such that the pH was stabilized over the course of 2 d
at the target pH. For temperature studies, the medium was adjusted to pH 7.1, and flasks were
equilibrated at various temperatures for about an hour prior to inoculation. All flasks were
maintained at a selected temperature over the course of the studies.
To initiate studies, 1.0 ml of 2N NaOH was added to each sidearm of the biometer flasks to
trap 14CO2. The inoculum was prepared from a cell concentrate (48-hr growth on complex
medium LB, harvested, washed, and resuspended in 0.05 M phosphate buffer) and added to
obtain an initial optical density of 0.5 at 600 nm (about 3 to 5 x 108 cells/ml). Flasks were run
in duplicate, and killed-cell controls were also used. Flasks were shaken at 120 rpm at 30°C
in darkness for up to 8 days. NaOH samples were collected intermittently and analyzed by
liquid scintillation the same day.
Identification of Inhibitory Creosote Constituents
Biometer flasks were again used to monitor 14CO2 evolution from 14C-PAH in the presence of
various concentrations of creosote and its acid-, neutral-, and base-extractable fractions to study
the effect of phenols, PAHs, and neutrally extractable heterocycles (carbazole, dibenzothiophene,
dibenzofuran, and thianaphthene) and other N-, S-, and O-containing heterocycles, respectively
(4,5). Synthetic mixtures of each of these fractions were prepared as defined in Table 1 to more
accurately evaluate the effect of each of these mixtures (6). An "artificially weathered" (heating
the neutral fraction at 65°C ± 5°C for 24 hr), creosote-neutral fraction was also analyzed to
examine the effect of low molecular-weight PAHs (i.e., those containing two fused rings). A
killed-cell control was run for each different substrate, and a positive mineralization control (no
creosote) was run with each set of incubations.
The incubation medium was prepared as described above. Bushnell-Haas, however, was
supplemented with 0.03-percent Triton X-l 00 to facilitate study of constituents at concentrations
above their natural water solubilities. For consistency, Triton X-l 00 was added to each flask.
The appropriate amount of creosote, or some fraction thereof, was added via glass gas-tight
syringe.
Flasks were shaken 120 rpm at 30°C in darkness for 10 d. NaOH samples were collected daily
and analyzed by liquid scintillation the same day. At the conclusion of these studies, flasks
144
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exhibiting inhibition were cultured for the determination of viable cells. The remaining contents
of each flask were subsequently extracted and analyzed for the concentration of creosote
constituents by gas chromatography/flame ionization detector (GC-FID) (4,5).
Results and Discussion
Average (n=2) percent release of 14CO2 from 14C-fluoranthene by strain EPA505 was essentially
identical for pH values of 6, 7, 8, and 9 (Figure 1). In these flasks, postincubation pH was
lowered by 0.5 to 1 pH unit. The pH-5 flasks quickly reached a plateau, after which
mineralization ceased. This was not characteristic of any of the other pH treatments. The
postinoculation pH of this flask was 4.6. Absence of extensive mineralization in the pH-4 and
pH-10 flasks correlated with the absence of the characteristic color change (colored degradation
intermediates) normally associated with fluoranthene mineralization by this bacterium (1,7,8).
Strain EPA505 was active at all temperature ranges tested to date (Figure 2), although rates and
extents of mineralization decreased with decreasing temperature. At the 25°C incubation
temperature, mineralization extent was reduced compared with 30°C and 37°C but might
eventually reach that seen with the higher temperatures given incubation times beyond 200 hr.
At 18°C, mineralization rates appeared to be leveling off at values below those seen at higher
temperatures, and it does not appear that continued incubation beyond 200 hr will increase
mineralization much further. We are currently evaluating activity of this strain at a wider range
of temperature and incubation times. The effects of pH and temperature on the mineralization
of 14C-phenanthrene by strain CRE-7, a low molecular-weight PAH degrader, are currently under
study.
Of the creosote fractions assessed, the acid-extractable (phenolic) and base-extractable
(heterocyclic) fractions were the most inhibitory to the activity of strain EPA505. At 50 mg/L, the
phenolics fraction slowed the onset of mineralization; at 70 mg/L, no mineralization was
observed (Figure 3). The base-extractable fraction (mostly heterocycles) was inhibitory at 35
mg/L (data not shown). Whole creosote was inhibitory at >200 mg/L. The neutrally extracted
fraction and the weathered neutral fractions were not inhibitory at any concentration tested
(>210 mg/L).
The basis of this inhibition is not known but could be the result of direct toxicity to the cells or
isotope dilution caused by the use of more readily degradable substrates, or could be an effect
of decreased availability of the radiolabeled substrate. Studies are currently in progress using
synthetic mixtures of all fractions to decipher the inhibitory mechanism and more accurately
identify inhibitory constituents and concentrations. These studies will also identify individual
creosote constituents most inhibitory to this strain. Similar studies with strain CRE-7 are in
progress. In addition, the results of pure culture studies will be compared with results from
studies using natural microbial communities that have been enriched to degrade creosote.
Summary and Conclusions
If the isolated strains of bacteria under study represent the potential activities of bacteria in
contaminated site material, then environmental conditions may have to be manipulated, in some
cases, to provide optimal activity. Where low temperature and pH extremes are encountered
145
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in the field, substantial effects on PAH mineralization can be expected. In addition, if
bioaugmentation is considered as a biotreatment strategy, inoculants may have to be carefully
selected to be effective under these suboptimal conditions.
These data further support implementation of creosote bioremediation via a two-stage process
(patent pending) employing co-inoculation (e.g., bacterial strain to degrade the "toxic" phenolic
and heterocyclic fractions) and secondary biotreatment of more recalcitrant constituents (e.g.,
strain EPA505 to treat high molecular-weight PAHs) (9).
References
1. Mueller, J.G., P.J. Chapman, and P.M. Pritchard. 1989. Action of a
fluoranthene-utilizing bacterial community on polycyclic aromatic hydrocarbon
components of creosote. Appl. Environ. Microbiol. 55:3,085-3,090.
2. Mueller, J.G., S.E. Lantz, R.J. Colvin, D. Ross, D.P. Middaugh, and P.M. Pritchard.
1993. Strategy using bioreactors and specially selected microorganisms for
bioremediation of ground water contaminated with creosote and pentachlorophenol.
Environ. Sci. Technol. 27:691-698.
3. Mueller, J.G., S.M. Resnick, M.E. Shelton, and P.M. Pritchard. 1992. Effect of
inoculation on the biodegradation of weathered Prudhoe Bay crude oil. J. Indust.
Microbiol. 10:95-105.
4. Mueller, J.G., S.E. Lantz, B.O. Blattmann, and P.J. Chapman. 1991. Bench-scale
evaluation of alternative biological treatment processes for the remediation of
pentachlorophenol- and creosote-contaminated materials: Solid-phase bioremediation.
Environ. Sci. Technol. 25:1,045-1,055.
5. Mueller, J.G., S.E. Lantz, B.O. Blattmann, and P.J. Chapman. 1991. Bench-scale
evaluation of alternative biological treatment processes for the remediation of
pentachlorophenol- and creosote-contaminated materials: Slurry-phase bioremediation.
Environ. Sci. Technol. 25:1,055-1,061.
6. Mueller, J.G., P.J. Chapman, and P.M. Pritchard. 1989. Creosote-contaminated sites:
Their potential for bioremediation. Environ. Sci. Technol. 23:1,197-1,201.
7. Lin, J.-E., J.G. Mueller, S.E. Lantz, and P.M. Pritchard. 1994. Influencing mechanisms
of operational factors on the degradation of fluoranthene bySph/ngomonas paucimobilis
strain EPA505. Biochem. Eng. Internal review.
8. Mueller, J.G., P.J. Chapman, B.O. Blattmann, and P.M. Pritchard. 1 990. Isolation and
characterization of a fluoranthene-utilizing strain of Pseudomonas paucimobilis. Appl.
Environ. Microbiol. 56:1,079-1,086.
9. Mueller, J.G., J.-E. Lin, S.E. Lantz, and P.M. Pritchard. 1993. Recent developments in
cleanup technologies: Implementing innovative bioremediation technologies.
Remediation (summer issue), pp. 369-381.
146
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Table 1. Composition" of Synthetic Mixtures of Creosote Constituentsb Used in
Mineralization Inhibition Studies
Neutral Fraction
(PAr
Naphthalene
2 -Methyl na phthalene
1 -Methylnaphthalene
Biphenyl
2,3-Dimethylnaphthalene
2,6-Dimethylnaphthalene
Acenaphthene
Acenaphthylene
Fluorene
Phenanthrene
Anthracene
2-Methylanthracene
Anthraquinone
Fluoranthene
Pyrene
Benzo[a]anthracene
Chrysene
2,3-Benzofluorene
Benzo[a]pyrene
Acidic Fraction
(Phenolicsl
Phenol
o-Cresol
m-Cresol
p-Cresol
2,5-Xylenol
3,5-Xylenol
2,3-Xylenol
2,4-Xylenol
2,6-Xylenol
3,4-Xylenol
2,3,5-Trimethylphenol
Basic Fraction
(Heterocvclics)
Quinoline
Isoquinoline
Carbazole
Acridine
2-Methylquinoline
4-Methylquinoline
Dibenzothiophene
Dibenzofuran
"Composition of fractions based on data reported by Mueller et al. (6)
bCompounds listed in order of elution during gas chromatography according to methods
previously described (4,5)
147
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o
N
Zj
QC
ID
70
60-
50
40
30
20
10
0
-10
50 100 150
INCUBATION TIME (hrs)
200
pH4
pH5
pH6
pH7
pH8
pH9
pHlO
Figure 1. Effect of media pH on 14C-fluoranthene mineralization by strain EPA505.
70
o
P
a.
1U
s
8«
60-
50-
40-
30-
20-
10-
37 C
30 C
25 C
18C
50 100 150
INCUBATION TIME (hrs)
200
Figure 2. Effect of incubation temperature on "C-fluoranthene mineralization by strain EPA505.
148
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o
H
N
OC
III
20-
10-
246
INCUBATION TIME (days)
Figure 3. Mineralization of 14C-fluoranthene by strain EPA505 in the presence of the acid-
extractable fraction of creosote (phenol ics).
149
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Molecular Genetic Approaches to the Study of the Biodegradation of Polycyclic
Aromatic Chemicals
Richard W. Eaton and Peter J. Chapman
U.S Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
James D. Nitterauer
Technical Resources, Inc., Gulf Breeze, FL, and
University of Arkansas for Medical Sciences, Little Rock, AK
Petroleum, coal, and their derivatives are composed of a variety of chemicals, including
polycyclic aromatic hydrocarbons (PAHs), heterocyclics, and alkyl-substituted aromatics. As these
chemicals increase in size and complexity, bacteria have more difficulty metabolizing them. In
addition, their catabolic pathways are lengthy and often branched, making it more difficult to
study them.
The approach that we are taking to study the biodegradation of individual xenobiotic chemicals
involves a variety of strategies; foremost among these are molecular genetic techniques such as:
1) cloning genes that encode enzymes that catalyze reactions of interest and 2) isolating
transposon-induced mutants that lack enzymes of a metabolic pathway. These approaches
allow an individual enzyme-catalyzed reaction or set of reactions to be studied in the absence
of other reactions that complicate analysis. This approach obviously allows the simultaneous
study of both the enzymes and the genes that confer on an organism its metabolic capabilities.
Naphthalene and benzothiophene are simple, fused-ring compounds that can serve as models
for more complex polycyclic aromatic chemicals (PACs) in biodegradation studies. The pathway
for the bacterial metabolism of naphthalene (Figure 1) was characterized (1,2) using
recombinant bacteria containing genes cloned from the naphthalene catabolic plasmid NAH7
(Figure 2). Bacteria carrying the plasmid, pRE657, which contains a 10-kb EcoRI-C/al fragment
on which the genes nahA, nahB, and nahC are located, converted naphthalene (Figure 1, I) to
a mixture of two chemicals, 2-hydroxychromene-2-carboxylate (HCCA, Figure 1, VI) and trans-o-
hydroxybenzylidenepyruvate (tHBPA, Figure 1, VII). The initial product, HCCA, and tHBPA
spontaneously isomerize in aqueous solution to form an equilibrium mixture of the two
compounds, making their identification difficult. Separation was possible, however, using
column chromatography on Sephadex G-25 with water as solvent; this allowed the rigorous
identification of these compounds using 'H- and 13C-NMR spectroscopy and gas
chromatography/mass spectrometry (GC/MS). Subclones pRE701 and pRE718 were obtained
that encode the enzymes tHBPA hydratase-aldolase (Figure 1, E) and HCCA isomerase (Figure
1, D), respectively, and act on these intermediates. These two intermediates, and the enzymes
that degrade them, are characteristic of pathways for the degradation of aromatic compounds
with two or more rings. The genes that encode these enzymes (nahE and nahD) may thus have
value as specific probes for environmental microorganisms that degrade PACs; this serves as
part of the justification for the recently completed sequencing of these genes (3).
150 1994 Symposium on Koremediation of Hazardous Wastes
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The sulfur-containing heterocycle benzothiophene is transformed by isopropylbenzene-degrading
bacteria to a mixture of products. One of these strains, Pseudomonas putida RE204, and its
Tn5-generated mutant derivatives (4) were used to study these biotransformations (5). Three
products were formed from benzothiophene by the isopropylbenzene-induced wild-type strain
RE204: frans-4-(3-hydroxy-2-thienyl)-2-oxobut-3-enoate (Figure 3, XIII), 2-mercap-
tophenylglyoxalate (Figure 3, XV), and 2'-mercaptomandelaldehyde (Figure 3, XVI). The latter
was identified following its conversion to its isomer, frans-2,3-dihydroxy-2,3-
dihydrobenzothiophene. Once again, chromatography on Sephadex G-25 was employed to
separate these relatively unstable chemicals, which were subsequently identified by 'H- and >3C-
NMRspectroscopy and GC/MS. A 2,3-dihydroxy-2,3-dihydroisopropylbenzene dehydrogenase
(Figure 3, enzyme B') deficient mutant strain, RE213, converted benzothiophene to a's-4,5-
dihydroxy-4,5-dihydrobenzothiophene (Figure 3, XII) and 2'-mercaptomandelaldehyde,
presumably formed by the spontaneous opening of the thiohemiacetal c/s-2,3-dihydroxy-2,3-
dihydrobenzothiophene (Figure 3, XVI). Neither XIII nor XV was formed by this mutant;
apparently the dihydrodiol dehydrogenase is required for the formation of these compounds.
The complex mixture of products formed by the fortuitous metabolism of the relatively simple
chemical benzothiophene illustrates the problems that can occur in any bioremediation treatment
in which co-metabolism occurs. While substrates disappear, products that often are not targeted
by the analytical method employed may form and accumulate.
References
1. Eaton, R.W., and P.J. Chapman. 1991. Degradation of naphthalene, PAHs, and
heterocyclics. In: U.S. EPA. Symposium on bioremediation of hazardous wastes:
EPA's biosystems technology development program (abstracts), pp. 81-83.
2. Eaton, R.W., and P.J. Chapman. 1992. Bacterial metabolism of naphthalene:
Construction and use of recombinant bacteria to study ring cleavage of 1,2-
dihydroxynaphthalene and subsequent reactions. J. Bacteriol. 174:7,542-7,554.
3. Eaton, R.W. 1994. Sequence of the DNA encoding 2-hydroxychromene-2-carboxylate
isomerase and frans-o-hydroxybenzylidenepyruvate hydratase-aldolase from the NAH7
plasmid. In preparation.
4. Eaton, R.W., and K.N.Timmis. 1986. Characterization of a plasmid-specified pathway
for cataholism of isopropylbenzene in Pseudomonas putida RE204. J. Bacteriol.
168:123-131.
5. Eaton, R.W., and J.D. Nitterauer. 1994. Biotransformation of benzothiophene by
isopropylbenzene-degrading bacteria. J. Bacteriol. Submitted.
151
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IX
Figure 1. Pathway for the bacterial metabolism of naphthalene to salicylate.
nart genes
B F C ED
^ "5 = £ 5 o"o 2
xS 3 3JS
XV
XVI
Figure 3. Biotransformations of benzothiophene.
152
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Comparison of Sulfur and Nitrogen Heterocyclic Compound Transport in
Creosote-Contaminated Aquifer Material
Ean Warren and E. Michael Godsy
U.S. Geological Survey, Menlo Park, CA
Introduction
Commonly, ground-water solute transport model inputs are generated from chemical and
ground-water properties that are not comparable with those at the site of contamination. Care
must be taken when assuming that chemicals with similar molecular structures or characteristics
possess equivalent transport properties. In addition, ground-water characteristics, such as pH,
must be compared with ionization constants (pKa) to determine the influence of the sediment's
ion exchange capacity. Simulated transport will not be accurate if the parameter determined
at one pH differs from that of the ground water.
In this paper, we compare the values of partition coefficients and retardation factors for the
sulfur and nitrogen heterocyclic compounds benzothiophene, dibenzothiophene, quinoline,
2(1 H)-quinolinone, acridine, and carbazole on low organic carbon content, low ion exchange
capacity aquifer material. Column breakthrough curves (BTCs) were modeled using the local
equilibrium assumption (LEA) for compounds with a log octanol-water partition coefficient (log
Kow) of less than 2.5 and the nonequilibrium assumption (NEA) for compounds with a log Kow
greater than 2.5.
Background
The column material is taken from sediment adjacent to an abandoned wood-preserving plant
within the city limits of Pensacola, Florida (1). The wood-preserving process consisted of steam
pressure treatment of pine poles with creosote and/or pentachlorophenol (PCP). For more than
80 years, a large but unknown quantity of waste water, consisting of extracted moisture from
the poles, cellular debris, creosote, PCP, and diesel fuel from the treatment processes, was
discharged to unlined surface impoundments that were in direct hydraulic contact with the sand-
and-gravel aquifer. The ground water, at a pH of 5.9 and moving at approximately 1 m/d, is
continually dissolving the more soluble compounds found in creosote, creating an extended
contamination plume. The aquifer material for the laboratory columns consisted of a low
organic carbon content (0.024 percent organic carbon), low ion exchange capacity (2 meq/
100 g) clay-like sand from the approximate centroid of the plume (Table 1).
BTCs of the water-soluble heterocyclic compounds in laboratory columns can be described by
the convection-dispersion equation using the LEA as proposed by Hashimoto et al. (2),
1994 Symposium an Bioremediation of Hazardous Wastes 153
-------
where 0 is the porosity (-),/>b is the bulk density of the aquifer material (g/m3), Kd is the partition
coefficient (m3/g), C is the aqueous concentration (g/m3), f is the time (d), D is the dispersion
coefficient (m2/d), x is the distance (m), and v is the linear velocity (m/d).
Transport of hydrophobic chemicals commonly must be modeled using the NBA as proposed
by van Genuchten and Wierenga (3), which accounts for a readily mobile fraction and a
stagnant or immobile fraction of water in the aquifer matrix (subscripts m and /m, respectively),
(2)
(3)
where f is the fraction of sorption sites in the mobile region (-) and a defines the transfer rate
of the solute between mobile and immobile water (d"1). As described by van Genuchten (A), the
variables, f and a, from equations 2 and 3, can be related to two fitted, dimensionless
parameters, respectively: ft, the fraction of the sites in the mobile region where sorption is
instantaneous, and (a, the ratio of hydrodynamic residence time to characteristic time of sorption
(5). The NEA model is based on the assumption that convection and dispersion govern
transport in the mobile water, and that diffusion controls the transfer of contaminant between
mobile and immobile water.
Both models assume a linear isotherm. Retardation factors, R, which describe the movement
of contaminants relative to a conservative tracer, can be related to partition coefficients, bulk
densities, and porosity by
(4)
Parameters were fit to BTCs using nonlinear regression analysis by the computer programs
HASHPE (6), to determine R for LEA, and CFITIM (4), to determine R,0, and (a for NEA. The
dispersion parameter for all model simulations was determined from CaCl2 breakthrough.
Brusseau and Rao (7) suggest that, for values of a> less than approximately 10, the NEA should
be used instead of the LEA to account for the observed tailing. The values of 0) for
benzothiophene, dibenzothiophene, carbazole, and acridine (compounds with log K^ > 2.5)
are well below 10 (Table 2), justifying the use of the NEA model. The NEA model determined
that the values of
-------
Results and Discussion
Fitted parameters and original coefficients for benzothiophene, dibenzothiophene, quinoline,
2(1H)-quinolinone, carbazole, and acridine using the models are given in Table 2. The
chemical structures are shown in Figure 1. The retardation factors for benzothiophene,
quinoline, 2(1 H)-quinolinone are quite similar to each other. 2(1 H)-Quinolinone, with a pl^of
5.29, is approximately 20-percent ionized, and quinoline, with a pKgOf 4.9, is approximately
9.1 -percent ionized. Zachara et at. (8) have shown that sorption of quinoline is dominated by
ion exchange up to 2 pH units above its pK,,. 2(1H)-Quinolinone, like quinoline, should be
retained by both ion exchange and organic sorption. Benzothiophene, however, is nonionic and
subject to organic sorption alone.
The values of /? for the sulfur heterocycles agree with each other but are greater than those for
the nitrogen heterocycles, suggesting a larger percentage of sites at which there is instantaneous
sorption for the sulfur heterocycles. The value of (o for the sulfur heterocycles is much less than
that for the nitrogen heterocycles, indicating that the characteristic time of sorption contributes
more to the retardation of nitrogen heterocycles, and to acridine transport in particular.
The retardation of acridine is much greater than that of dibenzothiophene and carbazole,
despite the fact that all have two benzene rings fused to a sulfur or nitrogen heterocyclic ring
(Figure 1 and Table 2) and have similar log K,,w; dibenzothiophene and carbazole, however, are
subject to organic sorption alone, whereas acridine is subject to both organic sorption and ion
exchange. The pK,, of acridine is 5.6 and of carbazole is -5.7 (Table 2). Thus, at pH 5.9, the
ionized fraction of acridine is 0.33, but carbazole is completely unionized. The degree of affinity
(the selectivity) of acridine to charged functional groups on the aquifer material and the extent
of ionization as well as the sediment's cation-exchange capacity contributes to the retention
capacity. With an acridine concentration of 18 g/m3 (0.10 meq/L), the column capacity due
to ion exchange is 160. The column capacity is based on the assumption of total sorption of
the ionized fraction of acridine to the aquifer material and complete displacement of calcium
ions.
Transport of organic chemicals in ground water must be modeled using parameters similar to
those at the site of interest. Assumptions about solute transport based on chemical and physical
properties of similar but not identical compounds, aquifer sediments, and ground water are not
always valid. Field conditions, such as pH, flow velocity, and chemical properties (such as
selectivity and pKJ, must be taken into consideration to effectively model solute transport.
References
1. Godsy, E.M., D.F. Goerlitz, and D. Grbic-Galic. 1992. Methanogenic biodegradation
of creosote contaminants in natural and simulated ground-water ecosystems. Ground
Water 30(2):232-242.
2. Hashimoto, I., K.B. Deshpande, and H.C. Thomas. 1964. Peclet numbers and
retardation factors for ion exchange columns. Ind. Eng. Chem. Fundam. 3:213-218.
3. van Genuchten, M.T., and P.J. Wierenga. 1976. Mass transfer studies in sorbing
porous media. I. Analytical solution. Soil Sci. Soc. Amer. Proc. 40:473-480.
155
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4. van Genuchten, M.T. 1981. Nonequilibrium transport parameters from miscible
displacement experiments. U.S. Department of Agriculture. U.S. Salinity Laboratory
Research Report 119:88.
5. Brusseau, M.L, and M.E. Reid. 1991. Nonequilibrium sorption of organic chemicals
by low organic-carbon aquifer materials. Chemosphere 22(3-4):341-350.
6. Oravitz, J.L. 1984. Transport of trace organics with one-dimensional saturated flow:
Mathematical modeling and parameter sensitivity analysis. M.S.C.E. thesis. Michigan
Technological University, Department of Civil Engineering.
7. Brusseau, M.L, and P.S.C. Rao. 1989. The influence of sorbate-organic matter
interactions on sorption nonequilibrium. Chemosphere 18(9-10): 1,691-1,706.
8. Zachara, J.M., et al. 1986. Quinoline sorption to subsurface materials: Role of pH
and retention of the organic cation. Environ. Sci. Technol. 20:620-627.
Table 1. Aquifer Material and Column Characteristics
Median particle diameter (m)
Percent organic carbon (-)
Cation exchange capacity (meq/100 g)
Column
Length (m)
Diameter (m)
Porosity (-)
Bulk density (g/m3 x 10'6)
Flow rate (mVdxlO6)
0.000375
.024
1.6
0.354
.025
.449
1.361
140
Table 2. pK,,, log K^, Partition Coefficients, Retardation Factors, and Nonequilibrium
Assumption Parameter Values for Benzothiophene, Dibenzothiophene, Quinoline,
2(1 H)-Quinolinone, Carbazole, and Acridine
Benzothiophene
Dibenzothiophene
Quinoline
2(1 H)-Quinolinone
Carbazole
Acridine
pKo
4.90
5.29
-5.70
5.60
logK^
3.12
4.38
2.03
1.26
3.29
3.40
Partition
Coefficient,
mVgxlO*
0.184
0.789
0.133
0.231
1.01
4.56
Retardation
Factor
1.74
3.84
1.32
1.54
4.34
39.6
/*
0.90
0.93
0.0
0.0
0.60
0.61
(O
0.48
0.23
160
1300
2.6
1.2
156
-------
Benzothiophene
Dibenzothiophene
H
Carbazole
2( lH)-Quinolinone
Figure 1. Chemical structures of benzothiophene, dibenzothiophene, quinoline, 2(1 H)-
quinolinone, carbazole, and acridine.
157
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Modeling Steady-State Methanogenic Degradation of Phenols in Ground Water at
Pensacola, Florida
Barbara A. Bekins, E. Michael Godsy, and Donald F. Goerlitz
Water Resources Division, U.S. Geological Survey, Menlo Park, CA
Introduction
The study site is an abandoned wood treatment facility in the extreme western end of the Florida
Panhandle within the city of Pensacola. For about 80 years, creosote-derived contaminants and
pentachlorophenol from unlined waste-disposal ponds entered the ground water in the
underlying sand and gravel aquifer. Concentrations of phenol and 2-, 3-, and 4-methylphenol
have been monitored at the study site for more than 12 years. The data indicate that a
nonaqueous-phase source below the ponds provides a constant input of dissolved phenols that
are then degraded within 200 m downgradient. Figure 1 is a generalized geologic section
along a flow line down the axis of the plume together with contours of total phenolic compound
concentration. The degradation process appears to be at steady state because the
concentration profile has not changed over the last 12 years. The aquifer consists of
approximately 90 m of poorly sorted fine to coarse grained deltaic sand deposits interrupted by
discontinuous silts and clays. Ground-water flow is generally horizontal and southward toward
Pensacola Bay. Flow velocities range from 0.3 m/d to 1.2 m/d (1).
Model Description
Godsy et al. (2) determined methanogenic utilization rates for four phenolic compounds in
microcosms containing aquifer sediments. They fit the change in concentration with time and
the associated microbial growth to the equations for Monod growth and substrate utilization.
Their results, given in Table 1, were used in a model describing transport and degradation at
the field site.
The modeled profile is 6 m below the surface in the methanogenic part of the contaminated
zone, below the depth at which recharge and floating hydrocarbon at the water table affect
concentrations and above the clay lenses. A one-dimensional model was used because the flow
direction is primarily horizontal and perpendicular to a wide contaminant source. Acridine
orange direct counts (AODC) indicate that the bacteria population is spatially uniform and low
(5 x 103 to 7.6 x 107 AODC/g dry weight of sediment) relative to subsurface enumerations at
other sites (3). The existence of a steady-state degradation profile of each substrate, together
with a low, uniform bacteria density, indicates that the bacterial population is exhibiting no net
growth (4). Thus, the bacteria concentration in the model is held constant in time and uniform
in space.
We assume that the substrate profile at a depth of 6 m satisfies the one-dimensional transport
equation with a Monod reaction term:
158 1994 Symposium on Roremediation of Hazardous Wastes
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where S is the substrate concentration (mg/L); f is time (d); x is distance downgradient from the
first observation well (m); D is the dispersion coefficient (m2/d); v is average linear velocity (m/d);
fim is maximum growth rate (d"1}; Y is yield (mg bacteria per mg S); 8 is the concentration of the
active degrading bacteria (mg/L); 9 is porosity; and Ks is the half-saturation constant (mg/L).
This equation was solved using a computer code described by Kindred and Celia (5), with
boundary and initial conditions given by:
S (0,1) = S0;
d* x-250
= 0; S (x,0) = S0; (2)
where 50 is the contaminant concentration 6m below the ground surface at Site 3, the closest
site to the source.
Model Results
Two predicted steady-state substrate profiles, along with the measured phenolic-compound
concentrations at 6 m below land surface at each sample site, are shown in Figure 2. The
computed profiles are steady-state solutions to a one-dimensional advective-dispersive equation
with a biological reaction term (Equation 1). The upper curve predicts the field profile that
would result from the phenol degradation rate that was measured in the lab, whereas the lower
curve corresponds to the rate measured for 2-methylphenol. These two rates were used because
they have the smallest associated errors and bracket the rates for the other two compounds
(Table 1). The values for bacteria concentration were varied to obtain the best match to the
data. The parameters used in the solution for phenol and 3-methylphenol were: S0 = 26.0
mg/L, fim = 0.111 d'1, Y = 0.013, Ks = 1.33 mg/L, B = 1.5 x 10'2 mg/L, v = 1.0 m/d, D =
1.0 m2/d. For 2- and 4-methylphenol, the values used were: S0 = 13.5 mg/L, fim = 0.044
d-', Y = 0.022, K. = 0.25 mg/L, B = 3.0 x 1O'2 mg/L, v = 1.0 m/d, D = 1.0 m2/d. A steady-
state solution was obtained for all \ > 1,000 days.
The model profiles indicate that the rates measured in the microcosm simulations accurately
represent the degradation process taking place in the field. The validity of the Monod kinetics
expression for the degradation rate is apparent from the field data because the rate of decrease
in the phenol concentration changes dramatically around Site 40 (located about 90 m
downgradient from Site 3). When the substrate concentration is high (upgradient of Site 40),
the degradation kinetics can be approximated by a zero-order reaction term consistent with the
low values of Ks observed in the lab studies. When the substrate concentration is close to the
value of Kit the degradation rate drops as predicted by Monod kinetics. The fitted bacteria
concentration for the upper curve is twice that for the lower curve, because the yield value for
phenol is half that for the other compounds.
Recall that to obtain a steady-state solution for the concentrations, it was necessary to assume
no net growth for the bacteria. To investigate how this may happen, we used the values in
159
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Table 1 and the computed phenol concentrations from Figure 2 to predict the growth rate that
is consistent with the observed degradation rate. Figure 3 shows predicted bacterial growth
rates with and without the effect of toxic inhibition. In the curve with no inhibition, the peak
growth rate is roughly 0.1 day'. To maintain zero net growth, the plotted growth rate must be
balanced by an equivalent decay rate. A decay rate of 0.0192 day'1 for methanogens was
found by Suidan et al. (6) in a continuous reactor. This value, shown as a horizontal line in
Figure 3, is almost an order of magnitude too low to balance the predicted growth.
Furthermore, in theory, the functional form of the positive growth curve cannot be balanced by
a constant decay rate. When the toxicity of phenol is accounted for using a Haldane (7)
inhibition model, the predicted growth is about 50 percent lower but still much higher than the
published decay rate.
Summary and Conclusions
We have created a model of methanogenic degradation of phenolic compounds for a sand and
gravel aquifer at Pensacola, Florida. The model verifies that field disappearance rates of four
phenols match those determined in batch microcosm studies performed by Godsy et al. (2). The
degradation process appears to be at steady state because a sustained influx of contaminants
over several decades has been continuously disappearing within 150 m downgradient of the
source. Goerlitz et al. (8) concluded that sorption was insufficient to explain the observed loss.
The existence of a steady-state degradation profile of each substrate, together with a low
bacteria density in the aquifer, indicates that the bacterial population is exhibiting no net growth.
This is possibly because of the oligotrophic nature of the bacteria population indicated by the
low value for K$. A low K, causes growth and utilization to be approximately independent of the
phenolic-compound concentration for most of the concentration range. Thus, a roughly
constant bacteria growth rate should exist over much of the contaminated area. This growth
could be balanced by an unusually high decay or maintenance rate caused by hostile conditions
or predation. Alternatively, the loss of bacteria by transport downgradient is being investigated
with column studies.
References
1. Franks, B.J. 1988. Hydrogeology and flow of water in a sand and gravel aquifer
contaminated by wood-preserving compounds, Pensacola, Florida. U.S. Geological
Survey Water-Resources Investigations Report 87-4260. p. 72.
2. Godsy, E.M., D.F. Goerlitz, and D. Grbic-Galic. 1992. Methanogenic degradation
kinetics of phenolic compounds in aquifer-derived microcosms. Biodegradation
2:211-221.
3. Godsy, E.M., D.F. Goerlitz, and D. Grbic-Galid. 1992. Methanogenic biodegradation
of creosote contaminants in natural and simulated ground-water ecosystems. Ground
Water 30:232-242.
4. Bekins, B.A., E.M. Godsy, and D.M. Goerlitz. 1993. Modeling steady-state
methanogenenic degradation of phenols in ground water. J. Contam. Hydrol.
14:279-294.
160
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5. Kindred, J.S., and M.A. Celia. 1989. Contaminant transport and biodegradation 2.
Conceptual model and test simulations. Water Resour. Res. 25:1,149-1,159.
6. Suidan, M.T.,I.N. Najm,J.T. Pfeffer, and Y.T. Wang. 1989. Anaerobic biodegradation
of phenol: Inhibition kinetics and system stability. J. Environ. Eng. 114:1,359-1,376.
7. Haldane, J.B.S. 1930, Enzymes. London, New York: Longmans, Green.
8. Goerlitz, D.F., D.E. Troutman, E.M. Godsy, and B.J. Franks. 1985. Migration of wood
preserving chemicals in contaminated ground water in a sand aquifer at Pensacola,
Florida. Environ. Sci. Technol. 19:955-961.
Table 1. Kinetic Constants From Microcosm Studies for Each of the Phenolic
Compounds Tested (2)*
Compound
Phenol
2-Methylphenol
3-Methylphenol
4 -Methyl phenol
Growth Rate
A*. Id'1)
0.111 ±0.005
.044 ± 0.001
.103 ±0.078
.099 ±0.110
Half Saturation
*. (mg/U
1 .33 ± 0.07
.25 ± 0.82
.55 ± 6.67
3.34 ±11.1
Yield
Y (mg/mg)
0.013
.022
.026
.025
*Yield values were obtained from protein determinations before and
after substrate ulitization.
