EPA/600/9-91/001
January, 1991
Biological Remediation of Contaminated
Sediments, with Special Emphasis
on the Great Lakes
Report of a Workshop
Manitowoc, Wisconsin
July 17-19, 1990
Edited by C.T. Jafvert and J.E. Rogers
Co-Chairmen:
Chad T. Jafvert and John E. Rogers
Environmental Research Laboratory
U.S. Environmental Protection Agency
Athens, Georgia 30613
Support was provided by the U.S. Environmental Protection Agency's
Great Lakes National Program Office, through the Assessment and
Remediation of Contaminated Sediments (ARCS) Program, by
Environment Canada, and by the U.S. Environmental Protection
Agency's Biosystems Technology Development Program.
Environmental Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Athens, Georgia
U.S. Environmental Proteeliod
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th Floor
Chicago, IL 60604-3590
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NOTE
This document was originally published in January, 1991. This copy is from a second
printing made in January, 1994. Copies are also available through the National
Technical Information Service (NTIS), 5285 Port Royal Road, Springfield, Virginia
22167, phone (703) 487-4650.
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DISCLAIMER
The information in this document has been funded wholly or in part by the United States
Environmental Protection Agency. It has been subject to the Agency's peer and administrative
review, and it has been approved for publication as an EPA document. Mention of trade names
or commercial products does not constitute endorsement or recommendation for use by the U.S.
Environmental Protection Agency.
U.S. Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th Floor
Chicago, IL 60604-3590
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FOREWORD
As environmental controls become more costly to implement and the penalties of judgement
errors become more severe, environmental quality management requires more efficient analytical
tools based on greater knowledge of the environmental phenomena to be managed. As part of this
Laboratory's research on the occurrence, movement, transformation, impact, and control of
environmental contaminants, research is performed on the biological remediation of contaminated
sediments.
The Assessment and Remediation of Contaminated Sediments (ARCS) Program is a major
activity of the U.S. Environmental Protection Agency that evaluates and demonstrates remediation
alternatives for contaminated sediments within the Great Lakes Basin and associated risk
assessments. In the summer of 1990, more than 60 scientists from the United States, Canada,
and The Netherlands participated in a special workshop to present the current state-of-the-science
concerning the biodegradation of polychlorinated biphenyls and polyaromatic hydrocarbons and the
biological treatment of metal species. This proceedings provides a synopsis of the information
exchanged at that workshop.
Rosemarie C. Russo, Ph.D.
Director
Environmental Research Laboratory
Athens, Georgia
ill
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ABSTRACT
These proceedings describe a workshop held July 17-19, 1990 in Manitowoc, WI, at which
biological remediation of contaminated sediments was discussed. For the purpose of the workshop,
contaminated sediments of primary interest were those within six of the Areas of Concern (AOC)
identified by the U.SVCanada International Joint Commission's Great Lakes Water Quality Board;
five of which are priority concerns of the U.S. Environmental Protection Agency's Assessment and
Remediation of Contaminated Sediments (ARCS) program.
The workshop was organized around four topic areas: (1) Overview of the Areas of
Concern; (2) Biological degradation of PCBs; (3) Biological degradation of PAHs; and (4) Biological
treatment of metal species. For the first topic area, presentations were made describing site
characteristic of the Ash tabula River, OH; Buffalo River, NY; Sheboygan River, WI; Grand Calumet
River, IN; Saginaw River and Bay, MI; and Hamilton Harbor, Ontario, Canada. For the remaining
topic areas, presentations were made by investigators actively involved in either bench, pilot, or
full-scale studies concerning these areas. In this document extended abstracts written by the
presenters are given, as well as brief summaries of the presentations and discussion sessions.
IV
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CONTENTS
1 Introduction 1
2 Summary
\
2.1 Areas of Concern 3
S2.2 Polychlorinated Biphenyls (PCBs) 7
\ 2.3 Polycyclic Aromatic Hydrocarbons (PAHs) 11
v 2.4 Metals 13
2.5 Conclusions 15
Abstracts
3 Areas of Concern
~ 3.1 Buffalo River Remedial Action Plan Strategy
J. C. McMahon 17
' 3.2 Fields Brook Superfund Site/Ashtabula River Area
P. Sanders 29
' 3.3 Coal Tar Contamination Near Randle Reef, Hamilton Harbor
T. Murphy, H. Brouwer, M. E. Fox, E. Nagy,
L. McArdle, and A. Moller 36
3.4 Indiana Harbor/Grand Calumet River AOC
R. Bunner 38
^3.5 Saginaw River/Bay AOC
G. Goudy 42
"3 6 Sheboygan River and Harbor, Sheboygan, Wisconsin
B. L. Eleder 50
4 PCBs
4.1 Aerobic Biodegradation of PCBs
R. Unterman 55
""4.2 Anaerobic Dechlorination and the Bioremediation of PCBs
J. F. Quensen, S. A. Boyd, and J. M. Tiedje 59
~-4.3 Dechlorination and Biodegradation of Chlorinated Biphenyls in
Anaerobic Sediments
G-Y Rhee and B. Bush 73
"4.4 PCB Dechlorination in the Sheboygan River, Wisconsin
W. C. Sonzogni 75
"4.5 Anaerobic and Aerobic Biodegradation of Endogenous PCBs
D. A. Abramowicz and M. J. Brennan 79
4.6 Remediation Pilot Study in the Sheboygan River Wisconsin, USA
D. S. Foster 88
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Contents
5 PAHs
5.1 The Use of a Mycobacterium sp. in the Remediation of Polycyclic
Aromatic Hydrocarbons
v C. E. Cerniglia 91
5.2 Fungal Degradation of PAHs
x J. Glaser 108
5.3 Recent Studies on the Microbial Degradation of PAHs and Their
Relevance to Bioremediation
\ J. Mueller 110
5.4 Biological Remediation of Contaminated Sediments in the Netherlands
H. J. van Veen and G. J. Annokkee 113
6 Metals
6.1 Bacterial Leaching of Metals form Various Matrices Found in
Sediments, Removing Inorganics from Sediment-Associated Waters
Using Bioaccumulation and/or Biofix Beads
P. Altringer and S. Giddings 127
6.2 Biological Treatment of Metal-Contaminated Water
H. Edenborn 145
6.3 Bioleaching of Ores
E. G. Baglin 148
6.4 Mechanisms of Bacterial Metals Removal from Solids
A. E. Torma and P. A. Pryfogle 159
6.5 Linking Biological and Hydrogeochemical Mechanisms of Sediment
Leaching
R. H. Lambeth and B. C. Williams 166
Appendix I - Program 173
Appendix II - List of Attendees 177
VI
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List of Figures
2.1.1 Areas of Concern 6
3.1.1 Buffalo river area of concern location map 28
3.2.1 Vicinity map, Fields Brook 33
3.2.2 Fields Brook site map 34
3.2.3 Design investigation sequence 35
3.5.1 Location of the Saginaw River/Bay Area of Concern 47
3.5.2 Spatial distribution of PCB in surficial sediments of the Saginaw River 48
3.5.3 Vertical distribution of PCB in sediments near Bay City WWTP 49
4.2.1 Capillary gas chromatograms showing the anaerobic dechlorination of
700-ppm Aroclor 1242 after 16 weeks of incubation 68
4.2.2 Decrease in the average number of chlorines by position at three
Aroclor 1242 concentrations as a result of dechlorination by Hudson
River microorganisms 69
4.2.3 Effect of incubation temperature on the dechlorination of
Aroclor 1242 by Hudson River microorganisms 70
4.2.4 Decrease in the average number of chlorines for four Aroclors
as a result of dechlorination by Hudson River microorganisms 71
4.2.5 Comparison of the dechlorination rates of 3,3',4,4'-CB, 2,3)3',4,4'-CB,
and selected tetra- and penta- CBs present in Aroclor 1242 72
4.5.1 Acceleration of the reductive dechlorination of PCBs upon addition
of nutrients (8 week timepoint). A) autoclaved control; B) includes
distilled water; C) includes RAMM minimal medium. All samples
contain 500 ppm PCB (70% Aroclor 1242, 20% Aroclor 1254,
10% Aroclor 1260) inoculated with sediments from the Hudson River 83
4.5.2 Dechlorination patterns observed under different conditions (18 week
timepoint). A) autoclaved control; B) includes RAMM (pattern M); C)
includes RAMM + cysteine hydrochloride at 1 g/L (pattern Q) 84
4.5.3 Dechlorination of endogenous PCB contamination in Hudson River
sediments with sediments with RAMM (18 week timepoint)
A) autoclaved control; B) experimental 85
4.5.4 Dechlorination of endogenous PCB contamination in South Glens
Falls soil with 25% Hudson River sediment (23 week timepoint).
A) autoclaved control; B) experimental 86
4.5.5 Sequential Anaerobic/Aerobic treatment of endogenous PCB
contamination in Hudson River sediments. A) Aroclor 1242;
B) environmentally dechlorinated Aroclor 1242; C) B+ aerobic
treatment (1 OD cells; 1 day timepoint) 87
5.1.1 The structures and chemical and toxicological characteristics
of polycyclic aromatic hydrocarbons 99
5.1.2 Schematic representation of the environmental fate of
polycyclic aromatic hydrocarbons 100
5.1.3 Major pathways of bacterial oxidation of polycyclic
aromatic hydrocarbons 101
5.1.4 Photograph of Mycobacterium sp. colonies on MBS agar containing
low-levels of nutrients and coated with pyrene. The clear
zones around the bacterial colonies indicate pyrene utilization 102
5.1.5 Mineralization of naphthalene, phenanthrene, pyrene, fluoranthene,
1-nitropyrene, 6-nitrochrysene and 3-methylcholanthrene by the
Mycobacterium sp 103
5.1.6 The pathways utilized by the Mycobacterium sp. for the oxidation
of pyrene 104
Vli
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List of Figures
5.1.7 The pathways utilized by the Mycobacterium sp. for the oxidation
of naphthalene 105
5.1.8 The pathways utilized by the Mycobacterium sp. for the oxidation
of fluoranthene 106
5.1.9 The pathways utilized by the Mycobacterium sp. for the oxidation
of 1-nitropyrene 106
5.1.10 Mineralization of phenanthrene, 2-methylnaphthalene, pyrene and
benzo[a]pyrene in microcosms from De Gray Reservoir sediments and
water with and without Mycobacterium inoculation 107
5.3.1 Tri-phasic treatment approach 112
5.4.1 Hydrocyclone 123
5.4.2 Hydrocyclone results 124
5.4.3 Volume reduction by dewatering 125
5.4.4 Intensive versus extensive treatment (Geulhaven Rotterdam) 126
6.1.1 CN removal in single-pass 3-column trickling reactor 142
6.1.2 Metal sorption using BIO-FIX beads 143
6.1.3 Conceptual configuration for bioleaching sediments 144
6.3.1 Shake-flask bioleaching of Three Kids ore, 5 pet. factory molasses 157
6.3.2 Column bioleaching of Three Kids ore, 3 pet. food-grade molasses 158
Vlll
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List of Tables
3.1.1 Great Lakes water quality agreement impairment indicators 21
3.1.2 Summary of impairments, causes and sources 27
3.2.1 Priority pollutants found in sediment at the Fields Brook site 31
3.2.2 ARI - Main stem river sediment samples selected parameters -
statistical data presented on dry weight basisdocations
12201 through 20502) 32
4.2.1 Maximal observed dechlorination rates (means with standard deviations)
of the Aroclors tested for microorganisms collected from the two sites 67
4.5.1 Effect of RAMM components on dechlorination rate 82
5.4.1 Results of practical hydrocyclone applications 121
5.4.2 Results of biodegradation for various sediment samples 122
6.3.1 Shake-flask bioleaching of Manganese ores 154
6.3.2 Abiotic leaching of Three Kids ore with organic acids 154
6.3.3 Column and heap bioleaching of Three Kids ore 155
6.3.4 Stillwater ore minerals 155
6.3.5 Bio-oxidation of stillwater flotation concentrate 156
6.3.6 Cyanidation of bioleached and As-received stillwater concentrate 156
IX
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ACKNOWLEDGEMENT
We gratefully acknowledge the efforts of all those individuals who contributed in one form or
another to the origination of this report. The Workshop and this report, truly, were group projects.
Recognition is extended to David Cowgill and Paul Horvatin of E.P.A.'s Great Lakes National
Program Office (GLNPO) and members of GLNPO's Engineering and Technology Workgroup, and
its Chairman, Steve Yaksich of the U.S. Army Corp of Engineers, Buffalo District, for their support
and planning input. Also, we deeply appreciate the support and planning input provided by Griff
Sherbin and Ian Orchard of Environment Canada. Direction by all these individuals has enhanced
this report considerably by their endeavor to assure its applicability to contaminated sediment
scenarios within the Great Lakes. Appreciation is given to Paulette Altringer of the Bureau of
Mines, Salt Lake City Research Center, who was instrumental in organizing the Metals session.
Janice Heath of Technology Applications Inc. and Patricia Van Hoof of The University of Georgia
provided indispensable assistance in making Workshop arrangements and coordinating activities
during the Workshop. In particular, we wish to acknowledge Janice Heath for her singular effort
of synthesizing the many diverse forms of material submitted by the speakers into a consistently
formatted and understandable document.
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1 INTRODUCTION
The current state-of-the-science of biological remediation of contaminated sediments was
discussed in a workshop held July 17 - 19, 1990, in Manitowoc, WI. Special emphasis was
devoted to remediation alternatives for sediments within the Great Lakes Basin. The workshop
was supported by the U.S. EPA's Great Lakes National Program Office, through the
Assessment and Remediation of Contaminated Sediments (ARCS) Program, by Environment
Canada, and by EPA's Biosystems Technology Development Program. More than 60 scientists
from state and federal agencies, academia, and the private sector from the United States,
Canada, and The Netherlands participated.
For the purpose of the workshop, the sediments of primary interest were those within the
Areas of Concern identified by the U.SVCanada International Joint Committee's Great Lakes
Water Quality Board. Most of the 42 Areas of Concern are located in harbors, bays, or river
mouths; 25 are located within U.S. waters, 12 within Canadian waters, and 5 within inter-
national channels. Remedial Action Plans currently are being developed for these areas under
the 1987 revision of the Great Lakes Water Quality Agreement. A major purpose of EPA's
ARCS Program is to evaluate remediation alternatives for the cleanup of these sites with
special emphasis given to five sites. These five are Ashtabula River, OH; Buffalo River, NY;
Sheboygan River, WI; Grand Calumet River, IN; and Saginaw River and Bay, MI. Two of
these five overlap EPA Superfund sites to some extent.
The Workshop was organized around four topic areas:
I. Overview of the Primary Areas of Concern
II. Biological Degradation of PCBs, Laboratory and Field Studies
III. Biological Degradation of PAHs, Laboratory and Field Studies
IV. Biological Treatment of Metal Species
For the first topic area, presentations were made describing site characteristics of the five
primary U.S. Areas of Concern and for Hamilton Harbour, Ontario. Major contaminants within
these and other areas include polychlorinated biphenyls (PCBs), polycyclic aromatic
hydrocarbons (PAHs) and various heavy metal species. The toxicity and recalcitrant nature of
these compounds have caused serious environmental concern. Moreover, these classes of
contaminants present serious and rather complex treatability problems for essentially all
remediation technologies (including biological processes).
For the remaining topic areas, presentations were made by investigators actively involved
in either bench-, pilot-, or full-scale studies within these topic areas. To focus dialogue on the
Workshop intent, the participants were asked to address or keep in mind the following general
questions during presentations and discussion periods:
1. What stage of development have specific bioremediation technologies reached (e.g.,
laboratory research, laboratory-field development, or full-scale operation)?
2. Which development directions are logical continuations for the specific laboratory studies
(e.g., above ground reactor treatment, in situ treatment, CDF modification, land
farming, or other)?
3. What level of development is necessary before a full-scale application of this technology
is feasible?
4. What are the rate limiting factors controlling the optimization of the laboratory or field
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2 Introduction
process? These factors may be site characteristic considerations, process operation
considerations, or both.