161
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Well Sites
0 100 meters
I I
I r
0 300 feet
Vertical Exageration lOx
Sand
Sandy clay
Clay
Figure 1. Generalized geologic section along a flow line down the center of the plume.
Contours of total phenols are shown in mg/L.
162
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30
§ 20
Z
P 15
10
8
0
0 -
\
N \
| . . . 1 . f — .S*- L • . , 1
0 40 80 120 160 200
DISTANCE FROM WELL SITE 3 (METERS)
Figure 3. Theoretical growth rate computed from the phenol concentration, the Monod growth
expression, and the growth parameters measured in the microcosm simulations. The
two curves are computed with and without the effect of Haldane inhibition.
163
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Anaerobic Biodegradation of 5-Chlorovanillate as a Model Substrate for the
Bioremediation of Paper-Milling Wasle
B.R. Sharak Genthner and B.O. Blattmann
Avanti, Corp., Gulf Breeze, FL
P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
5-Chlorovanillate (5CV; 5-chloro-4-hydroxy-3-methoxybenzoic acid) was selected as a model
compound for studying the biodegradation of paper mill effluents. This compound contains the
methoxy-, chloro-, and carboxyl side groups often present on aromatic chlorinated compounds
released in paper mill effluents. The major pathway of 5CV degradation was previously
determined to be stepwise demethoxylation to 5-chloroprotocatechuate (5CP; 5-chloro-3,4-
dihydroxybenzoic acid), decarboxylation to 3-chlorocatechol (3CC; 3-chloro-l,2-dihydro-
xybenzene), and dechlorination to catechol, which was completely degraded (Figure 1). The
current research further investigates theanaerobic bacterial species responsible forthe individual
transformation steps. Once obtained in pure culture, studies can be performed investigating
individual transformation steps with reduction in toxicity of paper mill waste.
Selective media containing guaiacol (2-methoxyphenol), protocatechuate (dihydroxybenzoic acid)
and catechol as the sole energy source were inoculated with the original 5CV culture.
Transformation of target compounds in these enrichment cultures was followed using high-
performance liquid chromatography analyses. Immediately upon completing the transformation
of interest, the cultures were passed to fresh medium. The guaiacol, protocatechuate, and
catechol cultures were sequentially transferred through their respective media several times,
followed by several refeedings of the target compound to enrich for the bacterial species of
interest. These enrichments were then diluted in the respective media to obtain bacterial cultures
responsible for demethoxylation (Figure 2), decarboxylation (Figure 3), and catechol (Figure 4)
degradation. The data indicate that the demethoxylating and decarboxylating bacterial species
were more numerous by three orders of magnitude than the catechol-degrading bacterial
species. The transforming and degrading activity in these cultures has been sustained for several
months and through several transfers, indicating that the activity is stable—a condition necessary
for bioremediation applications. The demethoxylating and decarboxylating cultures continued
to transform guaiacol and protocatechuate in the presence of fairly high concentrations of
catechol. Demethoxylation rates begin to decline above 3 mM catechol (Figure 2B), while
decarboxylation rates did not decline significantly at 10 mM catechol (Figure 3B). Because
paper mill waste contains other phenolic compounds, applied bacterial cultures must tolerate
other toxics while performing the desired transformation. Photomicrographs of these cultures
show apparently pure cultures. Purity of these cultures is currently being confirmed.
The initial dechlorination of 5CV was investigated using a 3-chlorobenzoate-dechlorinating
anaerobic co-culture, which dechlorinated 5CVto vanillate and then demethoxylated vanillate
to protocatechuate. Protocatechuate was not further metabolized. Asulfate-reducing bacterium
was isolated from this co-culture and identified as a new bacterial species, Desu/fom/crob/um
escambium (1). Initial investigations with the pure culture of D. escambium showed a decline
in the concentration of 3-chlorobenzoate (3CB) in defined pyruvate/3CB medium, which
164 1994 Symposium on Bioremediation of Hazardous Wastes
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depended upon the presence of pyruvate. Because reductive dechlorination has been shown
to be very specific for halogen position (2,3), and 3CB and 5CV are both meta-chlorinated, the
basis for the decline in 3CB by D. escambium was further investigated.
Further studies indicated that D. escambium transformed not only 3CB but 3-bromobenzoate
(3BB) and benzoate as well (Figure 5). Again, the decline was dependent upon the presence
of pyruvate. Lactate, formate, ethanol, and hydrogen, which are used by D. escambium as
electron donors for sulfate reduction, did not support the transformation of these three
compounds. The similarity in transformation rates between benzoate and the two halogenated
benzoates suggested that the transformation being observed was not dehalogenation. After
derivitization, gas chromatography analysis revealed the presence of two unknown compounds
in each culture. Further investigation using gas chromatography/mass spectrometry (GC/MS)
analysis indicated that 3CB, 3BB, and benzoate were being reduced to their respective alcohols
without dehalogenation (Figure 5).
During GC/MS analysis, the second unknown peak was identified as succinate. Under
anaerobic conditions, succinate can result from the carboxylation of pyruvate. Afollowup study
showed that benzoate was not reduced in medium containing a gas phase of 100-percent
nitrogen. The requirement for both pyruvate and carbon dioxide indicates that the reduction of
the benzoate compounds to their respective alcohols by D. escambium is dependent upon
carboxylation of pyruvate to succinate. If sulfate is added to the pyruvate/benzoate medium,
sulfate is reduced, benzoate does not decline, and pyruvate is degraded to acetate and carbon
dioxide. Apparently, the reducing equivalents in this case are diverted from the reduction of
benzoate to the reduction of sulfate, energetically a more favorable reduction. If reductive
dechlorination competes similarly for reducing equivalents, the presence of sulfate would be
unfavorable for detoxification of paper mill waste.
Because D. escambium reduces but does not dechlorinate 3CB in pure culture, attempts are
currently under way to isolate the second member of the 3CB-dechlorinating co-culture. This
bacterial species may be responsible for dechlorination of 3CB and 5CV by the co-culture or
may provide a factor that enables D. escambium to divert reducing equivalents to the
dechlorination of 3CB or 5CV.
References
1. Sharak Genthner, B.R., G. Mundfrom, and R. Devereux. 1994. Characterization of
Desu/fom/crobium escambium sp. nov. and proposal to assign Desu/fov/brio
desulfuricans strain Norway 4 to the genus Desu/fomicrobium. Arch. Microbiol. In
press.
2. Boyd, S.A., and D.R. Shelton. 1984. Anaerobic biodegradation of chlorophenols in
fresh and acclimated sludge. Appl. Environ. Microbiol. 46:50-54.
3. Suflita, J.M., A. Horowitz, D.R. Shelton, and J.M. Tiedje. 1982. Dehalogenation: A
novel pathway for the anaerobic degradation of haloaromatic compounds. Science
218:1,115-1,117.
165
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OH
o
Cl-"\x-- "OH
OH Cl
5-CHLOROVANILLIC 5-CHLOROPROTOCATECHUIC 3-CHLOROCATECHOL
OH
C02 + CH4
CATECHOL
Figure 1. Pathway for the complete degradation of 5-chlorovanillic acid.
166
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3.0
JE,
2
o
t—
cr
i—
2
UJ
O
o
o
E,
~z.
o
h-
2
LJ
O
2
O
O
2.0--
1.0--
• •GUAIACOL
O OCATECHOL
0.0
4.0
3.0--
2,0-
1.0.
0.0<
0
10 20 30 40 50 60
• •GUAIACOL
O OCATECHOL
00
25 50
TIME (DAYS)
B
Figure 2. Enrichment for demethoxylating anaerobic bacterial species (A) and
demethoxylating activity in highest active (10"7) dilution of a demethoxylating (B)
anaerobic bacterial consortium.
167
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2.5
z
LJ
O
O
O
0
10 20
30
8»
§ 6 +
4-
2-
O-
PROTOCATECHUATE
C?'
0
O
00
.0—0
00°°
(S)
O
70
€>
/
,^
OO
0 25 50 75 100 125
TIME (DAYS)
150
Figure 3. Enrichmentfordecarboxylating anaerobic bacterial species (A) anddecarboxylating
activity in highest (1 0"7) active dilution (B) of decarboxylating anaerobic consortium.
168
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o
o
<
o
o
I
o
UJ
£
o
500
400--
300--
200
100-
0
800
600-
400
•'
200--
0
25
50
100 125 150
B
^
0 25 50 75 100 125 150 175
TIME (DAYS)
Figure 4. Enrichment for catechol-degrading anaerobic bacterial species (A) and catechol-
degrading activity in highest (10"4) active dilution (B) of a catechol-degrading
anaerobic consortium.
169
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600
400--
o
a:
o 200
o
o
600
0
10 20
TIME (DAYS)
--400
^200
Figure 5. Reduction of 3CB (O), 3BB (A), and benzoate (D) to 3-chloro- (V), 3-bromo- (X),
and benzyl alcohol (0) by desulfomicrobium escambium strain ESC1. Symbols:
Open, 0.2-percent pyruvate; closed, minus pyruvate.
170
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Characterization of a 4-Bromophenol Dehalogenating Enrichment Culture:
Conversion of Pentachlorophenol to Phenol by Sediment Augmentation
Xiaoming Zhang
National Research Council, National Academy of Sciences, Washington, DC
W. Jack Jones and John E. Rogers
U.S. Environmental Protection Agency, Environmental Research Laboratory, Athens, GA
Introduction
Pentachlorophenol (PCP), a carcinogen and ionophore (energy transfer inhibitor), is included
on EPA's list of priority pollutants. Reductive dechlorination was found to be a significant
reaction mechanism for the anaerobic degradation of PCP. The sequential removal of chlorines
from PCP and its intermediate products may lead to less toxic products. In this abstract, we
present data to demonstrate PCP transformation to phenol in sediment slurries inoculated with
cells from a 4-bromophenol (4-BP) dehalogenating enrichment culture. We also describe partial
characterization of the 4-BP-dehalogenating enrichment.
Methods
Sediment samples were collected from a freshwater pond in Cherokee Trailer Park, near Athens,
Georgia. Sediment slurries were adapted to degrade 3,4-dichlorophenol (3,4-DCP) by the
sequential addition of 3,4-DCP (61 ^M) to the slurries immediately following the disappearance
of the previous addition of 3,4-DCP (every 2 to 3 weeks). After 12 months, the 3,4-DCP-
adapted sediment slurry was transferred (1:1 vol/vol) to a mineral medium containing 0.1
percent yeast extract, 0.2 mM 3,4-DCP, and 50-percent (vol/vol) site water according to Zhang
and Wiegel (1). The pH of the medium was adjusted to pH 7.2 to 7.3 with HCI. Transfers were
made when the 3,4-DCP was redudively dechlorinated to at least 3-chlorophenol (3-CP). The
3,4-DCP dechlorination activity could also be maintained by substituting 4-BP for 3,4-DCP. The
4-BP (0.5 mM to 0.8 mM) maintained culture was used in subsequent experiments to examine
the dechlorination of PCP and its intermediate products. These experiments were performed
using: 1) 4-BP inoculated cultures in yeast extract-containing mineral medium, 2) washed cell
suspensions prepared from cells grown with 4-BP in the mineral medium, and 3) sediment
slurries amended with the 4-BP washed cell suspension.
Results and Discussion
2,3-DCP, 2,4-DCP, or 3,4-DCP, added to mineral medium and inoculated (20 percent v/v log
phase culture) with cells grown on 4-BP, were dechlorinated to monochlorophenols (MCPs).
Under the same conditions, PCP (18.8 ftM to 37.5 //M) was not dechlorinated. 4-BP was
dehalogenated to phenol in the control culture (plus 4-BP grown cells) supplemented with only
4-BP but not in the culture supplemented with both 4-BP and PCP, indicating that PCP inhibited
growth and/or activity of the dehalogenating culture.
1994 Symposium on Boremediation of Hazardous Wastes 171
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4-BP grown cells that were harvested from a late log culture, washed, and resuspended in
phosphate buffer to concentrate cells 40- to 100-fold, exhibited dehalogenating activity in the
presence of pyruvate. All chlorophenols tested (19 congeners), except the three MCPs, were
dechlorinated at orfho, mefa, or para positions in the presence of chloramphenicol, which
inhibited any further production of dehalogenating enzymes. As examples, 2,4-DCP was
dechlorinated to 2-CP and 4-CP, and 3,4-DCP was dechlorinated to 3-CP, which was not
further transformed. These results are consistent with a previous observation that all six
dichlorophenol isomers were dechlorinated in 3,4-DCP-adapted sediments (2).
Although PCP (300 juM) was preferentially dechlorinated at the orfho position by the 4-BP grown
cell suspension (concentrated 40-fold), dechlorination of meta and para chlorines was also
observed. 2,3,4,5-, 2,3,4,6-, and 2,3,5,6-tetrachlorophenol (TetCP) were identified as
intermediate products using a combination of high-performance liquid chromatography, gas
chromatography, and gas chromatography/mass spectrometry analyses. Addition of either
hydrogen, formate, or ethanol did not stimulate the dechlorination activity. Heat-treated (10 min
at 90°C) or solvent-permeated (toluene-treatment) cells lost dehalogenating activity. Sulfite,
thiosulfate, and sulfide inhibited the orfho and para dechlorination of 2,4-DCP. The addition of
sulfate or sodium chloride had no effect.
In a 4-BP grown cell suspension assay prepared in 99.9-percent deuterium oxide, 2,3,4-
trichlorophenol (2,3,4-TCP) was transformed to DCPs and MCPs containing one and two
deuterium atoms, respectively. This verified the identity of the proton source (water) for the
dechlorination of 2,3,4-TCP and its intermediates. This phenomenon has also been observed
for the reductive dechlorination of 2,5-dichlorobenzoate and 2,3,4,5,6-pentachlorobiphenyl
(3,4).
PCP (28 ^M) was dechlorinated to phenol (about 90-percent stoichiometric conversion) in 5
days in sterilized (autoclaved) and nonsterilized freshwater sediment slurries inoculated
(equivalent to 8-percent inoculation) with a washed cell suspension prepared from a 4-BP
dehalogenating enrichment culture. 2,3,4,5-TetCP, 3,4,5-TCP, 3,5-DCP, and 3-CP were
detected as transient intermediates (Figure 1). In addition, small peaks with retention times
similar to those found for 2,3,4,6-TetCP and 2,3,5,6-TetCP were also detected. In sterilized
and in nonsterilized, noninoculated control slurries, PCP was not transformed. The PCP
transformation pathway identified in this study was somewhat different than the pathway reported
by Bryant et al. (2) for 3,4-DCP-adapted sediment slurries (or a combination of 2,4-DCP- and
3,4-DCP-adapted sediments) prepared from the same site. 2,3,5,6-TetCP and 2,3,4,5-TetCP,
either alone or together, have been detected as products of PCP transformation in samples from
other ecosystems (5).
Specific experimental conditions were modified to identify factors affecting PCP transformation
in nonsterilized sediment slurries inoculated with the 4-BP enrichment culture. In these studies,
the PCP transformation rate was dependent on the concentration of added 4-BP grown cells,
pH, and temperature. Addition of potential electron donors, including pyruvate, formate, and
yeast extract, did not stimulate the transformation of PCP, suggesting that the concentration of
electron donor in the sediment slurry was not a rate-limiting factor for PCP transformation. The
presence or absence of 4-BP (0.15 mM) in these experiments did not significantly affect PCP
transformation. The rate of PCP transformation in an estuarine sediment slurry amended with
4-BP grown cells was 25 percent of the rate observed in the freshwater sediment slurry.
172
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In a previous study, Mikesell and Boyd (6) demonstrated that by inoculating PCP-adapted
sewage sludge into soil, PCP was dechlorinated to TCPs, DCPs, and MCPs in 28 to 35 days.
In our study, PCP was converted to phenol (90-percent recovery) within 5 days when a cell
suspension of the 4-BP dehalogenating enrichment culture was added to freshwater sediment
slurries. Taken together, these results suggest that bioaugmentation (and possibly induction) of
microbial populations may provide an alternative method of bioremediating PCP-contaminated
soils and sediments.
References
1. Zhang, X., and J. Wiegel. 1990. Isolation and partial characterization of a Clostridium
spec, transforming para-hydroxybenzoate and 3,4-dihydroxybenzoate and producing
phenols as the final transformation products. Microb. Ecol. 20:103-121.
2. Bryant, P.O., D.D. Hale, and J.E. Rogers. 1991. Regiospecific dechlorination of
pentachlorophenol bydichlorophenol-adapted microorganisms in freshwater, anaerobic
sediment slurries. Appl. Environ. Microbiol. 57:2,293-2,301.
3. Nies, L, and T.M. Vogel. 1991. Identification of the proton source for the microbial
reductive dechlorination of 2,3,4,5,6-pentachlorobiphenyl. Appl. Environ. Microbiol.
57:2,771-2,774.
4. Griffith, G.D.J.R. Cole, J.F. Quensen, III, and J.M.Tiedje. 1992. Specific deuteration
of dichlorobenzoate during reductive dehalogenation by Desulfomonile tiedjei in D2O.
Appl. Environ. Microbiol. 58:409-411.
5. Larsen, S., H.V. Hendriksen, and B.K. Ahring. 1 991. Potential forthermophilic (50"C)
anaerobic dechlorination of pentachlorophenol in different ecosystems. Appl. Environ.
Microbiol. 57:2,085-2,090.
6. Mikesell, M.D., and S.A. Boyd. 1986. Complete reductive dechlorination and
mineralization of pentachlorophenol by anaerobic microorganisms. Appl. Environ.
Microbiol. 52:861-865.
173
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o
o
2,3,4,5,6-PCP
2,3,4,5-TetCP
3,4,5-TCP
3,5-DCP
3-CP
Phenol
Incubation time [days]
Figure 1. Dechlorination of PCP to phenol in a nonsterile and unadopted sediment slurry
inoculated with cells harvested from a 4-BP dehalogenating enrichment culture.
174
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Stimulating the Miffobiol Dedilorination of PCBs; Overcoming Limiting Factors
John F. Quensen, III, Stephen A. Boyd, and James M. Tiedje
Michigan State University, East Lansing, Ml
John E. Rogers
U.S. Environmental Protection Agency, Environmental Research Laboratory, Athens, GA
Introduction
The discovery that polychlorinated biphenyls (PCBs) can be red actively dechlorinated by
microorganisms under anaerobic conditions has stimulated interest in the development of a
sequential anaerobic/aerobic biotreatment process for their destruction. While the aerobic
degradation of PCBs is generally limited to congeners with four or fewer chlorines, the anaerobic
process can dechlorinate more highly substituted congeners, producing products that are
aerobically degradable. Indeed, all products from the anaerobic dechlorination of Aroclor 1254
(1) have been shown to be aerobically degradable by one or more strains of aerobic bacteria
(2). Also, the high proportion of monochlorinated biphenyls that can accumulate as a result of
anaerobic PCB dechlorination may serve to induce PCB-degrading enzymes in aerobic
microorganisms (3). More highly chlorinated congeners can be aerobically co-metabolized but
are not inducing substrates (4).
A greater understanding of the factors controlling the anaerobic dechlorination of PCBs is
necessary before a successful sequential anaerobic/aerobic biotreatment process can be
developed for PCBs. In particular, how to stimulate more rapid and complete PCB
dechlorination in areas where the natural rate and/or extent of dechlorination is limited is
important to determine. The general approach we have taken is to identify site-specific factors
limiting in situ PCB dechlorination, then to apply treatments to alleviate the limitation(s). During
the first year of this project, our research focused on enhancing the dechlorination of PCBs in
soil and in sediments from the River Raisin in Michigan.
Drag Strip Soil Experiment
Factors most likely limiting PCB dechlorination in soils are a high redox potential, lack of
available organic carbon availability, and absence of PCB-dechlorinating microorganisms. To
determine how to alleviate limitations due to these three factors, we conducted an experiment
with PCB-contaminated drag strip soil from Glens Falls, New York. Alternate means tested for
achieving low redox conditions were to use a chemical reductant (Na2S) or to provide carbon
so that microbial activity would consume all oxygen present. The effectiveness of defined and
complex carbon sources were compared. Methanol was chosen as a defined carbon source
because it has been shown to enhance microbial dechlorination of PCBs (5). Trypticase soy
broth (TSB) was used because it is a complex carbon source used for the general culture of
anaerobic microorganisms. Inocula consisted of PCB-dechlorinating microorganisms eluted
from upper Hudson River sediments.
1994 Symposium on Roramediation of Hazardous Wastes 175
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Materials and Methods
The procedure followed was to first weigh 2 g of sieved soil into each anaerobic culture tube.
Depending on the treatment, sterile liquid medium or inoculum (10 ml) was then added while
flushing with O2-free N2:CO2 (80:20). Sterile (autoclaved) nonreduced media consisted of: 1)
minimal salts, 2) minimal salts plus 0.1 percent methanol, or 3) minimal salts plus 0.1 percent
TSB. Sterile reduced media consisted of these same three media but purged of oxygen with
nitrogen and amended with Na2S (0.24 g/L). All media were buffered at pH 7. Inocula were
prepared by eluting PCB-dechlorinating microorganisms from Hudson River sediments with each
of these six media. After adding the proper inoculum to each tube, the tubes were sealed with
Teflon-lined rubber stoppers and aluminum crimps. Controls were autoclaved 1 hrat 121°C.
Triplicate samples were analyzed every 4 weeks for 24 weeks. The entire contents of a culture
tube were extracted for each observation, and a congener-specific PCB analysis was performed
by capillary gas chromatography with electron capture detection.
To determine if the time required to achieve anaerobic conditions was related to the lag time
before dechlorination or to the subsequent extent of dechlorination, monitoring the redox of the
cultures was necessary. The redox indicator indigo disulfonate was added to parallel treatments
for this purpose, reduced to a colorless form at an Eh of -125 mV. The concentration of the
oxidized form was monitored photometrically during the first month of incubation.
Results
Dechlorination occurred only in inoculated treatments that received a carbon supplement
(methanol or TSB) (Figure 1). The lag time was slightly less (8 weeks) in the TSB/Na2S treatment
than in the other dechlorinating treatments (12 weeks), possibly because reduced conditions
were maintained more effectively (Figure 2). By the end of 24 weeks, about 0.69 and 0.62
meta plus para chlorines (m & p Cl) per biphenyl had been removed in the methanol and TSB
treatments without reductant (Na2S). The addition of Na2S and methanol gave more extensive
dechlorination (an average loss of 0.87 percent m & p Cl after 24 weeks) than methanol alone,
butNa2S did not stimulate further dechlorination with TSB. Thus, both inoculation and a carbon
supplement were necessary to initiate PCB dechlorination in this soil. Apparently, indigenous
microorganisms capable of PCB dechlorination were not abundant enough to express
dechlorination activity within the 24 weeks the experiment lasted.
It is interesting to note that the extent of dechlorination achieved in the inoculated treatments was
not simply related to the rate at which reducing conditions were achieved, as indicated by the
reduction of indigo disulfonate (Figure 2). Whether or not Na2S was used, the inoculated
methanol treatments took significantly longerto reduce all of the indigo disulfonate than the TSB
treatments did. Without Na2S, the same extent of dechlorination was achieved with each carbon
source, but with Na2S greater dechlorination was achieved with methanol.
River Raisin Sediment Experiment
We are conducting a similar experiment to determine the minimal amount of manipulation
necessary to dechlorinate PCBs in River Raisin sediments collected near Monroe, Michigan. In
a previous research project, we found that little in situ dechlorination of PCBs had occurred in
these sediments. PCB-dechlorinating microorganisms, however, exist in the sediments, the
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sediments support dechlorination in laboratory assays, and the PCBs are bioavailable because
they were dechlorinated under conditions of our treatability assay. In fad, individual congeners
in the contaminated sediment decreased 30 percent to 70 percent in 24 weeks at rates nearly
identical to rates for the same congeners freshly spiked into noncontaminated sediments. The
treatability assay was conducted using air-dried River Raisin sediments. They were slurried with
an equal weight of air-dried non-PCB-contaminated sediments and reduced anaerobic mineral
medium (RAMM). The slurry was then inoculated with microorganisms eluted from Hudson River
sediment, and 2',3,4-trichlorobiphenyl (2-34-CB) in a small volume of acetone was added to
a concentration of 500 fig/g sediment. The noncontaminated sediment was added to provide
a source of undefined nutrients. The medium included essential minerals and the chemical
reductant (Na2S) to lowerthe initial redox potential. Inoculation assured that PCB-dechlorinating
microorganisms were present. The 2-34-CB was added because the addition of a single PCB
(or poly bromi noted biphenyl) can somehow "prime" the dechlorination of PCBs already present
in a contaminated sediment (6).
The question thus becomes, What aspects of our treatability assay are necessary to dechlorinate
the PCBs present in the River Raisin sediments? We are conducting separate experiments with
wet and air-dried River Raisin sediments to answer this question. With the air-dried sediments,
the factors being considered are: 1) addition of 2-34-CB, 2) addition of the mineral salts in
RAMM, 3) addition of Na2S, and 4) addition of the non-PCB-contaminated sediments. All
treatments with the air-dried sediments were inoculated with microorganisms eluted from Hudson
River sediments. These same four factors are also being addressed in the experiment with wet
(i.e., never air-dried) River Raisin sediments. In this case, the necessity of inoculating with
Hudson River microorganisms is also being tested. These experiments are still in progress, and
data are not yet available.
References
1. Quensen, J.F., III, S.A. Boyd, and J.M. Tiedje. 1990. Dechlorination of four
commercial polychlorinated biphenyl mixtures (Aroclors) by anaerobic microorganisms
from sediments. Appl. Environ. Microbiol. 56:2,360-2,369.
2. Bedard, D.L., R.E. Wagner, MJ. Brennan, M.L Haberl, and J.F. Brown, Jr. 1 987.
Extensive degradation of Aroclors and environmentally transformed polychlorinated
biphenyls byA/ca/igenes eufrophus H850. Appl. Environ. Microbiol. 53:1,094-1,102.
3. Masse, R., F. Messier, L. Peloquin, C. Ayotte, and M. Sylvestre. 1984. Microbial
biodegradation of 4-chlorobiphenyl, a model compound of chlorinated biphenyls. Appl.
Environ. Microbiol. 41:947-951.
4. Furukawa, K., F. Matsumura, and K. Tonomura. 1978. Alca//genes and Ac/nefobacfer
strains capable of degrading polychlorinated biphenyls. Agric. Biol. Chem.
42:543-548.
5. Mies, L., and T.M. Vogel. 1990. Effects of organic substrates on dechlorination of
Aroclor 1242 in anaerobic sediments. Appl. Environ. Microbiol. 56:2,612-2,617.
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6. Bedard, D.L., H.M. Van Dort, R.J. May, KA DeWeerd, J.M. Principe, and LA Smullen.
1992. Stimulation of dechlorination of Aroclor 1260 in Woods Pond sediment. In:
General Electric Company research and development program for the destruction of
PCBs, 11th progress report. Schenectady, NY: General Electric Corporate Research
and Development, pp. 269-280.
w
0)
O
6
CD
cd
SH
2.0-
1.5-
1.0-
o—o Autoclaved
•—• Unamended
a—B Reductant
•—• Methanol
A—A TSB
Reductant Methanol
Reductant TSB
0
-I 1 1 r
5
10 15 20
Incubation Time (Weeks)
25
Figure 1. Dechlorination of PCBs in drag strip soil expressed as the decrease in the average
number of mefa and para chlorines over time.
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50
I 40^
CO
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O
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Potential Surfactant Effects on the Microbial Degradation of Organic Contaminants
Stephen A, Boyd, John F. Quensen, III, Mahmoud Mousa, and Jae Woo Park
Michigan State University, East Lansing, Ml
Shaobai Sun and William Inskeep
Montana State University, Bozeman, MT
Introduction
The biodegradation of poorly water soluble compounds in soil or sediment systems is believed
to be limited by low bioavailability due to strong sorption of the compounds to natural organic
matter (1-4). The use of surfactants to increase aqueous concentrations of these types of
compounds, and therefore their bioavailability, has often been suggested as a way of
overcoming this problem (5,6). Significant solubilization of the target compounds, however,
usually occurs only above the critical micelle concentration (CMC) of the surfactant, a
concentration often toxic or inhibitory to bacteria (7).
Petroleum sulfonate oil (PSO) surfactants are different from conventional surfactants in that they
form stable microemulsions in water rather than micelles, thereby enhancing solubilization at low
concentrations without apparent toxic effects to bacteria (5,8). We recently reported a 60-fold
decrease in the apparent soil sorption coefficient (K*) of 2,2',4,4',5,5'-hexachlorobiphenyl at
a PSO aqueous concentration of only 30 ppm, and a 200-fold decrease in K* at a 1 70 ppm
PSO (4). We therefore propose to investigate the use of this class of surfactants in enhancing
the anaerobic microbial dechlorination of polychlon'nated biphenyls (PCBs).
Although conventional surfactants are ineffective at enhancing HCH solubility at concentrations
below the CMC, evidence exists for stimulatory effects on biodegradation of aromatic
hydrocarbons in soils even when surfactant-induced disassociation from soil was not significant,
i.e., at concentrations below the CMC (9). For example, mineralization of phenanthrene was
substantially enhanced in a muck soil in the presence of 10/ig of nonionic surfactant per gram
of soil (10 ppm). Similar effects on biphenyl mineralization were not observed, and surfactant
concentrations of 100 ppm were either less stimulatory or inhibited mineralization.
A few reports indicate that sub-CMC concentrations of surfactants may enhance anaerobic
dechlorination of aromatic compounds. Dechlorination of pentachlorobenzene in sediment
slurries was stimulated by Tween 80 concentrations of 0.06 fig/ml to 100 jug/ml and SDS
concentrations of 0.3 /*g/mL to 40 fig/ml (10), while Tween 80 at concentrations below the
CMC slightly enhanced the dechlorination of hexachlorobenzene (11). Triton X-705 at 600
ppm decreased the lag time before PCB dechlorination took place in Hudson River sediment
slurries but did not affect the subsequent rate (12). Concentrations of other surfactants tested
(sodium dodecyl benzene sulfonate, Triton X-l 00, and X-045) were all at or above their CMCs
and inhibited dechlorination.
1994 Symposium on Roremediution of Hazardous Wastes
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Because these secondary stimulatory effects can occur at surfactant concentrations below the
CMC, they do not appear to be related to contaminant solubility enhancement. We are
attempting to establish the stimulatory effects on PCB dechlorination of surfactant concentrations
below the CMC for major types of nonionic, anionic, and PSO surfactants (Table 1) and to
attribute these effects to either solubility enhancement or secondary mechanisms. The
physiological or physical nature of such secondary mechanisms is being investigated.
Results
The surfactants used in this study are listed in Table 1. These include several nonionic
surfactants that were selected to provide a range of CMC values, and because previous studies
have shown thatthey provide beneficial effects on biodegradation as described above. We have
also included a twin-head anionic surfactant to minimize surfactant sorption to soils.
One of the major objectives of this research is to evaluate the effectiveness of sub-CMC
concentrations of surfactants in increasing the rate and extent of PCB dechlorination. To
determine what the exact aqueous phase concentration in soil- or sediment-water slurries is and
whether this concentration is above or below the CMC, we need to measure surfactant sorption
(i.e., obtain sorption isotherms) by the soils and sediments. To accomplish this, we will use a
batch equilibration technique, where the amount sorbed is determined from the difference
between the initial (added) and final (after sorption) aqueous phase surfactant concentrations.
The following three methods for measuring aqueous phase surfactant concentrations have been
evaluated: 1) tensiometer, 2) UV-absorption, and 3) total organic carbon. Sorption isotherms
developed using Method 1 indicated higher surfactant uptake by sediment then those obtained
using Methods 2 and 3. We suspect that the presence of dissolved or suspended organic matter
from the sediment may be influencing the surface tension measurement, and hence we have
elected not to use this method. Methods 2 and 3 resulted in essentially identical sorption
isotherms for Triton X-100 by Hudson River sediments. Method 3 is universally applicable to
all the surfactants listed in Table 1, whereas Method 2 is only applicable to surfactants with the
appropriate UV absorption properties. Hence, Method 3 is currently being used to obtain
sorption isotherms for all the surfactants listed in Table 1. This information will quantitate the
aqueous phase surfactant concentrations in our sediment slurries and determine whether these
are above or below the CMC.
To separate solubility enhancement effects of surfactants (which could increase bioavailability
and hence biodegradation rates) from the secondary effects of surfactants on biodegradation
rates, we are evaluating the sorption of PCBs in sediment-water-surfactant systems above and
below the CMC. We have now observed the effect of Triton X-100 on the sorption of
2,2',4/4',5,5'-PCB by soil by measuring the apparent sorption coefficient K* at different aqueous
surfactant concentrations (CJ. At Ch values below 200 ppm (approximately the CMC of Triton
X-100), K* values increased from ~ 500 to 1,200 with increasing surfactant concentration.
This is because in this concentration range, the added surfactant is strongly sorbed by soil, and
the soil-bound surfactant in turn enhances PCB sorption. At higher C^s (above the CMC), K*
decreases rapidly and substantially due to the formation of surfactant micelles in solution that
effectively dissolve PCBs and raise the apparent aqueous phase PCB concentration. These
preliminary results strongly suggest that the enhanced contaminant biodegradation rates
observed previously at low (below the CMC) surfactant concentrations are not due to increased
bioavailability associated with solubility enhancement effects. Thus, other indirect or secondary
181
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effects may be responsible for the stimulating biodegradation rates at surfactant levels below the
CMC. These mechanisms will be investigated in the future.
References
1. Ogram, A.V., R.E. Jessup, L.T. Lou, and P.S.C. Rao. 1985. Effects of sorption on
biological degradation rates of 2,4-dichlorophenoxy acetic acid in soil. Appl. Environ.
Microbiol. 49:582-587.
2. Steen, W.C., D.F. Paris, and G.L. Baughman. 1980. Effects of sediment sorption on
microbial degradation of toxic substances. In: R.A. Baker, ed. Contaminants and
sediments, Vol. 1. Ann Arbor, Ml: Ann Arbor Science, pp. 447-482.
3. Weissenfels, W.D., HJ. Klewer, and J. Langhoff. 1992. Adsorption of polycyclic
aromatic hydrocarbons (PAHs) by soil particles: Influence on biodegradability and
biotoxicity. Appl. Microbiol. Technol. 36:689-696.
4. Guerin, W.F., and S.A. Boyd. 1992. Differential bioavailability of soil sorbed
naphthalene to two bacterial species. Appl. Environ. Microbiol. 58:1,142-1,152.
5. Sun, S., and S.A. Boyd. 1993. Sorption of nonionic organic contaminants in soil-water
systems containing petroleum sulfonate-oil surfactants. Environ. Sci. Technol. 27:1,340-
1,346.
6. Laha, S., and R.G. Luthy. 1991. Inhibition of phenanthrene mineralization by nonionic
surfactants in soil-water systems. Environ. Sci. Technol. 25:1,920-1,930.
7. Kile, DA, and C.T. Chiou. 1989. Water solubility enhancement of DDT and
trichlorobenzene by some surfactants above and below the critical micelle concentration.
Environ. Sci. Technol. 23:832-838.
8. Kile, D.T., C.T. Chiou, and R.S. Helburn. 1990. Effects of some petroleum sulfonate
oil surfactants on the apparent water solubility of organic compounds. Environ. Sci.
Technol. 24:205-208.
9. Aronstein, B.N., Y.M. Calvillo, and M. Alexander. 1991. Effect of surfactants at low
concentrations on the desorption and bioavailability of sorbed aromatic compounds in
soil. Environ. Sci. Technol. 25:1,728-1,731.
10. Mousa, M.A., and J.E. Rogers. 1993. Enhancement of pentachlorobenzene
dechlorination by surfactant addition. Abstract Q-155. Presented at the 93rd General
Meeting of the American Society for Microbiology, Atlanta, GA.
11. Van Hoff, P.L., and C.T. Jafvert. 1991. Influence of nonionic surfactants on
hexachlorobenzene degradation. Abstract 498. Presented at the 12th Annual Meeting
of the Society of Environmental Chemistry and Toxicology, Seattle, WA.
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12. Ambramowicz, DA, MJ. Brennan, H.M. Van Dort, and E.L Gallagher. 1993. Factors
influencing the rate of polychlorinated biphenyl dechlorination in Hudson River
sediments. Environ. Sci. Technol. 27:1,125-1,131.
13. Schick, MJ. 1966. Nonionic surfactants. New York, NY: Marcel Dekker.
14. Rouse, J.D., and D.A. Sapatini. 1993. Minimizing surfactant losses using twin-head
anionic surfactants in subsurface remediation. Environ. Sci. Technol. 27:2,072-2,078.
Table 1. Surfactants Proposed for This Study
Surfactant CMC (mg/L)
Triton X-l 00 130(7)
Triton X-405 620 (7)
Triton X-705 625
TweenSO 13(13)
Alfonic810-60 275(9)
C14DPDS (Dowfax 8390) (14) 4,000
Petroleum sulfonate oil NA
NA = not applicable. These products form stable microemulsions in water and do not exhibit
a CMC. They consist of petroleum sulfonate (61 percent to 63 percent) and mineral oil (33
percent).
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Enhanced Dechlorination of PCBs in Contaminated Sediments by Addition of
Single Congeners of Chloro- and Bromobiphenyls
W. Jack Jones and John E. Rogers
U.S. Environmental Protection Agency, Environmental Research Laboratory, Athens, GA
Rebecca L Adams
Technology Applications, Inc., Athens, GA
Introduction
Bioremediation has been suggested as a technology that may be useful to decrease the level
of pollutants at contaminated sites. Forpolychlorinated biphenyl (PCB) contaminated sediments,
reductive dechlorination reactions (anaerobic) preferentially transform the more highly
chlorinated PCB congeners to less chlorinated derivatives, which are more amenable to aerobic
degradation. In this instance, the anaerobic and subsequent aerobic processes are
complementary and result in a reduction of toxic (higher chlorinated, coplanar) PCB congeners
and possibly the biological destruction of PCBs through subsequent aerobic oxidations. Before
using this method at a remediation site, it is important to assess the ability of indigenous
microorganisms from the site to transform the pollutants, to understand factors that control the
dechlorination reactions, and to develop techniques to enhance microbial activities.
PCB transformation in anaerobic environments, such as sediments of lakes and rivers, could be
inferred in the mid-1980s from the studies of Brown and coworkers (1). These investigators
noted that historically contaminated sediments from the Hudson River exhibited an altered PCB
congener profile compared with the congener profile of the original contaminating Aroclor. The
alterations were characterized by a reduction in the concentration of the more highly chlorinated
PCB congeners, with selective or preferential removal of mefa and para chlorines, and an
increase in the concentration of the more lightly chlorinated and orf/io-substituted congeners.
Thus, dechlorination of the more highly chlorinated PCB congeners was proposed to be
catalyzed by anaerobic microorganisms residing in the contaminated sediment. The biologically
mediated reductive dechlorination of PCBs from contaminated sediments was subsequently
demonstrated in several laboratory investigations (2-4). In some studies, the microbial inoculum
was obtained by "washing" PCB-contaminated sediments with anaerobic medium and collecting
the supernatant (4,5). Recently, reductive dechlorination of PCBs was suggested to be enhanced
when PCB-contaminated sediments are amended with PCB mixtures (Aroclors) or specific
PCB/polybrominated biphenyl (PBB) congeners (6).
To date, only a limited number of studies have attempted to understand the factors that affect
the reductive dechlorination of PCBs in historically contaminated sediments. Abramowicz et al.
(7) reported that addition of inorganic nutrients enhanced reductive dechlorination of
endogenous PCBs in laboratory incubations of Hudson River sediments. In a recent study using
methanogenic sediment slurries contaminated with Aroclor 1260, Bedard and Van Dort (2)
reported that addition of bromobiphenyl congeners stimulated the reductive dechlorination of
endogenous (historical) PCBs. In an earlier study, Bedard and coworkers (8) reported that
amendment of Woods Pond sediment with a high concentration (approximately 1 mM) of either
184 1994 Symposium on Bioremediotion of Hazardous Wastes
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2,3',4',5-CB or 2,3,4,5,6-CB stimulated reductive dechlorination of endogenous PCBs and that
transformation of congeners with para chlorines was especially evident.
The primary objectives of this study were to determine the reductive dechlorination potential of
PCB-contaminated sediments from the Sheboygan and Ashtabula Rivers and to further test the
hypothesis that addition of PCB and PBB congeners enhances the reductive dechlorination of
endogenous (historical) PCBs by indigenous microbial populations.
Materials and Methods
PCB-contaminated sediments were collected from the Sheboygan River, near Sheboygan Falls,
Wisconsin, and from the Ashtabula River, near Ashtabula, Ohio. Grab samples of sediments
were collected from all sites. Initial gas chromatography data indicated that a significant shift
in the PCB congener profile had occurred since the time of PCB deposition, suggesting previous
reductive dechlorination activity.
Biotransformation experiments were prepared by combining one volume of PCB-contaminated
sediment with one volume of anoxic (N2 sparged) site water and the mixture was stirred for
approximately 5 min. Aliquots of the sediment slurry (equivalent to 5 g dry sediment) were
dispensed into amber serum vials, and the sediment slurry was amended with various
chlorobiphenyl or bromobiphenyl congeners dissolved in acetone. Initial experiments were
conducted with PCB-contaminated Sheboygan River sediment (approximately 180 ppm total
PCBs) and Ashtabula River sediment (approximately TOO ppm total PCBs) and were amended
with penta-, hexa-, hepta-, or octa-chlorobiphenyl congeners. Additional experiments were
performed with more heavily contaminated Sheboygan River sediment (approximately 1,000
ppm total PCBs) and were amended with either a di- or tetra-chlorobiphenyl congener or the
corresponding di- or tetra-bromobiphenyl congener (final concentration of 1 mM). Autoclaved
controls and nonautoclaved, unamended controls were also included in the study. Triplicate
samples were analyzed at 4- to 8-week intervals for congener-specific PCBs using capillary gas
chromatography and electron capture detection.
Results
Enhanced Dechlorination Using Specific PCB Congeners
Initial dechlorination experiments were conducted with PCB-contaminated (~180 ppm)
Sheboygan River sediments amended with 20 ppm to 80 ppm of 2,2',3,3',4,5,6,6'-
octachlorobiphenyl (octa-CB). The most prominent PCB homologues detected in the
contaminated Sheboygan River sediments were trichlorobiphenyls andtetrachlorobiphenyls. The
percentages of octa-CB remaining in the samples after anaerobic incubation for 8 months were
35 percent, 20 percent, and 10 percent, respectively, for sediments amended with 20/ig/g, 40
fig/g, and 80/^g/g. In all sediment experiments amended with octa-CB, there was a decrease
in the concentration of hepta-, hexa-, penta-, tetra- and tri-CB congeners and an increase in
the concentration of di- and mono-CB congeners. The mole percentage of mono-CBs was less
than 1 percent at the onset of the experiment (Figure 1 A, Week 1); after anaerobic incubation
for 30 weeks/ this homologue group accounted for approximately 8 percent of the total PCB
congeners in sediments amended with 20 mg/g of octa-CB (Figure 1B). The major products
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of reductive dechlorination were di-CB congeners; this homologue group increased from 2.5
to 40 mole percent after 30 weeks of incubation. The most prominent di-CB peak detected in
octa-CB amended sediments consisted of two orf/)o-substituted congeners (2,2'-CB and 2,6-
CB). Two additional homologue groups, tri- and tetra-CBs, initially accounted for approximately
80 percent of the total PCB homologues in the contaminated Sheboygan sediments but were
reduced to less than 50 percent of the total following 30 weeks' incubation in octa-CB amended
experiments. The average number of chlorines per biphenyl (total of endogenous plus amended
PCBs) decreased from 4.2 to 2.8 (± 0.1 ), 2.5 (± 0.3), and 2.2 (± 0.3), respectively, in
experiments amended with 20 ftg/g, 40 ftg/g, and 80/ig/g of octa-CB.
PCB-contaminated Ashtabula River sediments were spiked with 2,3,3/,4,4/-pentachlorobiphenyl
(penta-CB), 2,3,3',4,4',5-hexachlorobiphenyl (hepta-CB), or 2,2',3,4,5,6,6'-
heptachlorobiphenyl (hepta-CB) or combinations thereof and incubated anaerobically.
Dechlorination of the added congeners was observed after lag periods of 5, 4, and 3 months
for experiments amended with either the penta-CB, hepta-CB, orhexa-CB, respectively. Addition
of the chlorobiphenyl congeners singly or as mixtures resulted in enhanced reductive
dechlorination of endogenous PCB congeners in a manner similar to that observed for
Sheboygan River sediment amended with octa-CB. Appreciable decreases in the mole
percentages of endogenous PCB homologue groups (tetra-CB and penta-CB) were coupled with
increases in the mole percentages of mono-, di-, and tri-CB congeners. The average number
of chlorines per biphenyl decreased from approximately 5.2 to 2.7 in Ashtabula River sediments
amended with any of the three congeners tested. No significant changes in the distribution of
the PCB homologue groups were noted in control experiments.
Dechlorination in the Presence of PBB/PCB Congeners
Recently, experiments have been initiated to test the hypothesis that amendment of PBB
congeners enhances the dechlorination of PCBs in contaminated sediments. Highly contaminated
(1,100 ppm PCBs) sediments from the Sheboygan River were amended with dibromo- or
dichlorobiphenyl congeners, or with tetrachloro- or tetrabromobiphenyl congeners, and
dehalogenation was followed over the course of 6 months incubation. After 6 months of
incubation, no enhancement of dechlorination of endogenous PCBs has been detected in
sediments amended with 2,2',4,5'-tetrabromobiphenyl or 2,2',4,5'-tetrachlorobiphenyl
compared with controls. Both mefa and para denomination of the added 2,2',4,5'-PBB
congener, however, was evident after 1 month of incubation, with 2,2'-dibromobiphenyl
observed as the major product. Approximately 25 percent of the parent 2,2',4,5'-PBB remained
after 6 months' incubation. Dehalogenation of the amended 2,2',4,5'-PCB congener was more
rapid than debromination of the corresponding PBB congener; more than 70 percent of the
2,2',4,5'-PCB was transformed to 2,2',4-PCB after 1 month's incubation. As with the added PBB
congener, however, enhanced dehalogenation of the endogenous PCBs was not evident.
In a separate set of experiments, 2,4-, 2,5-, or 2,6-dibromobiphenyl or dichlorobiphenyl
congeners were added to PCB-contaminated Sheboygan River sediments. Greater than 85
percent of the amended 2,4- and 2,5-dibromobiphenyl were debrominated at the para and
mefa positions, respectively, within the initial 3 months of incubation. No evidence of
debromination of the amended 2,6-dibromobiphenyl was noted. Further, addition of the
dibromobiphenyl congeners has not yet had an effect on the extent of dechlorination of the
endogenous PCBs compared with controls. Of the dichlorobiphenyls examined, significant loss
(40 percent) of only 2,5-dichlorobiphenyl has been observed. Dechlorination at the mefa
chlorine was accompanied by an increase in 2-chlorobiphenyl. Although results are only
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preliminary/ there appears to be a moderate reduction in the average number of me/a plus para
chlorines for endogenous PCBs in this data set.
The results from the present study demonstrate the dechlorination capacity of PCB-contaminated
Sheboygan River and Ashtabula River sediments. No appreciable dechlorination of endogenous
PCBs was observed in unamended sediment slurries. Several explanations are proposed for the
stimulation of reductive dechlorination of endogenous PCBs in sediments by addition of specific
PCB congeners: 1) the bioavailability of PCBs was enhanced, thus providing an available
electron acceptor for oxidation reactions; 2) the growth of indigenous PCB dechlorinating
microorganisms was stimulated; or 3) amended PCB congeners induced dechlorinating activity
of indigenous microbial populations. Additional strategies should be considered for PCB
bioremediation and may include increasing the physical-chemical availability of PCBs bound to
sediments (for example, the addition of surfactants) or cycling between anaerobic and aerobic
conditions.
References
*
1. Brown, J.F., R.E. Wagner, H. Feng, D.L. Bedard, M.J. Brennan, J.C. Camahan, and R.J.
May. 1987. Environmental dechlorination of PCBs. Environ. Toxicol. Chem. 6:579-
593.
2. Bedard, D.L, and H.M. Van Dort. 1992. Brominated biphenyls can stimulate reductive
dechlorination of endogenous Aroclor 1260 in methanogenic sediment slurries.
Presented at the 92nd General Meeting of the American Society for Microbiology, p.
339.
3. Quensen, J.F., III, J.M. Tiedje, and S.A. Boyd. 1988. Reductive dechlorination of
polychlorinated biphenyls by anaerobic microorganisms from sediments. Science
242:752-754.
4. Quensen, J.F., III, S.A. Boyd, and J.M. Tiedje. 1990. Dechlorination of four
commercial polychlorinated biphenyl mixtures (Aroclors) by anaerobic microorganisms
from sediments. Appl. Environ. Microbiol. 56:2,360-2,369.
5. Nies, L, and T.M. Vogel. 1990. Effects of organic substrates on dechlorination of
Aroclor 1242 in anaerobic sediments. Appl. Environ. Microbiol. 56:2,612-2,617.
6. Van Dort, H.M., and D.L. Bedard. 1991. Reductive ortho- and mefa-dechlorination of
a polychlorinated biphenyl congener by anaerobic microorganisms. Appl. Environ.
Microbiol. 57:1,576-1,578.
7. Abramowicz, D.A., M.J. Brennan, and H.M. Van Dort. 1990. Anaerobic and aerobic
biodegradation of endogenous PCBs. In: General Electric Company research and
development program for the destruction of PCBs, 9th progress report. Schenectady,
NY: General Electric Corporate Research and Development, pp. 55-69.
187
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8. Bedard, D.L., S.C. Bunnell, and H.M. Van Dort. 1990. Anaerobic dechlorination of
endogenous PCBs in Woods Pond sediment. In: General Electric Company research
and development program for the destruction of PCBs, 9th progress report.
Schenedady, NY: General Electric Corporate Research and Development, pp. 43-54.
Sheboygan River Sediment (Control)
60.00
40.00
20.00
0.00
MONO 01 TRI TETRA PENTA HEXA HEPTA OCTA NONA DECA
HOMOL06 GROUPS
Sheboygan River Sediment + 20 PPM OCTA-CB
B
0.00
MONO Dl TRI TETRA PENTA HEXA HEPTA OCTA NONA DECA
HOMOLOG GROUPS
Figure 1. Profile of amended and endogenous PCB biotransformation in (A) unamended
control sediments and (B) 20 ppm 2/2',3,3//4,5/6/6'-octachlorobiphenyl (octa-CB)
amended sediments.
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Effect of Heavy Metal Availability and Toxichy on Anaerobic Transformations of
Aromatic Hydrocarbons
John H. Pardue, Ronald D. DeLaune, and William H. Patrick, Jr.
Wetland Biogeochemistry Institute, Louisiana State University, Baton Rouge, LA
Introduction and Background
The existence of co-contaminants (e.g., heavy metals and toxic organics) in impacted sediments
has created concern over the potential for biodegradation to assist in remediating these sites.
Heavy metals can be inhibitory to microorganisms and microbial processes, including the
decomposition of organic matter and other biogeochemical processes (1). The characteristics
of this inhibition for biodegradation of toxic organics are poorly understood because of the large
number of variables involved. This study was initiated to determine the effect of heavy metals on
reductive dechlorination of chlorinated aromatic organics. Experiments are being conducted with
two metals, cadmium (a single valence [+2] transition metal) and chromium (a multivalence [+6
and +3] transition metal), and two chlorinated aromatics, hexachlorobenzene (HCB) and 2,3,4-
trichloroaniline (2,3,4-TCA). The reductive dechlorination of these compounds has been
demonstrated, and the degradation pathways are generally understood (2,3).
The interactions between metals and organic-degrading microbes or consortia are complex
because the observed effects are largely a function of the bioavailability of both the metals and
the organic compound. Studies have been conducted on aerobic biodegradation processes
(4,5), but inhibition of anaerobic biodegradation is not understood. At present, the best
information indicates that the soluble fraction of the co-contaminants is the "available" fraction
to the microorganisms (6). Under anaerobic conditions, metals may be precipitated as sulfides
or present as reduced forms of lower toxicity. Solubility and speciation of metals is strongly
dependent on the redox potential and pH of the sediment. An excellent example is the solubility
of chromium, which exists in two valence states with large differences in solubility—CrfVI) and
Cr(lll)—depending on the redox potential of the sediment (7). Other metals with single valence
states (e.g., Cd2"1", Zn2+) adsorb onto redox-sensitive surfaces (e.g., iron and manganese oxides)
and form various complexes under different redox conditions.
Results and Discussion
Experiments are being conducted to determine the effect of cadmium on reductive dechlorination
of 2,3,4-TCA in previously uncontaminated anaerobic freshwater sediment environments,
including a rice paddy soil, a cypress swamp soil, a bottomland hardwood soil, and a freshwater
marsh soil. These soils differ widely in sediment properties, including the organic matter
concentration, which ranges from 2.9 percent in the rice paddy soil to 74 percent in the
freshwater marsh. 2,3,4-TCA is a particularly useful model compound because chlorine
substituents are present atorfho, meta, and para chlorine positions. Representative results from
several soils are discussed here. Microcosms, with continuous monitoring of the Eh and pH,
were constructed using sediment slurries under anaerobic conditions. Sediments were amended
with 2,3,4-TCA (200 mg/kg soil) and varying concentrations of Cd2+ (control, 10 mg/kg soil,
1994 Symposium on Roremediation of Hazardous Wastes 189
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100 mg/kg soil, and 1,000 mg/kg soil). Periodically, subsamples of microcosms were removed
for quantification of metals and 2,3,4-TCA. Gas chromatography/mass spectrometry was used
to identify lower chlorinated aniline metabolites.
Degradation of 2,3,4-TCA in rice paddy soil is presented in Figure 1. Data are from
representative replicates. When no Cd was added, dechlorination proceeded rapidly by removal
of the ortrio chlorine to form 3,4-dichloroaniline (3,4-DCA). 3,4-DCA appeared only transiently
and was rapidly dechlorinated to 3-chloroaniline (3-CA). No further dechlorination was
observed. When 10 mg/kg Cd was added, dechlorination also proceeded rapidly but by the
removal of the para chlorine to form 2,3-DCA. Two monochloroanilines (2-CAand 3-CA) were
subsequently formed in nearly equal amounts. When cadmium was added at higher
concentrations (100 mg/kg and 1,000 mg/kg), no dechlorination was observed. Daily mass
balance of chloroanilines for the microcosms in Figure 1 averaged 103 percent ± 33 percent.
This general trend has also been observed in the cypress swamp soil and freshwater marsh soil,
despite wide differences in the degree of sorption of metals and organics in these soils. Studies
are ongoing in the fourth soil (bottomland hardwood soil). The observed pattern is ortho
dechlorination when no cadmium is added, para dechlorination when a critical level of
cadmium is reached, and complete inhibition at another critical level of cadmium. The trend
is poorly predicted by the total concentration of cadmium but appears to be well predicted by
"soluble" cadmium (measured as porewater cadmium passing through a 0.45-mm filter). Of
the three soils examined, orfho dechlorination occurred when soluble cadmium concentrations
ranged from less than 20 mg/L to 32 mg/L. Para dechlorination occurred when soluble
cadmium concentrations ranged from 0.15 mg/L to 0.2 mg/L. Complete inhibition occurred
when soluble cadmium concentrations ranged from 0.2 mg/L to 7.4 mg/L. Further experimental
replication may refine these ranges more accurately. These results are surprising in light of
differences in pore water chemical composition between these flooded soils. MINTEQ, a
geochemical speciation model, is being used to estimate concentrations of cadmium complexes,
which may shed further light on these results.
Preliminary batch studies have also been performed to determine the effect of Cr(VI) on 2,3,4-
TCA dechlorination in the bottomland hardwood soil. Results indicate that Cr(VI) additions affect
the dechlorination of 2,3,4-TCA by increasing the lag time necessary for degradation to occur
(Figure 2). Addition of CrfVI) at 20 M, 50 M, 75 M, and 175 M all increased the lag time for
dechlorination from approximately 2 weeks to 10 weeks. Following the lag time, apparent rates
of dechlorination of 2,3,4-TCA were unaffected by the initial chromium addition.
Biogeochemistry of chromium in the bottomland hardwood soil has been previously investigated
(7). Addition of CnVI) under low Eh conditions is followed by rapid (< 1 min) reduction to
Cr(lll), followed by precipitation/sorption of Cr(lll) from the soil solution. A critical Eh for the
reduction process has been identified, +300 mV, below which the reaction proceeds rapidly.
In the batch study (Eh = -200 mV), Cr(VI) was undetectable in solution (detection limit 5 ppb)
immediately following addition, and only low concentrations of Cr(lll) (< 50 ppb) were detected.
Methanogenesis, as indicated by the accumulation of CH4 in the vial headspace, was unaffected
by additions of Cr(VI). The mechanism by which chromium inhibits dechlorination is unclear,
although results suggest an initial toxic effect on the degrading population that requires time to
overcome (lengthening lag time). This effect could be direct (mortality of some microbial
population) or indirect (oxidation of some key redudant crucial to dechlorination).
190
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References
1. Capone, D.G., D.D. Reese, and R.P. Kiene. 1983. Effect of metals on
methanogenesis, sulfate reduction, carbon dioxide evolution, and microbial biomass in
anoxic salt marsh sediments. Appl. Environ. Microbiol. 45:1,586-1,591.
2. Kuhn, E.P., and J.M. Suflita. 1989. Sequential reductive dehalogenation of
chloroanilines from a methanogenic aquifer. Environ. Sci. Technol. 23:848-852.
3. Fathepure, B.Z., J.M. Tiedje, and S.A. Boyd. 1988. Reductive dechlorination of
hexachlorobenzene to tri- and dichlorobenzenes in anaerobic sewage sludge. Appl.
Environ. Microbiol. 54:327-330.
4. Said, W.A., and D.L. Lewis. 1991. Quantitative assessment of the effects of metals on
microbial degradation of organic chemicals. Appl. Environ. Microbiol. 57:1,498-
1,503.
5. Springael, D., L. Diets, L. Hooyberghs, S. Kreps, and M. Mergeay. 1993. Construction
and characterization of heavy metal-resistant haloaromatic-degrading Alcal/genes
eufroph/s strains. Appl. Environ. Microbiol. 59:334-339.
6. Duxbury, T. 1985. Ecological aspects of heavy metal responses in microorganisms.
Adv. Microbiol. Ecol. 8:185-235.
7. Masscheleyn, P.M., J.H. Pardue, R.D. DeLaune, and W.H. Patrick, Jr. 1992.
Chromium redox chemistry in a lower Mississippi valley bottomland hardwood wetland.
Environ. Sci. Technol. 26:1,217-1,226.
191
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ti
o
• I— t
-+J
cd
CD
O
£
O
o
CD
cd
O
o
250
200 -
150 -
Control
O 2,3,4-trichloroaniline
• 3,4-dichloroaniline
V 3-chloroaniline
100 -
Days
250
Cd (10 mg/kg)
r i i
O 2,3,4-trichloroaniline
• 2,3-dichloroaniline
v 3—chloroaniline
T 2 —chloroaniline
12
15
18
21
Days
Figure 1. Dechlorination of 2,3,4-trichloroaniline in a control (no cadmium added) and
cadmium amended (10 mg/kg soil) microcosm constructed from a rice paddy soil
(Crowley silt loam). Soluble cadmium was < 20/tg/Lforthe control and 0.19 mg/L
for the cadmium-amended microcosm. Soil Eh ranged from -200 to -250 mV.
192
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Biodegradation of Petroleum Hydrocarbons in Wetlands Microcosms
Rochelle Araujo and Marirosa Molina
U.S. Environmental Protection Agency, Environmental Research Laboratory, Athens, GA
Dave Bachoon
Department of Microbiology, University of Georgia, Athens, GA
Lawrence D. LaPlante
Technology Applications, Inc., Athens, GA
Introduction
In the aftermath of several major environmental oil spills, it became apparent that spill
preparedness did not include an up-to-date inventory of bioremediation strategies or adequate
methods for assessing the efficacy of bioremediation under field conditions. Thus, field trials of
bioremediation (1) preceded rigorous laboratory- and pilot-scale experimentation. A
cooperative effort to develop protocols for evaluating bioremediation strategies has led to the
adoption of a system of tiered assays for determining efficacy and environmental toxicity of
products that might be applied to spilled oil. Protocols include 1) analytical methods for
determining the extent of biodegradation, 2) toxicity assays for aquatic and sediment organisms,
3) flask experiments to determine potential for biodegradation, 4) and laboratory-scale
microcosms for assessing the potential for degradation in prototype environments, including
open water, beaches, and wetlands.
Development and Testing of Microcosm Protocols
Oil extraction, refining, and transshipment facilities are often located in coastal regions, putting
wetlands ecosystems at risk for exposure to spilled oil. The inaccessibility of sites and the fragile
nature of the ecosystems preclude mechanical cleanup of oil, making bioremediation a preferred
option for wetlands. Moreover, the high level of indigenous microbial activity suggests a
potential for biodegradation, especially if environmental nutrient limitations can be relieved by
fertilizer additions.
Results
Sediment microcosms were constructed from homogenized marsh sediments from Sapelo Island,
Georgia, and were flushed on a tidal basis with seawater adjusted to the salinity of the collection
site (20%o). The tidal cycle was continued until a clear boundary distinguished the aerobic and
anaerobic layers (3 mm to 5 mm) of the microcosm. Then, oil (521 fraction of Alaska North
Slope crude, 0.5 mm depth/3.93 mL) was applied to the sediment surface. The numbers of
hydrocarbon-degrading bacteria in the sediment prior to construction of the microcosms was
in the range 103 to 104 cells/g, which is consistent with nonpristine coastal areas (2). Products
to be tested were applied 1 day after the application of oil. The types of products submitted for
testing in protocol development included microbial cultures, nutrients, surfactants, sorbents, and
combinations thereof.
194 1994 Symposium on Bioremediation of Hazardous Wastes
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Figure 1 shows the composition of the oil, as determined by gas chromatography/mass
spectrometry, after a 6-week incubation. The alkane constituents of the oil (Figure 1A) were
appreciably degraded in all treatments relative to the original composition of the oil. The
degradation in the nutrient treatment (Product D) was slightly greater than in the nonfertilized
control. The addition of nutrients plus microbial inoculum (Product J) resulted in significant
degradation of the full range of alkanes (CIS to C35); that degradation was primarily biological
is indicated by the reduced ratios of Cl 7:pristane and Cl 8:phytane. Nevertheless, pristane and
phytane were reduced in concentration, indicating that they are also subject to biodegradation,
although at a slower rate than the normal alkanes. Thus, oil constituents that are more resistant
to biological degradation than are pristane and phytane are more suitable for use as internal
indices in longer incubation experiments; both hopanes (3) and C2-chrysenes have been
proposed for this application.
Degradation of the aromatic constituents of oil was negligible; only the naphthalene series
differed in concentration between treated microcosms and controls after 6 weeks' incubation.
The lack of degradation of aromatics in the continued presence of alkane constituents suggests
that degradation of the two classes of compounds may be sequential, although Foght et al. (4)
concluded that degradation of aliphatics and aromatics could occur concurrently if adapted
organisms are present. To test whether alkane degradation goes to completion before the onset
of degradation of aromatics, the length of the microcosm incubation period in subsequent
experiments was increased from 6 weeks to 3 months.
Factors Influencing the Persistence of PAHs in Sediments
In light of the relative degradability of the alkane constituents of petroleum and the toxicity and
carcinogenicity associated with the more recalcitrant polycyclic aromatic hydrocarbons (PAHs),
the effectiveness of a remediation effort in reducing ecological risk depends largely on the
degree to which the latter are degraded. Moreover, PAHs of industrial origin are of
environmental concern as soil and sediment contaminants in their own right. Thus, the
persistence of PAHs in the microcosms can be considered a shortcoming of bioremediation
measures.
Several explanations have been proposed to explain the persistence of PAHs in the environment.
Intrinsic controls on the rate of degradation include low solubility, toxicity, and interactions
between PAH compound classes; extrinsic controls include environmental factors such as salinity,
temperature, nutrient concentrations, and interactions between PAHs and other classes of
compounds, including natural organic matter. Interactions between PAHs and other compounds
may include co-metabolism, the competitive utilization of alternative substrates, orthe absence
of required inducer compounds.
Bauer and Capone (5) noted that preexposure of marine sediments to single PAHs enhanced
subsequent degradation of those compounds and that cross acclimation occurred between select
PAHs. Similarly, Kelley and Cerniglia (6) reported an interaction between fluoranthene and
pyrene and concluded that the catabolism of fluoranthene, pyrene, and phenanthrene was
catalyzed by a common enzyme system. Other researchers (7) observed that a mixed microbial
community was required for the complete utilization of some PAHs.
195
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Results
We tested the interactions between PAHs of different size classes to determine if interactions
between PAHs were responsible for the persistence of those compounds in sediments. The
presence of other PAHs, either grouped by size classes or as a mixture of 16 compounds, did
not affect the mineralization of pyrene by an acclimated microbial culture introduced into
sediment slurries with inorganic nutrients (Figure 2A). The same culture degraded pyrene more
slowly when four-, five- and six-ring PAHs were present in mineral medium enriched with
sediment organic extract (Figure 2B), and did not degrade pyrene at all when five- and six-ring
PAHs were present in mineral medium (Figure 2Q. We concluded that large PAHs are inhibitory
to the activity of organisms capable of degrading pyrene, but that the inhibition is removed when
the high molecular-weight compounds are sorbed to sediments or complexed with organic
matter. Toxicity due to large PAHs, therefore, probably did not explain the persistence of PAHs
in the microcosm trials.
Sediments that were inoculated with a culture that had not been recently exposed to PAHs
adapted to degrade pyrene after a lag of 1 day, unless protein synthesis was inhibited with
chloramphenicol (Figure 3). When the culture was preexposed to pyrene, the addition of
chloramphenicol did not appreciably inhibit degradation upon subsequent exposure. Similarly,
the antibiotic did not inhibit degradation of pyrene by a culture preexposed to phenanthrene,
although protein synthesis was necessary for pyrene degradation by cultures preexposed to
naphthalene. Therefore, we concluded that the cells shared a common enzyme system for
phenanthrene and pyrene, and another for naphthalene.
Ongoing Research
Current research includes the isolation and characterization of a Mycofaocferium sp. capable of
degrading pyrene as a sole carbon source. The isolate will be introduced into the mixed
microbial community of the sediment microcosm to assess survival and impact on the
degradation of PAHs. The microbial diversity in impacted and nonimpacted sediments will be
assessed by whole genome hybridization, and specific probes will be used to compare the
activities of oil degraders and lignocellulose degraders under various nutrient and surfactant
treatments.
References
1. Pritchard, H.P., and C.F. Costa. 1991. Environ. Sci. Technol. 25:372-379.
2. Munkin-Phillips, G.J., and J.E. Stewart. 1973. Distribution of hydrocarbon-utilizing
bacteria in Northwestern Atlantic waters and coastal sediments. Can. J. Microbiol.
20:955-962.
3. Prince, R.C., D.E. Elmendorf, J.R. Lute, C.S. Hsu, C.E. Haith, J.D. Senius, G.J. Dechert,
G.S. Douglas, E.L Butler. 1994. 17a(H),21/?(H)-Hopane as a conserved internal
marker for estimating biodegradation of crude oil. Environ. Sci. Technol. 28:142-145.
196
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4. Foght, J.M., P.M. Fedorak, and D.W.S. Westlake. 1989. Mineralization of
[14C]hexadecane and [^Qphenathrene in crude oil: Specificity among bacterial isolates.
Can. J. Microbiol. 36:169-175.
5. Bauer, J.E., and D.G. Capone. 1988. Effects of co-occurring aromatic hydrocarbons
on degradation of individual polycyclic aromatic hydrocarbons in marine sediment
slurries. Appl. Environ. Microbiol. 54:1,649-1,655.
6. Kelley, I., and C.E. Cemiglia. 1991. The metabolism of fluoranthene by a species of
Mycobacterium. J. Ind. Microbiol. 7:19-26.
7. Mueller, J.R.,P.J. Chapman, and P.M. Pritchard. 1989. Action of fluoranthene-utilizing
bacterial community on polycyclic aromatic hydrocarbon components of creosote. Appl.
Environ. Microbiol. 55:3,085-3,090.
197
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m PRODUCTj
D PRODUCT D
D CONTROL
B
dPRODUCTJ
D PRODUCT D
D CONTROL
Figure 1. Abundance of selected aliphatic (A) and aromatic (B) constituents of Alaska North
Slope crude oil after treatment for 30 days with a nutrient (Product D) and a
microbial (Product J) bioremediation product in wetlands microcosms. Peak areas
for individual hydrocarbons are referenced to hopane, an internal marker.
198
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Figure 2. Mineralization of pyrene (6 fig/ml) by an enrichment culture in the absence of other
PAHs (•) and in the presence of two- and three-ring PAHs (D), four-ring PAHs (^),
five- and six-ring PAHs (0), and a mixture of 16 PAHs (A) in sediment slurries
amended with organic nutrients (A), minimal medium containing organic sediment
extract (B), and minimal medium (C). The enrichment was previously acclimated in
sediment slurries to a mixture of 16 PAHs. No mineralization occurred in sterile
controls.