5. What types of costs are or will be associated with the development of proposed
treatment (e.g., capital, labor, maintenance)? How is this cost dependent on site
location and characteristics?
6. What other waste streams may be generated? What losses to the environment will
result from specific treatment alternatives? What contaminant residues will result?
7. What concerns you the most regarding the application of specific bioremediation
technologies to the problems associated with Great Lakes sediments?
8. Given the dissimilarity between bioremediation technologies and other physical or
chemical treatment technologies, how should one compare the environmental and
financial costs associated with each?
These questions were intended to be used as a guideline. Answers to some were
addressed in detail for specific bioremediation alternatives and are addressed in the Summary
sections and in several of the Abstracts. The answers to others were only alluded to or are
presently unknown. To a large extent this is because biological remediation to treat
contaminated sediments may take several forms. Each form (or process design) has its own
list of factors or parameters associated with it that must be considered when optimizing
treatment. Hence, there are generally no simple answers to questions regarding the feasibility
of biological remediation alternatives. Sediments are generally not contaminated with single
compounds or even classes of compounds. Additionally, the interactions among the various
organisms responsible for the decomposition of anthropogenic compounds and the sediment
matrix are unknown in many cases. Such intricacies make a concise summary of this diverse
workshop difficult; however, several general conclusions can be drawn. We hope this
Proceedings will benefit scientists and engineers who must make choices among diverse
treatment technologies. A brief summary of the Proceedings of this workshop has been
published by C. T. Jafvert (J. Great Lakes Res. 16(3):337-338, 1990).
Chad T. Jafvert
John E. Rogers
September 1990
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2 SUMMARY
2.1 Areas of Concern
Janice K. Heath
Technology Applications, Inc.
c/o Environmental Research Laboratory
U.S. Environmental Protection Agency
Athens, GA 30613
The locations of the 42 Areas of Concern (AOC) identified by the U.S7Canada International
Joint Commission's Great Lakes Water Quality Board are illustrated in Figure 2.1.1.
Environmental characteristics of the five U.S. AOC whose names are given in this figure were
described by either the State AOC Remedial Action Plan Coordinator or the Superfund Site
Coordinator for the adjacent Superfund site. A description of Hamilton Harbour, Ontario, was
given by Thomas Murphy of Environment Canada.
John McMahon, of the New York State Department of Environmental Conservation (DEC),
presented information on the Buffalo River AOC and the Remedial Action Plan Strategy. The
Buffalo River, located in western New York State, flows into Lake Erie near the mouth of the
Niagara River. Historically, the Buffalo River was used by industries as a transportation
channel, a source of cooling water, and a means of disposing of wastewater. These industries
were involved in chemical manufacturing (dyes and acids), coke and steel production, and oil
refining. Only two of these facilities are still in operation and they are under strict pollution
control regulations. Over the years, however, the pollution these industries generated
contaminated the river sediments and left hazardous waste on the banks. The bottom
sediments contain PAHs, PCBs, and heavy metals, which continue to be a source of
contamination to the Buffalo River, as are hazardous waste sites along its banks. Another
source of pollution to the river are combined sewer overflows that release dilute sewage and
associated contaminants into the river during storm events. In order to restore the Buffalo
River's integrity, a Remedial Action Plan (RAP) strategy was devised. The short term goal is
to restore the river's ecological system, while the long term goal is to eliminate the sources of
pollutants to the river. Presently, the DEC has committed to several initial actions
recommended by the RAP for dealing with the sources of contaminants and remediation of the
area.
An overview of the Fields Brook Superfund site and the Ashtabula River AOC was given by
Pete Sanders of the U. S. Environmental Protection Agency, Region V. The area involved is
located in northeast Ohio. Fields Brook flows into the Ashtabula River about 8000 feet from
the point at which the river empties into Lake Erie. The Fields Brook site has been on the
National Priorities list since the first list was established under Superfund in 1983.
Contamination of sediments in this area has resulted from a variety of chemical manufacturers
located along Fields Brook. The sediment contaminants include a variety of organic compounds
and heavy metals. Clean up and remediation efforts for the Superfund site will involve
excavating, dewatering, and either landfilling or thermally treating the contaminated sediment.
The option to landfill or thermally treat the sediment will be decided after investigating the
mobility of the contaminants, the toxicity and concentration of the contaminants, and the
concentration of PCBs. Thermal treatment was indicated in the Record of Decision (ROD)
signed by the U.S. EPA in 1986. The ROD also advised a Remedial Investigation
(RD/Feasibility Study (FS) to recognize current sources of contamination to Fields Brook and to
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4 Summary
examine the extent of contamination to the Ashtabula River. The Ashtabula River
investigation included sediment, water, and fish sampling, and started late in 1989. A plan is
being developed by the Army Corps of Engineers to dredge the upper portion of the
contaminated sediment from the river and place it in a confined disposal area.
Robert Bunner of the Indiana Department of Environmental Management gave an overview
of the Indiana Harbor/Grand Calumet River AOC. He showed the initial segment of a video
tape entitled "The Grand Calumet River, A River of Contradictions." A copy of this tape can
be obtained by contacting Robert Bunner, Indiana Department of Environmental Management,
105 S. Meridian Street, Indianapolis, Indiana 46225. The geographical region associated with
the harbor and river has had a long history of industrial activity, beginning in the early part
of this century. In fact, over the decades, this portion of the Grand Calumet River has been
modified dramatically from its preindustrial state. The major industrial complexes associated
with this site over this time period are steel plants. Currently, the dredging of contaminated
sediments is proposed in the harbor primarily for navigational purposes, and in the river for
remediation purposes. Within this system, deposition of sediments is to the point where it is
no longer safe to navigate large ships. The sediments within the river and harbor are
contaminated with PCBs, PAHs, and heavy metals including cadmium, chromium, and lead.
To give some historical context as to the industrial nature of this area, it is estimated that the
land extending one mile from Lake Michigan in the harbor area consists of fill generated from
the steel industries over the decades.
Bonnie Eleder of the U.S. Environmental Protection Agency, Region V, presented an
overview of the Sheboygan River and Harbor Area of Concern including the Superfund site.
The Superfund site includes about 14 miles of the river from the dam at Sheboygan Falls,
Wisconsin, east to the harbor on Lake Michigan, including the flood plain of this part of the
river. The Area of Concern includes the entire watershed of the Sheboygan River. From 1950
until 1969, the Army Corps of Engineers dredged the lower river and harbor for navigation
purposes. The dredging was stopped when heavy metals were found in the sediment. After
more testing and sampling, high levels of PCBs also were found. In 1986, the area was added
to Superfund's National Priorities List. Three potential sources of contamination were named,
and after negotiations, one of the potentially responsible parties agreed to undertake a
Remedial Investigation/Feasibility Study (RI/FS) to determine the extent of contamination and
look at potential remedial alternatives to deal with the contamination. An engineering firm
was hired to conduct the RI/FS. Certain remediation alternatives and associated alternatives
are now being assessed including: biological treatment within a pilot confined treatment facility,
sediment removal, in situ armoring, and monitoring programs.
Greg Goudy of the Michigan Department of Natural Resources presented a summary of the
Remedial Action Plan for the Saginaw River and Bay AOC. The Saginaw River empties into
Saginaw Bay, located along the eastern shore of Michigan's lower peninsula. As he stated, the
water quality of the Bay and River have improved over the last 20 years, but problems still
remain. Three primary water quality problems have been recognized in this area. The first is
eutrophication which has lead to extensive algal blooms causing taste and odor problems with
drinking water from the bay. Second is bacterial contamination caused by combined sewer
overflows that discharge raw sewage into the Saginaw River during heavy rains. Finally, there
is contamination by anthropogenic compounds such as PCBs and chlorinated dioxins. These
have been found in fish tissue and have resulted in public health advisories against fish
consumption. The intent of the RAP is to restore the river and bay area to a water quality
that is safe so that the areas can once again be used as originally intended without risk to
human or environmental health.
Tom Murphy from Environment Canada, discussed the Hamilton Harbour AOC. Hamilton
Harbour is located in Hamilton, Ontario on the western bank of Lake Ontario. The main
pollutants in the harbour are PAHs, coal tar, and heavy metals. These chemicals have led to
the unhealthy fishery, which is of great concern to the public, who have formed a citizens
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Summary 5
action group. Source controls imposed on the industries in this area have reduced air and
water contamination; however, some contaminated "hot spots" still exist. In the areas with low
metal contamination, there has been natural degradation of the PAHs and coal tar.
Pretreatment methods to make the metals less bioavailable so the bacteria can more easily
degrade the PAHs and coal tar are being tested. Another concern, however, is oxygen
availability in the sediments. For the biological degradation of PAH compounds, oxygen is
necessary; however, the sediments are largely under anoxic conditions. At this time, the
recommendation is to dredge and treat the "hot spots" while continuing to study remedial
alternatives to this problem.
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Summary
Figure 2.1.1. Great Lakes Areas of Concern
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Summary
2.2 Polychlorinated Biphenyls
Chad T. Jafvert
Environmental Research Laboratory
U. S. Environmental Protection Agency
Athens, GA 30613
The congener mixtures of polychlorinated biphenyls (PCBs), produced by Monsanto, were
sold under the trade name Aroclor, and contained from 30 to 60 individual congeners
(chlorinated analogs of the parent biphenyl of 209 congeners theoretically possible). The last
two digits of the number specifying each Aroclor mixture, i.e., 1248 relate the percent chlorine
content by weight of that mixture. Hence, Aroclors 1242 and 1248 are generally referred to as
the lower (molecular weight) Aroclors and contain mostly di-, tri-, and tetrachlorobiphenyls,
whereas Aroclors 1254 and 1260 are referred to as the higher Aroclors and contain mostly
penta-, hexa-, and heptachlorobiphenyls. Recent evidence, much of which was presented during
this session, shows that the complete microbial degradation of Aroclors is possible. However,
the complexity of the microbial processes responsible for degradation, the complexity of the
compounds themselves, and the complexity of sediment interactions with microbes and
individual congeners makes this class of compounds one of the greatest challenges to
bioremediation technologies.
Ronald Unterman presented information regarding the aerobic biodegradation of PCBs.
Under aerobic conditions, PCB biodegradation is a cometabolic process in which another
substrate, such as biphenyl, is required as a carbon and energy source. Because no advantage
may be gained by the indigenous microorganisms in degrading PCBs (no energy is gained), the
introduction of exogenous organisms, specifically isolated for their PCB degrading abilities, may
facilitate this process. He noted that Envirogen, Inc. is actively involved in isolating bacterial
strains with PCB-degrading capabilities, elucidating the biochemical pathways by which these
compounds degrade, and isolating the genes responsible for the various steps involved in this
degradation. Only the lower chlorinated congeners (i.e., mono-, di-, tri-, tetra-, and some penta-
) are amenable to aerobic degradation. As the number of chlorine substituents increases on the
biphenyl moiety, aerobic degradation is reduced. The positional selectivity of PCB-degrading
strains was also noted, suggesting that the use of several strains may result in the widest
range of degradation of all congeners. Several key parameters that must be evaluated when
optimizing aerobic degradation in the field include bioavailability, temperature, and utilization
of proper microbial strains. He stressed that experiments purporting to show biodegradation of
PCBs by simply quantifying total GC peak areas must be carefully evaluated.
John Quensen presented results of laboratory experiments designed to elucidate the
anaerobic biodegradation processes of PCBs. He stressed that anaerobic reductive
dechlorination occurs only for the more heavily chlorinated PCB congeners. Several of the
mono- and di-chlorinated congeners do not appear to be dechlorinated to any extent and
represent terminal products of the higher chlorinated congeners. Reductive dechlorination may
be of selective advantage to microorganisms in that it can result in a gain in energy for the
organisms and can serve as a terminal electron sink. Terminal electron acceptors are often
limiting for microbial growth in anaerobic systems. Drs. Quensen, Boyd, and Tiedje have
developed a method of transferring PCB-degrading organisms from acclimated sediment to clean
or sterilized sediments. Such transfers of activity have now been made for over 10 serial
passes. He discussed the difference in dechlorination patterns within sediments from various
locations historically exposed to different Aroclor mixtures. In all studies, however,
accumulation of ortAo-substituted products was observed. The extent of dechlorination was
shown to be concentration dependent. This may result both from decreased bioavailability of
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8 Summary
compound at lower concentrations and/or from increased growth of organisms at higher
compound concentrations. He related these studies to the potential for bioremediation of
contaminated sites, suggesting either that anaerobic biodegradation alone will reduce sediment
toxicity, or that anaerobic/aerobic sequential treatment may reduce the total concentration of
PCB congeners. Site assessment should involve evaluation of the presence of dechlorinating
microorganisms, in situ dechlorination patterns, sediment type, nutrient and organic carbon
concentrations, inhibitor concentrations, and the bioavailability of the PCBs.
G-Yull Rhee reported on laboratory studies of the anaerobic dechlorination of Aroclor 1242
and a single congener (2,3,4,2',4',5>-hexachlorobiphenyl) in Hudson River sediment. In the
Aroclor 1242 studies, dechlorination patterns were investigated as a function of Aroclor
concentration (100 to 1500 ppm on a sediment dry weight basis, and reducing conditions
(sulfide-reduced synthetic medium). After 3 months, significant changes in congener patterns
were evident, especially at 300 and 500 ppm Aroclor 1242 with mono-, di-, and
trichlorobiphenyls comprising 98% of the total remaining PCBs. Ort/io-substituted congeners
showed the most significant increases. After 6 months of incubation, congener profiles for the
100 and 800 ppm concentrations showed significant dechlorination, whereas no difference was
observed at 1200 and 1500 ppm. Similar to the results of others, no biodegradation other than
dechlorination was found. Anaerobic incubation of the single hexachlorobiphenyl produced
congeners with two to five chlorines per molecule. The relative concentration of these products
varied with incubation time.
William Sonzogni addressed the issue of whether PCBs are being biologically dechlorinated
in the Sheboygan River under ambient conditions. The contamination in this river is believed
to be primarily from Aroclor 1248 and 1254. Total PCB concentration ranged from 1586 ug/g
downstream from the site of contamination, to 0.04 ug/g upstream from the site, with the
highest PCB concentrations found in areas of sediment deposition. He presented strong
evidence that biological dechlorination was occurring in the river. This evidence included the
following observations: a shift in congener profiles (compared to 1248 and 1254) from the
higher chlorinated to the lower chlorinated congeners exists in sediment samples; meta- and
para- chlorinated congeners were depleted more than ortAo-chlorinated congeners; several
specific congeners were found in abundance; and finally congener patterns were found to be
PCB-concentration dependent with only samples with greater than 50 ug/g total PCB showing
these patterns. The physical and chemical processes that affect congener distribution were also
discussed. Abiotic degradation was ruled out because of the extreme conditions (temperature,
pH) necessary for this to occur over a reasonable time frame. Similarly, preferential sorption
of the more hydrophobic congeners would not result in the observed patterns. Laboratory
experiments with river sediments have yet to confirm these patterns. He also reported on a
multidimensional gas chromatography technique used to resolve congeners which normally co-
elute with conventional gas chromatographic methods. This analytical method is useful in
analysis of co-planar PCBs (those with dioxin-like toxic properties). Concentrations of these
congeners represent a fractional percentage of Sheboygan River PCBs.
Daniel Abramowicz presented results of laboratory studies in which the rate of anaerobic
dechlorination of PCB mixtures was enhanced by the addition of either nutrients, a complex
carbon source, a reducing medium, or surfactant. Additionally, he presented information
regarding the aerobic treatment of Hudson River sediments that had been previously
dechlorinated in the environment. The addition of minimal medium to Hudson River sediment
slurries was shown to increase the rate and magnitude of anaerobic dechlorination. Addition of
trace metals (at concentrations of less than 0.02 ppm) also increased the rate of PCB
dechlorination. The addition of the minimal medium and a chemical reducing agent (cysteine
hydrochloride) resulted in different patterns of dechlorination, indicating growth of different
microbial populations. Dechlorination was shown to occur in numerous, aged PCB-
contaminated sediments, including those from the Hudson River, the South Glens Falls
dragstrip (amended with Hudson River sediment), and Woods Pond. Aerobic treatment of
Hudson river sediments that had previously undergone extensive dechlorination of the higher-
chlorinated congeners (>85% mono- and dichlorobiphenyl remaining) resulted in greater than
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Summary 9
70% reduction of PCB concentration after one day of treatment.
Dawn Foster reported on the Sheboygan River and Harbor Remedial Investigation/
Feasibility Study Program. In the first phase of this program, the contaminants of concern
were identified to include PCBs and eight metals. This investigation led Tecumseh Products
Company (one of three potentially responsible parties) to propose an Alternative Specific
Remedial Investigation, which consists of pilot-scale studies to investigate various
bioremediation alternatives and bench-scale studies to investigate other alternatives. The
primary objectives included: evaluation of the potential to enhance biodegradation within a
confined treatment facility (CTF); evaluation of in situ armoring and the anaerobic
biodegradation of PCBs associated with these capped sediments; evaluation of mechanical
dredging methods and monitoring of the impact of these activities on the water column; and
bench-scale tests of other innovative technologies. The pilot-scale CTF constructed for the
enhancement studies has a capacity of 2500 cubic yards and has four cells that can be used to
test various treatment scenarios. In addition, various schemes will be examined for the
treatment of the cell effluent. Bench-scale studies are currently underway at the University of
Michigan that will provide information for the design of CTF enhancement studies by the
addition of various amendments. Armoring of in-place sediments was accomplished by placing
a geotextile material over the in-stream sediments followed by successive layers of bank run off
material (6 inches), another geotextile layer, and a final layer of stones and gabions. Sampling
ports through these layers will allow for the monitoring of the natural biodegradation process.
Questions and comments during the PCB discussion sessions encompassed a number of
issues; some related and others very specific and unique. The topics dealt with in some detail,
in order of their deliberation, included the following.
Development of a Sediment Testing Protocol. The speakers described many laboratory
experiments which all have a common theme - that of measuring biodegradation of PCB
compounds in sediment systems, and amending these systems to enhance rates of
transformation. However, no standardized testing protocol exists to facilitate testing by other
scientists or engineers for assessing the feasibility of bioremediation at other sites. It was
suggested that such a protocol be developed, and could be used as a guideline, as opposed to a
methods document, simply because of the continuously developing nature of this science, and
the rapidly expanding data base. It was mentioned that the EPA's Biosystems Technology
Development Program is currently developing a testing protocol for contaminated aerobic soils,
and that much could be learned from this other effort in developing one for PCB-contaminated
sediments. Such a document would be of value to Regional (Superfund) scientists and
engineers who must evaluate and oversee bench-scale and pilot-scale studies, and to Remedial
Action Plan coordinators who must develop remedial options for contaminated sites.
Deposition of Other Contaminated Sediments on Armored Material. Several questions were
asked concerning the integrity of the armored sediments and/or the possibility of re-
sedimentation of other contaminated sediments on the armored areas, necessitating re-armoring
of the Sheboygan sediments. In response, the pros and cons of armoring were discussed.
Basically, armoring can only be evaluated as an option in areas where (1) dredging of
sediments is not necessary, and (2) high currents will not disturb the armoring material. In
the case of the Sheboygan sediments, re-sedimentation of contaminated sediments should not
occur because of the elimination of the source (basically, the sediments are the current source).
Bioaccumulation of PCBs in Lower Organisms. It was asked whether the trends in
bioaccumulation of PCBs in lower organisms should coincide with those found in higher
organisms (i.e., fish). The discussion that followed addressed the issues of both chemical phase
distribution and chemical metabolism. From a thermodynamic standpoint, the potential to
bioaccumulate (normalized to organism lipid content) in higher and lower organisms is the
same. Factors limiting the kinetic uptake and depuration of these compounds in these
organisms, however, may differ. In addition, the ability of some organisms to metabolize these
compounds may result in body burdens less than those found in other organisms that can not
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10 Summary
metabolize them. The thermodynamic potential, the limiting kinetic factors (including such
things as migratory patterns), and the organism's ability to metabolize the compounds must all
be factored into the observed environmental bioaccumulation of these compounds.
Natural Substrates of the Aerobic Pathways of PCB Degradation. Because the aerobic
degradation of PCBs occurs through a cometabolic pathway, a question was raised concerning
the identity of the natural substrates for which this metabolic pathway exists, and the natural
distribution of the organisms containing the responsible enzymes. The point was raised that,
in natural systems, either the concentration of the final electron acceptor and/or carbon source
is sometimes the growth limiting factor, not the energy source, per se. It was suggested that
diagenetic humic material, which contains a considerable amount of aromatic structure and
already contains fairly reduced carbon, is the natural substrate. This would account for the
relatively ubiquitous distribution of PCB-degrading organisms in the environment.
Mass Balance Accounting of PCBs in the Environment. Part of the problem in identifying
natural PCB degradation is that the historical mass loadings of PCBs into various river and
harbor systems is not known, and therefore mass balance estimates on losses cannot be easily
made. From sampling exercises on the Hudson River between the mid '70s and '80s, it
appears that half of the PCBs estimated to be present from the first sampling period
(approximately 500,000 Ibs) have been lost from the system. It was suggested that long-term
sampling programs be initiated in areas where physical transport mechanisms are minimized
and where good mass balances can be measured to get a better idea of the extent to which
natural biological decay processes are occurring. Such studies may be possible using existing
confined disposal facilities.
Effects of Toxic Metals on PCB Degradation Rates. In most of the Areas of Concern, when
PCBs are present, heavy metal contamination coexists to some extent. Very little information
is available, however, concerning the toxicity of various metal species to PCB degraders. Also,
it should not be assumed that high concentrations of metals will decrease degradation rates or
are responsible for low degradation rates. Speciation and redox state is important, as well as
how the metals are associated with the sediment material. It was generally agreed that metal
toxicity should be addressed to some extent in bench-scale studies as metal toxicity will be very
site-specific.
Questions of Scale-Up and Number of Pilot Studies. The basic question "where do we go
from here" was asked. Do we start new studies, and at what level of effort should these
studies proceed (i.e., bench, pilot)? The general consensus of the group seemed to be that
currently we are working with a fairly small data base. Several studies have shown positive
results, and several have so far been negative. All the effects of, and relationships among, the
various controlling factors are not known; hence, the clearest path to site-specific optimization
is not always obvious. It was generally agreed that as the results of more studies become
available, biological treatment technologies will be refined, and the limits of these technologies
will become clearer. Because each level of scale-up involves different aspects of treatability, the
design of more pilot-scale studies, based on the results of bench-scale studies was suggested.
Acceptable Clean-Up Concentrations. A concern was raised regarding the fact that there is
generally good success at high PCB concentrations (> 50 ppm) and poor success at low
concentrations (< 50 ppm): Whereas, even single digit ppm concentrations of PCBs in sediment
(dry weight basis) may relate to significant concentrations in fish species. The suggestion was
made that engineered systems should focus on these lower concentrations where other chemical
or physical destruction technologies do not appear to be economically feasible as final
remediation remedies. The point was raised that this phenomenon (concentration dependence)
may be largely a consequence of reaction kinetics (including mass transfer limitations) or
microbial induction. Both of these causes can be assessed at the bench-scale level.
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Summary 11
2.3 Polycyclic Aromatic Hydrocarbons
Patricia L. Van Hoof
University of Georgia
Athens, GA 30613
Polycyclic aromatic hydrocarbons (PAHs) are a major class of environmental contaminants
that are byproducts of 1) burning of fuel, 2) generation of synthetic fuels from fossil fuels, and
3) wood treatment. This class of compounds exhibits a wide range of toxicity, hydrophobicity,
and recalcitrance in aquatic systems. While biodegradation of low-molecular-weight PAHs by a
wide variety of microorganisms is well documented, there is limited information on the
microbial utilization of the more recalcitrant and toxic PAHs consisting of four or more fused
rings. In order for bioremediation to be considered a viable treatment of PAH-contaminated
sites, the organisms, the processes, and the environmental conditions necessary for the
degradation of these compounds must be identified. The speakers in this session address this
challenge.
Carl Cerniglia discussed the use of a Mycobacterium sp. in the remediation of PAH wastes.
The pyrene-degrading bacterium was isolated by direct enrichment from sediment taken from
an oil field in Port Aransas, Texas. The bacteria were found to be quite versatile, degrading
both low and high-molecular-weight PAHs possessing up to five fused rings. In microcosm
studies, the organism was able to compete against bacteria indigenous to a variety of
environments (freshwater, marine, pristine, polluted), and enhanced the mineralization of PAHs.
He noted that the rates of degradation were dependent on compound structure and site history.
Lower-molecular-weight PAHs were degraded faster than higher-molecular PAHs, and
contaminated sites (freshwater and estuarine) demonstrated higher degradation rates than
pristine ones. Low levels of organic nutrients were reported to be necessary to initiate growth,
suggesting the degradation process is co-oxidative. In addition, inorganic nutrient supplements
(N and P) enhanced PAH degradation. He pointed out that the mechanism of oxidation is
unique as the Mycobacterium has both mono-and dioxygenases to catalyze PAH degradation.
H.J. van Veen addressed the problem of contaminated sediments (oil, PAHs, and metals) in
the Netherlands. These sediments are of particular concern not only because of their
environmental impact, but also because of the need for frequent dredging of the country's many
waterways. The speaker gave a survey of the current state of full-scale sediment remediation
and the development of biological treatment. Volume reduction of dredged sludge consists of a
combination of two techniques: hydrocyclones and dewatering. The "heavier" sand fraction is
separated from the finer and often more highly-contaminated fraction using a hydrocyclone,
which utilizes tangential flow and centrifugal force. He stressed that this operation will not
benefit cleanup of dredged sediment consisting mainly of fine particles or with a high organic
carbon content. After separation, the fines fraction is dewatered with a belt press, filter press
or decanter using polyelectrolytes. The results of a number of practical cases demonstrate that
this type of treatment is fairly successful; however, a couple of problems were pointed out.
First, the composition of the sludge often deviates from that expected based on preliminary
investigation. Second, in some cases, the sand fraction has high PAH concentrations. When
all size fractions are contaminated, a sludge can only be treated intensively in a bioreactor.
Whereas sludges that can be fractionated are more effectively treated extensively, i.e. the sand
fraction can be land-farmed and the fine fraction can be considered a waste liquid and treated
in an aeration basin. Intensively treating PAH-contaminated sediments was shown to be faster
than the extensive treatment of the fractionated material; however, over longer time periods
both processes were equally effective. He noted that practical considerations, such as material
volume, rates of degradation, space and cost will determine whether intensive (bioreactors) or
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12 Summary
extensive processes are required.
John Rogers presented the work of James Mueller and colleagues on the microbial
degradation of PAHs and their relevance to bioremediation. The efforts of this group have
been focused on the isolation of microorganisms capable of degrading high-molecular-weight
(HMW) PAHs. Mixed bacterial cultures capable of utilizing HMW PAHs as sole sources of
carbon and energy for growth have been isolated. He described how they are making use of
these microorganisms in a recently developed tri-phasic sequential treatment system for the
remediation of creosote and similarly contaminated soil and water. Under a Federal
Technology Transfer Act, they were able to transfer some of their biotechnology to an
engineering firm which provided separation technology. The steps in this remediation process
include conventional soil washing, membrane extraction, and biodegradation of extracted
pollutants. Each step in this process results in the volume reduction of contaminated material.
Depending on the type of starting material, soil washing can reduce the contaminated volume
to as little as 10% of the initial value. While the soil washing process reduces the volume of
material requiring treatment, the process generates large amounts of contaminated wash water
along with accumulated fine particles. To address this problem, reverse osmosis hyperfiltration
through porous stainless steel membranes is applied to dewater and concentrate pollutants.
While the effectiveness of this system on soil wash water is currently being evaluated, they
have demonstrated that >99% of creosote components present in contaminated groundwater are
removed. The speaker emphasized the potential capabilities of the membranes to : 1)
fractionate mixtures of chemicals to increase degradation efficiencies or reduce toxicity (e.g.
metals), and 2) recycle surfactants used in soil washing. Finally, the wash water is fed to
specially enriched microbes housed in continuous flow bioreactors. The ability of these
organisms to degrade artificial creosote mixtures has been demonstrated. Field demonstrations
of this sequential treatment system are currently being evaluated.
John Glaser discussed the use of white rot fungi (Phanerochaete chrysoporium and P.
sodida) to degrade a variety of target pollutants, including PAHs, in a variety of media.
Phanerochaete sp. grow quite rapidly on decaying wood. Consequently, this fungus possesses
great potential to degrade aromatic components of hazardous waste, based on its ability to
degrade lignin. The enzymes of this fungus are extracellular, extremely strong oxidizers,
largely non-specific, and not commonly found in other organisms. The speaker pointed out that
the non-specificity of these enzymes provides this organism with a capacity to degrade a wide
range of substrates (e.g. PAHs, PCBs, pesticides, and dyes). Two types of media have been
recently tested, liquid treatment using rotating contactors, and soil treatment. The liquid
treatment shows promise and is currently under pilot-scale evaluation to better control pH and
the mixing domain within the reactor. The application of wood chips inoculated with
Phanerochaete chrysosporium and P. sodida to soil contaminated with pentachlorophenol
resulted in 82% and 85% reduction, respectively, after 46 days. He noted that this fungus does
not grow naturally in soil and is non-pathenogenic to plants and animals. The required field
conditions (e.g. target compound and oxygen levels, temperature, reactor configuration) for
optimal biodegrading activities are currently being investigated.
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Summary 13
2.4 Metals
Paulette B. Altringer
U.S. Bureau of Mines
729 Arapeen Drive
Salt Lake City, Utah 84108
To summarize the metals session, a brief overview of the Bureau of Mines and the related
areas of research it is involved with are given. This is followed by a summary of the session
presentations which addressed the research ongoing at the Bureau of Mines and associated
research at the Department of Energy's Idaho National Engineering Laboratory (INEL) related
to this area and the possible application of this research to the remediation of inorganic-
contaminated sediments. The presenters stressed that all the remediation answers to metal-
contaminated sediments do not currently exist, but rather that some interesting possibilities in
this area, analogous to other current ongoing research in the field of mining and metallurgy,
show potential applicability.
The Bureau of Mines was established in 1910 as a Federal Agency in the Department of
the Interior. The Bureau is a relatively compact and mature agency by Washington standards.
The Bureau employs 2,200 people and is organized into three main directorates: Finance and
Management, Information and Analysis, and Research. The research component of the Bureau
is the largest element of the Bureau's overall program, employing about 1,300 people, with nine
dedicated laboratories located across the country. The Bureau is different from most Federal
agencies in that the Bureau performs its research in house instead of contracting it out: the
one exception to the inhouse research is a healthy program in concert with the Department of
Energy's Idaho National Engineering Laboratory (INEL), where two of the sessions speakers
(Arpad Torma and Peter Pryfogle) are employed. Bureau of Mines research is targeted at
three main areas: (1) Health, Safety, and Mining Technology, (2) Minerals and Materials
Science, and (3) Environmental Technology. The Bureau is responsible for a number of major
activities related to the minerals industry. Among these responsibilities is the performance of
research on mining and metallurgical technologies. This research has led to a number of major
developments that have benefitted the industry and the people of this country.
The 75 years of research and technical experience have also resulted in the Bureau
becoming the government's principal expert in the area of selective extraction of inorganic ions.