199
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pyr adopted
cells
pyr adapted
cells +
chloramp
non-adapted
cells
non-adapted
cells +
chloramp
naph adapted
cells
naph adapted
cells +
chloramp
phe adapted
cells
phe adapted
cells +
chloramp
Figure 3. Mineralization of pyrene (5 jug/ml) in sediment slurries (10 percent w/v) by an
enrichment culture acclimated by preexposure for 9 days to 50 fig/ml of the
indicated PAHs. D'jring the 9 days, the added PAHs were completely degraded.
200
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Biodegrodotion of Petroleum Hydrocarbons in Wetlands:
Constraints on Natural and Engineered Remediation
John H. Pardue, Andrew Jackson, and Ronald D. DeLaune
Wetland Biogeochemistry Institute, Louisiana State University, Baton Rouge, LA
Introduction
Sensitive wetland ecosystems are susceptible to impact from spilled and discharged oils. Major
oil recovery and processing operations are located in wetland ecosystems, including Louisiana,
where 40 percent of the U.S. coastal wetlands and 15 percent of U.S. crude oil production are
located. Understanding the responses of these wetland ecosystems to oil-related impacts is
critical for the design of remediation strategies. Bioremediation is particularly attractive because
mechanical cleaning or washing operations are usually impossible due to the sensitivity of these
systems. At present, however, little information is available on the constraints on bioremediating
spilled oils in wetland ecosystems.
Coastal marshes are wetland ecosystems in the Gulf Coast region where oil production and
transshipment are concentrated. Marsh soils differ from typical bottom sediments in fundamental
ways that will affect bioremediation in these systems: 1) highly organic marsh soils store large
amounts of nutrients but very little in readily available forms; 2) marshes are heavily vegetated
with macrophytes that can serve as conduits for O2 diffusion, dramatically increasing aerobic
surface area in the rhizosphere of marsh soils; and 3) marshes are characterized by periods of
flooding and drying, which expose a larger volume of porous soil to the atmosphere. Because
these features of marshes and other wetland types are unique, this study was recently initiated
to determine the constraints on natural and engineered oil biodegradation in wetlands. The
project is a cooperative agreement with the EPA Environmental Research Laboratory in Athens,
Georgia.
Background
Biodegradation of oil components in wetlands has been demonstrated (1) but rates of
degradation are strongly dependent on environmental conditions. These conditions include
temperature, salinity, Eh, pH, sorption, and the oxygen and nutrient status of the environment.
Studies have documented changes in microbial populations in wetlands in response to spilled
oils (2,3). These responses were generally increases in total microbial populations and increases
in the ratio of oil degraders to total heterotrophs.
In general, wetlands are dominated by anaerobic processes: methanogenesis in freshwater
wetlands and sulfate reduction in brackish and saline wetlands. Several novel microbial
processes have been identified that degrade oil components under anaerobic conditions (4).
Aerobic processes, however, are recognized to act on a broader spectrum of compounds and
are more rapid and complete (e.g., mineralization to CO2 and H2O). In marshes, aerobic
heterotrophic activity is concentrated at the sediment-water interface in a small (several
millimeters) aerobic layer and around the rhizosphere of rooted marsh macrophytes. High
1994 Symposium on Ronnwdiation of Hazardous Waste 201
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sediment oxygen demand, created by a sequence of events leading from organic matter
diagenesis, prevents further O2 penetration.
The maintenance of this aerobic layer is critical to microbial degradation of petroleum
hydrocarbons. In oil-impacted wetlands, petroleum components provide an additional
overwhelming carbon source and potentially serve as a physical barrier for O2 diffusion. Some
of this limitation may be overcome by passive diffusion of O2 through marsh plants, although
the relative supply and demand of this process has not been calculated. Flooding/drying cycles,
either tidal or seasonal, will also control O2 supply to marsh soils. In addition to oxygen
limitation, essential nutrients such as nitrogen may become limiting due to disruption of natural
biogeochemical cycles and competition from highly productive macrophytes. Availability of
nutrients such as nitrogen depends on microbial mineralization processes that convert nutrients
to usable forms, which are rapidly assimilated by plants and microorganisms. This "tight" internal
cycling is characteristic of marshes, where externally supplied nutrients are only a fraction of
those required for observed plant (and microbial) growth. Fertilization may be required to
maximize a microbial response to oil.
Preliminary Results
Study sites that have been selected in the Barataria Basin, Louisiana, include a freshwater marsh
and a salt marsh located along a salinity gradient extending toward the coast. Seasonal samples
are being taken from these sites, and numerous nutrient, microbial, and geochemical analyses
are being conducted relating to bioremediation potential. For example, samples taken during
January/February 1994 were evaluated for aerobic biodegradation potential of two oil
components, phenanthrene and hexadecane, using radiorespirometry. Surface marsh samples
were removed from the marsh using thin-walled aluminum cores, homogenized, and dispensed
in center-well respirometry vials. Slurries were amended with the labeled hydrocarbons in an oil
matrix (~ 1 percent to 2 percent South Louisiana "sweet" crude, v/v), and 14CO2 was quantified
using liquid scintillation. Treatments included controls, killed controls, and fertilization (with
nitrogen, phosphorus, and iron). Results indicate that fertilization can increase the extent of
mineralization of hexadecane and phenanthrene. Fertilization approximately doubled the extent
of hexadecane mineralization in both the salt and fresh marshes (Figure 1). Fertilization effects
on phenanthrene were significant in the salt marsh but within the experimental error in the fresh
marsh (Figure 2). Nutrient availability in the winter months are generally highest due to the lack
of competition from growing plants; therefore, fertilization may have more dramatic effects in
otherseasons. Most probable numbers of oil-degrading microorganisms in the fresh marsh (103)
were several orders of magnitude higher than in the salt marsh (101). This may explain observed
higher rates of phenanthrene mineralization. Results will be contrasted with seasonal data taken
over the next year.
Current work is also being conducted on other aspects of oil degradation in wetlands. The
application of stable isotope techniques is being investigated as a method of measuring oil
biodegradation in marshes. Marsh soils have characteristic
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technique for determining oil biodegradation in spill situations. Additional studies are being
conducted on oil degradation using core and controlled Eh-pH microcosms. Variables being
investigated include tidal and flooding regime, fertilization, vegetation density, and soil oxygen
demand. Gas chromatography/mass spectrometry analysis of crudes is being used to quantify
50 to 60 oil components, including alkanes, polycyclic aromatic hydrocarbons, naphthenes, and
isoprenoids.
References
1. Hambrick, G.A., III, R.D. DeLaune, and W.H. Patrick, Jr. 1980. Effect of estuarine pH
and oxidation-reduction potential on microbial hydrocarbon degradation. Appl. Environ.
Microbiol. 40:365-369.
2. Hood, MA, W.S. Bishop, Jr., F.W. Bishop, S.P. Meyers, and T. Whelam. 1975.
Microbial indicators of oil-rich salt marsh sediments. Appl. Microbiol. 30:982-987.
3. Kator, H., and R. Herwig. 1977. Microbial responses after two experimental oil spills
in an eastern coastal plain ecosystem. In: Proceedings of the 1979 Oil Spill
Conference. API Publ. No. 4284. Washington, DC: American Petroleum Institute, pp.
517-522.
4. Milhelcic, J.R., and R.G. Luthy. 1988. Microbial degradation of acenaphthene and
naphthalene under denitrification conditions in soil-water systems. Appl. Environ.
Microbiol. 54:1,188-1,198.
5. Aggarwal, P.K., and R.E. Hinchee. 1991. Monitoring the in situ biodegradation of
hydrocarbons by using stable carbon isotopes. Environ. Sci. Technol. 25:1,1 78-1,1 80.
6. DeLaune, R.D. 1986. The use of <513C signature of C-3 and C-4 plants in determining
past depositional environments in rapidly accreting marshes of the Mississippi River
deltaic plain, Louisiana. Chem. Geol. 59:315-320.
7. Kennicutt, M.C., II. 1988. The effect of biodegradation on crude oil bulk and
molecular composition. Oil Chem. Poll. 4:89-112.
Table 1. (513C (%o) of Marsh Soils of Louisiana Coastal
Region and of Petroleum Products (6,7)
Source <513C (°/J
Fresh marsh (Panicum hemitomon) -27.9
Intermediate marsh (Sagittaria fa/cataj -26.6
Brackish marsh (Sparfina patens) -14.9
Salt marsh fSpart/na altemiflora) -16.5
Crude oil -30.6
203
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Salt Marsh
O Fertilized
• Unfertilized
3 6 9 12 15 18 21 24 27 30 33
Fresh. Marsh
O Fertilized
• Unfertilized
12 15 18 21 24 27 30 33
Figure 1. Mineralization of 14C-hexadecane (in an oil matrix) in fertilized and unfertilized salt
marsh and fresh marsh soils in coastal Louisiana (soil samples taken in February
1994).
204
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V
.2 so
u
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Anaerobic Biotransformation of Munitions Wastes
Deborah J. Roberts and Farrukh Ahmad
Department of Civil and Environmental Engineering, University of Houston, Houston, TX
Don L. Crawford and Ronald L. Crawford
Center for Hazardous Waste Remediation Research, University of Idaho, Moscow, ID
Introduction
An environmental problem associated with U.S. military facilities is the presence of soil,
sediment, surface water, and ground water contaminated with toxic explosive compounds. With
the current emphasis on demilitarization and returning land to the private sector, the remediation
of the contaminants from these sites has become important. Several types of remediation
procedures are under investigation for the removal of munitions from soils and water.
Incineration has been demonstrated to be an effective process for the remediation of soils from
these sites. The physical process of wet air oxidation of munitions contaminants is under
investigation, as well as several biological remediation procedures. Kaplan (1) reviews the
literature concerning the biological degradation of munitions compounds and shows that under
aerobic conditions the compound 2,4,6-trinitrotoluene (TNT) is degraded by a reductive process
and is not mineralized but merely transformed, producing dinitrotoluenes and azoxy compounds
as the products of metabolism. This suggests that a process that is reductive in nature (i.e.,
anaerobic) would be the best approach to the treatment of soils contaminated with TNT. Under
anaerobic conditions, reductive processes would occur at a faster rate, so lower amounts of the
hydroxylamino intermediates would be produced and thus lower amounts of the azoxy dimers
and polymers.
Current studies show that an aerobic treatment might be a possibility, using Phanerocheate
chrysosporium (2-4). Boopathy et al. (5, 6) have published findings concerning the anaerobic
degradation of TNT by a sulfate reducing bacteria. Many investigations are currently under way
concerning the biological degradation of TNT, but the procedure outlined below is the first pilot-
scale application of a biological technology for munitions degradation that has been
demonstrated.
Background
A procedure for the anaerobic remediation of munitions compounds including TNT, hexahydro-
1,3,5-trinitro-l,3,5-triazine (RDX), and 1,3,5,7-tetranitro-l,3,5,7-tetraazocine (HMX) from
contaminated soil has been developed (7-10) and is being demonstrated at Weldon Springs,
Missouri. The procedure, first developed and demonstrated for the removal of Dinoseb from
soils, involves flooding the soil with water and adding a carbon source with a high oxygen
requirement (such as starch) (11-13). Aerobic heterotrophs deplete the oxygen from the
aqueous phase while utilizing the starch. The aqueous/soil mixture will then be anaerobic,
allowing the degradation of TNT, RDX, and HMX to occur. The procedure requires that the pH
be controlled to between 6.5 and 7 and that the temperature be in the mesophilic range (8).
206 1994 Symposium on Roremediotion of Hazardous Wastes
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The pathway for TNT reduction as seen under anaerobic conditions is initially a reductive one,
where first 4-amino-2,6-dinitrotoluene (4A), then 2,-4-diaminotoluene (24DA), and finally 2,4,6-
triaminotoluene (TAT) are produced. TAT very rarely accumulates in the cultures but is rapidly
converted to 2,4,6-trihydroxytoluene (methylphloroglucinol, MPG) by some unknown mechanism.
This is followed by dehydroxylation reactions, leading ultimately to p-cresol, which can undergo
ring cleavage either anaerobically or aerobically (14,15). Although the latter compounds do
not accumulate in the soil during regular treatment procedures, they have been detected in
laboratory cultures degrading TNT when yeast extract was added as a nutrient supplement for
cultures enriched from soil.
A proposed improvement to the anaerobic remediation strategy is to implement an aerobic stage
after the reductive stage of the procedure is complete. This would ensure mineralization of the
carbon to CO2 rather than a fermentation to several short chain fatty acids. This requires that
the addition of starch at the beginning of the procedure be reexamined, as there is always
excess starch when the treatment is complete; thus, oxygenating the system is very hard (7). To
do this, the use of external carbon sources that were more defined and thus easier to control
than starch were investigated. The use of a commercial soluble starch, glucose, and acetate
was compared with the insoluble starch supplied by J.R. Simplot Co. (Boise, Idaho).
Laboratory experiments were conducted to determine the soil loading rates for the treatment of
a soil from Umatilla, Oregon, contaminated with 12,000 mg TNT/kg soil, 3,000 mg RDX/kg
soil, and 300 mg HMX/kg soil. These led to experiments designed to determine the effect of
the reduced intermediates on the reduction of TNT and on the metabolism of the intermediates.
All experiments were performed using a 1 percent (w/v) addition of a soil that had been
contaminated with Dinoseb and treated using the anaerobic procedure as an inoculum.
Experiments to determine the effects of carbon source additions were performed using 4 percent
(w/v) Umatilla soil in phosphate buffer. Experiments to determine the effects of 4A on
metabolism were performed in cultures spiked with TNT and 4A at the levels indicated in Figure
3. Analyses were performed using narrow-bore high-performance liquid chromatography, as
described by Ahmad (16).
Results
The results of the experiments with various carbon sources led us to glucose as the carbon
source of choice (Figure 1). Acetate was not utilized as a carbon source for oxygen depletion
in these cultures. The reason is unknown, but the contaminants in the soil possibly either
inhibited some reaction in the TCA cycle or the glyoxylate shunt, the two main pathways for the
utilization of acetate. Commercially available soluble starch did not serve as a carbon source
either, probably due to the absence of starch-degrading organisms in the soil inoculum. The
insoluble starch was used as a carbon source for oxygen depletion in these experiments, as had
been demonstrated previously (8,11). This starch contains its own microbial component (11),
thus the presence of starch-degrading organisms in the soil was unnecessary. Cultures fed
glucose reduced the redox potential to the lowest values and showed the fastest initial
degradation of TNT.
When the amount of soil used in the treatment procedure was increased from 1 percent (w/v)
to 4 percent (w/v), the first intermediate (4A) accumulated to an extent that had not been seen
207
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before (50 mg/L) (Figure 2). This accumulation was accompanied by a reduction in the rate
and extent of reduction of TNT. To further examine this observation, experiments were
conducted to determine the effects of 2A on the reduction of TNT and on the degradation of
2A. The results show that when 2A was spiked into the media containing TNT, a reduction in
the rate and extent of degradation of TNT and 2A occurred (Figure 3).
Summary and Conclusions
Glucose was successfully used as an external carbon source, allowing an accurate calculation
of the oxygen demand and a determination of the amount to add that would allow consumption
of the oxygen present initially and maintenance of anaerobic conditions for a specified time.
Calculations show that 28.8 mg/L of glucose must be supplied to remove all initial dissolved
oxygen (DO) and keep the aqueous phase free of DO, assuming an initial DO of 9.08 mg/L,
a reaeration rate of 0.908 mg/L, and an incubation time of 24 d. The calculation assumed that
all glucose was used for oxygen consumption, and no fermentation of the glucose occurred.
To correct for this, a figure of 100 mg/L glucose could be used as a conservative starting point.
Future experiments at the University of Houston will determine whether this is sufficient to allow
the creation of and to sustain anaerobic conditions for the required period, and whether the
institution of an aerobic stage is beneficial to the procedure.
The process must be engineered towards rapid removal of intermediates rather than only rapid
removal of TNT. This will ensure that buildup of toxic intermediates will not occur and that the
process may be performed reliably in the field. The development of more efficient inocula that
will ensure efficient removal of intermediates produced during TNT degradation is currently
under investigation at the University of Idaho and the University of Houston. The effects of the
intermediates on the growth and metabolic activities of the organisms involved is also being
investigated at the University of Houston.
References
1. Kaplan, D.L. 1990. Biotransformation pathways of hazardous energetic organo-nitro
compounds. In: Kamely, D., A. Chakrabarty, and G.S. Omenn, eds. Biotechnology
and biodegradation. TX: Portfolio Publishing Company, p. 155.
2. Fernando, T., and S.D. Aust, eds. 1991. Biodegradation of munition waste, TNT
(2,4,6-trinitrotoluene), and RDX (hexahydro-l,3,5-trinitro-l,3,5-triazine) by
Pnanerochaefe chrysosporium. In: Emerging technologies in hazardous waste
management. American Chemical Society, p. 214.
3. Fernando, T., and S.D. Aust. 1991. Biological decontamination of water contaminated
with explosives by Phanerochaefe chrysosporium. Proceedings of the IGT Symposium
on Gas, Oil, Coal and Environmental Biotechnology III. pp. 193-206.
4. Fernando, T., J.A. Bumpus, and S.D. Aust. 1990. Biodegradation of TNT (2,4,6-
trinitrotoluene) by Phanerochaefe c/irysosporium. Appl. Environ. Microbiol. 56:1,666-
1,671.
208
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5. Boopathy, R., and C.F. Kulpa. 1992. Trinitrotoluene (TNT) as a sole nitrogen source
for a sulfate reducing bacterium Desulfovibrio sp. (B strain) isolated from an anaerobic
digester. Curr. Microbiol. 25:235-241.
6. Boopathy, R., M. Wilson, and C.F. Kulpa. 1992. Biotransformation of 2,4,6-
trinitrotoluene (TNT) by a sulfate reducing bacterium (B strain) isolated from an
anaerobic reactor treating furfural. Abstract Q143. Presented at the American Society
for Microbiology 92nd General Meeting, New Orleans, LA.
7. Funk, S.B., D.L Crawford, D.J. Roberts, and R.L. Crawford. 1994. Two stage
bioremediation of TNT contaminated soils. In: Schepart, B.S., ed. Bioremediation of
pollutants in soil and water. ASTM STP 1235. Philadelphia, PA: American Society for
Testing Materials.
8. Funk, S.B., D.J. Roberts, D.L. Crawford, and R.L. Crawford. 1993. Initial-phase
optimization for bioremediation of munition compound-contaminated soils. Appl.
Environ. Microbiol. 59:2,171-2,177.
9. Funk, S.B., DJ. Roberts, and R.A. Korus. 1992. Physical parameters affecting the
anaerobic degradation of TNT in munitions-contaminated soil. Abstract Q142.
Presented at the American Society for Microbiology 92nd General Meeting, New
Orleans, LA.
10. Roberts, D.J., S.B. Funk, D.L. Crawford, and R.L. Crawford. 1993. Anaerobic
biotransformation of munitions wastes. In: U.S. EPA. Symposium on bioremediation
of hazardous wastes: Research, development, and field evaluations (abstracts).
EPA/600/R-93/054. Washington, DC (May).
11. Kaake, R.H., DJ. Roberts, T.O. Stevens, R.L. Crawford, and D.L. Crawford. 1992.
Bioremediation of soils contaminated with 2-sec-butyl-4,6-dinitrophenol (Dinoseb).
Appl. Environ. Microbiol. 58:1,683-1,689.
12. Roberts, D.J., R.H. Kaake, S.B. Funk, D.L. Crawford, and R.L Crawford. 1992.
Anaerobic remediation of Dinoseb from contaminated soil: An onsite demonstration.
Appl. Biochem. Biotechnol. 39:781-789.
13. Roberts, D.J., R.H. Kaake, S.B. Funk, D.L. Crawford, and R.L. Crawford. 1 992. Field
scale anaerobic bioremediation of Dinoseb-contaminated soils. In: Gealt, M., and M.
Levin, eds. Biotreatment of industrial and hazardous wastes. New York, NY: McGraw-
Hill.
14. Roberts, D.J., and D.L. Crawford. 1991. Anaerobic degradation of TNT. Abstract
Q160. Presented at the American Society for Microbiology 91st General Meeting,
Dallas, TX.
15. Roberts, D.J., S.B. Funk, and R.A. Korus. 1992. Intermediary metabolism during
anaerobic degradation of TNT from munitions-contaminated soil. Abstract Q136.
Presented at the American Society for Microbiology 92nd General Meeting, New
Orleans, LA.
209
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16. Ahmad, F., and D.J. Roberts. 1994. The use of narrow bore HPLC-diode array
detection to identify and quantitate intermediates during the biological degradation of
2,4,6-trinitrotoluene. J. Chromatog. In press.
a. Redox potential with glucose, insoluble starch or
soluble starch as external carbon sources.
i
i
400
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800
too
0
-100
200
-aoo
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O IniolubU Bt*rcli
1C.
TII.H- (.))
b. Redox potential with acetate or glucose as external carbon sources.
Figure 1. The effect of external carbon sources on redox potential and TNT degradation in
cultures containing 5 percent Umatilla soil/phosphate buffer and inoculated with
treated soil.
210
-------
c. TNT concentration with glucose, insoluble starch or
soluble starch as external carbon sources..
IS 20
Ttmt (d)
d. TNT concentration with acetate or glucose as external carbon sources.
Figure 1 (continued).
211
-------
100 -
BO -
^ 80 -
60
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7 8°-
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0 5 10 15 20 25
Time (d)
'5
30
Figure 2. Concentrations of TNT and its metabolic intermediates during the anaerobic
remediation of Umatilla soil in cultures inoculated with treated soil.
O 0 mg/L 4A2.8DNT
A 20 mg/L 4A8.6DNT
0 40 mg/L 4A8.6DNT
14
Figure 3. Concentrations of TNT in aqueous cultures inoculated with treated soil degrading
100 mg/L TNT in the presence of 4-amino-2,6-dinitrotoluene.
212
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Covalent Binding of Aromatic Amines to Natural Organic Matter:
Study of Reaction Mechanisms and Development of Remediation Schemes
Eric J. Weber and Dalizza Colon
U.S. Environmental Protection Agency, Environmental Research Laboratory, Athens, GA
Michael S. Elovitz
National Research Council, Environmental Research Laboratory, Athens, GA
Introduction
Aromatic amines comprise an important class of environmental 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 or carcinogenic to animals. Aromatic
amines can enterthe environment from the degradation of textile dyes, munitions, and numerous
herbicides. Because of their importance as synthetic building blocks for many industrial
chemicals, the loss of aromatic amines to the environment may also result from production
processes or improper treatment of industrial waste streams. The high probability of
contamination of soils, sediments, and ground-water aquifers with aromatic amines necessitates
the development of innovative, cost-effective in situ remediation techniques for their treatment.
Numerous studies have demonstrated that aromatic amines become covalently bound to the
organic fraction of soils and sediments through oxidative coupling or nucleophilic addition
reactions (1 -4). It is generally accepted that once bound, the bound residue is less bioavailable
and less mobile than the parent compound. Thus, procedures for enhancing the irreversible
binding of aromatic amines to soil constituents could potentially serve as remediation
technologies.
Model studies suggest that oxidative enzymes derived from soil microorganisms play a significant
role in catalyzing the formation of bound residues (5,6). Stimulation of these naturally occurring
enzymes could provide an effective in situ method for the treatment of soils, sediments, and
ground-water aquifers contaminated with aromatic amines (7). For example, Berry and Boyd
(8) were able to enhance the covalent binding of the potent carcinogen 3,3'-dichlorobenzidine
(DCB) in a 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 overall level of oxidative coupling in the soil was increased, which lead
to enhanced incorporation of DCB.
To gain a more in-depth understanding of the enzyme-mediated binding of organic amines to
soils and sediments, we have studied the effects of enzyme amendments to sediments treated
with aromatic amines such as aniline, reduction products of TNT and atrazine, and metabolic
reaction products of atrazine.
1994 Symposium on Bioremediotion of Hazardous Wastes 213
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Results and Discussion
Initially, experiments were conducted to determine the limiting factors controlling the binding of
aniline to amended sediments. Figure 1 illustrates the effect of the addition of various
combinations of horseradish peroxidase, H2O2, and ferulic acid to Beaver Dam sediment-water
systems treated with aniline at an initial aqueous concentration of 5 x 10'5 M. In each case, the
amendments were added 24 hr after the addition of aniline.
It is apparent from the data in Figure 1 that the binding capacity of the sediment for aniline was
limited prior to the addition of the amendments. Only 10 percent of the initial concentration
of aniline was irreversibly bound to the untreated natural sediment. All amendments tested
greatly enhanced the removal of aniline from the aqueous phase of the Beaver Dam sediment-
water systems, as the concentration of aniline in the aqueous phase was below detectable limits
in a matter of hours. The observation that the addition of H2O2 alone catalyzed the removal
of aniline suggested that the sediment was not limited in peroxidase activity or oxidizable
substrates.
To determine the effect of H2O2 on the binding of aniline in a sediment with no peroxidase, we
monitored the aqueous concentration of aniline in both a nonsterile and a heat-sterilized Beaver
Dam sediment with and without the addition of H2O2 (Figure 2). The aqueous concentration
of aniline was measured for 24 hr prior to the addition of H2O2. As before, the control study
(no addition of H2O2) demonstrated the limited binding capacity of the sediment for aniline.
Surprisingly, the addition of H2O2 24 hr after the initial addition of aniline had a significant
effect on the aqueous concentration of aniline in both the sterile and nonsterile sediment-water
systems.
Because our initial assumption was that heat sterilization would destroy peroxidase activity, the
observation that treatment of the heat-sterilized Beaver Dam sediment-water system greatly
enhanced the removal of aniline suggested that a mechanism other than peroxidase activation
may exist. The high iron content of the sediment may have resulted in the iron-mediated
reduction of H2O2 to form hydroxyl radicals (Fenton's reaction), which could subsequently react
with aniline via hydrogen abstraction and ring addition. Recently, the chemical oxidation of
chlorinated organics by addition of H2O2 to sand containing iron has been demonstrated by
Ravikumar and Gurol (9).
In an attempt to determine if the iron-mediated reaction was occum'ng, two Beaver Dam
sediment-water systems were treated with H2O2 24 hr prior to the addition of aniline. We
hypothesized that if Fenton-type reactions were occurring, the extremely reactive hydroxyl radicals
would react quickly with the organic matter and subsequently would not be available to react
directly with aniline upon its addition 24 hr later. Surprisingly, at both concentrations of H2O2
studied, the binding capacity of the Beaver Dam sediment for aniline was increased by treatment
with H2O2 24 hr prior to the addition of aniline. These findings suggest that hydoxyl radicals,
like activated peroxidase, may react with organic matter to produce binding sites for compounds
such as aromatic amines (Figure 3).
In summary, we feel that hydrogen peroxide treatment of soils and sediments contaminated with
aromatic amines and other classes of reactive chemicals shows promise as a remediation
method. We are currently extending this remediation technology to other aromatic amines of
interest, such as TNT reduction products and atrazine and its metabolites, whose contamination
214
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of soils and sediments has been reported. Experiments are also in progress to further our
understanding of the mechanisms by which H2O2 enhances the covalent binding of aromatic
amines.
References
1. Baughman, G.L., E.J. Weber, R.L. Adams, and M.S. Brewer. 1992. Fate of colored
smoke dyes. Army Project No. 88PP8863. U.S. Department of the Army, Frederick,
MD.
2. Graveel, J.G., L.E. Sommers, and D.W. Nelson. 1985. Sites of benzidine, a-
naphthylamine, and p-toluidine retention in soils. Environ. Toxicol. Chem. 4:607-613.
3. Paris, G.E. 1980. Covalent binding of aromatic amines to humates. 1. Reactions with
carbonyl groups and quinones. Environ. Sci. Technol. 1-4:1,099-1,105.
4. Scheunert, I., M. Mansour, and F. Andreux. 1992. Binding of organic pollutants to soil
organic matter. Intern. J. Environ. Anal. Chem. 46:189-199.
5. Bollag, J., and W.B. Bollag. 1990. A model for enzymatic binding of pollutants in the
soil. J. Environ. Anal. Chem. 39:147-157.
6. Claus, H., and Z. Filip. 1990. Enzymatic oxidation of some substituted phenols and
aromatic amines, and the behavior of some phenoloxidases in the presence of soil
related adsorbents. Water Sci. Tech. 22:69-77.
7. Bollag, J. 1992. Decontaminating soil with enzymes. Environ. Sci. Technol. 26:1,876-
1,881.
8. Berry, D.F., and S.A. Boyd. 1985. Decontamination of soil through enhanced
formation of bound residues. Environ. Sci. Technol. 19:1,132-1,133.
9. Ravikumar, J.X., and M.D. Gurol. 1994. Chemical oxidation of chlorinated organics
by hydrogen peroxide in the presence of sand. Environ. Sci. Technol. 28:394-400.
215
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10
20 30 40
Time (hr)
50
60
Figure 1. Effect of amendments on the aqueous phase concentration of aniline in Beaver Dam
sediment-water system: (0) control, no treatment; (|) ferulic acid, peroxidase, and
H2O2; (*) ferulic acid and H2O2; and (Q H2O2.
OE+0
0 5 10 15 20 25 30 35
Time (hours)
Figure 2. Effect of hydrogen peroxide treatment on the aqueous concentration of aniline in a
Beaver Dam sediment-water system: (|) nonsterile control, no H2O2 treatment; (^)
nonsterile sediment treated with H2O2 at t=24 hr; and (*) heat-sterilized sediment
treated with H2O2 at t=24 hr.
216
-------
ot-o
2 5E-5*
0)
1 4E-5
0.
M
3 <
8 3E-5
I •
^v
~ 2E-5
0)
c
1 1E-5
nc.n
* • No HaOz Addition
-
•
»
* [HaOs] = 3.6x10-3M
-
^
»
[HzOz] = 3.6 X 10-2 M
^ •» i •*' • ' • '
10
20
30
40
50
Figure 3. Effect of H2O2 treatment of a Beaver Dam sediment-water system 24 hr prior to the
addition of aniline: initial [aniline] = 5.5 x 10'5 M.
217
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Kinetics of Anaerobic Biodegradation of Munitions Wastes
Jiayang Cheng and Makram T. Suidan
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Albert D. Venosa
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
2,4-Dinitrotoluene (2,4-DNT) is formed during the manufacture of propellent and is commonly
found in munitions wastewater. It has been found to be mutagenic in bacterial and mammalian
assays and carcinogenic in animal studies (1). Because of its toxic nature and large-scale use,
2,4-DNT is listed as a priority pollutant by EPA. Early studies on the biodegradation of 2,4-DNT
suggested that 2,4-DNT was resistant to biological treatment in aerobic processes such as
activated sludge systems (2). Recently, some investigators reported complete degradation of
2,4-DNT by a pure aerobic culture (3,4). Industrial application of the aerobic biodegradation
of 2,4-DNT, however, reveals that it is very difficult to achieve compliance with EPA discharge
limits. Under anaerobic conditions, 2,4-DNT can be completely transformed to 2,4-
diaminotoluene (2,4-DAT) with ethanol serving as the primary substrate (5). Subsequently, 2,4-
DAT can be easily mineralized aerobically (5).
In this study, the anaerobic biotransformation of 2,4-DNT with ethanol serving as the primary
substrate was investigated. The culture was acclimated in a chemostat with 2,4-DNT and
ethanol as substrates. The pH and the temperature in the chemostat were kept at 7.2 and
35°C, respectively. The hydraulic retention time in the chemostat was 40 days. Biochemical
methane potential (BMP) tests with 2,4-DNT and ethanol as substrates were conducted using an
anaerobic respirometer with the culture from the chemostat serving as an inoculum. Sodium
sulfide and L-cysteine hydrochloride were used to maintain a reducing environment for the BMP
tests. The impact of the reducing agent on the biotransformation of 2,4-DNT and ethanol was
studied. The effect of 2,4-DNT, the biotransformation intermediates, and 2,4-DAT on the
byconversion of ethanol was also investigated.
Results and Discussion
After steady-state operation was established in the chemostat (i.e., the effluent composition, the
volumetric gas production rate and composition, and the biomass concentration in the
chemostat had been constant for over 120 days), mixed culture from the chemostat was used
as an inoculum for the BMP tests. The culture was transferred into the BMP reactors in an
oxygen-free anaerobic chamber at 35°C. The pH and the temperature in the BMP reactors were
kept the same as those in the chemostat. Different initial 2,4-DNT concentrations were used in
the BMP tests, while the initial concentration of ethanol was the same in all of the reactors.
Figure 1 illustrates the biotransformation process of 2,4-DNT, in the presence of ethanol and
50 mg/L sodium sulfide hydrate and 100 mg/L L-cysteine hydrochloride as the reducing agents.
2,4-DNT was completely transformed to 2,4-DAT, with 2-amino-4-nitrotoluene (2-A-4-NT) and
218 1994 Symposium on Roramediation of Hazardous Wastes
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4-amino-2-nitrotoluene (4-A-2-NT) appearing as intermediates. The initial transformation rate
decreased with increasing initial 2,4-DNT concentrations (Figure 1 a). Note that at a low initial
concentration of 2,4-DNT, a greater buildup of 4-A-2-NT occurred compared with 2-A-4-NT
(Figures Ic and Id). As the initial 2,4-DNT concentration increased, more 2,4-DNT was
transformed via 2-A-4-NT (Figures 1 c to 1 f). A higher concentration of 2-A-4-NT than 4-A-2-
NT was formed at the high initial 2,4-DNT concentration (Figure If). The results suggest two
pathways leading to the complete biotransformation of 2,4-DNT to 2,4-DAT (Figure 2), with
pathway a occurring faster at high initial 2,4-DNT concentrations and pathway b occurring faster
at low initial 2,4-DNT concentrations.
Another BMP test was conducted under similar conditions except, in this instance, the reducing
agent was 200 mg/LNa2S'9H2O. The rate of biotransformation of 2,4-DNT was much higher,
and 2,4-DNT exhibited much less inhibition to its biotransformation as a result of the presence
of a higher concentration of sulfide. The presence of the higher concentration of sulfide
provided a more reducing environment, which was favorable to the biotransformation of 2,4-
DNT. An abiotic test was conducted to evaluate the potential for chemical reduction of 2,4-
DNT. Results suggest that 2,4-DNT is chemically reduced to 2,4-DAT via 2-A-4-NT or 4-A-2-
NT in the presence of high concentrations of sulfide and minerals.
The bioconversion of ethanol was also affected by the reducing agent used in the BMP test. L-
cysteine hydrochloride is widely used as a reducing agent in anaerobic experiments. When L-
cysteine (100 mg/L) and Na2S (50 mg/L) were used as reducing agents in the co-metabolic
biodegradation of 2,4-DNT, propionate was formed during the bioconversion of the primary
substrate ethanol when the initial concentration of 2,4-DNT was lower than 6 mg/L (6). No
such propionate production, however, was observed when sulfide (200 mg/L) was the sole
reducing agent. L-cysteine hydrochloride may contribute to the formation of propionate during
the fermentation of ethanol in the presence of 2,4-DNT.