This includes technology to extract low concentrations of metals and other inorganic materials
from their host environment, solid or liquid. This capability includes another relatively new
technique: "biotechnology", which is the use of bacteria to treat metal-contaminated solids and
liquids. The "newness" really refers to the use of biotreatment, under controlled conditions, as
part of a metallurgical treatment process; nature has employed this approach for millions of
years. These mechanisms have been and are being employed in the minerals industry on a
daily basis as part of leaching operations, for example, for the production of copper. Bacteria
were enhancing copper leaching long before man was aware of the bacterial leaching
interaction. This same basic mechanism, operating on an uncontrolled basis, contributes to
acid drainage from coal mines. Leaching inorganics from solids can be enhanced using bacteria
and, alternatively, other types of bacteria can precipitate metals and destroy toxic inorganic
processing chemicals in solutions. Both aerobic and anaerobic microorganisms are involved in
these processes. This biotechnology can be applied beyond the minerals industry to the field of
Superfund and RCRA remediation. Biotechnology often produces a lower level of contaminants
in the treated material than is possible to achieve using conventional physical and chemical
treatments. In some cases, combinations of biotechnical, chemical, and beneficiation techniques
might be the only way to achieve the low level of contaminants in treated materials required
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14 Summary
by environmental legislation.
Almost half of the Bureau researchers are involved in research that can be generally
described as "metallurgical" in nature. Research on extractive processes - selective capture of
one or more elements from host materials that are either natural or recycled materials --
represents a large component of this part of the Bureau's program. Four of the nine Bureau of
Mines laboratories have ongoing projects involving bioextraction of metals and INEL is actively
involved in associated biotechnical research.
In the first presentation, I presented the Salt Lake City Research Center's work involving
the bioaccumulation of elements such as arsenic, cadmium, lead, mercury and selenium, from
solution using both viable bacteria and biomass immobilized in what we call "BIOFIX" beads.
In addition, the destruction of cyanide in process streams using viable bacteria was discussed.
This research may have direct application to inorganics removal from sediment-associated
waters. This research is being expanded at the Salt Lake Research Center to include
bioleaching of inorganic contaminants from sediments and mine tailings using bacteria. The
nature of these low-level, high-volume wastes makes most processing options extremely
expensive. Bacterial leaching in situ or on heap pads may provide an answer to this wide-
spread problem.
Hank Edenborn, from the Pittsburgh Research Center, reported on biotechnology for the
remediation of acid mine drainage from coal mines. He described the use of "wetlands"
technologies for this purpose, and how this technology may be directly applicable to sediment
remediation. He also described the use of bactericides to inhibit bacterial leaching in the event
that sediments should have to be dredged from waterways immediately, but could not be
treated for a period of time. Bactericides would prevent the biologically mobilized inorganic
contaminants from leaching from the sediments and entering the surface or groundwater during
storage prior to treatment.
Betty Baglin reported on research at the Reno Research Center on the bacterial leaching
of manganese, platinum and gold ores as a means of improved leaching technology. She
related the applicability of this work to remediation of contaminated sediments.
The Department of Energy's Idaho National Engineering Laboratory (INEL) has been
studying the mechanisms of bacterial metals removal from solids and the application of these
results in conjunction with the Bureau of Mines. Arpad Torma from INEL discussed
biochemical possibilities of inorganic sediment remediation and Peter Pryfogle provided
information on INEL's research capabilities.
Robert Lambeth from the Spokane Research Center presented information on linking
biological and hydrogeochemical mechanisms (models) of sediment leaching. This is a complex
research area and involves (1) field and laboratory data requirements, and (2) computer model
requirements. The Bureau's Spokane Research Center has been using geochemical computer
models to interpret hydrogeochemical mechanisms of mine tailings and sediment leaching.
Recently, personnel from the Spokane and Salt Lake City Research Centers conducted a joint
sampling trip to a copper-gold tailings impoundment in Washington State in the hope of
linking biological to hydrogeochemical mechanisms of inorganic leaching. Currently a "cook-
book" for predicting contaminant fate at new sites does not exist, but rather the presentation
focused on an approach to developing techniques for predicting contaminant fate at new sites
based upon knowledge gained from sites that have already been studied.
The research presented during this session and described in the abstracts has great
potential for biotreatment of inorganics in sediments. Successful development of the
biotechnical techniques may provide on-the-shelf technology for environmental problems
untreatable with conventional technology today.
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Summary
2.5 Conclusions
During the past decade, a great deal has been learned regarding biological processes that
act to transform or mineralize anthropogenic pollutants, including those discussed in detail
during the Workshop. The ability of microorganisms to degrade or transform chlorinated
organic compounds such as the PCBs, polycyclic aromatic hydrocarbons (PAHs), and metal
species is now well documented. Yet, an understanding of how these mechanisms function in
environmental systems, to the extent that we can consistently optimize them for bioremediation
purpose, is not totally understood. Two general areas in which information gaps can be
grouped for the problem at hand include: (1) The specific processes and mechanisms controlling
observed degradation rates and patterns, and (2) issues associated with extrapolation of bench-
scale studies to pilot or full scale field studies. A majority of the specific questions and issues
that were discussed during the workshop fell into these two areas.
Clearly, a significant amount of information on the biological transformations of pollutants
already is known from process research. Much of this research is at the phenomenological
level. The results have helped identify empirically, or allude to mechanistically, the
interactions among microorganisms, pollutants, and the sedimentary and aqueous media in
which they exist. These interactions can be rather complex, even for rather simple systems,
such as the transformation of a single compound by a pure microbial culture in an
homogeneous solution. In this simple system, characterization of the degradation process
requires an understanding of nutrient and growth requirements, the kinetics of transformation
reactions, degradation pathways, pollutant concentration dependencies, effects of alternative
substrates and electron acceptors, temperature dependencies, the effects of metabolic inhibitors,
and in some cases, the effects of varying carbon sources. The additional complexity associated
with investigating the same microbial decay process in natural or manipulated sediments is
obvious. Additional consideration must be given to organic and inorganic inhibitor availability,
combined inhibitory effects, pollutant bioavailability and the kinetics of this availability, and
microorganism competition or cooperation of the indigenous bacteria. Although a complete
understanding of how these processes interact at specific sites would result in the most obvious
approaches to treatability, a comprehensive understanding may not always be necessary. In
many cases, biological treatment efficiency may be significantly enhanced (above background
levels) by regulating a few critical factors limiting activity. These factors must be identified at
the bench-scale level through simple process studies. In many cases, differences in these
controlling factors are reasons for the site (or sediment) specific nature of biological treatability
successes. Clearly, while much is known, a better definition of the chemical, physical, and
biological processes (or factors) controlling observed transformation rates and pathways in
natural and manipulated sediments will enhance the frequency and degree of bioremediation
successes.
On the other hand, the extrapolation of results from bench-scale studies to pilot or full
scale studies is largely untested for remediation of sediments contaminated with the pollutants
of concern. Examples of extrapolation were presented during the Workshop. They include
technologies developed for the separation or removal of metal species from mine tailings or
drainage, and other bioremediation technologies evolving the remediation of soils or liquid
waste streams containing organic contaminates. Also, applied bioremediation may take many
forms, from simple low energy in situ (in place or CDF) systems to highly engineered, high
energy systems. Each form has its own list of design factors or parameters that must be
considered when optimizing treatment. As more field-scale efforts become realities, however,
systems obviously will be refined, and a clearer connection between bench-scale methods (and
treatment efficiencies) and applied field scale processes will become evident.
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16 Summary
This page is provided for your notes:
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ABSTRACTS
3 AREAS OF CONCERN
3.1 Buffalo River Remedial Action Plan Strategy
John C. McMahon
New York State Department of Environmental Conservation
600 Delaware Avenue
Buffalo, New York 14202
Abstract
In February 1987, the Buffalo River Citizens Committee was formed to assist tbe New York
State Department of Environmental Conservation in the preparation of a Remedial Action Plan
(RAP) for the Buffalo River. The goal of the plan is to restore and maintain the chemical,
physical, and biological integrity of the Buffalo River ecosystem in accordance with the Great
Lakes Water Quality Agreement (GLWQA). The GLWQA lists conditions that indicate
impairments of environmental quality. Scientific data and professional opinions were used to
confirm the impairments and link them to causes. The RAP addresses the river's
environmental concerns through a remedial action strategy to address contaminants and their
sources in the Buffalo River.
Introduction
As a tributary to the Great Lakes, the largest freshwater basin in the world, the Buffalo
River watershed feeds one of the most important ecosystems in New York State. Conditions
that impact the water quality of the Buffalo River may affect the water quality of the
downstream international waters of the Niagara River, Lake Ontario, and the St. Lawrence
River. As a result, pollutants added to the Buffalo River ecosystem may contribute to
impairments of these downstream waters that are part of the Great Lakes system.
Improvements to the environmental integrity; of the Great Lakes can best start with its
harbors and tributaries, such as the Buffalo River, where pollutants are concentrated before
they disperse throughout the lakes.
The high concentration, of past industrial discharges to the Buffalo River has polluted the
river and its sediments. The area exhibits environmental degradation and some beneficial uses
of water and biota are impaired.
The United States-Canada International Joint Commission (IJC) designated the Buffalo
River as one of 42 Areas of Concern (AOC) where pollution problems may affect the health of
the Great Lakes ecosystem. The IJC requested that the responsible jurisdiction prepare plans
for remediation of the AOCs.
The 1987 amendments to the United States-Canada Great Lakes Water Quality Agreement
(GLWQA) specify requirements for "remedial action plans" (RAPs) for the Areas of Concern.
The RAPs are to define environmental problems and identify actions needed to restore
beneficial uses of the waterbody. Plans are to embody a systematic, comprehensive, ecosystem
approach to restoring and protecting the biota and water quality. They should set time
schedules, name responsible agencies, and describe processes to monitor the AOC environment
and track implementation. The lead agency for a RAP should work closely with citizens to
17
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18 Areas of Concern
develop an ecosystem-based plan that represents the concerns of the local community.
The Buffalo River RAP was developed by the New York State Department of Environmental
Conservation (DEC) in cooperation with citizens concerned about the river's revitalization. In
1987 a group of interested citizens was appointed by DEC as the Buffalo River Citizens'
Committee (BRCC) comprising 21 environmental, small business, university, community, and
local government representatives. BRCC representatives and key DEC staff created a
10-member steering committee that directed the development of the Buffalo River RAP. The
steering committee established the goals of the RAP, mapped out a project workplan, defined
responsibilities, and developed and reviewed data summaries and document drafts.
This document summarizes the Buffalo River Remedial Action Plan that resulted from this
cooperative endeavor. More detailed information about problems and sources affecting the
Buffalo River, remediation programs, recommendations, and agency commitments is contained
in the full RAP report.
Setting
To understand the problems of the Buffalo River and the remedial actions needed to resolve
these problems, it is important to understand several things about the river: (1) where it is
located and the general character of its surroundings (the geography); (2) the uses of the river
from which benefits are derived (beneficial uses); (3) the occurrence, distribution, and movement
of water (hydrology) and sediments in the AOC and its watershed that carry pollutants and
constrain remedial actions; and (4) the water quality of the three tributary creeks that drain
into and affect the AOC.
The following describes the Buffalo River AOC and watershed area and sets the scene for
the discussion of remedial actions. It describes geography, beneficial uses, and hydrology and
bottom sediments, first for the AOC, and second for the watershed. A description of the water
quality in the tributaries is also included.
AREA OF CONCERN
Geography
The Buffalo River AOC is located in the City of Buffalo, Erie County, in Western New York
State (Figure 3.1.1). It extends about six miles from the mouth of the Buffalo River to the
eastern border of the City of Buffalo. In this area, the water level of the river is influenced by
the level in Lake Erie. The river flows from the east and enters Lake Erie near the head of
the Niagara River.
The river is dredged to just below the junction of Cazenovia Creek, and is used as a
transportation channel. It passes through an industrial area characterized by some active
industries, but also by many abandoned buildings, junkyards, and trash-littered areas that give
it the appearance of an industrial wasteland.
Beneficial Uses
Industrial
The Buffalo River historically served the industries along its banks as a convenient
transportation corridor, a source of process and cooling water, and a receptacle for wastewater.
The major industries include two grain milling firms, General Mills and Pillsbury, two chemical
companies, Buffalo Color and PVS Chemicals (both formerly Allied Chemical), coke and steel
manufacturing by Donner-Hanna Coke and Republic Steel (both no longer operating), and a
Mobil Oil Company refinery (currently functioning only as a storage terminal). The Buffalo
River Improvement Corporation (BRIG) was formed in the late 1960s to supply water from the
Buffalo Harbor on Lake Erie to these industries (except for the grain milling firms) for process
and cooling purposes. Because of industrial plant closures and process shutdowns, current
BRIG pumpage and discharge is down from 120 million gallons per day in the late 1960s to 18
million gallons per day.
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J.C. McMahon 19
Currently industries are operating under strict pollution control regulation by the state.
However, their past operations have created a legacy of contaminated sediments on the river
bottom and abandoned hazardous waste deposits along its banks.
Combined Sewer Overflows (CSOs)
The City of Buffalo discharges excess water collected during times of runoff through a
combined sewer overflow system. The system was designed to collect and transport both
sanitary sewage and wet-weather storm flow in the Buffalo area. CSOs prevent sewers from
backing up and flooding city streets during storms. However, their existence also means
untreated sewage is discharged during some storms and this has been a problem in the AOC.
The Buffalo Sewer Authority currently is reevaluating the CSO system.
Commercial Shipping
The US Army Corps of Engineers maintains the river within the AOC (by periodic
dredging) as a transportation corridor for commercial freight vessels. Dredging disturbs bottom
life and the bulkheading and dock construction by private interests along the river bank have
removed wetlands and shallow areas which were once habitat for fish and wildlife.
Recreation
People use the AOC for recreation. A few people fish the river, although the state health
department advises against consuming certain species of fish taken there. Fishing use is
restrained also because of limited land access points, a perception that the river is polluted,
and the ready availability of nearby alternative fishing sites. Small, powerboats travel the
river in the AOC for recreational purposes primarily near the mouth or the river.
Swimming is not a common activity, probably because Lake Erie is more accessible and
more aesthetically pleasing.
River Hydrology and Bottom Sediments
The US Army Corps of Engineers dredges the river to maintain it at a depth of 22 feet
below low lake level for navigation purposes. Dredging the Buffalo River slows the river flow
and increases the volume of backflow from Lake Erie. When the flow is high, the river has a
"riverine" (one directional) character. Under low flow conditions, the river takes on an
"estuarine" (two directional) character. When this occurs, the river is influenced by lake level
variations associated with the passage of storms through Lake Erie and by seasonal thermal
differences between lake and river waters. The river and lake waters do not remain separate,
but mix at varying rates depending on relative water temperatures.
Studies of bottom sediments show that the river traps all sand particles until its flow
exceeds 20,000 cubic feet per second, which occurs only rarely. The finer clay and silt particles
pass through the river during the high flows associated with most storms, but are retained
during periods of normal and lower flow. The wider portions of the river trap the most
particles.
In addition to natural river flow, the river is augmented by water pumped from Lake Erie
by BRIG.
WATERSHED
Geography and Beneficial Uses
The watershed of the Buffalo River has a drainage area of 446 square miles and is fed by
three tributaries: Cazenovia Creek, Buffalo Creek, and Cayuga Creek.
Wastewater Discharges: Industrial, Municipal
Our society is still dependent on waterbodies as receptacles for treated industrial and
municipal waste. Cayuga, Buffalo, and Cazenovia Creeks receive treated discharges from a
number of industries and municipal treatment plants, as well as sewer system overflows. The
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20 Areas of Concern
quality of the Buffalo River is influenced by the flow from these tributaries.
Land Uses: Agricultural, Woodland, Residential
Farmland and wooded areas dominate the upland areas of Cayuga Creek and the land
areas adjacent to Buffalo Creek and Cazenovia Creek. Several park and recreational areas are
located along these waterways.
The lower reaches of Cayuga Creek pass through residential communities, as do parts of
Buffalo and Cazenovia Creeks. Along their way to the Buffalo River, these creeks receive
runoff from agricultural, suburban, and urban lands that contains sediments and pollutants
picked up from city street, farms, and from rain and snow. The magnitude of this pollutant
load and its effect on the Buffalo River are not known.
Recreation
The Buffalo River drainage basin supports a variety of fish habitats. Conditions range from
brook trout habitat in some upper streams to warm water species habitat in the lower, urban
areas. To enhance recreational opportunity, DEC stocks trout and pan fish. Salmon, black
bass, and northern pike are among the many species found in the Buffalo River and its
tributaries.
Watershed Hydrology and Current Water Quality
The three major tributaries of the Buffalo River are generally fast-flowing streams with
many rapids and low waterfalls that serve to aerate the water. Water quality monitoring
stations on the three tributaries show high water quality. Comparison with Class A standards
(the best use is classified as drinking water) indicates that the three tributaries meet the
established standards for all conventional parameters and metals except iron. In addition,
analysis of volatile organic compounds in 1987 revealed virtually no volatile organics, further
indicating a high quality of water in these streams.
THE RAP GOALS AND THE PLANNING PROCESS
The goals for remediation were identified at the beginning of the process jointly by DEC
and BRCC.
Short Term Goal
The short-term goal of the Buffalo River RAP is to restore and maintain the chemical,
physical, and biological integrity of the Buffalo River ecosystem in accordance with the
GLWQA. To meet this goal, this plan takes steps toward the restoration of water quality
which provides for propagation of fish, shellfish, and wildlife, and for recreation in and on the
water, consistent with state law, rules, and regulations as they continue to evolve.
This goal is called "short-term" because, given a funding commitment, it could likely be
accomplished within 15 years.
Long Term Goal
The long-term goal is to eliminate the discharge of pollutants to the Buffalo River. This
includes, but goes beyond, the GLWQA policy of the virtual elimination of discharges of
persistent toxic substances.
The immediate intent of this RAP is to address the short-term goal. As remedial action
moves toward the short-tenn goal, the long-term goal will also be approached. In addition, the
various statewide program activities driving New York State toward pollution elimination, such
as technology-based discharge permit limits, will continue to operate. Because these are
statewide activities, the Buffalo River RAP includes them in the plan by reference only. The
RAP focuses on the immediate objective - attainment of the short-term goal, through actions
specific to the Buffalo River.
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J.C. McMahon 21
Ways of Determining if the Short-Term Goal is Being Met
NYS Stream Classification
Impairments to the short-term goal are ultimately determined by criteria derived from the
NYS stream classification system, which classifies every waterbody in New York State
according to the public's desired "best use" of the water resource. The classification takes into
account such factors as the character of bordering lands, stream flow, water quality, and
present, past, and desired future uses of the water, after a formal public participation process,
including public hearings, DEC assigns to each fresh surface waterbody one of the following
classifications. Each class includes all the best uses for classes below it.
Class Best Use
AA, A Drinking Water
B Primary Contact Recreation
C Fishing and Fish Propagation
D Fishing
Each designated classification has a set of standards defining the type and quantity of
substances the water can contain and still be used as intended. Classifications are subject to
review every three years. Public input is an important part of this process. The Buffalo River
is currently classified D. Proposals to change that classification are under consideration by
DEC in its statewide review of water classifications. The Buffalo River Citizens' Committee
has requested a change to a B classification.
Great Lakes Water Quality Agreement
The GLWQA (Annex 2) lists 14 impairment indicators to be examined by the RAP process.
These are presented in Table 3.1.1. For the Buffalo River, those indicators that relate to the
best use of fishing (class D) are the ones that are important for determining whether or not
impairments exist. These GLWQA indicators are: restrictions on fish and wildlife consumption,
tainting of fish and wildlife flavor, degradation of fish and wildlife populations, fish tumors or
other deformities, bird or animal deformities or reproduction problems, degradation of benthos,
eutrophication or undesirable algae, degradation of aesthetics, degradation of phytoplankton and
zooplankton populations, and loss of fish and wildlife habitat.
If the waters were classified as B (best use swimming) the additional GLWQA impairment
indicator "beach closings" would be also used. If the waters were classified as a (best use
drinking water supply) then the GLWQA impairment indicator "restrictions on drinking water
consumption, or taste and odor problems" would be used.
TABLE 3.1.1
GREAT LAKES WATER QUALITY AGREEMENT IMPAIRMENT INDICATORS
(i) Restrictions on fish and wildlife consumption;
(ii) Tainting of fish and wildlife flavor;
(iii) Degradation of fish and wildlife populations;
(iv) Fish tumors or other deformities;
(v) Bird or animal deformities or reproduction problems;
(vi) Degradation of benthos;
(vii) Restrictions on dredging activities;
(viii) Eutrophication of undesirable algae;
(ix) Restrictions on drinking water consumption, or taste and odor problems;
(x) Beach closings;
(xi) Degradation of aesthetics;
(xii) Added costs to agriculture or industry;
(xiii) Degradation of phytoplankton and zooplankton populations; and
(xiv) Loss of fish and wildlife habitat.
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22 Areas of Concern
Two GLWQA impairment indicators are anomalous because they do not have any
counterpart in the New York State classification system. These are: restrictions on dredging
activities, and added costs to agriculture or industry.
The RAP addresses all 14 indicators, but the overall impairment is related to the best uses
according to New York State's stream classification system.
RAP STRUCTURE
The process of developing the RAP proceeded as follows:
* Identify goals
* Assess Impairments - The short-term goal is addressed by examining information on
water quality, sediments, and aquatic life that shows whether or not the best uses are
impaired. The 14 specific indicators provided by the GLWQA helped determine these
impairments. The impairments are determined by the New York State stream
classification system.
* Identify Pollutants or Disturbances - When an impairment indicator suggests an
impairment, all available information is examined to determine the cause of the
impairment. In some cases, definite causes cannot be assigned with a high degree of
certainty.
* Identify Sources of Pollutants or Disturbances - The points of entry of pollutants or
the origin of disturbances are determined.
* Describe Remediation Strategy and Commitments - The overall remedial strategy
identifies actions to address the sources of pollutants and disturbances causing
impairments. Where information is not sufficient to recommend remedial action, the
strategy identifies investigations needed to obtain this information.
* Describe Monitoring Program - Measurements and examinations of the ecosystem
reveal whether or not the remedial actions work as planned, and whether or not the
indicators of use impairment show recovery.
* Describe Tracking - Progress reports and periodic RAP updates, both with
participation of the concerned public, provide a process for tracking plan
implementation.
Impairments, Causes and Sources
The Buffalo River and its sediments have been polluted by past industrial and municipal
discharge and disposal of waste. Fishing and survival of aquatic life within the Area of
Concern have been impaired by PCBs, chlordane, and polynuclear aromatic hydrocarbons
(PAHs). Fish and wildlife habitats have been degraded by navigational dredging of the river
and by bulkheading and other alterations of the shoreline. Low dissolved oxygen and DDT are
likely causes of aquatic life degradation, but they have not yet been definitely established as
such. In addition, metals and cyanides in the sediment prevent open lake disposal of bottom
sediments dredged from the river.
Contaminated bottom sediments are the one certain source of pollutants causing
impairments. Other sources have been identified as potential sources because the pollutants
causing impairments are known to exist at these locations, but the link between the source and
the impairment has not been clearly established. The potential sources include inactive
hazardous waste sites, combined sewer overflows, and other point and nonpoint sources of
pollution. A summary of impairments, causes and sources is shown in Table 3.1.2.
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J.C. McMahon 23
Remedial Objectives and Recommendations
A comprehensive and focused strategy has been developed to:
remediate the bottom sediments;
establish a river monitoring program that will determine whether potential sources
contribute to impairments;
continue the on-going programs that remediate inactive hazardous waste sites, control
point source discharges, and manage nonpoint sources; and
improve fish and wildlife habitat.
The recommended program is:
Remediate Bottom Sediments
Objective:
Correct the impairments to the Buffalo River's fishery and aquatic life caused by
contaminated sediments.
Recommendation:
1. Develop a model of sediment flow and deposition in the Buffalo River in order to
determine the potential for armoring layers to be established over the
contaminated sediments in certain sections of the river.
2. Develop sediment criteria that will allow decisions to be made about which
particular bottom sediments are causing impairment of the fishery and aquatic
life.
3. Assess the river sediments based on criteria to determine specific areas of the
river where remedial work is needed.
4. Evaluate removal/armoring alternatives and then carry out appropriate remedial
work.
Improve Stream Quality Monitoring
Objective :
Ensure that all sources have been addressed in the remedial action plan.
Recommendation:
1. Establish an automated sampling station on the Buffalo River so that the
amounts of contaminants of concern can be accurately determined.
2. Develop models to relate amounts of contaminants in the river to their potential
for harming fish or aquatic life.
Objective:
Determine whether low dissolved oxygen in the Buffalo River is likely to impair the
fishery.
Recommendation:
Carry out an intensive dissolved oxygen study.
Remediate Inactive Hazardous Waste Sites
Objective:
Prevent inactive hazardous waste sites from contributing contaminants to the river.
Recommendation:
Continue the on-going program for remedial work in the Buffalo River drainage area
with particular attention to protecting the Buffalo River itself.
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24 Areas of Concern
Remediate Other Nonpoint Sources as Necessary
Objective:
Prevent the nonpoint sources from adversely affecting the river. [Nonpoint sources are
sources that do not discharge to the river at well-defined points such as through a
pipe.]
Recommendation:
1. Use stream water quality monitoring to determine whether or not these sources
are making a significant contribution to the amount of pollutants in the river.
2. If nonpoint sources are important, determine which ones require remedial action.
3. Select and carry out appropriate control or remedial actions.
Maintain Controls on Municipal and Industrial Wastewater Facilities
Objective:
Insure that municipal and industrial point sources do not significantly contribute to
impairment of the fishery or aquatic life. [Point sources are sources that discharge to
the river at well-defined points, such as through a pipe.]
Recommendation:
1. Renew permits, as they expire, incorporating current technology and water
quality based limits.
2. Carry out monitoring of industrial and municipal discharges and compliance or
enforcement actions as needed.
Improve Combined Sewer Overflow Systems
Objective:
Insure that combined sewer overflows do not significantly contribute to impairment of
the fishery or aquatic life. [Combined sewer overflows are used to relieve the flow to
sewage treatment plants during storms when surface runoff would cause the flow in the
sewers to exceed the capacity of the system.]
Recommendation:
1. Carry out system modeling to determine where improvements can be made to
increase flow within the system and minimize overflow.
2. Design and carry out improvements as necessary.
Remediate Other Point Sources as Necessary
Objective:
Insure that other point sources do not significantly contribute to impairment of the
fishery or aquatic life.
Recommendation:
1. If stream water quality shows that other point sources are likely to be a
problem, then identify these sources.
2. Design and carry out remedial work as required.
Restore Fish and Wildlife Habitat
Objective:
Improve fish and wildlife habitat in and along the river.
Recommendation:
1. Carry out an assessment of habitat conditions and the potential for improvement
in the Area of Concern.
2. Develop a habitat improvement plan.
3. Acquire the necessary land.
4. Design and carry out specific habitat improvement projects.
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J.C. McMahon 25
Commitments and Future Actions
The Department of Environmental Conservation has committed to a number of initial
actions in this plan where funding is available. All of these initial actions are to be completed
in 1990. As further funding becomes available, further commitments can be made. DEC has
made commitments for specific actions to begin the remediation strategy:
REMEDIAL ACTION COMMITMENTS
A. Stream Water Quality Monitoring
1. Establish a flow-activated sampling station
DEC will have a station in place in 1990 that will allow sample collection to be
correlated with flow, so that loadings of chemicals transported by the river can be
measured. The next step will be to collect water quality samples from this station and
in the upper basin and compare the results to determine loadings from sources along
the river.
2. Carry out comprehensive dissolved oxygen measurements on the Buffalo River
DEC will carry out dissolved oxygen measurements on the Buffalo River to determine
whether lack of dissolved oxygen is impairing best uses and, if it is, the causes of
decreased dissolved oxygen. The next step, if needed, will be to propose remedial
actions.
B. Bottom Sediments
1. Develop requirements for a sediment model improvement
DEC will develop the requirements for a model that will allow prediction of scouring
and deposition. The next step will be to contract, develop, and implement the model.
2. Develop methods to determine sediment criteria
DEC will urge EPA to develop national sediment criteria. Criteria should relate
directly to environmental effects of sediment so decisions can be made on the need for
remedial work. The next step will be to apply the criteria to the sediments in the
Buffalo River in order to map the portions of the river that are contributing to use
impairments.
C. Inactive Hazardous Waste Sites
1. Conduct Phase I site investigations
DEC will continue Phase I investigations for each site in the Buffalo River Basin. All
Phase I studies will be completed in 1990. The next step will be to conduct phase II
investigations.
2. Conduct Phase II site investigations
DEC will conduct nine Phase II investigations. The next step will be to prepare and
conduct Remedial Investigation/Feasibility Studies at these sites when required.
3. Conduct Remedial Investigation/Feasibility Studies (RI/FS)
DEC will conduct two RI/FS at hazardous waste sites. These studies will be completed
in 1990. The next step will be to design remedial measures at these sites.
D. Municipal and Industrial Wastewater Facilities
Continue discharge permit monitoring
DEC will continue this ongoing program for all permitted discharges. Permits will be
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26 Areas of Concern
reissued every five years based on current technology requirements and water quality
standards.
E. Combined Sewer Overflows
Evaluate the combined sewer model
BSA is responsible under its State Pollutant Discharge Elimination System permit for
developing and evaluating the model of their CSO system. This work is underway and is
expected to be completed in 1990. The next step will be to use the model to simulate
alternatives for minimizing overflows. Then, remedial measures will be planned based on
the model simulation results.
F. Fish and Wildlife Habitat
Develop a plan for assessment of habitat conditions
DEC will develop a plan for the assessment of habitat conditions and improvement potential
by March, 1990. The next step will be to carry out the assessment according to the plan.
A continuing process, based on annual status reports and workplans, has been established
for reporting on remedial progress, for making commitments as funding becomes available, and
for revising the remedial action plan as new information develops.
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J.C. McMahon
27
TABLE 3.1.2
Summary of Impairments, Causes and Sources
Impairments and
No. Impairment Indicators Impairments
Likely Causes Known Sources Potential Sources
1. Restrictions of fish
and wildlife consumption
Yes
Polychlorinated Bottom
biphenyls Sediments
Chlordane
2.
3.
4.
5.
6.
7.
Tainting of fish and
wildlife flavor
Degradation of fish
& wildlife populations
Fish tumors and other
deformities
Bird or animal
deformities or
reproduction
Degradation of benthos
Restrictions on dredging
activities
Likely
Likely
Yes
Likely
Yes
Yes
Polynuclear
aromatic
hydrocarbons
Low dissolved
oxygen1
Polynuclear
aromatic
hydrocarbons
Polychlorinated
biphenyls
DDT and
metabolites
None Identified
Metals and
cyanides
Bottom
sediments
Bottom
sediments
Bottom
sediments
Not applicable
Bottom
sediments
8. Eutrophication or No
undesirable algae
9. Restrictions on drinking No
water consumption or taste
and odor problems
10. Beach closings
11. Degradation of aesthetics
applicable
12. Added costs to agriculture
or applicable industry
13. Degradation of
phytoplankton & applicable
zooplankton population
14. Loss of fish and wildlife Yes
habitat
N/A
N/A
N/A
N/A
Inactive hazardous waste sites
Bottom" sediments
Inactive hazardous waste sites
Combined sewer overflows
Bottom sediments
Inactive hazardous waste sites
Combined Sewer overflows
Other point sources
Other nonpoint sources
Inactive hazardous waste sites
Combined sewer overflows
Inactive hazardous waste sites
Bottom sediments
Not applicable (N/A)
Inactive hazardous waste sites
Combined sewer overflows
Other nonpoint sites
Other point sites
N/A
N/A
No
No
No
No
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
Physical Bulkheading
disturbances Dredging
Steep bank slopes
River channelization is also a potential factor
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28
Areas of Concern
Buffalo River
Buffalo1 >rea of Concern Map
Figure 3.1.1. Buffalo river area of concern location map
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P. Sanders 29
3.2 Fields Brook Superfund Site/Ashtabula River Area of Concern
Peter Sanders
U.S. EPA
230 South Dearborn Street
Chicago, niinois 60604
The Ashtabula River and Fields Brook are located in extreme northeast Ohio, in Ashtabula
County, approximately 55 miles east of Cleveland, Ohio (Figure 3.2.1). The Ashtabula River
drainage basin covers an area of approximately 137 square miles. The drainage basin is
predominantly rural and agricultural, with the city of Ashtabula as the only significant
urbanized area. The major tributaries include Fields Brook, Hubbard Run, and Ashtabula
Creek. Most of the industrial development is concentrated around Fields Brook.