References
1. Ellis, H.V., C.B. Hong, C.C. Lee, J.C. Dacre, and J.P. Glennon. 1 985. Subchronic and
chronic toxicity study of 2,4-dinitrotolune. Part I. Beagle dogs. J. Am. Coll. Toxicity
4(4):233-242.
2. McCormick, N.G., J.H. Cornell, and A.M. Kaplan. 1978. Identification of
biotransformation products from 2,4-dinitrotoluene. Appl. Environ. Microbiol. 35:945-
948.
3. Spanggord, R.J., J.C. Spain, S.F. Nishino, and K.E. Mortelmans. 1991.
Biodegradation of 2,4-dinitrotoluene by a Pseudomonas sp. Appl. Environ. Microbiol.
57:3,200-3,205.
4. Valli, K., B.J. Brock, D.K. Joshi, and M.H. Gold. 1992. Degradation of 2,4-
dinitrotoluene by the lignin-degrading fungus Phanerochaete chrysosporium. Appl.
Environ. Microbiol. 58:221-228.
219
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5. Berchtold, S.R. 1993. Treatment of 2,4-dinitrotoluene using a two-stage system:
Fluidized bed anaerobic GAC reactors and aerobic activated sludge reactors. Master's
thesis. University of Cincinnati, Cincinnati, OH.
6. Cheng, J., M.T. Suidan, and A.D. Venosa. 1993. Co-metabolic biodegradation of
2,4-dinitrotoluene using ethanol as a primary substrate. In: U.S. EPA. Symposium on
bioremediation of hazardous wastes: Research, development, and field evaluations
(abstracts). EPA/600/R-93/054. Washington, DC (May), pp. 145-148.
220
-------
J2 0.20
o
a
a
- 0.15 -
a
o
>P4
JJ
a
d
a
o
u
a
i
•*
N
o
a
B
(4
a
a
a
d
o
u
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o
I-*
O
a
G
a
o
o
o
a
o
20
(e)
40 60 780 800
Time, Hrs
0 20 40 60 80 100 780 800
(f) Time, Hrs
(b), (c). (d). (e), and (f): O . • 2.4-DNT. -
D . • 4A2NT, Average 4A2NT; A , A 2A4NT,
V , T 2.4-DAT, Average 2,4-DAT.
Average 2,4-DNT;
- Average 2A4NT;
Figure 1. Anaerobic biotransformation of 2,4-DNT with ethanol as primary substrate.
221
-------
CH
NH,
NO,
CH.
NH.
NH
2
2,4-DAT
Figure 2. Pathway of anaerobic biotransformation of 2,4-DNT.
222
-------
Biodegradotion of Chlorinated Solvents
Sergey A. Selifonov, Lisa N. Newman, Michael E. Shelton, and Lawrence P. Wackett
Department of Biochemistry and Institute for Advanced Studies in Biological Process
Technology, University of Minnesota, St. Paul, MN
Introduction and Background Information
Haloorganics comprise the largest single group of chemicals on the EPA list of priority pollutants
(1) because many of these industrially important compounds have been demonstrated to be
mutagenic and carcinogenic in mammals. Successful application of chlorinated solvent
bioremediation requires extensive knowledge of underlying molecular mechanisms of
biodegradation. Such knowledge will allow a rationale for selection of organisms and treatment
schemes, and prevent slow, costly empirical approaches to bioremediate every different site.
Microbial action on chlorinated solvents often involves co-metabolism or cases of fortuitous
metabolism, which provide no net benefit to the organism involved. An example of this is the
bacterial degradation of trichloroethylene (TCE), a widespread ground-water pollutant.
Gratuitous metabolism of TCE has been observed to be catalyzed by a number of different
oxygenases: toluene dioxygenase (2,3), toluene-4-monooxygenase (4), ammonia
monooxygenase (5), soluble methane monooxygenase (sMMO) (6), propane monooxygenase
(7), toluene-2-monooxygenase (8), phenol hydroxylase (9), and isoprene oxygenase (10).
Currently methanotrophs expressing sMMO oxidize TCE most rapidly in small-scale laboratory
studies. In practice, the use of methanotrophs suffers from 1) inactivation of sMMO resulting
from alkylation by acyl chlorides derived from TCE oxidation, 2) formation of toxic chloral
hydrate as a TCE byproduct, 3) cooxidation of co-contaminants to more toxic materials (i.e.,
chlorobenzene to chlorophenols), 4) inhibition with methane, and 5) inability to maintain sMMO
under field conditions.
In light of the above, other TCE-degrading organisms might outperform methanotrophs, or
toluene dioxygenase-expressing strains, over sustained periods and under field conditions. One
of our experimental models is the strain of Pseudomonas cepacia G4 (8,11), whose TCE-
degrading ability is based on co-metabolic action of the toluene-2-monooxygenase system. The
performance and safe application of TCE-biodegraders necessitates a greater understanding of
the mechanisms of oxygen addition to TCE and rigorous determination of the final recoverable
products. Purification of TMO activity from P. cepacia G4 will facilitate determination of the
complete product stoichiometry of TCE oxidation. These questions are important in the context
of understanding the physiological basis by which P. cepacia (toluene-2-monooxygenase, TMO)
is less influenced by toxic effects resulting from TCE oxidation than are Pseudomonas putida Fl
(toluene dioxygenase, TOO) and other organisms.
Understanding the biochemical basis of advantages of TMO over other chloroethene-degraders
may open new, direct approaches for search of more effective strains and enzymes.
1994 Symposium on Bioremediation of Hazardous Wastes 223
-------
Physiology and Biochemistry of TCE Oxidation by P. Cepacia G4
In Vivo Studies With P. Cepacia G4
Generally, in vivo studies have focused on measuring the disappearance of chlorinated
compounds. Supplementing this information, however, with a deeper knowledge of the products
obtained from chlorinated solvent oxidation is crucial. TCE oxidation has been investigated most
extensively, but only substoichiometric accounting of products has been accomplished. The
present study addresses possible formation of epoxides from chloroethenes and of products
arising from chloride migration during oxygen addition.
Identification of TCE Biodearadation Products. In experiments with TCE, 200 f*M was essentially
quantitatively degraded by P. cepac/a G4. At that time, culture filtrates were extracted and
analyzed by gas chromatography (GC) for the presence of the possible chloride rearrangement
products 2,2,2-trichloroacetaldehyde and 2,2,2-trichloroethanol. Neither compound was
detected above the level of 0.25 percent of the total TCE transformed (less than 0.5 /^M).
Analysis of culture filtrates obtained in experiments with [14C]-TCE and washed cell suspensions
of P. cepacia G4 was performed by high-performance liquid chromatography (Bio-Rad Aminex
organic acid column). The major detectable metabolite, in all cases, comigrated with authentic
glyoxylate and accounted for 2.5 percent, 29 percent, and 19 percent of the added TCE at 0
min, 30 min, and 60 min of incubation, respectively. (Zero time control contained live induced
cells centrifuged with TCE, so several minutes elapsed before the cells were actually removed
from the culture supernatant fluid.) In subsequent experiments with 10 mM glyoxylate added as
cold trap, more than 60 percent of the products were accounted for as glyoxylate. The data
indicate that glyoxylate is a likely major product and is further metabolized by P. cepac/a G4.
Two minor products were also observed transiently; one of them may be formate, the identity
of other is unknown. These analyses provided no evidence for the formation of trichloroacetate,
dichloroacetate, oxalate, and glycolate by P. cepac/a G4 from [14C]-TCE.
Evidence of Epoxide Formation From Chloroethenes by P. Cepoa'o G4. Production of glyoxylate
infers the formation of TCE-epoxide as precursor. While TCE-epoxide is unstable in water (ty2< 1
min), frans-l,2-dichloroethylene epoxide undergoes hydrolysis and isomerization relatively slowly.
frans-1,2-Dichloroethylene (traris-1,2-DCE) was used as a model compound to obtain evidence
for epoxide formation, frans-1,2-DCE was readily oxidized by P. cepacia G4 induced with
toluene vapor; at a starting concentration of 200 /*M, 85 percent of frans-1,2-DCE was
transformed after 60 min. Only 3 percent of the transformed frans-1,2-DCE was recovered,
however, as its colored epoxide adduct with 4-(p-nitrobenzyl)-pyridine (12). Noninduced P.
cepacia GA showed no significant production of material forming the colored 4-(p-nitrobenzyl)-
pyridine adduct.
GC/mass spectrometry (MS) and GC/Fourier transfer infrared (FTIR) was used to analyze
pentane extracts of cell supernatants after incubation of P. cepacia G4 with frans-1,2-DCE. A
compound was found with the same R,, mass and infrared spectra as synthetic frans-1,2-DCE
epoxide. Synthetic 2,2-dichloroacetaldehyde showed different R, on the GC column used, and
its MS fragmentation and infrared spectrum differed from that of the epoxide. These data
indicate that frans-1,2-DCE epoxide is the major pentane extractable product formed.
224
-------
Purification of Toluene-2-Monooxygenase (TMO)
Conditions have been established to obtain active crude extracts and to achieve partial
purification of the components for this multicomponent enzyme system. Active crude extracts
were obtained in a buffer consisting of 25 mM MOPs, pH 7.5, 200 /iM Fe(NH4)2(SO4)2, and
5 mM cysteine. Partial purification of an NAD(P)H oxidoreductase, with an apparent molecular
weight of 38,000 daltons on SDS-PAGE, has been accomplished through the use of ion
exchange chromatography at different pHs. The reductase is capable of reducing cytochrome
c and supports reconstituted toluene monooxygenase activity. In addition, the reductase from
phenol hydroxylase of Pseudomonas sp. SF600 is capable of complementing the toluene-orfho-
monooxygenase system, indicating a possible similarity between these enzyme systems.
Oxidation of Structural Analogues of Perchloroethylene (PCE) and TCE
Presently known aerobic biodegradation processes for chlorinated ethenes are based on the co-
metabolic action of oxygenase enzymes involved in catabolic pathways providing effective
utilization of compounds showing little or no structural relationship to TCE or PCE (e.g., toluene,
isopropylbenzene, methane, camphor, isoprene). Analogues of chlorinated ethenes with one or
more chlorine atoms replaced by methyl groups can be used for studies of biochemical
mechanisms involved in oxidation of TCE, and of factors limiting activities of oxygenases on
PCE. It is more important, however, that they can serve as potential structural surrogates of PCE
and TCE that may support growth of bacterial strains and be useful carbon and energy sources
for enrichment cultures to search for organisms and enzymes, effectively metabolizing PCE and
TCE themselves. This approach requires prior synthetic work to obtain such surrogate substrates
because the most promising compounds, such as 1,1 -dichloro-2-methyl-1 -propene or 1,1,2-
trichloro-1-propene, are not available commercially. Neither TDO- or TMO-expressing strains
are capable of oxidizing these compounds. The feasibility of use of methylated analogs of PCE
and TCE as enrichment substrates will be analyzed.
Summary and Conclusions
The production of glyoxylate as a major TCE oxidation product differs from previous
observations of in vivo and in vitro TCE oxidation catalyzed by toluene dioxygenase and
methane monooxygenase. In the previous studies with TDO and sMMO, glyoxylate formation
is a minor pathway. With P. cepac/a G4 and TMO, the pathway in Figure 1 appears to be more
prominent. This could be due either to possible enzyme participation in C-CI bond cleavage
reactions or to a different intracellular environment that promotes glyoxylate formation from
chemical decomposition of TCE-epoxide, thereby avoiding formation of toxic or alkylating
intermediates.
Compared with TMO of P. cepacia G4, enzymes such as TDO or methane monooxygenase are
inactivated in vivo by reactive intermediates generated during TCE oxidation; cells expressing
these activities experience cytotoxicity from oxidizing TCE (13). Generally, most known TCE
oxidation reactions are characterized by low reaction rates and formation of harmful metabolites.
With respect to TCE (or PCE) co-metabolism, the bacteria cannot help themselves to select
against or for such fortuitous reactions. These reactions provide no net benefit to cells as energy
225
-------
and carbon sources. Counter argument would point out that TCE and PCE are not natural
products, and are only recently found in soil and water, so natural selection has not had time
to select against this deleterious co-metabolism.
Using surrogate carbon and energy sources may offer a practical solution to finding
microorganisms that 1) are capable of not forming toxic substrates and 2) have higher reaction
rates of TCE and PCE oxidation comparable with the conversion rates for growth (catabolic)
substrates. Either direct dihydroxylation or a monooxygenation/hydration sequence would
produce intermediates (Figure 2) capable of serving as carbon and energy sources. Therefore,
the enrichment culture approach may provide a selection tool for finding new biological
mechanisms capable of attacking the hindered double bond of PCE and TCE in an appropriate
electrophilic environment.
Neither TMO or TOO can oxidize such hindered compounds as l,l-dichloro-2-methyl-l-
propene or 1,1,2-trichloro-l-propene. The less hindered compound, l,1-difluoro-2,2-
dichloroethylene, however, is oxidized by TOO and sMMO. This fad indicates that strong steric
hindrance rather than the electrophilic environment of the double bond appears to be a limiting
factor determining the success of oxidative reactions on PCE and TCE.
This work is supported by Cooperative Agreement EPA/CR820771 -01 -0 between the U.S. EPA
Environmental Research Laboratory, Gulf Breeze, and the University of Minnesota.
References
1. Leisinger, T. 1983. Microorganisms and xenobiotic compounds. Experientia
39:1,183-1,191.
2. Nelson, M.J.K., S.O. Montgomery, and P.M. Pritchard. 1988. Trichloroethylene
metabolism by microorganisms that degrade aromatic compounds. Appl. Environ.
Microbiol. 54:604-606.
3. Wackett, L.P., and D.T. Gibson. 1988. Degradation of trichloroethylene by toluene
dioxygenase in whole cell studies with Pseudomonas putida Fl. Appl. Environ.
Microbiol. 54: 1,703-1,708.
4. Winter, R.B., K.-M. Yen, and B.D. Ensley. 1989. Efficient degradation of
trichloroethylene by a recombinant Escherichia coll. Biotechnology 7:282-285.
5. Arciero, D., T. Vannelli, M. Logan, and A.B. Hooper. 1989. Degradation of
trichloroethylene by the ammonia-oxidizing bacterium Nifrosomonas europaea.
Biochem. Biophys. Res. Commun. 159:640-643.
6. Oldenius, R., R.L. Vink, J.M. Vink, D.B. Janssen, and B. Witholt. 1989. Degradation
of chlorinated aliphatic hydrocarbons by Methylosinus frichosporium OB3b expressing
soluble methane monooxygenase. Appl. Environ. Microbiol. 55:2,819-2,826.
226
-------
7. Wackett L.P., G. Brusseau, S. Householder, and R.S. Hanson. 1989. Survey of
microbial oxygenases: Trichloroethylene degradation by propane oxidizing bacteria.
Appl. Environ. Microbiol. 55:2,960-2,964.
8. Folsom, R.R, P.J. Chapman, and P.H. Pritchard. 1990. Phenol and trichloroethylene
degradation by Pseudomonas cepac/a G4: Kinetics and interaction between substrates.
Appl. Environ. Microbiol. 56:1,279-1,285.
9. Montgomery, S.O., M.S. Shields, P.J. Chapman, and P.H. Pritchard. 1989.
Identification and characterization of trichloroethylene degrading bacteria. Abstract K-
68:256. Presented at the Annual Meeting of the American Society for Microbiology.
10. Ewers, J., D. Frier-Shroder, and H.J. Knackmuss. 1990. Selection of trichloroethylene
(TCE) degrading bacteria that resist inactivation by TCE. Arch. Microbiol. 154:410-
413.
11. Nelson, M.J.K., S.O. Montgomery, E.J. O'Neill, and P.H. Pritchard. 1986. Aerobic
metabolism of trichloroethylene by a bacterial isolate. Appl. Environ. Microbiol.
52:383-384.
12. Fox, B.G., J.B. Borneman, L.P. Wackett, and J.D. Lipscomb. 1990. Haloalkene
oxidation by the soluble methane monooxygenase from Mefhy/os/nus fricfiosporium
OB3b: Mechanistic and environmental applications. Biochemistry 29:6,419-6,427.
13. Wackett L.P., and S.R. Householder. 1989. Toxicity of trichloroethylene to
Pseudomonas puf/da Fl is mediated by toluene dioxygenase. Appl. Environ. Microbiol.
55:2,723-2,725.
Cl
Cl
Cl 02 HO
=( *• CI-^_
H or 1/2 O2 Cl
+ H 0
2
Major pathway for
P.cepacia G4
•
O 1
OH
L-CI
H
i
' O
~y-<~
Figur
Cl
e1.
H
-
^
Major pathway
for sMMO and
TOO organisms
Q
^.
HO
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;c=o +
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T
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(formate)
Q
H
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(carbon
monooxide)
(glyoxylate)
Figure 1.
227
-------
an
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>=< »
Cl CH,
3
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2
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apLJ ?
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n \ / rw
wi ^ •.^••^ wn<3 •^•u^
Cl CH3
0 P
* V< -
HO CH3
(pyruvate)
_ OH
^i ^^i i
— fc-
-------
Characterization of Bacteria in a TCE Degrading Biofiher
Alec W. Breen, Alex Rooney, Todd Ward, and John C. Loper
Department of Molecular Genetics, University of Cincinnati, Cincinnati, OH
Rakesh Govind
Department of Chemical Engineering, University of Cincinnati, Cincinnati, OH
John R. Haines
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
A trichloroethylene- (TCE-) degrading vapor phase biofilter was investigated to determine the
microbial population(s) mediating degradation. Initial observations suggested that ammonia-
oxidizing bacteria could be responsible for TCE degradation. The biofilter being studied had
been maintained in the presence of a gas stream containing methylene chloride, benzene,
ethylbenzene, toluene, and TCE. During operation, a microbial community was established that
could oxidize TCE when all other substrates were removed from the gas stream. Twenty to thirty
percent removal of TCE at an inlet concentration of 21 ppmv (0.113 mg/L) and a gas residence
time of 1 minute was experimentally observed. TCE degradation capability remained intact for
over 12 months. The standard OECD mineral salts solution with excess ammonia was trickled
over the biofilter. The fact that ammonia was present in the nutrient solution provided
circumstantial evidence that it could serve as a co-metabolite for nitrifying bacteria mediating
TCE degradation. The ammonia monooxygenase (AMO) system, responsible for the conversion
of ammonia to hydroxylamine, has been shown to carry out a co-metabolic oxidation of TCE
(1,2). Characterization of the biofilter community was undertaken to establish if ammonia-
oxidizing bacteria were responsible for TCE oxidation.
Background
Studies on the aerobic metabolism of TCE have shown that a diverse group of organisms can
oxidize this compound in a co-metabolic fashion (3). The initial observation by Wilson and
Wilson (4) demonstrating co-metabolism of TCE by methanotrophs was followed by reports of
TCE degradation by toluene oxidizers (5), propane oxidizers (6), and ammonia oxidizers (1).
Strategies for the treatment of TCE containing wastes often focus on the optimization of
degradation using the addition of a co-metabolite to the appropriate group of organisms.
Experimental System and Results
The presence of nitrifying bacteria was monitored by most probable number (MPN) methodology
and by gene probing with an AMO gene probe. The data generated showed that levels of
ammonia oxidizers were low, generally below the level of detection of the AMO probe and 102
1994 Symposium on Bioremediation of Hazardous Wastes 229
-------
to 104 per gram of biofilter biomass. Following gene probing and MPN analysis, TCE
degradation experiments were begun.
A series of TCE degradation experiments were conducted with biofilter biomass in a batch
degradation assay using 1l4C-TCE as a tracer and trapped 14CO2 as the ultimate product of
oxidation. The radiolabel experiments were conducted in 40.0-mL screw cap vials. The vials
were capped with Teflon-lined septa, allowing injection into the vial. An inner vial containing
0.4 N NaOH was placed inside the larger to serve as a CO2 trap. The trap was assayed by
scintillation counting. The vials, inoculated with biomass, contained 2.0 ml of media and 38.0
ml of head space. After the appropriate incubation period, vials were acidified with 0.2 ml of
2 N H2SO4 to drive off CO2. The sterile control values were subtracted from experimental
values when determining conversion to CO2. All data reported represent the mean value of
three vials. Mass balance calculation on sterile controls were conducted by assaying the
discharge per minute (dpm) in the NaOH trap, the aqueous phase, and a 2.0 ml hexane
extract. Greater than 85 percent of added TCE could be accounted for at the end of the
experimental incubation time. Counts in the sterile control were always less than 2 percent of
the total dpm added.
The initial phase of this study was designed to test the hypothesis that autotrophic ammonia-
oxidizing bacteria were responsible for TCE degradation. Figure 1 shows the effect of nitrapyrin,
an inhibitor of autotrophic ammonia oxidation, on TCE degradation (7,8). A number of batch
treatments on the biomass were carried out as part of this study. The effects of ammonia,
nitrate, phenol, and glucose, in both the presence and the absence of nitrapyrin, were
examined. None of the treatments tested, including those to which nitrapyrin was added, greatly
affect TCE mineralization. These results suggested that ammonia oxidizers were not responsible
for TCE mineralization.
A time course experiment was conducted over a range of TCE concentrations in both the
presence and the absence of ammonia. In this experiment, the oxidation of ammonia was
assayed by a colorimetric method to detect both nitrite and nitrate. For this experiment, three
TCE concentrations (0.021, 0.149, and 0.372 mg/L) and three time points (0, 20, and 44 hr)
were chosen. Ammonia supplemented (+ ammonia) and nitrate (- ammonia) batch tests were
inoculated with 0.003 mg of biofilter biomass. Data from this experiment are shown in Table
1. After 1 hr, no conversion of TCE to CO2 was observed at any TCE concentration, either with
or without ammonia. After 20 hr, TCE mineralization occurred at lower TCE concentrations.
No mineralization occurred at the highest TCE concentration at 20 hr or at 44 hr. Conversion
to CO2 in the vials at the lowest TCE concentration appeared to level off in 20 hr, showing little
increase after 44 hr. The 0.149 mg/L TCE concentration continued to demonstrate increased
TCE conversion at 44 hr. The effect of ammonia does not appear to be great at any
concentration. Aslight enhancement of mineralization in the ammonia-treated sample occurred
after 20 hr, and a slight decrease in the ammonia-treated sample occurred after 44 hr. Vials
from the 44-hr time point were assayed for nitrite and nitrate by colorimetric assay. No nitrite
or nitrate was detected in any vials, suggesting that little ammonia oxidation was occurring. The
nitrogen source had no effect on TCE mineralization. At this point, the biomass was examined
to determine which organisms were mineralizing TCE without co-metabolite addition.
The persistence of aromatic hydrocarbon oxidizers in the biofilter suggests that they may be
responsible for TCE oxidation. Enrichment cultures using biofilter biomass were incubated in
50.0 ml flasks in 10 ml of a mineral salts medium. These flasks were placed in 5-gal
desiccators and exposed to 0.5 mL of either toluene or benzene. These flasks grew to turbidity
230
-------
and produced a yellow metabolite indicative of aromatic ring cleavage. The yellow metabolite
was observed at the greatest dilutions tested (10"4). These enrichment cultures were tested for
mineralization in mineral salts in the absence of toluene or benzene, and showed high levels of
TCE mineralization. The predominant culture appearing on vapor phase plates appears to be
unique relative to previously described organisms and is being characterized. In contrast, TCE
mineralization assays of positive MPN cultures did not mineralize TCE.
Conclusions
Ammonia oxidizers are present in the biofilter, but at low levels.
Removal of ammonia from the medium did not effect TCE mineralization by the
biomass.
Addition of the inhibitor nitrapyrin did not effect TCE mineralization by the
biomass.
Nitrifier enrichment cultures from the biofilter did not mineralize TCE.
A high level of toluene/benzene oxidizers is present in the biofilter, and
enrichment cultures can mineralize TCE without addition of an organic co-
metabolite. These cultures are robust in the biofilter environment and have
persisted in the biofilter for over 1 year.
References
1. Arciero, D., T. Vanned!, M. Logan, and A.B. Hooper. 1989. Degradation of
trichloroethylene by the ammonia oxidizing bacterium Nitrosomonas europea. Biochem.
Biophys, Res. Commun. 159:640-643.
2. Hyman, M.R., R. Ely, S. Russell, K. Williamson, and D. Arp. 1993. Co-metabolism of
TCE by nitrifying bacteria. In: U.S. EPA Symposium on bio remediation of hazardous
wastes: Research, development and field evaluations (abstracts). EPA/600/R-93/054.
Washington, DC (May).
3. Ensley, B.D. 1991. Biochemical diversity of trichloroethylene metabolism. Ann. Rev.
Microbiol. 45:283-300.
4. Wilson, J.T., and B.H. Wilson. 1994. Biotransformation of trichloroethylene in soil.
Appl. Environ. Microbiol. 49:242-243.
5. Nelson, M.K.J., S.O. Montgomery, E.J. O'Neil, and P.M. Pritchard. 1986. Aerobic
metabolism of trichloroethylene by a bacterial isolate. Appl. Environ. Microbiol.
52:383-384.
231
-------
6. Wackett, L.P., G.A. Brusseau, S.R. Householder, and R.S. Hanson. 1989. Survey of
microbial oxygenases: Trichloroethylene degradation by propane oxidizing bacteria.
Appl. Environ. Microbiol. 55: 2,960-2,964.
7. Oremland, R.S.,and D.G. Capone. 1988. Use of "specific" inhibitors in biochemistry
and microbial ecology. Adv. Microb. Ecol. 10:285-383.
8. Powell, S.J., and J.I. Prosser. 1984. Inhibition of biofilm populations of Nitrosomonas
europea. Microb. Ecol. 24:43-50.
Table 1. TCE Mineralization by Biofilter Biomass With and Without Ammonia Addition
NH4 TCE
/*g
+ 0.4
+ 2.9
+ 7.25
0.4
2.9
7.25
TCE Mineralization
Ihr
%
0.0
0.0
0.0
0.0
0.0
0.0
/
-------
g
u
Figure 1. Nitrapyrin inhibition experiment. Biofilter biomass (0.01 mg biomass/vial) was used
to test the effect of an inhibitor on TCE oxidation in the presence of various inducer
compounds. Cultures were incubated in the presence of TCE (0.4 /ig/vial) for 5 days
prior to acidification. Results are reported as percent of added radiolabel recovered
as CO2: 1) heat-killed control, 2) time 0, 3) ammonia treated, 4) ammonia plus
nitrapyrin, 5) nitrate, 6) nitrate plus nitrapyrin, 7) phenol treated, 8) phenol plus
nitrapyrin, 9) glucose treated, and 10) glucose plus nitrapyrin.
233
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Bioremediotion of TCE: Risk Analysis for Inoculation Strategies
Richard A. Snyder and Malcolm S. Shields
Center for Environmental Diagnostics and Bioremediation, University of West Florida,
Pensacola, FL
P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Introduction
The introduction of non-native species has a colorful past for metazoan organisms. Controlled
introductions of non-native or genetically engineered bacteria to date have not been
documented to cause undesirable effects. The ubiquity of microorganisms has been largely
assumed, providing a rationale for the safe release of "non-native" bacteria. The ubiquitous
distribution argument assumes that all microorganisms have equal opportunity to occur in all
environments, and that selective pressures determining distribution and abundance will eliminate
introduced microorganisms that do not already occur in the target environment. The most
successful introductions have resulted from isolating an organism from the targeted environment,
modifying it, and returning it to its previous niche, e.g., Rhizobium spp. (1). An alternate
strategy that has proved effective is to modify the environment to provide a niche for the
phenotype of interest and to allow natural selective processes to occur (2). The success of these
strategies supports the ubiquity argument. The history of virulent pathogen distribution, however,
provides a model to warn us that the microbial world is not entirely homogeneous, and that
some environments may be subject to invasion by non-native microorganisms. With the
development of bacteria with potentially novel genetic combinations, we have a responsibility
to determine if released organisms will be constrained by the selective pressures of the target
environment.
Bacterial populations in nature are under constant selective pressures from physical and
chemical conditions, substrate availability for growth, competition between species, and
predatory/viral interactions. The balance of these forces determines both bacterial species
composition and individual species' abundance. The relative significance of the biological
factors (growth, competition, and predation) is determined by physical and chemical factors, as
the limits of individual species' tolerance are reached within trophic or contaminant gradients.
The addition of bacteria to environmental microbial communities may locally and temporarily
change the balance of selective pressures, but these cells would ultimately face the selective
forces of the target environment.
We have begun to address the abiotic and biological parameters for survival of Pseudomonas
cepac/a G4 PR-1 in laboratory microcosms utilizing ground water and sediment from the aquifer
beneath the Borden Canadian Armed Forces Base in Ontario, Canada. This is the site of a
proposed bioremediation test using a funnel-and-gate technique (3) to control ground-water flow
and force a trichloroethylene- (TCE-) contaminated plume through biocassettes colonized with
PR-1. This bacterium constitutively expresses a toluene orthomonooxygenase that mineralizes
TCE (4). The Borden aquifer is oligotrophic (3.5 to 6 mg DOC L"1), with a ground-water flow
of approximately 10 cm/day"1 through a well-sorted fine sand sediment (5). Determining the
234 1994 Symposium on Bioremediation of Hazardous Wastes
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transport of bacterial cells from a treatment zone as well as their survival necessitates the
development of field tracking methods for the organism and the plasmid that confers the ability
to mineralize TCE.
Approach and Preliminary Results
Results obtained from analysis of the behavior of PR-1 in aquifer material in laboratory tests will
be compared with the response at field scale during the release. This combination is hoped to
highlight basic biological characteristics of bacteria that can be assessed in the laboratory; in
this manner, future genetically engineered microorganism releases can be evaluated without
expensive testing of the organism in mesoscale or semicontained systems prior to release.
Characterization of Native Organisms
This initial phase is targeted towards identifying potential competitors, predators, and viruses in
the target environment. Selective plating and gene probing are employed to identify G4-like
organisms that may be displaced by the addition of PR-1 or that may contribute to the loss of
PR-1. Phenol-utilizing bacteria in the relatively pristine Borden aquifer represent about 3 percent
of the colony-forming units (CPUs) obtained on the ground-water medium R2-A. In contrast,
aquifer material from a TCE-contaminated site in Wichita, Kansas, had 62.6 percent of the R2-A
CPUs appearing on phenol plates. Whether these differences will affect PR-1 survival remains
to be determined.
We have enumerated protozoan predators of PR-1 in most probable number (MPN) growth
assays using PR-1 cells as the growth substrate. Both flagellates (391 gdw1), naked amoebae
(298 gdw"1), and testaceans (52 gdw"1) have been recovered that respond quickly and grow
very well on PR-1 cells. The species diversity and numbers of protozoans recovered by this
method are higher when sterile-filtered site ground water is used as a diluent rather than a
phosphate buffer (6) or sodium pyrophosphate as a mild surfactant.
Both viruses and competitive interactions between PR-1 and native bacteria isolated on plates
will be assayed using overlay plates with PR-1 cells and scoring for clearing zones. Native
viruses have not as yet been reported from aquifer environments, but their widespread
distribution in terrestrial and aquatic environments almost ensures their occurrence. Whether
there are active viruses against PR-1 cells in the target environment remains to be determined.
PR-1 Tracking
A monoclonal antibody has been prepared against the o-side chain of PR-1 IPS (7). We have
tested this monoclonal against a wide variety of bacteria, including other P. cepac/a strains and
isolates from the Borden aquifer, without evidence of cross reactivity. We have also tested the
use of the monoclonal by tracking survival of PR-1 in laboratory microcosms by direct
immunofluorescence and immunoblots of colonies from plates.
We are developing a polymerase chain reaction (PCR) detection assay for PR-1 utilizing the
unique junction sites of Tn-5 from the insertion mutagenesis in both the plasmid and the
genome. A set of three primers has been used to target an IS50 on the plasmid: two flanking
235
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primers and one asymmetrically situated in the interior sequence. This primer set yields a two
band "fingerprint" when the PCR product is run out on gels.
PR-1 Survival
Tests for survival of PR-1 in ground water, sediment slurries in shake flasks, and flow through
sediment columns are being conducted with the site material. Preliminary results suggest that
the abiotic conditions of the aquifer are not limiting to PR-1 survival. When we introduced 1
x 107 PR-1 ml"1 into sterilized ground water, no loss of PR-1 cells was observed by
immunofluorescent counts over 30 days, and plate counts dropped approximately an order of
magnitude and then stabilized for 25 days. Seven months later, both direct counts and plate
counts had dropped an additional order of magnitude each. In nonsterile ground water,
however, PR-1 was eliminated within 10 days, despite a stable population of total bacteria
determined by direct counts with the fluorochrome DAPI. In shaken sediment slurries, 2 x 107
PR-1 was eliminated within 4 days, and numbers of protozoa increased concomitant with the
decrease in PR-1, suggesting that predation may be an important mechanism for loss of the
bacterium from the system. Shifts in the bacterial community structure were apparent in the
slurries based on colony morphologies on the heterotrophic medium R2A.
Presterilized and nonsterile sediment columns were set up using 50 cm long by 2 cm diameter
tubes with 10 sampling ports sealed with silicone stoppers. A continuous culture of PR-1 set to
a generation time of approximately 100 hr and a cell yield of 6 x 107 cells ml"1 was used as
a source to feed to the top of the columns, with excess flow shunted off to a waste container.
Flow through the column was controlled by a pump at the column outflow and set to 10
cm/day1 as found in the aquifer. PR-1 cells were detected in the effluents with fluorescent
antibodies after one void volume passed through the column (4.5 days). After two void volume
replacements, the inflow of cells was stopped and switched to basal salts in an attempt to elute
PR-1 from the columns. As in the ground water and sediment slurries, PR-1 persisted at higher
levels in the sterile versus the nonsterile column, and we detected high numbers of bacterivorous
flagellates in the nonsterile system. Unlike the ground water and sediment slurries, PR-1
persisted through 22 days of elution in the presence of predators. Extraction of the sediments
with 0.1 percent sodium pyrophosphate at the termination of the experiment indicated that more
of the PR-1 cells in the nonsterile system were particle associated than free in the pore water
compared with the presterile system.
Conclusions
The preliminary results from our laboratory tests indicate that the abiotic conditions of the aquifer
will not affect the persistence of PR-1, but losses to biological vectors will be a major factor.
Cells free in the pore water will be quickly eliminated, but PR-1 may find refuge from predation
in association with sediment particles that will allow long-term persistence of the organism in the
target environment.
236
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Acknowledgments
This work was supported by EPA Cooperative Research Agreement CR822568-01-0. Steven
Franciscan! (NRC Post-Doc at USEPA GBERL) contributed sequencing data and probe design.
Thanks also to technicians Wendy S. Steffensen and Shiree Enfinger, and to undergraduate
assistants Margo Posten, John Millward, and Angela Andrews.