Fields Brook drains a 5.6 mile watershed (defined as the Superfund Site), including areas of
Ashtabula Township and the City of Ashtabula (Figure 3.2.2). The brook flows westerly
through an industrial area that is considered one of the largest and most diversified
concentrations of chemical plants in Ohio, then through a residential area in the City of
Ashtabula, to its confluence with the Ashtabula River. The Ashtabula River empties into Lake
Erie about 8,000 feet downstream of its confluence with Fields Brook.
Industrial sources have contaminated the sediment in Fields Brook with a variety of organic
and heavy metal pollutants (Table 3.2.1), consisting of numerous chlorinated compounds
including polychlorinated biphenyls, hexachlorobenzene, hexachlorobutadiene, 1,1,2,2,-
trichloroethane, and tetrachloroethene and inorganics including mercury, zinc, arsenic,
chromium, cadmium, and lead.
The Fields Brook site was included on the October 23, 1981 Interim Priority List and then
placed on the first National Priorities List on September 8, 1983. In March of 1985 the U.S.
EPA published a Remedial Investigation (RI) Report for the site and July of 1986 published the
Feasibility Study (FS) describing the remedial alternatives considered for site cleanup. A
Record of Decision (ROD) was signed by the U.S. EPA on September 30, 1986, which described
the selected alternative for the Sediment Operable Unit, which consisted of excavation of
contaminated sediments from the brook, temporary storage and dewatering and the thermal
treatment of a portion, approximately 16,000 cubic yards, and the solidification and landfilling
of the remainder, approximately 36,000 cubic yards, and subsequent water treatment. The
volume of the material to be thermally treated verses that which will be solidified and
landfilled is based on three factors 1) mobility of contaminants, 2) toxicity and concentration
and 3) PCB concentrations. The partition coefficient (K.J of compounds were considered when
determining mobility and the sediment ingestion rate represents a factor that provides a means
of quantitative accounting of both toxicity and concentration. A plot of volume of sediments
exceeding the 10"6 risk guideline verses the mobility was developed. It was determined that for
locations that have compounds with K,,,. values lower than 2,400 ml/g and where sediment
ingestion risk associated with the presence of these compounds is greater than the 10s level,
these sediments would be thermally treated. In addition, sediments containing greater than 50
parts per million PCB will be thermally treated. The ROD also proposed two subsequent
activities, including a RI/FS to identify any ongoing sources of contamination to Fields Brook
and a study to address the contamination in Ashtabula River. On March 22, 1989, the U.S.
EPA issued a Unilateral Administrative order to nineteen Potentially Responsible Parties
(PRPs) to perform the Sediment Operable Unit Remedial Design activities and the RI/FS to
identify sources of contamination. To date six PRPs have agreed to comply with this order.
On September 26, 1989, the U.S. EPA and Ohio EPA and five PRPs signed a Consent Order
for the PRPs to perform an investigative study of the Ashtabula River. In addition to the
objectives outlined under the Superfund ROD, the River Investigation was also conducted to
generate data to be used by the U.S. Army Corps of Engineers (COE) to design a dredging
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30 Areas of Concern
program for the river.
The Ashtabula River Investigation, which began in late 1989, includes collection and
analysis of sediment, water and fish samples. To date, only results from the sediment
sampling have been made available to the U.S. EPA for review. Samples were collected at a
total of 115 locations, consisting of two locations in Lake Erie, two locations in Ashtabula
Harbor, 103 locations along the main stem of the river and eight locations off the main stem of
the river. Samples were collected from sediments in the river using a boat mounted vibrocore
rig and in the harbor and lake using a Ponar dredge where vibrocoring was impractical or
unsuccessful. Over 450 sediment samples were analyzed, results for compounds included in the
"U.S. EPA Guidelines to Classify Sediments From Great Lakes Harbors" are summarized in
Table 3.2.2. In general, contamination in the river sediments is greatest below -8 Lake Erie
low water datum (LWD; elevation 568.6 feet above MSL at Father Point, Quebec). The COE
has tentatively developed a plan to dredge material from the navigation channel to a depth of -
6 LWD with an estimated volume of 18,000 cubic yards. This material will be disposed of in a
confined disposal area. It has been estimated that nearly 500,000 cubic yards of sediment
would have to be dredged from the Ashtabula River Area of Concern for proper clean up.
Work has begun at the Fields Brook Site on Phase I of the Source Control RI. This work
involves characterization of the regional ground water basin including piezometer installation,
geophysical surveys, soil borings and determination of ground water recharge and discharge
reaches; soil gas surveys to help determine the existence and extent of volatile organic
compounds (VOCs) in the soil ground water; and industrial outfall sampling, including both dry
and wet weather sampling. After Phase I is completed objectives of the RI will concentrate on
property or source-specific investigations. Upon completion of the RI, a FS will be carried out
to identify potential treatment technologies, pre-screen these technologies and assemble
alternatives for detailed analysis which will result in a determination by the U.S. EPA and
OEPA of a recommended alternative.
The Sediment Operable Unit design investigation will begin soon, this investigation has
been divided into five task investigations (Figure 3.2.3) which will culminate in the final
design. The task investigations include:
1. A sediment quantification investigation to better define the volume of sediments to be
handled by thermal treatment or solidification;
2. a thermal treatment design investigation involving several test burns (pilot scale),
evaluation of the characteristics of the ash generated and identification of Applicable or
Relevant and Appropriate Requirements (ARARs) for emissions and residues;
3. a solidification design investigation to develop measurements of treatment effectiveness,
guidelines for performance monitoring, refined estimates of the landfill capacity
requirements and identify ARARs;
4. a sediment dewatering and wastewater treatment design investigation to evaluate the
physical and chemical characteristics of aqueous waste streams that could require
treatment prior to discharge and to determine the relative "dewaterablility" of sediments
that will be needed for thermal treatment and solidification; and
5. a facility design investigation to identify potential sites for the dewatering, solidification,
thermal treatment facility, RCRA-type landfill, and temporary storage facility.
Currently, these five task investigations are scheduled to be completed by early 1992 and
results will be used to develop the final design for the Sediment Operable Unit.
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P. Sanders
31
Table 3.2.1.
Priority Pollutants Found in Sediment at the Fields Brook Site
Volatiles
Base/Neutrals
Benzene (C)
Chlorobenzene
1,1,1-Trichloroethane
1,1,2-Trichloroethane
1,1,2,2-Tetrachloroethane
Chloroform (C)
1,1-Dichloroethene (C) (I)
Trans-l,2-dichloroethene
Ethylbenzene
Methylene Chloride (C) (I)
Tetrachloroethene (C)
Toluene
Trichloroethene (C)
Vinyl Chloride (C)
Acids
2-Chlorophenol
Phenol
Pesticides
Heptachlor (C)
y-Hexachlorocyclohexane (C)
a-Hexachlorocyclohexane (C)
PCB 1016 (C)
PCB 1242 (C
PCB 1248 (C)
PCB 1254 (C)
Metals
Antimony
Arsenic (C)
Beryllium (C) (W)
Cadmium (C) (W)
Chromium (C) (W)
Copper
Cyanide
Lead
Mercury
Nickel (C) (W)
Selenium
Silver
Thallium
Zinc
Acenaphthene
Benzidine (C) (I)
1,2,4-Trichlorobenzene
Hexachlorobenzene (C)
Hexachloroethane (C)
1,2-Dichlorobenzene
1,3-Dichlorobenzene
1,4-Dichlorobenzene
Fluoranthene
Hexachlorobutadiene (C)
Isophorone
Naphthalene
Nitrobenzene
N-nitrosodiphenylamine (C)
Bis(2-ethyl hexyl) phthalate
Butylbenzyl phthalate
Di-n-butyl phthalate
Diethyl phthalate
Dimethyl phthalate
Benzo(a)anthracene
Benzo(a)pyrene (C)
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Chrysene
Acenaphthylene
Anthracene
Benzo(ghi)perylene
Fluorene
Phenanthrene
Dibenzo(a,h)anthracene
Indeno(l,2,3-cd) pyrene
Pyrene
C = Carcinogenic.
W = Carcinogenic based on human occupational exposure.
I = Carcinogenic based on animal inhalation studies.
-------
32
Areas of Concern
Table 322
ARI - Main Stem River Sediment Samples
Selected Parameters - Statistical Data
Presented of Dry Weight Basis
(Locations 12201 through 20502)
COMPOUND NAME UNITS
Arsenic mg/kg
Barium mg/kg
Cadmium mg/kg
Chromium mg/kg
Copper mg/kg
Iron mg/kg
Lead mg/kg
Manganese mg/kg
Mercury mg/kg
Nickel mg/kg
Zinc mg/kg
PCB's mg/kg
NO.OF
SAMP.
129
129
129
129
129
58
129
58
129
129
129
400
NO. OF
DET. SAMP.
129
129
129
129
129
58
129
58
129
129
129
321
AVE.
CONG.
812.92
402.33
2.76
402.81
44.26
30201.37
60.30
491.38
0.96
41.28
209.64
11.85
MIN.
CONG.
4.46
35.39
0.00
12.43
14.42
18441.56
9.89
124.40
0.00
13.61
62.47
0.00
MAX.
CONC.
31.06
2152.00
25.00
5739.91
414.02
48387.10
248.06
2900.43
11.32
142.00
1161.18
660.07
-------
P. Sanders
33
FIELDS SHOOK
tilt-
LAKE ERIE
FIELDS
SHOOK
S II
ICAltlNMlLff
OHIO
Figure 3.2.1. Vicinity map, Fields Brook
-------
34
Areas of Concern
SITE MAP
\_
\
s.*,*
Tf*11*TWWM
n,,,i
| OS TRIBUTARY
OC
I ,
I ,—'
**•
SCALE IN KET
•i«» ^__
o I /1> ity~^
\ £[ r?s ^-vmtM
"r,r: _Lr--<_.x-fv<:"1BUt
rs\^ ' \
/^ A1 \ M**t*o*d
x^» ^*^ \
L-*^ \
DETREX TRIBUTARY
S\
UNNAMED \
TRIBUTARY 22 N
t ROUTE 11 U
UNNAMED / ^ VRimiTARY-^ ?
TRIBUTARY 9'
'-n !
ki
Figure 3.2.2.
Fields Brook site map
-------
P. Sanders
35
SEDIMENT
QUANTIFICATION
DESIGN
INVESTIGATION
DEWATERMQ ft
WA&TEWATER
TREATMENT
DESIGN
INVESTIGATION
THERMAL
TREATMENT
DESIGN
INVESTIGATION
SOLIDIFICATION
DESIGN
MVE8TIGATION
FACILITY
SITING DESIGN
INVESTIGATION
PRELIMINARY
DESIGN
REPORT
Figure 3.2.3. Sediment operable unit investigations
-------
36 Areas of Concern
3.3 Coal Tar Contamination Near Randle Reef, Hamilton Harbor
T.P. Murphy*, H. Brouwer*, M.E. Fox*, E. Nagy*
L. McArdle*, and A. Moller*
861 Lakeshore Rd.
Burlington, Ont. L7R 4A6
*Redeemer College
Ancaster, Ontario, L9G 3N6
Abstract
To support the remedial action plan of Hamilton Harbor, and to determine the extent of
coal tar contamination in a toxic area of the harbor, 81 sediment cores were collected for
chemical and biological study. Approximately 55,000 m3 of sediments bounded by Randle Reef,
pier 15, and Stelco are contaminated with coal tar. The coal tar distribution is variable but
the highest concentrations are near the Stelco outfall pipe. The total concentration of the 16
polynuclear aromatic hydrocarbons (PAHs) in 48,3000 m3 of near-surface sediments exceeds 200
Ig/g. The concentration of PAHs that results in the death of 50% of Daphnia magna and
Hexagenia is less than 244 Ig/g and 329 Ig/g, respectively. Sediments containing more than 89
Ig/g of PAHs suppress at least half of the photoactivity of Photobacterium phosphoreum. The
acute toxicity of the sediments of all of Hamilton Harbor is significantly correlated to the PAH
concentration.
Management Perspective
Recommendations
A. Needing immediate action.
1. Adopt the following cleanup standard; the mean concentration of PAHs in sediments
resulting in the death of 50% of Daphnia, and Hexagenia, and the suppression of 50% of
the photoactivity of Photobacterium (200 Ig/g).
2. Use the best available safety procedures when handling the most contaminated
sediments.
3. Develop a cleanup protocol that includes advanced processing of the most contaminated
sediments, i.e., recycling, pyrolysis, but not a simple CDF.
4. Examine existing MOE data to confirm that industrial PAH discharges into combined
sewers will not continue to result in the formation of contaminated sediments.
5. Expand upon the current limited data set to confirm that PCBs are not a major
contaminant in the sediments of the hot spot.
B. Needing future action.
1. Determine the environmental variables restricting bacterial degradation of the PAHs in
Hamilton Harbor.
-------
T.P. Murphy, H. Brouwer, M£. Fox, E. Nagy, L. McArtle, and A. Moller 37
2. Develop a "finger print" assay to distinguish between coal tar and coal dust.
3. Determine if the black sediments at the northwest corner of Stelco contain high
concentrations of coal dust.
4. Determine the relative contribution of coal tar and coal dust to the elevated PAH
concentrations in the deep basin of Hamilton Harbor.
5. Determine the relative bioavailability of PAHs in coal tar and coal dust.
6. Determine the effect of coal tar and coal dust on the distribution of benthic
invertebrates.
-------
38 Areas of Concern
3.3 Advancement Towards A Remedial Action Plan for the Indiana
Harbor and Canal, the Grand Calumet River, and the Nearshore
Lake Michigan
Robert K. Bunner II
Remedial Action Plan Coordinator
Indiana Department of Environmental Management
105 S. Meridian St.
Indianapolis, Indiana 46225
-Alas- Indiana is making rapid advancement towards a Remedial Action Plan (RAP) for the
Indiana Harbor Canal, the Grand Calumet River, and the nearshore Lake Michigan.
Today, Indiana is preparing Stage One of the RAP. Stage One of the RAP process, in brief,
is an identification of the problem. We have committed to the completion of Stage One on or
before January 1, 1991. As of today, we're on schedule, and we will meet our target date.
Although we have only committed to Stage One of the RAP this year, we are rapidly advancing
towards implementation.
Before I discuss the progress toward implementation, it is important for you to understand
the scope of the problems we are confronted with in this International Area of Concern.
The Indiana Harbor and Canal and the Grand Calumet River are located about 20 miles
southeast of Chicago, Illinois, in the northwestern most part of the State of Indiana. The Area
of Concern is commonly referred to as "The Region'.
This Region produces more steel than any other region of comparable size in the United
States, with five active steel companies. It also contains four oil refineries, six crude oil
pipelines and 18 refined petroleum product companies.
Located within the Region are five Superfund Sites, 56 CERCLA Sites, 425XRCRA Sites, 23
TSDs, 9 hazardous waste landfills or surface impoundments, and 462 registered underground
storage tanks, 150 of which are reported to be leaking.
The Region is currently classified as non attainment of National Ambient Air Quality
Standards for particulate matter, ozone, carbon monoxide and sulfur dioxide.
The Region has major groundwater contamination from the petroleum companies and steel
industries. Often, because of the regions high water table, contaminated groundwater becomes
surface water and thus, causes large oil slicks to appear in the river and harbor.
Regarding surface waters, during the last 20 years considerable improvement has been
noted in the water quality of the river and the harbor. In the early 1960s a TV documentary
about the river and the harbor was entitled "Too Thick to Navigate, too Thin to Cultivate".
The documentary described how the river often caught fire because of the thick layer of
petroleum on the surface. Until the 1970's, not even algae lived in the River.
Although water quality has improved, there is much to be done. It is estimated that each
year 11 billion gallons of untreated wastewater enters the river and harbor through combined
sewer overflows.
Fish communities in the river and harbor are depressed. A combination of lack of food
resources, low dissolved oxygen, and toxic stress have resulted in a lack of a stable resident
fish community. The quality of biological habitat is poor. The aquatic community is adversely
impacted by both organic pollution and toxic stress. Water quality monitoring data have shown
problems in these waters with several parameters including ammonia, dissolved oxygen, total
phosphorus, chlorides, fluorides, sulfates, oil and grease, bacteria, cyanide, iron, lead, copper,
mercury and PCBs.
The 1990 Fish Advisory states that no fish should be eaten from the waters of the Grand
Calumet River and the Indiana Harbor Canal.
-------
RJL Bunner 39
An important environmental concern is a significant accumulation of contaminated
sediments in the river and harbor. Today a three-mile footprint of contaminated sediment
stretches into Lake Michigan from the Indiana Harbor. Infrared photos show water intake
pipes for the cities of Hammond, Whiting and East Chicago are within 1/2 mile of the
sediments. This means that there exists a potential threat to the drinking water supplies of
approximately 291,000 area residents.
The U.S. Army Corps of Engineers estimates that the cost of removing and treating the
sediments in the Harbor alone could be as much as one billion dollars. The cost of removing
and treating the sediments in the Grand Calumet River could be another billion dollars.
The cost of dredging and storing the sediments is estimated to cost much less...about 127
million dollars. The issue of where to dispose of the sediments lingers.
The magnitude of the environmental problems in the Region is staggering. But now, let's
look at what the new administration in Indiana is doing to address the many problems:
Water Quality Standards
Very significant to the overall success of the RAP is the adoption of the new Water Quality
Standards. Previously, the river and harbor were designated for 'industrial' use. A few
months ago, Indiana Governor Evan Bayh, signed into law the most stringent water quality
standards in the history of the State. The new standards upgrade the designated use of the
river and harbor to 'whole body contact recreation' waters. Although it will be several years
before the Indiana Harbor Canal and the Grand Calumet River become safe for whole body
contact recreation, the adoption of the standards provide the legal frame-work for repairing the
damage that one hundred years of industrialization has done to these waterways.
Control of Air Toxics
Indiana is currently holding public meetings throughout the State seeking public
participation as the Department moves toward the adoption of rules to control hazardous air
pollutants. These new rules will include provisions to address deposition of air toxics into
aquatic ecosystems, such as the Great Lakes.
Indiana intends to enact rules after reauthorization of the Clean Air Act in 1990 or after it
is clear that Congress will not reauthorize the Clean Air Act this year.
New Office - More Staff
The Indiana Department of Environmental Management will soon open a new office in the
heart of the Area of Concern. (In the past, it was necessary for department staff to travel
about 160 miles to the area.) The new office will be complimented by the staffing of 27
environmental engineers and scientists. Seventeen of those positions will be new. The new
office looks out over beautiful Lake Michigan with only the smoke stacks of U.S. Steel
obstructing the view.
Beefed Up Enforcement
Major advancements are being made toward enforcement of criminal and civil
environmental laws. A few months ago, the former Superintendent of the Hammond Sanitary
District agreed to plead guilty to four felony counts for having submitted falsified Discharge
Monitoring Reports to the State. The former operator has now become State's witness as the
investigation begins to broaden. The felony convictions of this wastewater treatment plant
operator would be only the tip of the iceberg as more prosecution of environmental offenders
advances.
As for civil litigation, the State and the U.S. EPA have court actions pending against
almost every major discharger in noncompliance in the Area of Concern.
This summer, the United States Steel Corporation, through a Consent Decree filed in
federal court, agreed to:
-------
40 Areas of Concern
-Spend at least $25 million to upgrade its wastewater treatment equipment and related
facilities.
-Spend $7.5 million to investigate and clean up contaminated sediments on the Grand
Calumet River bottom.
-Pay a $1.6 million civil penalty for past water pollution violations.
- Develop a comprehensive management plan by June 30 to treat coke plant wastewater.
- Design a corrective action plan to reduce the amount of ammonia, cyanide and phenols
in wastewater discharged from the coke plant.
-Improve overall system to collect and treat wastewater from the steel-making process at
the plant.
- Determine the makeup and toxicity of sediments in the riverbed and develop a plan to
remove or contain them by September 1995.
- Design a program to reduce the volume of oil and grease discharged from the steel
plant.
Many other cases are pending before the court and it is expected several dischargers will
soon join in the efforts for remediation.
CARE Committee
For the Remedial Action Plan to be successful, the Plan must come from the community
and the community must have a vision of its success. For that reason, Kathy Prosser, the new
Commissioner of the Indiana Department of Environmental Management appointed 12
community leaders to the new Citizen's Advisory for the Remediation of the Environment
(CARE Committee). The Committee is made up of the three mayors from the Area of Concern,
a senior union leader, a senior chamber of commerce official, a senior professor of a local
university, a senior petroleum company official, a CEO of a major steel corporation, and three
recognized environmental leaders from the community. The CARE Committee is chaired by
Commissioner Kathy Prosser.
The mission of the new CARE Committee is to advise the Indiana Department of
Environmental Management on the matters relating to environmental and recreational
restoration and revitalization of the area in and around the near shore Lake Michigan, the
Indiana Harbor Canal and the Grand Calumet River, specifically by:
1. Representing the interests of key organizations and constituents in the development of
the Remedial Action Plan.
2. Reviewing chapters of the remedial Action Plan as they are developed.
3. Initiating public education programs to:
a. develop widespread recognition of pollution as a cause of poor water quality and
reduced economic and environmental value in the area; and
b. promote a sense of responsibility for restoration of the area of concern,
acceptance of the remedial measures that are necessary to abate pollution
problems, and the motivation to implement these remedial measures.
4. Encouraging and assisting the public in participating in the remedial action planning
-------
Bunner 41
process, including the development of a vision for the Area of Concern, RAP goals,
objectives, remedial measures, and implementation measures.
5. Developing a strategy for implementing the remedial action recommendations in a
deliberate, vigorous, and timely manner and uniting the diverse and necessary interests
that are essential for successful implementation.
In fulfilling these responsibilities, the CARE. Committee is to meet the major objectives of
the Remedial Action Plan, to:
1. Develop an approach to reduce toxics from all significant sources, including in-place
pollutants, to levels that protect human health.
2. Recommend actions needed to reduce nutrient and sediment loadings to the river,
harbor and near shore areas to a level that eliminates unacceptable health risks in the
Area of Concern.
3. Recommend actions to protect and rehabilitate shorelands, improve land management,
provide for compatible recreational and commercial uses, and develop a framework for a
long-term dredge and dredge spoil disposal plan associated with the river, harbor, and
near shore areas.
4. Describe the measures necessary to bring about new and protect existing spawning
areas, reestablish critical aquatic habitats, and reestablish proper species diversity
among fish and other aquatic life.
5. Increase public awareness of the beneficial use potential of the Grand Calumet River,
the Indiana Harbor Canal, and the near shore Lake Michigan; and encourage public
participation in identification of problems and selection of remedial actions.
Conclusion
Do we have all of the solutions yet? No. Are we trying to find solutions? Indeed we are.
Are we making progress? A resounding yes!
There are no easy answers to the many environmental problems we are confronted with in
Northwest Indiana. But, alas, Indiana is rapidly advancing towards a Remedial Action Plan
for the Indiana Harbor Canal and the Grand Calumet River!
-------
42 Areas of Concern
3.5 Saginaw River/Bay AOC
Greg Goudy
Michigan Department of Natural Resources
P.O. Box 30028
Lansing, Michigan 48909
Background
The Saginaw River and Saginaw Bay have been defined as one of 42 Great Lakes Areas of
Concern (AOCs) by the International Joint Commission (IJC) because degraded water quality
conditions impair certain beneficial uses for which these waters are designated. Environmental
programs have produced substantial improvements in Saginaw River and Saginaw Bay water
quality over the past 20 years, but additional efforts are needed to address the remaining problems.
The most effective way of dealing with these issues is to design and implement site-specific
activities that are tailored to the Saginaw Bay area. This would provide a more focused effort than
would be possible solely with statewide or national programs.
Consequently, in July 1986, the Michigan Department of Natural Resources (MDNR) began the
development of a Remedial Action Plan (RAP) for the Saginaw River/Bay AOC. The RAP was
completed two years later in September 1988 with the additional assistance of a wide variety of
local, state and federal groups. The principal participants were the Saginaw Basin Natural
Resources Steering Committee, the East Central Michigan Planning and Development Region, and
the National Wildlife Federation. The RAP is viewed as an iterative document that will be
periodically updated and revised as more data are acquired, remedial measures are implemented,
and environmental conditions improve. Currently, a large number of activities identified in the
RAP are being implemented and it is anticipated that the RAP will be updated following the
completion of these efforts.
Environmental Setting
Saginaw Bay is a large and relatively shallow southwestern extension of Lake Huron located
midway along the eastern shore of Michigan's lower peninsula (Figure 3.5.1). The bay is 26 miles
wide at its mouth along a line drawn between Au Sable Point and Point Aux Barques at the
interface with open Lake Huron. From the midpoint of this transect to the mouth of the Saginaw
River the bay is 52 miles in length. The bay's surface area of 1,143 square miles is roughly 5% of
Lake Huron's total surface area.
The Saginaw Bay shoreline extends for 149 miles and constricts the bay to a width of 13 miles
between Point Lookout and Sand Point, approximately midway along the bay's length. This
constriction, along with a broad shoal area between Charity Island and Sand Point, divides the bay
into inner and outer halves with equal surface areas. The inner bay is much shallower than the
outer bay, having a mean depth of only 15 feet and a maximum depth of 46 feet versus mean and
maximum depths of 48 feet and 132 feet, respectively, for the outer bay. Consequently, the outer
bay contains about 68% of the total bay volume. The total bay volume of 6.8 cubic miles is about
0.8% of Lake Huron's total volume.
The Saginaw Bay watershed consists of 8,709 square miles, which is about 15% of Michigan's
total land area. Twenty-eight rivers, creeks or drains flow directly into Saginaw Bay from three
drainage areas - the East Coastal, West Coastal, and Saginaw River basins. The Saginaw River
basin is the largest of the three and the largest in Michigan, covering 6,276 square miles, which
includes 72% of the total Saginaw Bay watershed. The Saginaw River itself is relatively short,
extending only 22 miles to the south from the southern end of Saginaw Bay. Though short, the
Saginaw River has a large average flow of over 4,000 fP/sec, which is about 75% of the tributary
hydraulic input to Saginaw Bay. Most of the Saginaw River flow originates from its four major
-------
G. Goudy 43
tributaries - the Cass, Flint, Shiawassee and Tittabawassee rivers - with 50% of the flow coming
from the Tittabawassee River. All four rivers converge near the head of the Saginaw River.
Four major urban areas are located within the basin - Flint, Saginaw, Bay City and Midland -
along with 90 additional city or village municipalities. Two of the four major urban areas are
located directly on the Saginaw River, Bay City at its mouth and Saginaw at the head. Midland
and Flint are also located in the Saginaw River watershed on the Tittabawassee and Hint rivers,
respectively.
The physical boundaries of the Saginaw River/Bay Area of Concern are defined as extending
from the head of the Saginaw River, at the confluence of the Shiawassee and Tittabawassee rivers
upstream of Saginaw, to its mouth, and all of Saginaw Bay out to its interface with open Lake
Huron at the imaginary line drawn between Au Sable Point and Point Aux Barques. Areas outside
these physical boundaries, but within the Saginaw Bay basin, are considered in the RAP if they are
sources of contaminant materials delivered to the Saginaw River and/or Saginaw Bay.
Environmental Concerns
Saginaw Bay is an important and unique ecological, economic and recreational resource to the
state of Michigan. Water drawn from the bay is used as a source of drinking water for over
300,000 people, and for industrial water supply to an extensive industrial infrastructure. The
shallow, nutrient-rich waters support extensive coastal wetland areas, which provide important
spawning, nursery and feeding areas for many of the over 90 species of fish reported from
Saginaw Bay. The wetlands also provide important habitat to many waterfowl as the bay is
located on a major migratory flyway. The bay is used for extensive recreational boating and
commercial navigation. Sport and commercial fishing are important activities with sport fishing
taking place year-round, drawing anglers from other states and throughout Michigan. Saginaw Bay
sport fishing generally accounts for over 60% of the total Lake Huron catch in Michigan waters.
The Saginaw Bay shoreline provides important recreational opportunities for swimming, picnicking,
hiking and bird watching. Finally, the bay is important for its aesthetic qualities.
Unfortunately, past waste disposal practices and poor land use activities have degraded water
quality of the Saginaw River and Saginaw Bay. Anthropogenic inputs to Saginaw Bay have been
dominated by agriculture, which is the most extensive single category of land use in the watershed,
in the rural areas of the basin, and by industrial and municipal wastewater discharges from urban
areas.
Industry is quite diversified in the Saginaw Bay basin due to a wide range of natural resources,
a well developed transportation network, and the early establishment of automobile manufacturing
and related primary industries. The transportation equipment industry remains the largest
employer in the basin and is located almost entirely within the Saginaw River watershed cities of
Bay City, Saginaw and Flint. Other large industries include fabricated and primary metals,
nonelectric machinery, chemicals, electronic equipment, and food processing.
Three major water quality issues have been identified as causing degraded environmental
conditions and impairing designated uses in the Saginaw River/Bay system and these are cultural
eutrophication, bacterial contamination, and toxic material contamination. Excess nutrients in
Saginaw Bay have created eutrophic conditions with nuisance population levels of blue-green algae
which have caused taste and odor problems in public drinking water supplies at the point of water
intake. Eutrophication in the Saginaw River has also contributed to low dissolved oxygen levels in
the river. Combined sewer overflows in the city of Saginaw during wet weather events have
resulted in elevated bacterial counts in the Saginaw River and the issuance of public health
warnings on the river. Fish tissue contamination by PCB in Saginaw Bay fish, and by both PCB
and dioxin (2,3,7,8-TCDD) in Saginaw River fish, has resulted in the issuance of public health fish
consumption advisories.
The goal of the Saginaw River/Bay RAP is to restore all designated uses that are presently
impaired because of degraded water quality conditions. This goal is expressed in terms of three
specific objectives. The first is to reduce toxic material levels in fish tissue to the point where
public health fish consumption advisories are no longer needed for any fish species in the AOC.
Presently, there is an advisory warning against the consumption of carp and catfish in both the
Saginaw River and Saginaw Bay. The advisory also suggests that people restrict their consumption
of all game fish species in the Saginaw River, and limit consumption of lake, brown and rainbow
-------
44 Areas of Concern
trout in the bay. There are no advisories for Saginaw Bay on the two principal sport fish species,
these being walleye and yellow perch.
The second objective is to reduce toxic material levels in ambient water throughout the AOC to
those of Michigan's water quality standards. This is an ambitious and long-term objective for
certain toxicants such as PCB. For instance, right now Michigan's PCB water quality standard is 20
ppq, which is not only below the current analytical level of detection, but is also below PCB levels
presently measured in the open waters of the Great Lakes and in rain water. Ambient water PCB
concentrations measured at the mouth of the Saginaw River in 1989 were approximately 17-18 ppt
and remained elevated relative to upstream and bay water PCB concentrations that were on the
order of 3-5 ppt.
The third objective is to reduce eutrophication in Saginaw Bay to a level where the bay will
support a balanced mesotrophic biological community. Doing so should eliminate the nuisance
levels of blue-green algae populations, which are the source of the taste and odor problems in
drinking water supplies drawn from the bay. It may also allow the reestablishment of the
Hexagenia Umbata mayfly population, thereby providing important forage for bay fish populations.
Toxic Materials
Extensive efforts have been made to reduce the discharge of toxic materials to the Saginaw
River/Bay AOC. For example, Michigan has made great progress in stopping the discharge of
PCBs and has a goal of eliminating all point source discharges. Presently there are only three
known remaining point source discharges of PCBs in the Saginaw Bay watershed and these
contributed at a total of only 4.4 kg of PCBs during 1987. This is a relatively small amount — less
than 9% of the 53 kg/yr estimated as being contributed by atmospheric deposition in 1980, and
less than 1% of the 458 kg/yr load calculated at the mouth of the Saginaw River in 1980.