References
1. Pritchard, P.M. 1992. Use of inoculation in bioremediation. Curr. Opin. Biotechnol.
3:232-243.
2. Hopkins, G.D., J. Munakata, L. Semprini, and P.L. McCarty. 1993. Trichloroethylene
concentration effects on pilot field-scale in situ ground-water bioremediation by phenol
oxidizing microorganisms. Environ. Sci. Technol. 27:2,542-2,547.
3. Starr, R.C., J.A. Cherry, and E.S. Vales. 1992. A new type of steel sheet piling with
sealed joints for ground-water pollution control. Proceedings of the 45th Canadian
Geotechnical Conference, Toronto, pp. 75-1 - 75-9.
4. Shields, M.S., and M.J. Reagin. 1992. Selection of a Pseudomonas cepac/a strain
constitutive for the degradation of trichloroethylene. Appl. Environ. Microbiol.
58:3,977-3,983.
5. Sudicky, E.A. 1986. A natural gradient experiment on solute transport in a sand
aquifer: Spatial variability of hydraulic conductivity and its role in the dispersion process.
Water Resour. Res. 22:2,069-2,082.
6. Sinclair, J.L, and W.C. Ghiorse. 1987. Distribution of protozoa in subsurface sediment
of a pristine ground-water study site in Oklahoma. Appl. Environ. Microbiol. 53:1,157-
1,163.
7. Winkler, J., K.N. Timmis, and R.A. Snyder. Tracking survival of Pseudomonas cepac/a
introduced into aquifer sediment and ground-water microcosms. In preparation.
237
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Studies on the Aerobic/Anaerobic Degradation of Recalcitrant Volatile Chlorinated
Chemicals in a Hydrogel Encapsulated Biomass Biofilter
Rakesh Govind and P.S.R.V. Prasad
Department of Chemical Engineering, University of Cincinnati, Cincinnati, OH
Dolloff F. Bishop
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
Trichloroethylene (TCE) and tetrachlorethylene (PCE) are organic solvents most frequently
detected as ground-water contaminants. Both TCE and PCE undergo reductive dechlorination
in anaerobic environments. PCE is aerobically recalcitrant.
In an ongoing biofilter study, experimental work is being conducted to evaluate the potential of
gel-entrapped biomass for treating volatile chlorinated solvents, such as TCE and PCE, in the
gas phase. Entrapped biomass offers the possibility of aerobic/anaerobic environments in the
gel bead interior while aerobic conditions are maintained outside the bead. The reduced
environment allows contaminants such as TCE and PCE to be degraded in a biofilter column
packed with gel beads containing entrapped biomass.
Background
TCE degrades under anaerobic conditions, forming intermediates such as vinyl chloride,
dichloroethylenes, and ethylene (1). TCE also degrades under aerobic conditions usually as a
co-metabolite in the presence of a primary substrate. A number of compounds serve as primary
substrates for TCE degradation, including aromatics, such as toluene and phenol (2,3); alkanes,
such as methane and propane (4,5); and 2,4 dichlorophenoxyacetic acid (1). These
microorganisms degrade TCE because of the enzymes expressed in response to the primary
substrate; for example, toluene monooxgenase, which enables microorganisms to degrade
toluene and other aromatics, allows degradation of TCE. The primary metabolite-to-TCE ratio
has been found to be 2 g/g to 40 g/g in a recent study (6). Studies of TCE degradation (6)
were conducted in a gas-lift loop reactor. TCE concentrations of between 300 fig/L (60 ppmv)
and 3,000 /*g/L (600 ppmv) were degraded with 95 percent or better efficiency. Results of
another TCE study indicates that certain bacteria may be able to express the above enzyme even
in the absence of toluene or phenol (7). Recently, biofiltration studies with 25 ppmv gas-phase
inlet concentration of TCE in a celite-pellet packed bed have shown that TCE can be successfully
degraded with phenol present in the trickling nutrients (8).
238 1994 Symposium on Roremediotion of Hazardous Wastes
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Materials and Methods
Activated sludge biomass in an aqueous bioreactor was acclimated to toluene and TCE by
exposing the sludge to air contaminated with toluene and TCE for a period of 30 days. The
reactor was supplied with mineral nutrients, and the inlet and exit gas phase concentrations were
periodically analyzed. After acclimation was achieved, complete toluene conversion and about
30 percent TCE conversion were observed in the reactor. The biomass was then removed from
the reactor, mixed with k-Carragenan at 50°C, and extruded into 0.5 cm x 1.5 cm cylindrical
beads. The beads, once extruded, were quenched in a mineral medium and then packed in
a biofilter. The experimental biofilter consists of a 1 -in. diameter, 5-in. height bed packed with
k-Carragenan beads, with biomass encapsulated in each bead.
Contaminated air stream was obtained by injecting the substrate into the air stream by means
of a syringe pump (Harvard Apparatus, Model 11). The flow rate of air was controlled by an
MKS thermal mass flow controller (Controller 1259, Control Module 2-47). Because both air
flow rate and substrate injection rate were precisely controlled, uniformity of the substrate
composition in the air stream was ensured. The contaminated air stream was introduced at the
bottom of the biofilter to ensure uniform distribution. OECD nutrient solution was introduced at
the top of the biofilter bed at a flow rate of 300 ml/day, TCE concentrations were analyzed on
a Hewlett-Packard 571OA gas chromatograph with a 20-ft long, 1/8-in. diameter column
having the packing (PT 10-percent Alltech AT-100 on Chromosorb W-AW 80/100). Carrier gas
was nitrogen, and the detector was flame ionization (FID). Chloride ion concentrations in the
nutrient solution were measured by an Orion solid-state chloride ion combination electrode
(#9617BN) on an Accumet 1003 pH/mV/ISE meter. The pH of nutrient solutions was measured
by a combination pH electrode connected to the above meter. Ammonia-nitrogen concentration
in nutrient solution was measured by an Orion gas sensing ammonia electrode (#9512BN).
Nitrite ions in nutrient solutions were detected using a Hach NI-7 nitrite detection kit.
Results and Discussion
Separate studies were conducted with toluene at 300 ppmv inlet concentration at various gas
phase residence times. Figure 1 shows the removal efficiency as a function of gas phase
residence time for toluene. Toluene degrades aerobically in the biofilter, achieving 100-percent
removal efficiency at less than 1 min residence time.
Studies were also conducted with 25 ppmv inlet concentration of TCE at various gas phase
residence times. No toluene was present in the inlet gas stream. Complete mineralization of
TCE was observed at a gas residence time exceeding 4 min, suggesting a nonaerobic pathway.
Corresponding increases in chloride ion were observed in the liquid nutrient phase, which
demonstrated that TCE was mineralized to carbon dioxide and chloride ion. No partially
chlorinated byproducts were observed in the exit gas phase.
Studies are currently being conducted to 1) measure the dissolved oxygen concentration as a
function of depth in the hydrogel bead using a microsensor; 2) investigate the effect of bead size
on reactor removal efficiency for TCE (as the bead size decreases, the extent of the anaerobic
zone is expected to decrease); 3) develop a mathematical model for the hydrogel bead biofilter
and validate the model using the experimental data; and 4) extend the TCE study to other
chlorinated solvents, such as PCE.
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References
1. Marker, A.R., and Y. Kim. 1990. Trichloroethylene degradation by two independent
aromatic degrading pathways in Alcaligenes eutrophus JMP134. Appl. Environ.
Microbiol. 56:1,179-1,181.
2. Folsom, B.R., P.J. Chapman, and P.M. Pritchard. 1990. Phenol and trichloroethylene
degradation by Pseudomonas cepaa'a G4: Kinetics and interactions between substrates.
Appl. Environ. Microbiol. 56:1,279-1,285.
3. Wackett, L.P., and S.R. Householder. 1989. Toxicity of trichloroethylene to
Pseudomonas putida Fl is mediated by toluene dioxygenase. Appl. Environ. Microbiol.
55:2,723-2,725.
4. Wilson, J.T., and B.H. Wilson. 1985. Biotransformation of trichloroethylene in soil.
Appl. Environ. Microbiol. 49:242-243.
5. Kampbell, D.H., J.T. Wilson, H.W. Read, and T.T. Stocksdale. 1987. Removal of
volatile aliphatic hydrocarbons in a soil bioreactor. JAPCA 37:1,236-1,240.
6. Ensley, B.D. 1993. Biodegradation of chlorinated hydrocarbons in a vapor phase
reactor. Final report under contract no. 02112407. Springfield, VA: National
Technical Information Service.
7. Shields, M.S., R. Schaubhut, R. Gerger, M. Reagin, C. Somerville, R. Campbell, and J.
Hu-Primmer. 1993. Bioreactor and in situ applications of a constitutive
trichloroethylene degrading bacterium. Paper 97c. Presented at the AlChE Spring
National Meeting, Houston, TX (April).
8. Bishop, D.F., and R. Govind. 1993. Environmental remediation using biofilters.
Presented at Frontiers in Bioprocessing III, Boulder, CO (September 1 9-23).
240
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I
• PM
V
£
w
I
Toluene(300 ppmv)
TCE(25ppmv)
40
20
10 20 30 40
Residence Time(min.)
Figure 1. Plot of percent removal efficiency for toluene and TCE in the gel-bead biofilter with
encapsulated biomass. Toluene and TCE studies were conducted separately.
241
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Poster Session
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Pilot-Scale Evaluation of Nutrient Delivery for Oil-Contaminated Beaches
Michael Boufadel and Makram T. Suidan
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Albert D. Venosa
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
In situ bioremediation is emerging as an efficient and economical strategy for the cleanup of oil-
contaminated beaches. The mechanisms and routes of nutrient delivery in the presence of tides,
however, are not well understood. The main objective of this project is to investigate these
phenomena to identify the best nutrient application technology.
Results and Discussion
For this purpose, a pilot-scale beach simulation unit is being built. This unit will be 8 m long,
0.60 m wide, and 1.8 m tall, and will be equipped with a pneumatic wave generator. The unit
is intended to simulate waves that propagate perpendicularly to beaches. The height of the unit
was selected to permit investigation of tidal effects. Prior to construction of the pilot-scale unit,
a small bench-scale unit was constructed and tested to observe wave generation and beach
erosion. The results observed from the bench-scale unit were very encouraging. A periodic
wave was generated and sustained over several days.
The initial part of the study will investigate nutrient transport using tracer studies. A distributed
computer model will be developed in parallel. The model parameters will be estimated from
the results of tracer studies. Subsequently, the model will be evaluated at pilot scale and later
on real beaches. The experimental data will also be evaluated against the mathematical model
developed by Wise et al. (1).
References
1. Wise, W.R., O. Guvn, F.J. Molz, and S.C. McCutcheon. 1 994. Nutrient retention time
in a high-permeability oil fouled beach. In press.
1994 Symposium on Bioremediation of Hazardous Wastes 245
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Metabolites of Oil Biodegradotion and Their Toxitity
Peter J. Chapman
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Michael E. Shelton
University of Minnesota, Department of Biochemistry, St. Paul, MN
Simon Akkerman
University of West Florida, Center for Environmental Diagnostics and Bioremediation,
Pensacola, FL
Steven S. Foss, Douglas P. Middaugh, and William S. Fisher
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Development of strategies for the bioremediation of crude oil and refinery processed petroleum
must build on a basic understanding of microbial degradation of oil and its many chemical
constituents, as well as the limitations imposed on these processes by environmental factors.
Numerous studies document microbial activities on bulk oil and its components (1,2), yet little
is known of the formation, accumulation, and toxicity of compounds during oil biodegradation.
Recent reports of petroleum-derived oxidation products in ground water (3) and in the tissues
of mollusks (4) indicate the need to characterize products formed during crude oil
biodegradation and to assess their environmental effects. This work addresses some of these
questions.
Amounts of neutral and acidic materials recovered from different oil-degrading cultures (from
both marine and terrestrial sources) were significantly greater than from sterile controls.
Biologically generated neutral materials were toxic (100-percent mortality) to larvae of
Mysidopsis ban/a (5), to grass shrimp embryos (6), and to embryos of Menidia beryllina (7) at
concentrations matching those at which they were formed in cultures. Menidia embryos
exhibited developmental defects. Work is continuing to define the nature of the toxic
components of these neutral fractions, their precursors in oil, and the microorganisms and
processes that lead to their formation.
References
1. Atlas, R.M. 1984. Petroleum microbiology. New York, NY: Macmillan.
2. Leahy, J.G., and R.R. Colwell. 1990. Microbial degradation of hydrocarbons in the
environment. Microbiol. Rev. 54:305-315.
3. Cozzarelli, I.M., M.J. Baedecker, R.P. Eganhouse, and D.F. Goerlitz. 1994. The
geochemical evolution of low-molecular-weight-organic acids derived from the
degradation of petroleum contaminants in ground water. Geochim. Cosmochim. Acta
58:863-877.
246 1994 Symposium on Bioremediation of Hazardous Wastes
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4. Bums, K.A. 1993. Evidence for the importance of including hydrocarbon oxidation
products in environmental studies. Mar. Pollut. Bull. 26:77-85.
5. U.S. EPA. 1987. Short-term methods for estimating the chronic toxicity of effluents and
receiving waters to marine and estuarine organisms. EPA/600/4-87/028. Cincinnati,
OH. pp. 171-238.
6. Fisher, W., and S. Foss. 1993. A simple test for toxicity of number 2 fuel oil and oil
dispersants to embryos of grass shrimp, Pa/aemonefes pugio. Mar. Pollut. Bull. 26:385-
391.
7. Middaugh, D.P., R.L Thomas, S.E. Lantz, C.S. Heard, and J.G. Mueller. 1994. Field-
scale testing of a hyperfiltration unit for removal of creosote and pentachlorophenol
from ground water: Chemical and biological assessment. Arch. Environ. Contam.
Toxicol. 26:309-319.
247
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The Use of In Situ Carbon Dioxide Measurement To Determine
Bioremediation Success
Richard PJ. Swannell
Biotechnology Services, National Environmental Technology Centre, AEA Technology,
Oxon, United Kingdom
Francois X. Merlin
CEDRE, Plouzane, Brest, France
Introduction
Monitoring bioremediation success involves complex analytical chemistry and time-consuming
microbiology. Potentially, a more valuable tool for the oil spill treatment specialist would be one
that enabled the efficacy of a bioremediation strategy to be determined in real time in situ. This
poster describes preliminary research on a method for making in situ measurements of
bioremediation efficacy based on the estimation of CO2 evolution. These studies were conducted
in the field near Landevennec, France. The trial involved the oiling of six plots on a beach
consisting largely of shale on a clay base. Three plots were amended with a slow-release
inorganic nutrient, and three plots remained untreated as controls. Three plots were also
delimited on the same beach to act as unoiled controls.
Methods
Two sampling devices were made from stainless steel, consisting of a shallow cylinder (0.2 m
high and 1.1 m in diameter) sealed at one end with a base plate. The base plate was pierced
with two steel tubes connected to valves on the outside of the device. The samplers were pushed
gently into the beach surface, with the base plate facing upward and the valves open to the air.
The CO2 analyzer was then connected to the valves, and air from the sampler was circulated
through it, giving an initial CO2 reading. The CO2 level was then monitored periodically over
the next 5 to 20 min. Measurements were taken at the same coordinates on the oiled controls,
the unoiled controls, and the plots treated with oil and fertilizer. Readings were made 26, 1 1 6,
and 144 days after oiling. Nutrients were applied 11 days after oiling and monthly thereafter.
Results and Discussion
On each sampling day, the rate of CO2 evolution was enhanced on oiled plots treated with
fertilizer in comparison to oiled controls and unoiled controls. The largest difference was noted
15 days after nutrient addition when the rates increased from 3.1 to 4.0 ppm CO2.min"1 on the
oiled controls to 12.6 to 22.3 ppm CO2.min"' on the fertilized plots. The unoiled controls gave
values between 2.8 and 4.2 ppm CO2.min''. These data suggest that nutrient addition
stimulated the CO2 evolution rates when compared with untreated controls. The rates were
found to decrease in subsequent measurements of the fertilized plots but were still 1.5 to 2.0
times greater than the controls, suggesting the stimulation in CO2 production was sustained.
248 1994 Symposium on Bioremediation of Hazardous Wastes
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Conclusion
These preliminary data suggest that addition of fertilizer to oiled plots stimulates CO2 evolution.
Whether this stimulation reflects enhanced oil biodegradation, as we suspect, remains to be
proven absolutely using gathered chemical samples. Further, although the measured values are,
by their nature, relative rates and not absolute indicators of CO2 production, the results suggest
that this technique may provide useful data when examining the efficacy of bioremediation
strategies and products on contaminated shorelines. A second field trial conducted in the United
Kingdom in the summer of 1994, funded by EPA, will allow a more detailed evaluation of this
promising technique.
249
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Toxicant Generation and Removal During Crude Oil Degradation
Linda E. Rudd
North Carolina State University, Raleigh, NC
Larry D. Claxton, Virginia S. Houk, Ron W. Williams
U.S. Environmental Protection Agency, Research Triangle Park, NC
Jerome J. Perry
North Carolina State University, Raleigh, NC
Introduction
As microorganisms are promoted for environmental bioremediation efforts, the potential risk of
adverse effects of pollutant exposure to the microbes must be assessed. Although fungi (1,2) and
bacteria (3-5) degrade hydrocarbons, thegenotoxic consequences of degradation have not been
addressed. Bacterial species use enzyme systems to convert hydrocarbons to metabolites with
increased toxicity (6-8) or to mineralize toxic compounds during metabolism (9). This study
involves interactive use of microbial culture, analytical chemistry, and mutagenicity bioassays to
investigate the genotoxicity of the oil degradation process. Following degradation by two fungi,
Cunninghamella elegans and Penicillium zonatum (10,11), crude oils of low, moderate, and
high mutagenicity are tested for their resulting mutagenic activities.
Methods
C elegans ATCC 36112 or P. zonatum ATCC 24353 was inoculated into 500 mL L-Salts
medium (12) with 5 mL of crude oil. Flasks were incubated at 30°C for 4 to 30 days; at 2-day
intervals, flasks were sacrificed, and crude oil was extracted with methylene chloride by a
modification of the method used by Cemiglia (10,13). Oil mass determinations were calculated
from oil residue weights. Extracted oils were analyzed for conversion of straight chain
hydrocarbons by gas chromatography and for mutagenicity by the spiral Salmonella assay
(14,15). Controls included "weathered" (uninoculated) oil flasks and fungi grown on 2-percent
glucose to test for mutagenic products from fungal growth alone ("fungal mat controls").
Results
Pennsylvania and Cook Inlet Alaska crude oils' mycelial mat weights are directly proportional
to biologically linked oil degradation. The fungi consistently form sturdy mats with Pennsylvania
crude; the Cook Inlet mat, however, is more fragile. Mat weights are not proportional to West
Texas sour crude utilization; sturdy mats are not consistently produced by either organism even
though the oil is utilized as the sole carbon source. The loss of oil mass is evidenced by a
significant decrease in C7 to C20 hydrocarbons as incubation time increases. Weathered
samples of the three oils do not exhibit changes in mutagenic activity over time. The
250 1994 Symposium on Bioremediation of Hazardous Wastes
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mutagenicity of the most potent oil, Pennsylvania crude, is significantly reduced following
degradation by either fungus (Table 1). The activity of the weakly mutagenic West Texas crude
exhibits little change upon treatment (data not shown). The nonmutagenic Cook Inlet Alaska
crude oil becomes mutagenic when incubated with either fungus (Table 2).
Conclusion
The fungal species used in this study may convert crude oil hydrocarbons to products more
mutagenic than the original compound. Further studies in progress address effects of
oxygenation, nitrogen and phosphorus enrichments, and surfactant addition to the experimental
system.
References
1. Kirk, P.W., and A.S. Gordon. 1988. Hydrocarbon degradation by filamentous marine
higher fungi. Mycologia 80(6):776-782.
2. Jobson, A., F.D. Cook, and D.W.S. Westlake. 1972. Microbial utilization of crude oil.
Appl. Microbiol. 23(6):1,082-1,089.
3. Cemiglia, C.E. 1992. Biodegradation of polycyclic aromatic hydrocarbons.
Biodegradation 3:351-368.
4. Perry, J.J. 1968. Substrate specificity in hydrocarbon utilizing microorganisms. Antonie
van Leeuwenhoek 34:27-36.
5. Walker, J.D., L Petrakis, and R.R. Colwell. 1976. Comparison of the biodegradability
of crude and fuel oils. Can. J. Microbiol. 22:598-602.
6. Gibson, D.T., V. Mahadevan, D.M. Jerina, H. Yagi, and HJ.C. Yeh. 1975. Oxidation
of the carcinogens benzo[a]pyrene and benzo[a]anthracene to dihydrodiols by a
bacterium. Science 1 89:295-297.
7. Middaugh, D.P., S.M. Resnick, S.E. Lantz, C.S. Heard, and J.G. Mueller. 1993.
Toxicological assessment of biodegraded pentachlorophenol: Microtox™ and fish
embryos. Arch. Environ. Contam. Toxicol. 24:165-172.
8. Liu, D., R.J. Maguire, G.J. Pacepavicius, and E. Nagy. 1992. Microbial degradation
of polycyclic aromatic hydrocarbons and polycyclic aromatic nitrogen heterocyclics.
Environ. Toxicol. Water Qual. 7(4):355-372.
9. Burback, B.L., and J.J. Perry. 1993. Biodegradation and biotransformation of ground-
water pollutant mixtures by M/cobaderium vaccae. Appl. Environ. Microbiol.
59(4):1,025-1,029.
10. Cemiglia, C.E., and J.J. Perry. 1973. Crude oil degradation by microorganisms
isolated from the marine environment. Zeitschrift fur Allg. Mikrobiologie 13(4):299-306.
251
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11. Hodges, C.S., and J.J. Perry. 1 973. A new species of Eupenidllium from soil.
Mycologia 65(3):697-702.
12. Leadbetter, E.R., and J.W. Foster. 1958. Studies on some methane-utilizing bacteria.
Arch. Mikrobiol. 30:91-118.
13. Cemiglia, C.E. 1975. Oxidation and assimilation of hydrocarbons by microorganisms
isolated from the marine environment. Dissertation. Raleigh: North Carolina State
University.
14. Moron, D., and B.N. Ames. 1983. Revised methods for the Salmonella mutagenicity
test. Mutation Res. 113:173-212.
15. Houlc, V.S., S. Schalkowsky, and L.D. Claxton. 1989. Development and validation of
the spiral Salmonella assay: An automated approach to bacterial mutagenicity testing.
Mutation Res. 223:49-64.
Table 1. Pennsylvania Crude (+ + + highly mutagenic)
Incubation Mutagenic % Biological Mat
Organism (Days) Response Loss* Weight (g)
C elegans 2 +++ 7% 0
4 ++ 9% 0.2
6 ++ 17% 0.6
8 + 23% 0.6
10 + 42% 1.0
12 - 26% 0.4
14 + 32% 0.5
P. zonafum 2 ++ 8% 0
4 ++ 16% 0
6 - 21% 0.4
8 - 27% 0.5
10 - 33% 0.3
12 - 29% 0.4
14 - 18% 0.3
*Biological Loss = Amount of oil used by fungus (corrected for procedural nonbiological oil
loss)
252
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Table 2. Alaska Crude (- nonmutagenic)
Incubation Mutagenic % Biological Mat
Organism (Days) Response Loss Weight (g)
C. e/egans 2 - 4% 0
4 - 4% 0
6 - 19% 0.1
8+19% 0.1
10 + 18% 0.1
12 ++ 18% 0.1
14 ++ 16% 0.1
P. zonafum 2 5% 0
4 +/- 13% 0
6 + 16% 0.1
8 +/- 28% 0.2
10 +/- 24% 0.2
12 +/- 27% 0.2
14 + 24% 0.1
253
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Intrinsic Bioremediotion of JP-4 Jet Fuel Contamination at
George AFB, California
John T. Wilson, Michael L. Cook, and Don H. Kampbell
U.S. Environmental Protection Agency, Robert S. Kerr Environmental Research Laboratory,
Ada, OK
Intrinsic bioremediation is difficult to evaluate from monitoring well data. Depending on the
screened interval and the pumping rate, a well may produce water from an uncontaminated part
of the aquifer, resulting in a sample that is greatly diluted by clean water. In addition, a well
may miss the plume entirely. Both effects give the false impression that in situ biological
processes are attenuating the contaminants. A rigorous demonstration of intrinsic bioremediation
should include 1) information on the use of available electron acceptors and 2) information on
the concentration of a tracer associated with the plume that can be used to correct for dilution.
Ground water at George Air Force Base (AFB) was contaminated by a release of JP-4 jet fuel.
Well MW 24 is near the center of the spill. Well MW 25 is 500 ft from well MW 24 in a
direction that is perpendicular to ground-water flow. Wells MW 27, 29, and 31 are along a
flow path down-gradient of well MW 24. The plume velocity is near 100 ft/yr.
Oxygen and nitrate were depleted downgradient of the spill. The concentration of benzene was
reduced more than 300-fold, while the concentration of a more recalcitrant compound, 1,2,3-
trimethylbenzene, was only reduced three-fold. After correcting for dilution, benzene
concentrations were reduced at least 100-fold due to intrinsic bioremediation.
Table 1. Intrinsic Bioremediation of Benzene and Toluene
Location
Oxygen
Nitrate
Benzene
Toluene
1,2,3-
Trimethylbenzene
MW24
Center of oil
lens
<0.5
0.8
1,620
1,500
73
MW25
Edge of oil
lens
if
lr
8.0
3.7
t
v
194
604
39
MW27
700ft
away
MW29
1,200ft
away
MW31
1 ,800 ft
away
ng/liter)
0.6
0.4
<0.5
0.3
1.1
3.1
wg/liter)
80
<0.5
56
4.8
<0.5
20
<0.5
<0.5
<0.5
254
1994 Symposium on Bioremediation of Hazardous Wastes
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Field Treatment of BTEX in Yadose Soils Using Vacuum Extraction or Air Stripping
and Biofihers
Rakesh Govind
Department of Chemical Engineering, University of Cincinnati, Cincinnati, OH
E. Radha Krishnan and Gerard Henderson
International Technology Corporation, Cincinnati, OH
Dolloff F. Bishop
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
Spills of fuels and leaking fuel tanks represent a major source of vadose soil contamination. This
contamination, which includes the aromatic hydrocarbons benzene, toluene, ethylbenzene, and
thexylenes (BTEX), leaches through vadose soil into ground water. Aromatic hydrocarbons pose
health risks when ground water is used as a drinking water supply.
EPA's Risk Reduction Engineering Laboratory (RREL), in cooperation with the University of
Cincinnati, is developing engineering systems to bioremediate fuel-contaminated vadose soils
or ground water. Vacuum extraction of soils or air stripping of ground water, which transfers the
volatile organic compounds (VOCs) from the soils or ground water to air, is combined with air
biofiltration to achieve treatment.
Field Demonstration
Two types of air biofilters will be studied: 1) packed beds with ceramic pellets, 6-mm average
diameter (Celite, Manville Corporation), as the packing material; and 2) straight-passages
ceramic monoliths with 50 square passages per square inch (as shown in Figure 1). A schematic
of the experimental system is shown in Figure 2. The aerobic mixed cultures, from an activated
sludge treatment plant, are immobilized on the surface of the packing. Nutrient solution, needed
for microbial growth, is trickled down through the packed bed, with the contaminated air flowing
countercurrent to the nutrient flow. The gas residence time in each biofilter is varied between 1
and 3 minutes. Electricity and water are used to raise the temperature of the extracted air to
approximately 30°C and to prehumidify the air. A syringe pump is used during startup to
contaminate the air with jet fuel to establish the biofilms in the biofilters.
The biofilters will be constructed at EPA's Test and Evaluation (T&E) Facility in Cincinnati. The
system will include gas chromatography for analyses of the influent and effluent gas streams
from each biofilter. The biofilm on the support media will be preacclimated to jet fuel (JP-4)
hydrocarbons. The skid-mounted biofilters with acclimated biofilms will be transported to the site
for connection to the vacuum extraction or air stripping system.
1994 Symposium on ffioremediation of Hazardous Wastes 255
-------
The site for the field demonstration has not yet been selected but is likely to be an air force base
in Ohio. The performance of the integrated system will be characterized for approximately 3
months.
TREATED AIR
NUTRIENTS
BIOFILM
STRAIGHT-PASSAGES
MONOLITH
t
CONTAMINATED AIR
Figure 1. Schematic of the straight-passages monolith media.
256
-------
Blower
HXI-
ecycle SamP'f Porf
jt
Biofilter
Media
Sample Port Nutrje|
&
1 » 1 4 1
Sample Ports
Flow Control
Valves
.. ii i*^s\ t^s\ it
* 1 \ IXJ .
rtn...
n^i — 1 1 *
t
-«
-------
TCE Remediation Using a Plasmid Specifying Constitutive TCE Degradation:
Alteration of Bacterial Strain Designs Based on Field Evaluations
Malcolm S. Shields, Allison Blake, Michael Reagin, Tracy Moody, Kenneth Overstreet, and
Robert Campbell
Center for Environmental Diagnostics and Bioremediation, Department of Cellular and
Molecular Biology, University of West Florida, Pensacola, PL
Stephen C. Francesconi
National Research Council, U.S. Environmental Protection Agency, Environmental Research
Laboratory, Gulf Breeze, FL
P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
An integrated study was undertaken to determine the potential for field application of altered
strains of Pseudomonas cepacia G4 (PR123 and PR131) developed by us for the bioremediation
of trichloroethylene (TCE). The investigation demonstrated the ability of PR123 to degrade TCE
without inducer substrates via the constitutive expression of toluene orfbo-monooxygenase
(TOM). Two fundamental areas of research are detailed: 1) the effectiveness of the PR123
phenotype in a field bioreador and 2) laboratory transfer of the constitutive degradative
phenotype to two new bacterial strains selected for their capacity to colonize bioreactor matrices.
PR123 was field tested in a 100-L plugged flow reactor receiving contaminated water at 2 L/min
and a daily batch input of cells (6 L) for a period of 2 weeb. Under these conditions, PR123
was able to effectively degrade TCE and c/s-DCE in contaminated aquifer water at
concentrations up to 700/
-------
Dechlorinotion With o Biofilm-Elecfrode Reactor
John W. Norton and Makram T. Suidan
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Albert D. Venosa
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory,
Cincinnati, OH
Introduction
Pentachlorophenol (PCP) is a pesticide and bactericide that is widely used in the wood and
leather preserving industries (1). PCP, however, is a suspected mutagen and carcinogen (2),
and, in 1986, EPA set a maximum contaminant level of 0.001 mg/L Superfund documents
have reported PCP levels as high as several hundred milligrams per liter in contaminated ground
water.
According to Krumme (3), in systems without a carbon or energy source PCP has been shown
to be dechlorinated and mineralized to about 40 percent of the influent concentration (3). In
systems using a co-substrate, it has been demonstrated that PCP can be dechlorinated up to
99.9 percent (4). The addition of external carbon and energy sources, however, could pose
difficulties in both in situ and ex situ treatment of contaminated sites. Cell growth is enhanced
by the addition of these carbon and energy sources, and the disposal of the resultant sludge can
prove to be costly. In situ treatment of PCP can also pose problems; the addition of a carbon
and energy source into the ground might cause the formation of hazardous soluble compounds.
Thus, there is reason to examine methods of enhancing microbial activity that could reduce or
remove the need to provide external energy and carbon sources.
Results and Discussion
The objective of this project is to examine the dechlorination and mineralization of PCP under
anaerobic conditions using the electrolytic reduction of water to provide an external energy
source and hydrogen donor. Researchers have demonstrated that biological processes can be
enhanced when subjected to an electric current (5,6). These studies examined the role of
electrolytically produced hydrogen in the denitrification of wastewater. Islam et al. (6) found a
correlation between the applied current and the removal efficiency of the reactor system and
determined the optimum current to be 20 mA, for which the removal efficiency was greater then
98 percent.
The reactor is a fixed-film chemostat with trace salts and nutrients added. PCP dissolved in
ethanol is added at two different feed concentrations (5 mg/L and 50 mg/L), with a current of
15.0 mA across the junction. The flow rate is 5 L/day, with a hydraulic detention time of 0.44
days. The reactor was seeded with biomass from an anaerobic, expanded-bed, granular
activated carbon (GAC) reactor that had been successfully dechlorinating PCP. The gas
production of about 96 ml/day of methane and the intermediates in the effluent indicate the
presence of an active growing biofilm.
1994 Symposium on Bioramediation of Hazardous Wastes 259
-------
Good dechlorinotion of PCP was achieved, with about 0.24 percent of the influent PCP
remaining as PCP, 0.1 percent as tetrachlorophenol, 0.87 percent as trichlorophenol, 10.28
percent as dichlorophenol, and about 55 percent as monochlorophenol on a molar basis
(Figure 1). The remaining 33.5 percent was presumed to be mineralized to HCI, CO2, and
H2O. Currently, the feed alcohol concentration is being reduced stepwise as the biofilm
stabilizes to the operating concentration.
Work on this project is continuing; new data will be included in the poster presentation.
References
1. Crosby, D.G. 1981. Environmental chemistry of pentachlorophenol. Pure Appl. Chem.
53:1,051-1,080.
2. Keith, L.H., and W.A. Telliard. 1979. Priority pollutants, I. A perspective view. Environ.
Sci. Technol. 13:416-423.
3. Krumme, M.L, and S.A. Boyd. 1988. Reductive dechlorination of chlorinated phenols
in anaerobic upflow bioreactors. Water Res. 22 (2):171-1 77.
4. Guthrie, M.A., EJ. Kirsch, R.F. Wukasch, and C.P.L Grady, Jr. 1984.
Pentachlorophenol biodegradation. II. Anaerobic. Water Res. 18(4):451-461.
5. Mellor, R.B., J. Ronnenberg, W. Campbell, and S. Diekmann. 1992. Reduction of
nitrate and nitrite in water by immobilized enzymes. Nature 355(20):717-719.
6. Islam, S., J.R.V. Flora, M.T. Suidan, P. Biswas, and Y. Sakakibara. 1993. Paper No.
AC93-039-002. Proceedings of the Water Environment Federation 66th Annual
Conference and Exposition, pp. 217-225.
260
-------
POP and intermediates vs time
Bio-electrolytic reactor, 15.0 mamps
0.02
o
E 0.01
P o. PfW.
367
387
407
Days
427
447
phenol ° mcp
dcp
* tri
tetra
o pep • influent]
Figure 1. PCP and intermediates versus time (bioelectrolytic reactor, 15.0 mA).
261
-------
Degradation of a Mixture of High Molecular-Weight Polycydic Aromatic
Hydrocarbons by a Mycobaderium Species
I. Kelley, A. Selby, and Carl E. Cemiglia
U.S. Food and Drug Administration, National Center for lexicological Research,
Division of Microbiology, Jefferson, AR
A Mycobacterium sp., which was previously tested for its ability to mineralize several individual
polycyclic aromatic hydrocarbons (PAHs), simultaneously degraded phenanthrene, anthracene,
fluoranthene, pyrene, and benzo[a]pyrene in a six-component synthetic mixture. Chrysene,
however, was not degraded to any significant extent. When provided with a primary carbon
source, the Mycobacterium sp. degraded more than 74 percent of the total PAH mixture during
6 days of incubation. The Mycobacterium sp. appeared to degrade phenanthrene preferentially.
No significant difference in degradation rates was observed between fluoranthene and pyrene.
Anthracene degradation was slightly delayed, but, once initiated, degradation proceeded at
approximately the same rate. Benzofa]pyrene was degraded to a lesser extent. Additionally,
degradation of a crude mixture of benzene-soluble PAH components from sediments resulted
in a 47-percent reduction of the material in 6 days compared with autoclaved controls. Initial
experiments using environmental microcosm test systems indicated that mineralization rates of
individual [14C] labeled compounds were significantly lower in the mixtures than in equivalent
doses of these compounds alone. Mineralization of the complete mixture was estimated
conservatively to be between 49.7 percent and 53.6 percent in 12 weeks. Mineralization was
nearly 50 percent within 30 days of incubation when all compounds were radiolabeled. These
results strengthen the argument for the potential application of this Mycobacferii/m sp. in
bioaugmentation of PAH-contaminated wastes.
262 1994 Symposium on Boreimd'ntion of Hazardous Wastes
-------
Potentiation of 2,6-Dinitrotoluene Bioactivation by Atrazine in Fischer 344 Rats
S. Elizabeth George, Robert W. Chadwick, Michael J. Kohan, and Joycelyn C. Allison
U.S. Environmental Protection Agency, Health Effects Research Laboratory,
Research Triangle Park, NC
Sarah H. Warren and Ron W. Williams
Integrated Laboratory Systems, Research Triangle Park, NC
Larry D. Claxton
U.S. Environmental Protection Agency, Health Effects Research Laboratory,
Research Triangle Park, NC
Because of widespread use, pesticides are often found as co-pollutants at hazardous waste sites
and other sites contaminated by xenobiotics. The herbicide atrazine is used as a weed control
agent during the cultivation of food crops and is found frequently as a ground-water
contaminant. To study atrazine as a co-pollutant, this study explored the effect of atrazine
treatment on the bioactivation of the promutagen 2,6-dinrtrotoluene (2,6-DNT). For 5 weeks,
male Fischer 344 rats (21 d) were administered p.o. 50 mg/kg of atrazine. At 1, 3, and 5
weeks, both control and atrazine-pretreated rats were administered 75 mg/kg of 2,6-DNT by
gavage and were placed into metabolism cages for urine collection. Following urine
concentration, a microsuspension modification of the SalmoneMa assay with and without
metabolic activation was used to detect urinary mutagens. No significant change in mutagen
excretion was observed in atrazine-pretreated rats. A significant increase, however, was detected
in direct-acting urine mutagens from rats receiving atrazine and 2,6-DNT at Week 1 (359 ±
68 revertants/mL versus 621 ±96 revertants/mL) and Week 5 (278 ± 46 revertants/mL versus
667 ± 109 revertants/mL) of treatment. Urinary mutagenicity was accompanied by an increase
in small intestinal nitroreductase activity. At Week 5, elevations in large intestine nitroreductase
and 6-glucuronidase were observed. This study suggests that atrazine potentiates the
metabolism and excretion of the mutagenic metabolites of 2,6-DNT by modifying the intestinal
enzymes responsible for promutagen bioactivation. [This is an abstract of a proposed
presentation and does not necessarily reflect EPA policy.]
1994 Symposium on Koremediation of Hazardous Wastes 263
-------
Effects of Loctobocillus Reuteri on Intestinal Colonization of
Bioremedifltion Agents
Mitra Fiuzat
Department of Microbiology, North Carolina State University, Raleigh, NC
S. Elizabeth George
U.S. Environmental Protection Agency, Health Effects Research Laboratory,
Research Triangle Park, NC
Walter J. Dobrogosz
Department of Microbiology, North Carolina State University, Raleigh, NC
Introduction
Lactobadllus reuteri is the predominant heterofermentative species of Lacfobaa'//us inhabiting
the gastrointestinal (Gl) tract of humans, swine, poultry, rodents, and a number of other animals
(1). Studies on chicks and poults have shown that oral (probiotic) treatment of flocks at hatch
with viable, host-specific L reuferi prior to challenge at Day 1 posthatch with S. typhimurium
reduces mortality by 50 percent to 75 percent compared with untreated flocks (2). L reuteri is
unique among bacteria in its ability to produce and secrete the potent, broad-spectrum
antimicrobial agent reuterin when incubated in the presence of glycerol under physiological
conditions similar to those which exist in the Gl tract (3,4) Reuterin has been purified,
chemically characterized, and identified as an equilibrium mixture of monomeric, hydrated
monomeric, and cyclic dimeric forms of 3-hydroxypropionaldehyde (5,6).
The environmental release of naturally occurring, mutant, and recombinant microorganisms has
prompted questions concerning human health and environmental effects (7,8). To date, a variety
of microbes have been released into the environment for many uses. Currently, investigators are
engineering microorganisms, primarily pseudomonads, for their ability to degrade hazardous
environmental contaminants such as pentachlorophenol, 2,4,5-trichlorophenoxyacetate,
chlorobenzoates, andtrichloroethylene. Pseudomonasspp., however, have long been recognized
as opportunistic pathogens, readily occurring in serious secondary infections, and they have
been linked to major infections in immunosuppressed and leukemia patients as well as those
treated with antibiotics (9-11). Because of the clinical significance of Pseudomonas spp., their
potential health effects have been studied in terms of their ability to compete and survive in a
CD-I mouse model system (12,13). The effects of antibiotics on theirsurvival and translocation
to other organs have also been investigated. Results from these studies indicate that
environmental pseudomonads can survive in the Gl tract for up to 14 days, where they can alter
the normal microbiota. Their translocation to the spleen and/or liver also occurs, indicating the
potential fora systemic infection (14,15). This research was undertaken to determine if L reuteri
prophylaxis could mitigate the pathogenic effects of these Pseudomonas spp. in the mouse
model system.
264 1W4 Symposium on ffioremediation of Hazardous Wastes
-------
Materials and Methods
Bacterial Strains
Three Pseudomonas aeruginosa strains were used in this study. Strain BC16 degrades
polychlorinated biphenyl, strain AC869 degrades 3,5-dichlorobenzoate, and strain PAO is a
clinical isolate. Four mouse-specific L reuteri strains were used.
Animals
Thirty-day-old CD-I male mice were used in this study. These animals were administered L
reuteri (109 colony-forming units (CFU)/mL) in sterilized water daily for 5 days prior to
Pseudomonas administration by gavage (one group 108 CPU and the other group 109 CPU)
and thereafter during the entire experiment. Control mice were given only sterilized water. On
Day 2 and Day 7 after the Pseudomonas administrations, the animals were sacrificed, and their
livers and ceca were analyzed for presence of L reuteri and Pseudomonas spp.
Defection of L Reuteri and Pseudomonas spp.
Mice were sacrificed by CO2 asphyxiation. Ceca and livers were removed aseptically and
homogenized in 5 ml PBS buffer. Homogenate dilutions were made in buffer, and duplicate
platings were carried out on Lactobacillus selection (LBS) agar and Pseudomonas isolation agar
(PIA). The LBS medium was used to enumerate the total gut and liver population of lactobacilli.
The subpopulation of L reuteri colonies on appropriately diluted plates is identified based on
the ability of L reuteri colonies to convert glycerol to reuterin under anaerobic conditions. The
PIA plates were used for Pseudomonas spp. detection in livers and ceca.
Results and Discussion
Animals that were treated with P. aeruginosa strains BC16 and AC869 and L reuferi were
cleared of the infectious agent in 7 days. Of animals that were not treated with L reuteri, 55
percent and 33 percent remained infected at that time with P. aeruginosa strains BC16 and
AC869, respectively. When the mice were given 10' cells of P. aeruginosa AC869 by Day 7,
83 percent remained infected compared with a 50-percent infection rate in the L reuteri treated
group. Animals treated with P. aeruginosa PAO (109 cells per mouse) in the absence of L reuteri
were 75-percent infected by Day 7; those treated with L reuferi were only 50-percent infected.
Some indigenous lactobacilli have been shown to inhibit colonization of pathogenic bacteria,
particularly in the small intestine, by means of what has been termed colonization resistance (CR)
or competitive exclusion (CE) (16). Neither the mechanism(s) underlying this phenomenon nor
the protective effect of L reuteri on the Pseudomonas infections described in this report is fully
understood. Our research has indicated, however, that 1) L reuteri prophylaxis is beneficial to
the host animal's health and 2) this treatment could have applications concerning the protection
of animals against Pseudomonas spp. Preliminary studies (17) indicate that L reuteri's efficacy
in this regard could be based on its ability to stimulate a protective immune response to P.
aeruginosa infections.
265
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References
1. Kandler, O., and N. Weiss 1986. Regular gram-positive nonsporing rods. In: Sneath,
P.H.A., M.E. Sharps, and J.G. Holt, eds. Sergey's manual of systematic bacteriology,
Vol.2, pp. 1,208-1,234.
2. Casas, I.A., F.W. Edens, WJ. Dobrogosz, and C.R. Parkhurst. 1993. Performance of
GAIAfeed and GAIAspray: A Lacfobaa'Mus reuferi based probiotic for poultry. In:
Jensen, J.F., M.H. Hinton, and R.W.A.W. Mulder, eds. Prevention and control of
potentially pathogenic microorganisms in poultry and poultry meat products.
Proceedings 12, FLAIR No. 6. Probiotics and Pathogenicrty, DLO Centre for Poultry
Research and Informational Services. The Netherlands: Beekbergen. pp. 63-71.
3. Axelsson, L.T., T.C. Chung, S.E. Lindgren, and W.J. Dobrogosz. 1989. Production of
a broad spectrum antimicrobial substance by Lactobacillus reuferi. Microbial Ecol.
Health Dis. 2:131-136.
4. Chung, T.C., L.T. Axelsson, S.E. Lindgren, and W.J. Dobrogosz. 1989. In vitro studies
on reuterin synthesis by Lactobac/Wt/s reuferi. Microbial Ecol. Health Dis. 2:137-144.
5. Talarico, T.L., I.A. Casas, T.C. Chung, and W.J. Dobrogosz. 1989. Production and
isolation of reuterin: A growth inhibitor produced by Lactobacillus reuferi. Antimicrob.
Agents Chemother. 32:1,854-1,858.
6. Talarico, T.L., and WJ. Dobrogosz. 1989. Chemical characterization of an
antimicrobial substance produced by Lactobacillus reuferi. Antimicrobial. Agents
Chemother. 33:674-679.
7. Franklin, C.A. 1988. Modern biotechnology: A review of current regulatory status and
identification of research and regulatory needs. Toxicol. Ind. Health 4:91-105.
8. Rissler, J.F. 1984. Research needs for biotic environmental effect of genetically
engineered microorganisms. Recomb. DNA Tech. Bull. 7:20-30.
9. Guiot, E.F.L., J.W.M. van der Meer, and R. van Furth. 1981. Selective antimicrobial
modulation of human microbial flora: Infection prevention in patients with decreased
host defense mechanisms by selective elimination of potentially pathogenic bacteria. J.
Infec. Dis. 143:644-654.
10. Schimpff, S.C. 1980. Infection prevention during profound granulocytopenia: New
approaches to alimentary canal microbial suppression. Ann. Intern. Med. 93:358-361.
11. Bartlett, J.G. 1979. Antibiotic-associated pseudomembranous colitis. Rev. Infect. Dis.
1:530-538.
12. George, S.E., M.J. Kohan, D.B. Walsh, and L.D. Claxton. 1989. Acute colonization of
polychlorinated biphenyl-degrading pseudomonads in the mouse intestinal tract:
Comparison of single and multiple exposures. Environ. Toxicol. Chem. 8:123-131.
266
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13. George, S.E., MJ. Kohan, D.B. Walsh, A.G. Stead, and LD. Claxton. 1989.
Polychlorinated biphenyl-degrading pseudomonads: Survival in mouse intestines and
competition with normal flora. J. Toxicol. Environ. Health. 26:19-37.
14. George, S.E., MJ. Kohan, D.J. Whitehouse, J.P. Creason, and LD. Claxton. 1990.
Influence of antibiotics on intestinal tract survival and translocation of environmental
Pseudomonas species. Appl. Environ. Microbiol. In press.
15. George, S.E., D.B. Walsh, A.G. Stead, and LD. Claxton. 1989. Effect of
ampicillin-induced alterations in murine intestinal microbiota on the survival and
competition of environmentally released pseudomonads. Fund. Appl. Toxicol.
13:670-680.
16. Fuller, R. ed. 1992. Probiotics: The scientific basis. NY: Chapman and Hall.
1 7. Dobrogosz, W.J., HJ. Dunham, F.W. Edens, and I.A. Casas. 1992. Lactobaci/lus reuferi
immunomodulation of stressor-associated diseases in newly hatched avion species.
International Symposium on Intestinal Microecology, Helsinki, Finland, August 28-29.
267
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Bioavailobility Factors Affeding the Aerobic Biddegradation of Hydrophobic
Chemicals
Pamela J. Morris
Soil and Water Science Department, University of Florida, U.S. Environmental Protection
Agency, Environmental Research Laboratory, Gulf Breeze, FL
Suresh C. Rao
Soil and Water Science Department, University of Florida, Gainesville, FL
Semen Akkerman
Center for Environmental Diagnostics and Bioremediation, University of West Florida, U.S.
Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
Michael E. Shelton
Department of Biochemistry, University of Minnesota, U.S. Environmental Protection Agency,
Environmental Research Laboratory, Gulf Breeze, FL
Peter J. Chapman and P.M. Pritchard
U.S. Environmental Protection Agency, Environmental Research Laboratory, Gulf Breeze, FL
We are currently studying interactions between complex waste mixtures and microorganisms that
are capable of transforming organic components of these mixtures. Our goal is to integrate
methodologies used to study the abiotic behavior of hydrophobic organics in soil with the
biological degradation of the organics. Sorption of hydrophobic compounds, such as
polychlorinated biphenyls (PCBs), to soil represents a potential barrier to their degradation and
detoxification in the environment, and influences the relative accessibility of these compounds
to a number of physical, chemical, and biological processes. We find the concept of
bioavailability a unique opportunity to couple interesting basic research to applied
bioremediation problems. Our long-term objectives include 1) the study of the desorption of
PCBs from historically contaminated soils and sediments; 2) the determination of the influence
of co-contaminants, cosolvents, and surfactants on PCB desorption enhancement; and 3) the
coupling of PCB desorption and biodegradation kinetics. The soil that we are studying is from
a former racing drag strip in Glen Falls, New York, contaminated with Aroclor 1242. Previous
studies have shown that approximately half of the PCBs present in the soil are unavailable for
aerobic biodegradation. This surface soil, classified as a sand (95-percent sand, 4.2-percent
silt, and 0.8-percent clay), contains 1.9 percent organic carbon and 1.43 percent oil and
grease. Mineralogical analyses show that the soil minerals consist of 40 percent quartz, 45
percent chlorite, and 15 percent Ca-albite (all low internal surface-area minerals). Heavy metal
analysis suggests that only lead levels are somewhat high, averaging 190 ppm. Specific
surface-area analysis indicates a low value of 0.1444 m2/g. The total pore volume is 0.0016
cm3/g, and the average pore diameter is 443.78 A. We are also characterizing the following
drag strip soil fractions individually: medium sand (2.00 mm to 0.425 mm), fine sand (0.425
mm to 0.08 mm), and silt/clay (<0.08 mm). Studies on the biodegradation of PCBs found in
each of the three fractions suggest that biodegradation of PCBs from the silt/clay fraction is less
than biodegradation from the fine and medium sand fractions. Since the silt/clay fraction
represents the major reservoir for organic carbon, oil and grease, heavy metals, and PCBs due
268 1994 Symposium on Bioremediation of Hazardous Wastes
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to its high surface area, the release of PCBs from this fraction may be essential to enhancing
PCB biodegradation. The biodegradation of the PCBs found in this fraction is currently the focus
of our studies. We are using the traditional batch method to examine congener-specific
desorption from the drag strip soil and the three fractions. In addition, we will compare the
miscible displacement technique with results from batch studies. The miscible displacement
technique uses preparative high-performance liquid chromatography (HPLQ glass columns
packed with drag strip soil and high-precision HPLC pumps to provide a steady flow rate.
Column effluent fractions are collected after passage through a flow-through
variable-wavelength UV detector. Both the batch method and miscible displacement technique
allow us to examine the influence of cosolvents and surfactants (biological and synthetic) on PCB
desorption and mobility. Enhanced desorption and mobility may contribute to increased
availability to biodegradation processes. In addition, we are examining the biodegradation of
the oil and grease in the drag strip soil. Analysis of the oil and grease by column
chromatography shows the distribution of organics to be 81.9 percent hydrocarbons, 16.9
percent polars, and 1.2 percent asphaltenes. This oil is very weathered and contains few readily
biodegradable components. We are in the process of enriching for microorganisms capable
of transforming this oil matrix and will test whether biodegradation of the oil results in enhanced
availability and biodegradation of the PCBs present.
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Use of Sulfur Oxidizing Bacteria To Remove Nitrate From Ground Water
Michael S. Davidson, Thomas Cormack, Harry Ridgway, and Grisel Rodriguez
Biotechnology Research Department, Orange County Water District, Fountain Valley, CA
The chemoautotrophic bacterium Thiobadllus denitrificans is capable of effective removal of
nitrate from ground water under anoxic conditions. This microorganism is capable of deriving
metabolic energy from oxidation of inorganic sulfur compounds including elemental sulfur,
hydrogen sulfide, thiosulfate, metabisulfite, tetrathionate, and sulfite. All carbon required for
biosynthesis is derived from carbon dioxide, carbonate, and bicarbonate. The primary products
of autotrophic denrtrification are nitrogen gas, sulfate, water, and biomass. The potential
advantages of using elemental sulfur (in powdered, flaked, or prilled form) are as follows: 1) low
cost and wide availability of energy source, 2) low toxicity compared with other energy sources
(i.e., methanol or ethanol), 3) ease and safety of storage, 4} potential for development of water
treatment reactors capable of operating for long periods (months) at a time with little or no
maintenance or operator attention, and 5) potential for use in situ to remediate nitrate-
contaminated aquifers.
A column reactor (3.6 m long x 0.051 m ID) has been operated continuously for over 1 year
outdoors. The reactor was initially filled to a depth of 1.83 m with sulfur granules graded
-16/+30 Mesh (U.S. Standard Sieve). Well-water nitrate content could be consistently reduced
to less than 0.3 ppm from an influent level of 55 ppm with a reactor feed rate of 0.35 L/min.
Increasing flow to 0.45 L/min resulted in an effluent containing nitrate concentrations ranging
from less than 0.3 ppm to 5 ppm. Maintenance of constant bed volume for a given flow rate
required periodic replenishment of the bed with fresh sulfur granules. As denitrification
proceeds, the granules decrease in mass (i.e., are consumed) to the point that their mass is
insufficient to remain within the reactor. A novel fluidized bed reactor system has been designed
that will permit essentially complete utilization of the smaller particles.
A variety of heterotrophic (organotrophic) bacteria were found to become established in reactors
fed only inorganic energy sources (elemental sulfur or sodium thiosulfate). The first survey
involved 15 bacterial isolates recovered from a chemostat reactor operated with precipitated
sulfur slurry as the energy source and nitrate as the terminal electron acceptor. The isolates
were recovered by plating dilutions of water samples on R2A (an organic-based medium) under
aerobic conditions. Isolates were purified by restreaking on R2A and were subjected to a
proprietary identification system, API-NFT, designed to identify nonfermentative bacteria. Of 15
isolates, one isolate each was identified as Achromobacter sp., Pseudomonas stutzeri,
Flavobacterium sp., and Pseudomonas pufrefaci'ens. Seven of the isolates were Gram-negative
"nonidentifiable." The remaining four isolates were Gram-positive "nonidentifiable." The second
survey involved 19 isolates recovered from a chemostat reactor operated with sodium thiosulfate
as the energy source and nitrate as the terminal electron acceptor. Of these, one isolate each
was identified as Achromobacter sp., Pseudomonas pseudoalcaligenes, and Pseudomonas
paucimobilis. Twelve isolates were identified as Pseudomonas aeruginosa. Four isolates were
Gram-negative "nonidentifiable." The "nonidentifiable" designation refers to isolates that gave
biochemical reactions profiles uncharacteristic of the API-NFT database collection. Work in
progress should result in identification to the genus level.
270 1994 Symposium on Koremediation of Hazardous Wastes
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Sodium thiosulfate was tested as an energy source in a small, prototype fluidized bed reactor.
The pyrex column (40 cm long x 2.54 cm ID) contained a 16-cm deep bed of 0.10-mm
diameter silica spheres (settled bed depth under zero flow conditions). In this reactor
configuration, the silica spheres serve only as an inert support matrix. Sodium thiosulfate is
highly soluble in water and can be supplied in correct proportion with the aid of a metering
pump. The degree of bed expansion was easily controllable between 0 percent and 100
percent. The reactor demonstration involved recirculation of 14 liters of a defined mineral salts
solution containing 1,227 ppm nitrate and 2,252 ppm thiosulfate through the column.
Following inoculation, flow was set at 30 ml/min (equal to 25 percent bed expansion).
Approximately 7 percent of the nitrate was removed by Day 7. Nitrate removal had increased
to nearly,35 percent by Day 11. Runs conducted with varying concentrations of nitrate relative
to thiosulfate revealed that acceptable denitrification efficiency required careful control of the
relative proportions of the two reactants. While technically feasible, the level of control required
to reliably produce denitrified water on a practical scale might prove difficult. Thiosulfate also
suffers from the disadvantage of higher cost per unit of nitrate removed in comparison to
elemental sulfur.
Respirometric experiments were conducted using pure cultures of Th/obaa'/lus denifrificans.
Washed cells obtained from aerobic cultures with either thiosulfate ortetrathionate as the energy
source were unable to denitrify in short-term experiments. This demonstrates that, as is the case
with heterotrophic bacteria, denitrification is an inducible rather than a constitutive metabolic
capability. However, anoxically grown cells could tolerate exposure to oxygen without
immediate deterioration or loss of denitrification activity. On a practical level, this suggests that
a biological denitrification reactor could readily withstand periodic ingress of oxygen resulting
from periodic air-scour or high flow backwash procedures, as might be required to control
formation of excess biomass deposits. Rapid recovery of denitrification activity following such
treatments would be a decided advantage.
In conclusion, sulfur-mediated biological denitrification of ground water appears to be
technically feasible. A fluidized bed reactor containing granular sulfur has been operated for
over 1 year. Autotrophic sulfur bacteria and nonautotrophic (organotrophic) bacteria appear
to coexist stably. The nature of their relationship (possibly syntrophic or mutualistic) is under
further study. The use of readily soluble sulfosalts as thiosulfate or tetrathionate in reactors
containing an inert support material is less certain. This approach will require additional basic
research to determine the relationship between nitrate concentration and energy-yielding
substrate and their overall effect on denitrification rates and efficiency.
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Engineering Evaluation and Optimization of Biopiles for Treatment of Soils
Contaminated With Hazardous Waste
Carl L. Potter and John A. Glaser
U.S. Environmental Protection Agency, Andrew W. Breidenbach Environmental Research
Center, Cincinnati, OH
Biopile systems offer the potential for low-cost treatment of hazardous waste in soil. Biopiles
provide favorable environments for naturally occurring microorganisms to degrade soil
contaminants. The microbial environment can be manipulated to promote aerobic or anaerobic
metabolism. Air is supplied to the system by a plumbing network that forces air through the pile
by applying either pressure or vacuum.
Biopiles differ from compost piles in that bulking agents necessary for composting are not added
to biopiles. Some nutrients and exogenous microorganisms, however, may be added to a biopile
in the form of manure or other nutrient-rich material. Biopiles will normally produce less heat
than compost piles because less organic substrate is added, although significant aerobic
microbial activity will produce some heat. While heat production is often desired in compost
piles, we may wish to limit heat production in biopiles to avoid killoff of mesophilic organisms
involved in biodegradation of soil contaminants.
The goal of this project is to evaluate the potential of biopile systems to remediate soils
contaminated with hazardous chemicals. Pilot-scale reactors with a volume of 2 yd3 to 3 yd3
each are being constructed at EPA's Test and Evaluation (T&E) Facility in Cincinnati.
Contaminated field soil from selected sites will be brought to the T&E Facility for this research.
Depending on availability of soil, contaminants may include any or all of the following:
pentachlorophenol, creosote, munitions, and petroleum hydrocarbons.
Short-term work will focus on designing and constructing pilot-scale biopile reactors and defining
suitable operating conditions. Pilot-scale operations may permit collection of reliable data to
develop effective aeration strategies, document degradation rates and metabolic products of
hazardous chemicals, and identify metabolically active microbial species. Physical and chemical
data to be collected include heat production; density (g/cm3); fractions of solids, moisture, and
organics; pressure drop across sections of aerated biopiles; and pH changes in various reactor
locations. Subsequent studies will emphasize treatability of contaminated soils.
Future investigations will focus on the potential to enhance biodegradation by manipulation of
physical and biological parameters. For example, anaerobic treatment may be necessary to
initiate degradation of recalcitrant compounds via reductive metabolism. Following reductive
metabolism, toxicants may be amenable to aerobic biodegradation. Research may identify the
most effective combination of anaerobic/aerobic conditions for biodegradation of recalcitrant
substrates in biopile systems.
272 1994 Symposium on Koramediation of Hazardous Wastes
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Factors Affecting Delivery of Nutrients and Moisture for Enhanced In Situ
Bioremediation in the Unsaturated Zone
James G. Uber and Ronghui Liang
Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, OH
Paul T. McCauley
U.S. Environmental Protection Agency, Risk Reduction Engineering Laboratory, Water and
Hazardous Waste Treatment Research Division, Cincinnati, OH
Introduction
Successful in situ bioremediation in the unsaturated zone requires that water, oxygen, and trace
nutrients be available in appropriate amounts and correct locations. To enhance degradation
rates, some applications may require delivery of moisture, oxygen, or trace nutrients via
subsurface or surface application of fluids. Since the exact locations and geometry of
contaminated regions are unknown, a practical engineering approach is to design fluid delivery
systems to uniformly distribute the fluids to a subsurface region.
This project investigates limitations of engineered systems for delivery of nutrients, either liquid
or gas, to contaminated soils in the unsaturated zone. These limitations are derived from two
sources: 1) the basic design of fluid delivery systems (e.g., inherent limitations in using vertical
wells or surface irrigation systems to uniformly distribute and collect a fluid in an unsaturated
subsurface region) and 2) heterogeneity in porous media properties that affect fluid flow in the
unsaturated zone (e.g., spatial variability of saturated hydraulic conductivity).
Unfortunately, the design of common fluid delivery systems and the heterogeneity of hydraulic
soil properties work against achieving the goal of uniform fluid distribution. Vertical wells and
soaker hoses are two means of fluid delivery, but these are essentially point or line sources.
Thus, there exist important unanswered questions about the proper spacing of these devices to
achieve a uniform application rate. A potentially more difficult issue is the significant spatial
heterogeneity in the hydraulic properties of natural soils. This heterogeneity creates paths of
preferred flow on a variety of spatial scales; only a fraction of the porous media may contribute
to fluid flow, and thus an engineered system designed to deliver moisture, oxygen, or nutrients
could fail to achieve a uniform distribution. Thus, the conventional notion of, for example, a
well's "region of influence" is less clear and will be critically reexamined through experimental
and theoretical approaches.
Work in Progress
This poster presents findings from a review of soil science and in s/fu bioremediation literature,
focusing on the potential effects of preferential flow on in s/fu bioremediation effectiveness. This
review was initiated at the start of the project in January 1994 and is being used to guide the
design of experiments scheduled to begin later this year. Future plans regarding the experimental
investigations also will be presented.
1994 Symposium on Bioremediation of Hazardous Wastes 273
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The Bioremediation in the Field Search System (BFSS)
Fran V. Kremer
U.S. Environmental Protection Agency, Office of Research and Development, Cincinnati, OH
Linda B. Diamond, Susan P.E. Richmond, Jeff B. Box, and Ivan B. Rudnicki
Eastern Research Group, Inc., Lexington, MA
The Bioremediation in the Field Search System (BFSS) is a PC-based software application
developed by EPA's Bioremediation Field Initiative. BFSS provides access to a database of
information compiled by the Initiative on hazardous waste sites where bioremediation is being
tested or implemented, or has been completed. Sites include Comprehensive Environmental
Response, Compensation, and Liability Act (CERCLA) sites, Resource Conservation and Recovery
Act (RCRA) sites, Toxic Substances Control Act (TSCA) sites, and Underground Storage Tank
(UST) sites. The database currently contains information on approximately 160 sites, primarily
those under federal authority. This summer the Initiative plans to expand the database by
soliciting information from industry, contractors, and vendors—an effort that is expected to
double or triple the number of sites in the database.
BFSS contains both general site information and data on the operation of specific biological
technologies. General site information includes the location of the site, site contacts, the
predominant site contaminants, and the legislative authority under which the site is being
remediated. Technology-specific information includes the stage of operation, the type of
treatment being used, the wastes and media being treated, the cleanup level goals, and the
performance and cost of the treatment. Both ex situ and in situ technologies are represented,
including activated sludge, extended aeration, contact stabilization, fixed-film, fluidized bed,
sequencing batch, and slurry reactor treatments; aerated lagoon, pile, and land treatments; and
bioventing, air sparging, in situ ground-water treatment, and confined treatment facilities.
BFSS allows the user to search the system based on location, regulatory authority for cleanup,
media, contaminants, status of the project, and treatment utilized. Based on the search criteria
specified by the user, BFSS generates a list of qualifying sites. BFSS allows the user to view
on-line information about these sites and to print site reports based on information contained
in the database.
The Initiative established the BFSS database to provide federal and state project managers,
consulting engineers, industry personnel, and researchers with timely information regarding new
developments in field applications of bioremediation. BFSS data and the operation of the search
system have been reviewed by representatives of the target user community, including personnel
from EPA regional offices and other professionals in the field of bioremediation. Information
in the database is updated semiannually and is reported in EPA's quarterly Bioremed/at/on In
the Field bulletin, which is published by the Office of Research and Development (ORD) and the
Office of Solid Waste and Emergency Response (OSWER). The bulletin provides a valuable
information-sharing resource for site managers using or considering the use of bioremediation.
274 1994 Symposium on Roremediotion of Hazardous Wastes
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Version 1.0 of BFSS will be available by June 1994 on several EPA electronic bulletin
boards—Cleanup Information (CLU-IN), Alternative Treatment Technology Information Center
(ATTIC), and ORD bulletin board systems—and on diskette from the EPA Center for
Environmental Research Information.
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Poster Presentations Supported by EPA's
Hazardous Substance Research Center Program
The following research is being carried out under the auspices of EPA's Hazardous Substance
Research Center (HSRC) program. EPA established this program in response to provisions in the
1986 amendments to the Comprehensive Environmental Response, Compensation, and Liability Act
(CERCLA). These provisions authorized EPA to establish HSRCs with a mission to study all aspects of
the "manufacture, use, transportation, disposal, and management of hazardous substances" and
made the Agency responsible for the "publication and dissemination of the results of such research."
The program is managed by the director of EPA's Office of Exploratory Research (OER) in the Office
of Research and Development (ORD).
EPA has established five research consortia, with each serving two adjacent federal regions. These
include:
• Northeast Hazardous Substance Research Center: Region-Pair 1 and 2, which includes the
New England states, New York, New Jersey, and the territories of Puerto Rico and the U.S.
Virgin Islands. The lead institution is the New Jersey Institute of Technology, and the center's
director is Dr. Richard Magee. Other consortium partners include the Massachusetts Institute
of Technology, Tufts University, Rutgers University, Stevens Institute of Technology, Princeton
University, and the University of Medicine and Dentistry of New Jersey.
• Great Lakes and Mid-Atlantic Hazardous Substance Research Center: Region-Pair 3 and 5,
which comprises the Great Lakes states and the mid-Atlantic states of Virginia, West Virginia,
Maryland, Pennsylvania, and Delaware. This three-university consortium is headed by Dr.
Walter Weber of the University of Michigan; Michigan State University and Harvard University
are partner institutions.
• South/Southwest Hazardous Substance Research Center: Region-Pair 4 and 6, which is
made up of Gulf Coast and southern states. Louisiana State University heads this center, in
partnership with Georgia Institute of Technology and Rice University. The center's director is
Dr. Louis Thibodeaux of Louisiana State University.
• Great Plains and Rocky Mountain Hazardous Substance Research Center: Region-Pair 7
and 8, which includes the states on the eastern side of the Great Basin along with the Great
Plains states. This large consortium is run by Dr. Larry Erickson of Kansas State University.
The other six participating institutions are Montana State University and the Universities of
Iowa, Missouri, Montana, Nebraska, and Utah.
• Western Region Hazardous Substance Research Center: Region-Pair 9 and 10, which
includes the West Coast states along with Alaska, Arizona, Hawaii, and Idaho. Stanford
University and Oregon State University make up this consortium. Dr. Perry McCarty of
Stanford University is the center's director.
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In Situ Attenuation of Chlorinated Aliphatic in Glaciol Alluvial Deposits
Michael J. Barcelona, Mark A. Henry, and Walter J. Weber, Jr.
University of Michigan, Ann Arbor, Ml
The National Center for Integrated Bioremediation Research and Development (NCIBRD) has
located operations atthe recently decommissioned Wurtsmith Air Force Base (WAFB) in Oscoda,
Michigan. NCIBRD is dedicated to the evaluation of decontamination technologies for
hazardous wastes and remediation of spill and disposal sites. These activities are administered
by the University of Michigan and oversight is provided by a science advisory board comprised
of the directors of the Hazardous Substance Resource Centers, representatives of the EPA
Biosystems Group, and nationally recognized engineers and scientists from government and
private sectors.
WAFB is ideally suited for in situ bioremediation research activities. The 7-square-mile base is
bordered by the Au Sable River to the south and west, and by Van Etten Lake to the east. The
property sits on a 20-m bed of highly transmissive glacial sand underlain by a thick silty-clay
aquitard. The ground water is found at about 6 m throughout the study area. The U.S. Air
Force has been working with the U.S. Geological Survey (USGS) to characterize the extent of
contamination at WAFB for the past 12 years, resulting in a large database and an array of
approximately 600 permanent monitoring wells. An excess of 70 sites are tainted by a variety
of sorbed, dissolved, and nonaqueous-phase petroleum hydrocarbon mixtures, chlorinated
solvents, and heavy metals. Air Force remediation activities have been limited to the installation
of three conventional air strippers for the containment of the largest plumes. These systems will
provide the capture zone needed for the eventual controlled release of tracer chemicals,
allowing an in-depth field study of the fate and transport of contaminants.
The USGS database provided information indicating that natural bioattenuation of aromatic and
chlorinated aliphatic compounds was occurring at WAFB. A sampling program is currently
being implemented to study the process at two of these sites: FT-02 (a heavily used fire training
area) and OT-16 (a former jet engine test cell).
Fire training was conducted at FT-02 from 1952 to 1993. Typically, 8,000 L of jet fuel (and
some incidental chlorinated solvents) was pumped over a simulated aircraft structure, ignited,
and extinguished. Unfortunately, unburned fuel and solvents infiltrated into the aquifer. The
USGS and Air Force installed 49 monitoring wells in 1 7 clusters to track the movement of the
plume originating from this site. Preliminary well monitoring and solid borings have shown
evidence of a large plume, with total volatile organic compounds exceeding 1,000 mg/L, that
is undergoing natural biotransformation. Concentrations of these compounds in the aquifer
solids reflect co-metabolic transformations; in other words, upgradient vadose zone levels of
trichloroethylene (5 mg/kg), BTEX (600 mg/kg), and dissolved oxygen decrease and
concentrations of cis-1,2-dichloroethylene increase to 5 mg/kg downgradient from thesite. This
site is located approximately 300 m from OT-16 and is hydraulically connected; plumes from
these sites are believed to merge downgradient.
The jet engine test cell was used for a variety of test activities. Cleanup of this structure typically
involved washing solvents off the floor into an oil-water separator, which eventually failed,
1994 Symposium on Boraimdiation of Hazardous Wastes 279
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allowing the solvents to enter the aquifer. The plume contains high concentrations of BTEX (4
mg/L) and moderate amounts of chlorinated solvents (70 mg/L). The Air Force installed 19
wells downgradient of this site, but little sampling has been done. NCIBRD has just begun site
characterization efforts at this site.
Future work at these two sites will supplement existing physical-chemical information with
location and geophysical surveys, meteorological monitoring, additional borings and monitoring
well emplacements, soil gas surveys, permanently installed water level recorders, grain-size and
hydraulic conductivity determinations, as well as chemical property measurements (e.g.,
mineralogy, carbonate, organic carbon, metal and metal oxide content, cation-exchange
capacity, etc.). In addition, routine well sampling will document not only contaminant
concentrations but also changes in metabolic levels in the aquifer. This effect will support
experimental applications of in situ remediation technologies to be conducted by consulting,
private industry, and academic professionals.
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In Situ Bioremediotion of Chlorinated Solvent Ground-Water Contamination:
Scaling up From a Field Experiment to a Full-Scale Demonstration
Perry L. McCarty, Gary D. Hopkins, and Mark N. Goliz
Western Region Hazardous Substance Research Center, Stanford University, Stanford, CA
Studies conducted at an experimental field site at Moffett Naval Air Station have demonstrated
that trichloroethylene (TCE) can be effectively biodegraded co-metabolically through the
introduction into the subsurface of a primary substrate (such as phenol or toluene) and oxygen
to support the growth and energy requirements of a native population of microorganisms.
Additional preliminary experimental work at Moffett Field has now been conducted in
preparation for a full-scale demonstration.
A full-scale demonstration at a real hazardous waste site is likely to encounter a plume with
multiple contaminants. It was therefore desirable to determine how other contaminants which
could potentially be present might affect the rate and extent of TCE degradation. In particular,
previous laboratory studies at Stanford University have indicated that the degradation products
of 1,1 -DCE are toxic to methane-oxidizing bacteria. Follow-on field work conducted at Moffett
Field demonstrated that the presence of 1,1 -DCE inhibited TCE degradation by phenol-oxidizing
microorganisms. Thus, 1,1-DCE should not be present at the site selected for a full-scale
demonstration of this technology.
An effective method to provide the indigenous microorganisms with sufficient oxygen to oxidize
the primary substrate is needed for the field demonstration. In past studies at Moffett Field,
molecular oxygen has been used as an oxygen source. Molecular oxygen, however, is difficult
to transfer to solution. Hydrogen peroxide is an alternative oxygen source that has been used
in bioremediation of petroleum hydrocarbons and is much easier to apply to the subsurface than
molecular oxygen. Preliminary work at Moffett Field showed that hydrogen peroxide worked as
effectively as molecular oxygen in degrading TCE.
Another question that needs to be answered prior to full-scale implementation of this technology
is how best to mix a primary substrate, an oxygen source, and TCE and to deliver the mixture
to the microorganisms. At Moffett Field, mixing of these three components was accomplished
aboveground, with the mixture then introduced into the subsurface through an injection well.
In a full-scale demonstration, the TCE will, of course, already be in the ground water. A major
objective of this demonstration will be to investigate how a primary substrate and an oxygen
source can be efficiently mixed and transported to indigenous microorganisms, to promote co-
metabolic degradation of TCE. For the demonstration, a subsurface recirculation system similar
to that described by Herrling (1) and McCarty and Semprini (2) is expected to be used. The
remediation system will consist of a single well, screened at two depths. In operation, a
submersible pump installed between the two screens would draw TCE contaminated water into
the well at one screened interval. The primary substrate and oxygen will then be introduced into
the water through feed lines, and the water, which now contains TCE, primary substrate, and
oxygen, will be discharged into the aquifer from the second screened interval. In essence, an
in sifu treatment zone will be created in the aquifer around the discharge screen. Based on the
Moffett Field results, this treatment zone is expected to cover an area within approximately 1
day's ground-water travel distance out from the well.
1994 Symposium on Bioremediation of Hazardous Wastes 281
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Ultimately, these studies, in which the laboratory and the Moffett Field site are being used to
make predictions regarding processes and to help design systems at a Veal-world" site, will
hopefully help lead to a better understanding of how laboratory and field investigations can best
be scaled up to make better real-world predictions.
References
1. Herrling, B. 1991. Hydraulic circulation system for in situ bioreclamation and/or in situ
remediation of strippable contamination. In: Hinchee, R.E., and R.F. Olfenbuttel, eds.
Onsite bioreclamation. Boston, MA: Butterworth-Heinemann. pp. 173-175.
2. McCarty, P.L., and L. Semprini. 1993. Ground-water treatment for chlorinated
solvents. In: Norn's, R.D., et al., eds. Handbook of bioremediation. Boca Raton, PL:
Lewis Publishers, pp. 87-116.
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BioavailabilHy and Transformation of Highly Chlorinated Dibenzo-p-dioxins and
Dibenzofurans in Anaerobic Soils and Sediments
Peter Adriaens and Quingzhai Fu
Department of Civil and Environmental Engineering, University of Michigan, Ann Arbor, Ml
Polychlorinoted dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) are
introduced via several industrial and municipal channels into both aerobic and anaerobic
environmental compartments. Due to their high toxicity and uncertain genotoxic potential, their
determination and fate in environmental samples is of great interest. The fate of highly
chlorinated PCDD/PCDF congeners was studied in both high and low organic carbon anaerobic
microcosm incubations. The inocula were derived from historically contaminated anaerobic
environments such as polychlorinated biphenyl-contaminated sediments and creosote-
contaminated aquifer samples, and were amended with a mixture of aromatic and aliphatic
acids for methanogenic growth. The samples were analyzed and quantified using high
resolution gas chromatography coupled with an electron capture detector and a low resolution
mass selective detector operated in selected ion monitoring (SIM) mode ([M+], [M++2], and
[M++4] ions). Recovery efficiencies after soxhlet extraction and sample cleanup were 40
percent to 70 percent, based on 1,2,3,4-tetrachlorodibenzo-p-dioxin as an internal standard.
The long-term (> 2 years) removal patterns of sediment-sorbed PCDDs/PCDFs in both
sediments could be explained by labile and resistant PCDD/PCDF desorption components,
presumably due to intraparticle diffusion-controlled mass transfer limitations. Mass transfer
limitations were based on incubation time-dependent decreased extraction efficiencies of
PCDDs/PCDFs from inactive controls. The net first-order initial rate constants of disappearance
ranged from 0.30 to 0.75 (x 10'3) d'1 for aquifer sediments and from 0.46 to 1.87 (x 10'3) d'1
for high organic carbon Hudson River sediments. Moreover, the overall decrease in
PCDDs/PCDFs from the sediment particles in active microcosms sacrificed after 30 months was
as much as 20 percent greater compared with the autoclaved controls. Lesser chlorinated
congeners were found in all active microcosms analyzed. Isomer-specific analysis of the lesser
chlorinated congeners indicated that the 1,4,6,9-chlorines were preferentially removed, thus
enriching the medium in 2,3,7,8-substituted congeners and increasing the overall relative
toxicity. These observations contribute to our knowledge regarding the fate of PCDDs/PCDFs
in anaerobic soils and sediments, and indicate the importance of congener "fingerprinting"
during environmental source/fate analysis.
1994 Symposium on Koramediotion of Hazardous Wastes 283
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Localization of Tetrachloromethane Transformation Activity in
Shewonello Putrefoa'ens MR-1
Erik A. Petrovskis, Peter Adn'aens, and Timothy M. Vogel
University of Michigan, Department of Civil and Environmental Engineering, Ann Arbor, Ml
Investigations of pollutant transformation by pure cultures may enhance our understanding of
in situ natural attenuation processes in these environments. Shewanella putrefadens MR-1, an
Fe(lll)- and Mn(IV)-reducing facultative anaerobe, has been shown to dechlorinate
tetrachloromethane (CT) to chloroform (24 percent), after growth under nitrate- or Fe(lll)-
respiring conditions. Mass balance for carbon included 58-percent incorporation in biomass,
4.1-percent formation of nonvolatile products, and 5.5-percent mineralization. Product
distribution was independent of growth conditions. Amendment of MR-1 cell suspensions with
lactate, formate, or hydrogen increased CT transformation activity, while methanol did not. The
rate and extent of CT transformation increased for MR-1 cells grown with electron acceptors
having more positive half-reduction potentials (E°). Nitrate did not inhibit CT transformation.
In the presence of Fe(lll), reductive dechlorination was enhanced and resulted in the production
of dichloromethane (DCM), presumably by abiotic mechanisms involving Fe(ll).
In MR-1 cell extracts, NADH was the most effective electron donor for CT transformation.
Addition of FMN increased the activity 3- to 10-fold. Furthermore, CT transformation activity
has been localized primarily to membrane fractions (89 percent).
The effects of respiratory inhibitors on CT transformation activity have been examined.
Rotenone, an inhibitor of NADH dehydrogenase, reduced CT transformation activity in MR-1
whole-cell suspensions using lactate or NADH as an electron donor. Quinacrine, an inhibitor
of flavins, enhanced this activity. No significant effect was seen in the presence of pCMPS,
sodium azide, and sodium cyanide or in the presence of the cytochrome inhibitors HQNO and
Antimycin A. These results suggest that transformation of CT may be mediated by a nonheme
electron transfer agent.
Respiratory mutants of MR-1 have been screened for CT transformation activity. Rates of CT
transformation for MR-1 mutants in Fe(lll) reductase, Mn(IV) reductase, orfumarate reductase
were equivalent or greater than those for the MR-1 wild-type strain. MR-1 mutants that did not
synthesize menaquinones (MK) and so lost the ability to couple nitrate, Fe(lll), or fumarate
reduction for growth also lost 90 percent of CT transformation activity. When cell suspensions
of MK-deficient mutants were complemented with an MK precursor, CT transformation rates
returned to MR-1 wild-type levels. These results indicate that MK or another electron transfer
mediator reduced by MK but not a terminal reductase may be responsible for CT transformation
by MR-1.
284 1994 Symposium on Roremediation of Hazardous Wastes
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Formation and Transformation of Pesticide Degradation Products Under Various
Electron Acceptor Conditions
Paige J. Novak, Gene F. Parkin, and Craig L Just
University of Iowa, Iowa City, IA
Introduction
Pesticide contamination of ground-water supplies is a serious and growing problem in the United
States. More than 600 active chemicals exist that are used to protect crops from target pests
(1). Pesticides can remain in the environment for a long time, entering the air or ground-water
supply by partitioning to or diffusing through the soil column. Transformation of these chemicals
to one or more principal metabolites often occurs with unknown and unmonitored results. To
develop systems to destroy these contaminants and formulate intelligent policies to regulate or
restrict their use, an understanding of the reactions that these compounds undergo in the
environment is essential.
The herbicides alachlor and atrazine, the two most commonly used pesticides in the nation,
together account for 25 percent by weight of total pesticide use (2). These herbicides are also
the two most frequently detected pesticide contaminants in ground-water supplies in the Midwest
(2). Many xenobiotics can undergo mineralization to carbon dioxide and water by biological
means; alachlor and atrazine, however, undergo very little mineralization under typical
environmental conditions. Mineralization has been observed by only a few researchers,
generally at quantities of less than 5 percent of the initial herbicide concentration. As a single
exception, a recently completed study revealed that atrazine, when serving as the sole nitrogen
source for a microbial population, was mineralized at levels of greater than 80 percent of the
initial concentration, with a half-life of 0.5 to 2.0 days using a microbial consortium that had
undergone over 5 months of subculturing and enrichment in the laboratory (3). With little
natural mineralization occurring under typical environmental conditions, transformation
intermediates of alachlor and atrazine may be formed and may be accumulating in the soil and
ground water.
The specific objectives of this research project were to identify the transformation products of
alachlor and atrazine under four common electron acceptor conditions (aerobic, denitrifying,
sulfate-reducing, and methanogenic) and, to the extent possible, to determine kinetic coefficients
that describe the rate of formation and disappearance of these metabolites.
Experimental Design
Four9-L, fill-and-draw reactors were established to maintain specific environmental conditions.
Each reactor was fed a mineral nutrient solution typical of ground water under the redox
condition of interest. Temperature was maintained at 20°C in the dark to mimic environmental
conditions. Each of the reactors was fed acetate as the carbon and energy source, with some
of the batch denitrifying experiments carried out with citrate as an electron donor as well.
Alachlor and atrazine were fed at approximately 100/tg/L each, along with a phosphate buffer
1994 Symposium on Bioremediation of Hazardous Wastes 285
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to maintain a neutral pH. In addition, the specific electron acceptor for each system was added
in excess: O2 for the aerobic reactor, KNO3 for the denitrifying reactor, and MgSO4 • 7H2O
(at a high sulfate-to-organic ratio) for the sulfate-reducing reactor. The bacteria in each system
were acclimated to alachlor and atrazine prior to the start of the experiments.
Control experiments were set up to determine which physical and chemical means of alachlor
and atrazine transformation were important. The potential role of the phosphate buffer in
catalyzing chemical hydrolysis of alachlor and atrazine was studied with a phosphate control.
Reactions with resazurin, a color indicator of redox potential used in the denitrifying reactors,
were also studied using several control reactors with varying resazurin concentrations. A
mercuric-chloride-killed biological control was used to investigate sorption to biomass, and to
further assess the role of resazurin. Finally, a deionized water control was employed to identify
mixing problems, the significance of alachlor and atrazine sorption to the reactor itself, and
potential volatilization, chemical hydrolysis, or photolysis reactions.
All experiments were carried out in a batch format. An initial dose of alachlor and atrazine was
added to the reactor and allowed to mix for approximately 45 min, then samples for pesticide
analysis were taken from the reactor at various time intervals. The denitrifying and control
experiments were carried out in 2-L Pyrex bottles built like the larger 9-L reactors, so several
different conditions could be tested without affecting the stock enrichment culture. The
experiments involving the methanogenic and sulfate-reducing systems were carried out in the 9-L
reactors.
Results
Initially, alachlor and atrazine disappeared in batch reactors maintained under all terminal
electron acceptor conditions except aerobic conditions. Further experiments involving the
aerobic reactor were abandoned due to the absence of noticeable degradation of parent
compounds, Resazurin was added only to the denitrifying reactors to indicate whether the
proper conditions were maintained. This compound was found to be involved in the abiotic
transformation of alachlor and atrazine. Second-order degradation constants for alachlor and
atrazine transformation are given in Table 1; these constants are averaged values for four
experiments for each of the different terminal electron acceptor conditions. Each of these rate
constants has been corrected for the abiotic transformation of atrazine and alachlor in the
denitrifying reactors due to resazurin, and the abiotic transformation of alachlor in the
methanogenic and sulfate-reducing reactors due to the bisulfide ion. Therefore, the values given
in Table 1 represent only the biological transformation of alachlor and atrazine.
The standard deviation of these rate constants is relatively high, for two reasons. First, in the
denitrifying experiments duplicate reactors were used that contained different quantities of
biomass and most likely slightly varying microbial populations as well. A slight change in the
relative numbers of the different microorganisms present could result in the differences that were
observed in alachlor and atrazine transformation rates among the different reactors. For the
experiments involving the methanogenic and sulfate-reducing environments, one reactor was
used for the four experiments. Upon complete degradation of alachlor, 1 to 2 weeks were
allowed to pass with no pesticides added to the reactors while electron donor and acceptor
levels were maintained. At this point, alachlor was again dosed to the reactors, and the next
experiment was started. Over the course of the four experiments, the rate of alachlor
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transformation decreased considerably under both methanogenic and sulfate-reducing
conditions. At the end of the fourth experiment, no acetate utilization was observed in either
reactor, and no methane production occurred in the methanogenic reactor. At this point, 2 L
of fresh ground-water media was added to each of the reactors and the normal fill-and-draw
feeding was resumed, but no pesticides were added to either reactor. After 2 months, no
recovery of either population was observed. This effect on the microbes was thought to have
been a result of the buildup of nonmetabolizable and toxic alachlor or atrazine metabolites.
Several metabolites of alachlor were positively identified in these systems. Under denitrifying
conditions with resazurin and organisms present, aniline, m-xylene, acetyl alachlor, and diethyl
aniline were positively identified as products of alachlor degradation. Aniline, identified and
quantified bygaschromatography/mass spectrometry (GC/MS), appeared between Days 12 and
1 7 of the 45-day experiment and had degraded below detection limits by the last day. At the
maximum aniline concentration, 35 percent of the initial alachlor added had degraded to
aniline. Aniline formation and degradation constants are listed in Table 2; these rate constants
are based on the assumption that aniline is formed as a direct result of alachlor degradation
and thatbiomass remains constant throughout the experiment. Aniline formation was assumed
to have occurred to some maxima, at which point degradation began. Experiments are
presently underway to study the degradation of aniline in reactors fed only this compound. The
presence of aniline in ground water as a result of alachlor degradation is possible, but the high
rate of aniline removal by aerobic microorganisms makes the persistence of this substance for
a period of longer than a few days unlikely. However, under reducing conditions in an aquifer,
aniline may persist for a few weeks or conjugate to form compounds such as diphenylamine.
In the denitrifying reactors containing resazurin and acetate-utilizing organisms, m-xylene, a
suspected human carcinogen, appeared between Days 17 and 22 of the experiment and had
disappeared by Day 31. On Day 22, the highest m-xylene concentration was present in the
reactor sample and corresponded to approximately 9 percent of the initially fed alachlor.
m-Xylene was also detected in an abiotic reactor containing only resazurin, atrazine, and
alachlor under denitrifying conditions. On Day 45, the highest observed m-xylene concentration
was present in this reactor and accounted for 1 7 percent of the initial alachlor concentration.
Because this compound is also readily biodegradable, it is unlikely that m-xylene would persist
in ground water as a result of alachlor contamination and subsequent transformation. The role
of resazurin was not clearly defined. Biomass growth was observed in the reactor containing
only resazurin, alachlor, and atrazine, indicating that resazurin most likely served as an electron
donor for organism growth. Therefore, it is unclear whether resazurin itself or the organisms that
were capable of growth on only resazurin were responsible for the formation of m-xylene in this
reactor.
One of the denitrifying reactors contained only biomass; in this reactor, neither aniline nor
m-xylene was detected. Resazurin, or perhaps some compound that facilitates electron transfer,
such as vitamin B,2, may be required for at least one step in the degradation pathway that leads
to aniline and m-xylene production.
In the methanogenic and sulfate-reducing reactors, diethyl aniline and acetyl alachlor were
detected. Because these conditions are highly reducing, acetyl alachlor is an expected product
and is likely formed as a result of reductive dechlorination. Acetyl alachlor could not be
quantified because the sample received from Monsanto had evaporated to a residue. Diethyl
aniline is a product of further microbial attack of the ether and carbonyl groups of alachlor.
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At the highest observed concentration, diethyl aniline represented 9 percent and 20 percent of
the initial alachlor added to the system in the methane-genie and sulfate-reducing reactors,
respectively. Two unidentified metabolites, SMI and SM2, accumulated in both reactors,
perhaps causing the toxicity that eventually caused the organisms to stop their degradation of
acetate, alachlor, and atrazine.
Using the gas chromatograph with both an electron capture detector (GC/ECD) and a
nitrogen-phosphorous detector (GC/NPD), along with the GC/MS, many transformation
products were observed in all of the reactors yet could not be positively identified. By
preliminarily identifying these compounds using a spectra library from the National Bureau of
Standards on the GC/MS, an idea of the identity of some of these products was gained. Some
of the compounds were long, branched, saturated, and unsaturated hydrocarbon chains and
were probably caused by the breakdown and microbial metabolism of acetate and citrate.
Other compounds appeared to be caused by the conjugation or substitution of two or more
substances. Transformation products appeared to be formed by many different mechanisms,
such as dealkylation or reductive dechlorination, and had widely varying concentration profiles.
Compounds like acetyl alachlor in the denitrifying reactor appeared and disappeared in a few
days. Other compounds, such as diethyl aniline and the unknown metabolites SMI and SM2
detected in the methanogenic and sulfate-reducing reactors, were long-lived, persisting in the
reactor over a period of weeks.
No transformation products of atrazine were identified under any of the conditions investigated.
Since atrazine disappearance was measured in the denitrifying, methanogenic, and
sulfate-reducing systems, and complete mineralization to carbon dioxide and water was very
unlikely, metabolites should have been formed in these reactors. The C-18 solid-phase
extraction column used is reportedly not very effective at trapping polar substances. It is likely
that polar transformation products such as hydroxyatrazine were produced; the polar products
were probably lost during sample extraction because only those compounds that were
extractable by the use of the C-18 column were analyzed. Their loss is a possible explanation
for the lack of detected transformation products of atrazine. As new solid-phase extraction
columns are developed for effective extraction of pesticides and their polar metabolites, more
transformation products will be identified in these systems.
Summary and Conclusions
The speed and specific degradation steps followed in the transformation of alachlor and
atrazine, and the various degradation products that are formed as a result of this transformation,
are strong functions of environmental conditions, namely, the terminal electron acceptor
conditions present. In alachlor degradation, aniline and m-xylene were products detected only
in the denitrifying reactors. On the other hand, acetyl alachlor was identified under denitrifying,
methanogenic, and sulfate-reducing conditions. The product formation and transformation
patterns during alachlor degradation were very different in each of these systems. Analytical
limitations prevented the identification of likely polar products of atrazine degradation. Further
study is required to identify more of the metabolites that are formed and to try to formulate a
degradation pathway for alachlor and atrazine. The electron acceptors present, and
consequently the microbial population developed in these systems, affect the rate of herbicide
transformation, the pathway that this degradation takes, and the products that are formed that
may accumulate in the systems. The conditions under which herbicide degradation takes place
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also can result in the formation of compounds that are human health hazards and could be a
threat to ground-water supplies.
References
1. Somasundaram, L, J.R. Coats, K.D. Racke, and V.M. Shanbhag. 1991. Mobility of
pesticides and their hydrolysis metabolites insoil. Environ. Toxicol. Chem. 10:1 85-194.
2. Lynch, N.L 1990. Transformation of pesticides and halogenated hydrocarbons in the
subsurface environment. Ph.D. dissertation. University of Iowa, Department of Civil and
Environmental Engineering (May).
3. Mandelbaum, R.T., L.P. Wackett, and D.L. Allan. 1993. Mineralization of the s-triazine
ring of atrazine by stable bacterial mixed cultures. Appl. Environ. Microbiol.
59(6): 1,695-1,701.
4. Wilber, G.G. 1991. Kinetics of alachlor, atrazine, and chloroform transformation
under various electron acceptor conditions. Ph.D. dissertation. University of Iowa,
Department of Civil and Environmental Engineering (August).
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Table 1. Second-Order Degradation Constants forAlachlorand Atrazine Under Three Terminal
Electron Acceptor Conditions
Conditions
Denitrifying reactor
Methanogenic reactor
Sulfate-reducing reactor
Resazurin
Bisulfide ion (4)
Second-Order Degradation Constant
Alachlor
7.9xlO-5(±4.1 xlO'5)
L/mg VSS-day
2.9xlO'3(± 1.6 xlO'3)
L/mg VSS-day
1.5xlO'2(± 1.4 xlO'2)
L/mg VSS-day
5.0 xlO'2 (±5.4xlO'2)
L/mg res "day
1 .5 x 1 0-3 L/mg VSS'day
Atrazine
6.7xlO-5(±5.3xlO'5)
L/mg VSS-day
8.4 xlO'5 L/mg VSS-day
6.5 xlO"5 L/mg VSS-day
4.2xlO-2(±4.2xlO-2)
L/mg res-day
—
Table 2. Second-Order Formation and Degradation Constants for Aniline in the Reactor
Containing Both Resazurin and Denitrifying Organisms
Second-Order Formation Constant
8.4 xlO'5 L/mg VSS-day
Second-Order Degradation Constant
4.8 xlO'3 L/mg VSS-day
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Bioremediation of Aromatic Hydrocarbons at Seal Beach, California:
Laboratory and Field Investigations
Harold A. Ball, Gary D. Hopkins, Eva Orwin, and Martin Reinhard
Western Region Hazardous Substance Research Center, Stanford, CA
The objective of this study was to develop our understanding of processes that are important in
the anaerobic biodegradation of aromatic hydrocarbons in contaminated ground-water aquifers.
The focus of the investigation was a site at the Seal Beach Naval Weapons Station in Southern
California, where a significant gasoline spill resulted in contamination of the ground-water
aquifer. The project was divided into laboratory and field components, which were interrelated.
The goals of the laboratory experiments were to determine the capability of the aquifer microbial
community to transform aromatic hydrocarbon compounds under various anaerobic conditions
and to understand the effect of environmental factors on the transformation processes. Field
experiments were carried out on site at Seal Beach. The objectives of the field experiments were
to evaluate potential in situ application of anaerobic bioremediation processes and to attempt
to apply laboratory results to the field. The results from the field experiment will be used to
design a remediation proposal for the aquifer at the Seal Beach site.
Approach and Results
Laboratory Study
In a laboratory microcosm experiment, we evaluated several factors that were hypothesized to
influence field-scale bioremediation. Individual monoaromatic compounds (e.g., benzene,
toluene, ethylbenzene, and m-, p-, and o-xylene) were the primary substrates. To test the
influence of liquid-phase composition on the hydrocarbon degradation potential of Seal Beach
aquifer sediment, the sediment was placed in native ground water, native ground water with
nutrient amendments, and various other laboratory media formulations including denitrifying,
sulfate-reducing, and methanogenic media. In replicate bottles during the first 52 days of the
study, toluene and m+p-xylene (here, m-xylene and p-xylene were measured as a summed
parameter) were biotransformed in the unamended ground-water samples under presumed
sulfate-reducing conditions. Addition of nitrate to the ground water increased rates of toluene
biotransformation coupled to nitrate reduction, stimulated biotransformation of ethylbenzene,
and inhibited the complete loss of m+p-xylene that was observed when nitrate was not added
and sulfate-reducing conditions prevailed. Addition of the nutrients ammonia and phosphate
had no effect on either the rate of aromatics transformation or the distribution of aromatics
transformed. In contrast to nitrate-amended ground water, ethylbenzene was always transformed
first followed by toluene in the microcosms prepared with denitrifying media. In sulfate-reducing
media, lag times were increased, but toluene and m-xylene were ultimately transformed just as
in the microcosms with ground water alone. Although methane had been detected in the field,
there appeared to be no transformation activity in the methanogenic microcosms during the
period of the experiment.
1994 Symposium on KonmtdioHon of Hazardous Wastes 291
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Bioreactor Study
A pilot-scale facility consisting of 90-L reactors was constructed at the Seal Beach site. The
facility was designed for the operation of three anaerobic «n situ bioreactors. The reactors
consisted of aquifer sediment filled stainless steel cylindrical vessels with the capability to control
and monitor both hydrodynamic flow and supplements to the composition of the native ground-
water influent. Initial operation of the three anoxic/anaerobic reactors focused on evaluating
anaerobic bioremediation strategies foraromatic hydrocarbons underexisting (presumed sulfate-
reducing) and enhanced denitrifying conditions. Bioreactor results were consistent with the
laboratory microcosm experiments. Toluene and m+p-xylene were degraded in both the
unamended and nitrate-amended bioreactors. Degradation of ethylbenzene was stimulated by
nitrate addition. There was no evidence that benzene or o-xylene was transformed in either
reactor. The final percentage removal efficiency appeared to be higher in the unamended
bioreactor, where flow was slower.
Field Study
Field experiments have been conducted to assess anaerobic bioremediation of a test zone within
the contaminated aquifer at the Seal Beach site. A network of eight observation wells and one
extraction well was installed at the Seal Beach site. Hydrodynamic evaluation of the well field
indicated that two of the wells were satisfactory for further experimentation. Experiments have
been conducted using a slug test experiment design in which a single well was used for the
injection of the "slug" or test pulse and the same well was used to extract the test pulse. The
results of the experiments were inferred by differences measured in the samples collected during
extraction. Since the native ground water contained a variety of electron acceptors and the
water used for the injected pulses was water that had previously been extracted from the test
zone, the ground water was treated to control the concentration of all electron acceptors during
the injection of the test pulse. Before injection, the desired salts were added back to the
deoxygenated injection stream and the stream metered into the injection well. Sodium bromide
was added as a conservative tracer. Under this scenario, the different electron acceptors
investigated (e.g., nitrate and sulfate) could be added as desired. During initial tracer studies,
the injection water was organics free, and thus the source of the organics was desorption from
the in situ aquifer solids. In subsequent and ongoing bioremediation studies, benzene, toluene,
ethylbenzene, m-xylene, and o-xylene were added with the injection pulse at a concentration of
approximately 200 /tg/L each.
The initial bromide tracer data showed stable tracer concentrations and indicated no substantial
encroachment of native ground water detected in the first 0.4 pore volumes. There was a very
small hydraulic gradient at the site, hence recovery of the bromide mass from the test wells
ranged from 93 percent to 99 percent with the extraction of three pore volumes over a 103-day
period. During the tracer test, the equilibrium desorption concentrations for the aromatic
hydrocarbons when the electron acceptors nitrate and sulfate were absent from the ground water
were evaluated. Benzene, ethylbenzene, and o-xylene concentrations remained relatively stable
and thus appeared to be at an equilibrium. The toluene and m+p-xylene concentrations had
a downward trend relative to benzene once the native ground water encroached after
approximately 0.4 pore volumes, suggesting that the nitrate and sulfate concentrations available
in the native ground water supported some biological activity in the latter part of the experiment
for toluene and m+p-xylene removal.
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In a nitrate augmentation experiment, nitrate and dramatics were added to the injection pulse,
resulting in complete consumption of toluene and ethylbenzene followed by m-xylene within the
first 2 weeks. o-Xylene was slowly degraded, and its concentration approached zero by Day 60.
There was no apparent loss of benzene when compared with the inert tracer. The addition of
nitrate to the test region appeared to enhance the natural anaerobic denitrifying population.
This would confirm that there was already an active nitrate-reducing population in the aquifer
whose activity was enhanced by the addition of nitrate. With the exception of o-xylene
transformation, these results were comparable with those from the nitrate-amended microcosm
and bioreactor experiments, wherein toluene, ethylbenzene, and m-xylene were transformed
under denitrifying conditions.
During the tracer study, methane was detected in the test wells. With the encroachment of the
native ground water and associated increase in nitrate and sulfate concentrations, the methane
concentration decreased to values close to zero, suggesting that nitrate and sulfafe inhibit
methanogenesis at this site.
Additional experiments are underway to determine more precisely some of the kinetic constants
in the aquifer under denitrifying conditions and to evaluate rates and removal of aromatics
under sulfate-reducing and methanogenic conditions.
Acknowledgment
Funding for this study was provided by the EPA Office of Research and Development under
agreement R-815738-01 through the Western Region Hazardous Substance Research Center.
The content of this study does not necessarily represent the views of the agency. Additional
funding was obtained from the Chevron Research and Technology Company, Richmond,
California.
1994 Symposium on ffioreimdiation of Hazardous Wastes 293
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Pneumatic Fracturing To Enhance In Situ Bioremediation
John R. Schuring
Hazardous Substance Management Research Center, New Jersey Institute of Technology,
Newark, NJ
David S. Kosson and Shankar Venkatraman
Department of Chemical and Biochemical Engineering, Rutgers University, Piscataway, NJ
Thomas A. Boland
Hazardous Substance Management Research Center, New Jersey Institute of Technology,
Newark, NJ
In situ bioremediation is often limited by the transport rate of nutrients and electron acceptors
(e.g., oxygen, nitrate) to microorganisms, particularly in soil formations with moderate to low
permeability. An investigation is under way to integrate the process of pneumatic fracturing with
bioremediation to overcome these rate limitations. Pneumatic fracturing is an innovative
technology that utilizes high pressure air to create artificial fractures in contaminated geologic
formations, resulting in enhanced air flow and transport rates in the subsurface. The pneumatic
fracturing system can also be used to inject nutrients and other biological supplements directly
into the formation.
A project to investigate the coupling of these two technologies has been sponsored by EPA under
the Superfund Innovative Technology Evaluation (SITE) Emerging Technologies Program and is
scheduled for completion in the summer of 1994. Laboratory and field studies are being
carried out simultaneously to degrade BTX in gasoline. The laboratory studies are examining
the physical and biological processes at and near the fracture interfaces, including diffusion,
adsorption, and biodegradation. Both column and batch studies are being used to observe and
quantify the individual and combined effects of these processes. For the field portion of the
studies, a pilot demonstration is underway at an industrial site contaminated with gasoline that
is underlain by fill and natural claylike soils. First, a full-size prototype of the integrated
pneumatic fracturing/bioremediation system was developed. The site was then pneumatically
fractured, and periodic injections of nutrients are continuing over a period of 10 months. Off-
gases from the monitoring wells are being analyzed for BTX, oxygen, methane, and carbon
dioxide to evaluate process effectiveness. Preliminary results from the laboratory studies and
field demonstration available at the time of the conference will be presented.
•D.S. GOVHtNHENT PRINTING OF7ICE:1994-550-Oni/80391
294 1994 Symposium on Bioremediation of Hazardous Wastes
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