Nevertheless, fish consumption advisories remain in effect for the Saginaw AOC. Additionally,
recent studies suggest that toxic contaminants may be impacting the reproductive success of some
piscivorous birds. Since ambient water concentrations of toxic materials are quite low, it is thought
that sediments in the AOC, which were contaminated by historical discharges, may be acting as a
source of toxic materials to the aquatic food chain. Sediments in both the Saginaw River and
Saginaw Bay have elevated levels of PCBs, arsenic, cadmium, chromium, copper, lead, nickel and
zinc. Additionally, Saginaw River sediments may have elevated levels of dioxins and furans.
Sediments in the Saginaw River are most contaminated in, and immediately downstream of, the
two major urban centers of Saginaw and Bay City (Figure 3.5.2). The most heavily PCB
contaminated area is just downstream of the Bay City WWTP. In 1980, surficial sediment PCB
concentrations averaged about 10 ppm with a maximum of 23 ppm. The contamination covered
the entire width of the river in contaminated areas including the dredged navigation channel.
Contamination also extended somewhat upstream of source discharge points as a result of reverse
flow conditions, which can occur in the low gradient (approximately 1 inch/mile) Saginaw River
because of wind induced seiche conditions in Saginaw Bay. Reverse flows in the Saginaw River
have been noted all the way up to its confluence with the Tittabawassee and Shiawassee rivers, 22
miles upstream.
The 1980 data also showed that the PCB contamination was even greater in deeper sediments.
In a 25-30 cm deep section of a sediment core collected downstream of the Bay City WWTP, a PCB
concentration of 574 ppm was measured (Figure 3.5.2).
In Saginaw Bay, there is a large sediment depositional zone in the inner bay north of the
Saginaw River mouth. The surficial PCB concentrations in 1980 were generally in the 0.5-1.0 ppm
range. However, this area is so large that the amount of PCB estimated to be in the active
sediment layer in 1980 was 3.7 metric tons.
More recent sediment samples from the Saginaw River and Saginaw Bay were collected in 1988.
Over 200 ponar grab samples and 22 sediment cores were collected to assess present conditions
and the impact of a once-in-100 year flood, which occurred in the Saginaw River in 1986. The
final laboratory analytical results were just reported in April 1990 and consequently, data
interpretation has not yet been completed. However, initial data inspection indicates the average
surficial PCB concentrations have decreased about one order of magnitude since 1980 in both the
river and the bay.
The highest surficial PCB concentration observed in the Saginaw River was 4 ppm and the
-------
G. Goudy 45
highest observation from a core slice was 18 ppm in a 30-40 cm deep section. In Saginaw Bay, the
highest surficial PCB concentration was 0.5 ppm. Though PCB concentrations in deeper sediments
of Saginaw Bay were typically greater than corresponding surficial concentrations, they never
exceeded 0.8 ppm.
Several metals remain above the "heavily polluted" criteria under the U.S. EPA open lake
dredge disposal guidelines. These metals and their maximum 1988 surficial values (ppm) in the
Saginaw River are copper (76), lead (90) and zinc (540). As was the case with PCBs, metal
concentrations were higher in deeper sediments and maximum values in core slices were cadmium
(6), chromium (221), copper (174), lead (144), nickel (79) and zinc (898).
Sediment samples collected from the Saginaw River navigation channel by the U.S. Army Corps
of Engineers (ACOE) in 1988 indicated for the first time that there may be elevated levels of dioxin
in Saginaw River sediments. A previous 1983 survey had detected no dioxin in eight samples
analyzed at detection limits of 10-20 ppt. However, in 1988, duplicate analyses of a sample at a
site near the Zilwaukee Bridge (Interstate 75), just downstream of Saginaw, obtained 2,3,7,8-TCDD
concentrations of 110 ppt and 290 ppt. Two of the other nine samples also had measured
concentrations of 2,3,7,8-TCDD of near 100 ppt. Maximum concentrations of 2,3,7,8-TCDF were
about an order of magnitude higher at 1200-1500 ppt.
Recent sediment sampling in December 1989 and spring 1990, has been conducted in the lower
eight miles of the Saginaw River as part of the U.S. Environmental Protection Agency's Assessment
and Remediation of Contaminated Sediments program. Data are not yet available from this effort,
but it has been reported that many of the cores inspected visually have alternating strata of black
oily material and clay or sand. In general, the substrate of the Saginaw River is sand, with
depositional areas of fine clay and silt.
Because of the contamination of sediments in the navigation channel, sediments that are
dredged from this area are placed in a confined disposal facility (CDF) in Saginaw Bay
approximately one mile from the mouth of the Saginaw River. There has been concern that PCBs
may be leaking out of the CDF and contaminating the environment. In 1987 and 1988, the U.S.
EPA and U.S. ACOE conducted studies to measure any leakage. Results of this project indicated
that PCB leakage was negligible.
Eutrophication
Excessive phosphorus inputs to Saginaw Bay have impacted biological communities by creating
eutrophic conditions that favor nuisance species and inhibit more desirable species. Extensive blue-
green algae blooms created taste and odor problems in drinking water supplies drawn from the
bay as recently as the late 1970s. Of the four drinking water intakes on Saginaw Bay, the
Saginaw-Midland water intake, at Whitestone Point on the northwest side of the bay, accounts for
about 85% of the water drawn from the bay for potable use. In'1974, this intake had taste or odor
problems on 56 days, but by 1980 this had decreased to zero and there have been no reports of
taste and odor problems at this facility since then. However, the Bay City drinking water intake,
located in southern Saginaw Bay near the Saginaw River mouth, still has occasional taste or odor
problems during the summer months.
The decrease in taste and odor problems at the Saginaw-Midland water intake during the 1970s
indicated that the bay was becoming less eutrophic. Indeed, when the Saginaw Bay phytoplankton
community was last surveyed in 1980, blue-green algae population levels had decreased
substantially from those of the mid-70s. There was also a favorable shift in the phytoplankton
community with the almost complete disappearance of the nuisance blue-green algae,
Aphanizomenon and Anacystis. However, there remained a couple of problem areas related to
phosphorus sources and bay circulation.
The flow of water into, around, and out of Saginaw Bay varies with wind direction and
intensity. The most common circulation pattern is for flow to move south along the west side,
around the south end, and north along the east side. There is a shallow shoal area extending
across Saginaw Bay, approximately midway along its length, between Charity Island and Sand
Point. This shoal area, combined with constriction of the shoreline in this area, tends to divide the
Saginaw Bay water mass into two areas — an inner bay and an outer bay. As a result, there is a
secondary, counterclockwise gyre around the inner bay. Consequently, the Saginaw River
discharge tends to move east and north along the east side with some flow circulating back around
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46 Areas of Concern
the inner bay. As a result of these flow patterns, areas that still had populations of nuisance blue-
green algae were the Sebewaing/Wildfowl Bay area on the eastern shore, and along the eastern
shoreline north of Wildfowl Bay.
In addition to blue-green algae populations, other components of the phytoplankton and
zooplankton communities showed decreases in the population sizes of eu trophic organisms in 1980.
These community changes appear to have been the result of decreases in phosphorus loads to
Saginaw Bay, though phosphorus concentrations in bay water remain higher than anywhere else in
Lake Huron and, when last surveyed, the benthic macroinvertebrate community was composed
primarily of pollution tolerant forms such as the aquatic worms Limnodrilus and midges Chironomus.
Total phosphorus concentrations in Saginaw River water decreased 40% from 1974 to 1980, and
dropped another 25% from 1980 to 1986. Orthophosphorus concentrations (the bioavailable
fraction) declined even more dramatically, falling 70% by 1980 compared to the mid-70s. This has
resulted in declining phosphorus levels in Saginaw Bay, though not of as great a magnitude as in
the Saginaw River. It is thought that phosphorus concentrations in Saginaw Bay have not fallen
proportionally because of the periodic resuspension of bay sediments, and the associated sediment
bound phosphorus, from wind driven resuspension events.
The phosphorus concentration reductions in the Saginaw River have been brought about by
several actions including the 1977 state ban on the use of high phosphate detergents, reductions in
phosphorus discharges from industrial and municipal wastewater treatment plants due to facility
upgrades and better operation, and the implementation of various best management practices by
area agricultural producers. This resulted in a 79% decrease in phosphorus loads to Saginaw Bay
from these sources between 1974 and 1986, decreasing from 800 tonnes to 169 tonnes. Additional
reductions in phosphorus loads to the bay are needed, however, to further reduce eutrophic
conditions. Studies during the early 1980s indicated that roughly 55% of bay phosphorus loads
came from fertilizer runoff from cropland, while 17% originated from other nonpoint sources. This
supports the present phosphorus reduction strategy that includes major nonpoint source control
efforts.
Conclusion
The Saginaw River/Bay Remedial Action Plan describes a variety of actions that are needed to
further address the water quality problems just discussed. The cost of implementing the 101
actions identified in the RAP is estimated to be $170 million over the next ten years. This estimate
does not include any costs for sediment removal or treatment if needed.
The ARCS program is providing important information on the areal extent of sediment problem
areas, the toxicity of sediments, and contaminant bioaccumulation potential. The remediation
techniques being investigated under the ARCS program are of great interest to Saginaw RAP
participants. The potential for bioremediation is of particular interest because of the large areal
extent of the sediment contamination problem, particularly in Saginaw Bay, and the encouraging
recent findings with respect to biological degradation of PCBs in sediments.
-------
G. Goudy
47
ay so •>*
/Httu
- Source Ar*« of Concern
Figure 3.5.1. Location of Ihe Saginaw River/Bay Area of Concern
-------
48
Areas of Concern
Miles from Soginow Boy
T . T . ? . T . T . t . T . 1 . T . .
Gcntrtf
Mokw WWTP
Iron
Coiling
otrol Boy CHy
MotOrt WWTP
PCB Concwlrofton Range* (mgAg)
10-5.0 H &0-KM)[
Figure 3.5.2. Spatial distribution of PCBs in surficial sediments of the Saginaw River
-------
G. Goudy
49
0
04
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o
c.
• 3O
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40
50
PC8 lmg/kg]
20 40 60 80 K
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if **
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aiy WWTP
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PCB (mgAg)
20 40 80 BO 100 120 140
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— ' City WWTP
Figure 3.5.3. Vertical distribution of PCBs in sediments near Bay City WWTP
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50 Areas of Concern
3.6 Sheboygan River and Harbor, Sheboygan, Wisconsin
Bonnie L. Eleder
U.S. EPA
Office of Superfund
230 S. Dearborn
Chicago, Illinois 60604
I. Site Background/History
The SRH site is located near and on the western shore of Lake Michigan, approximately 55
miles north of Milwaukee in the State of Wisconsin. The site includes approximately 14 miles
of the Sheboygan River, and the Sheboygan Harbor, which is about 96 acres in size. In the
1950's, the ACOE began dredging the lower Sheboygan River and Harbor annually, depositing
the sediments in offshore waters of Lake Michigan. Dredging was discontinued in 1969 when
sampling of Harbor sediments revealed them to be contaminated with heavy metals. Due to
routine fish sampling undertaking by the WDNR in which fish were found to have elevated
concentrations of PCBs, U.S. EPA sampled sediments from both the River and Harbor in 1977
and found them to be contaminated with PCBs in concentrations exceeding 50 ppm. As a
result, ACOE plans for a CDF and any further dredging were put on hold due to concerns for
impacting public health and lack of an upland disposal site.
The SRH site was evaluated under the U.S. EPA Hazard Ranking System (HRS) due to the
PCB and heavy metal contamination of sediments and the PCB contamination of the fish.
Based on its HRS score, the SRH site was nominated for inclusion on the final NPL in May
1986, U.S. EPA and Wisconsin Department of Natural Resources signed a Consent Order with
Tecumseh Products Company, one of the three PRPs identified, requiring the Sheboygan Falls-
based company to conduct a Remedial Investigation/Feasibility Study (RI/FS). The contractor
for Tecumseh, Blasland & Bouck Engineers (B&B), began the RI/FS in the Spring of 1986.
II. RI/FS
The objectives of the RI/FS were to determine:
1. the hydraulic characteristics of the river;
2. sediment characteristics and horizontal and vertical distribution of contaminants;
3. sediment mobilization, diffusion and transport phenomena;
4. the affinity of the PCBs and other contaminants for various sediment particle sizes;
5. the level of contamination in the water column.
The RI/FS incorporated a unique approach, for that time. Called a "Phased Approach" to
conduct an RI/FS, the RI incorporated certain FS tasks early on by collecting only the amount
of information sufficient for making decisions concerning the development and screening of
potential remedial alternatives. This would allow for additional investigative efforts to be
performed during an Alternative Specific Remedial Investigation Feasibility Study phase.
These could include pilot studies, bench-scale studies, treatability studies, congener-specific PCB
analysis, biodegradation assessment, etc.
in. RI/ES Results
A draft Remedial Investigation/Enhanced Screening (RI/ES) Report was completed in 1988
and revealed that site sediments are highly contaminated with PCBs and a variety of toxic
metals including chromium, cadmium, lead, mercury, zinc, and nickel. The upper 2 3/4 mile
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B.L. Eleder 51
stretch of the river was determined to be the major source of PCBs to the site, with elevated
PCB concentrations found to be as high as 4500 ppm. Two dams in the Village of Kohler
restrict water flow, thereby causing sediments to drop out. As a result, the contaminated
sediments tend to be confined to this upper segment of the river. The next segment of the
river, from the second Kohler dam to the Pennsylvania Avenue bridge in Sheboygan, conversely
found PCBs to range from non-detect (ND) to less than 20 ppm. For the lower river and
harbor, the lower river (inner harbor) found the next greatest levels of contamination with
PCBs ranging from 0.03 to 0.220 ppm with all concentrations within the top 2 feet less than
21 ppm. The remainder of the harbor, the outer harbor, found PCBs ranging from ND to 1.1
ppm. Note: ND is 0.025 ppm. As to water column, the highest measured total (unfiltered)
PCB concentration was 0.27 ppb under moderate flow conditions (about 200 cfs). Other flow
regimes (low and low-moderate found maximum concentrations of PCBs tended to follow the
pattern of river sediments with the highs of 71 ppm and 30 ppm in the uppermost segment of
the river.
The Agency's review of the Endangerment Assessment concluded that there exist two
exposure scenarios posing unacceptable risks to human health (i.e. the calculated cancer risk
level exceeds 10 E4). These two scenarios are:
1. dermal exposure to river sediments;
2. ingestion of several species of fish and waterfowl.
The Sheboygan River can be easily accessed and is used by the public for a variety of
activities including canoeing, fishing, wading, and hiking along the shoreline. Fish
consumption advisories to not eat fish from the river and harbor have been issued by the
WDNR for over 11 years, while consumption advisories against eating waterfowl caught in the
area have been issued for the past 3 years.
The enhanced Screening (ES) segment identified potential remedial technologies and
constructed potential remedial alternatives. These alternatives were then evaluated and
screened based on effectiveness in reducing the contaminant toxicity, mobility, or volume;
technical feasibility; and administrative feasibility, including potential public acceptance. The
ES segment concluded with a listing of the remaining alternatives including in situ remedial
alternatives; sediment removal, treatment, and disposal alternatives; and the no action
alternative. The ASRI grew out of the RI/ES, specifically to address questions regarding the
feasibility of many of the innovative technologies identified as part of these remedial
alternatives.
IV. ASRI
The purpose of the Alternative Specific Remedial Investigation is to study and evaluate
innovative technologies which may be used in remediating PCB-contaminated sediments found
in the river. The goal of this study is to generate information to help determine an
appropriate course of remedial action at the SRH site. A side benefit to be realized from this
study is the minimization of human health and environmental risk due to the removal of the
most highly contaminated sediments from the river.
The ASRI is a multi-faceted study including:
1. A Pilot Confined Treatment Facility (CTF) to study enhanced natural biodegradation for
treatment of PCB-contaminated sediments removed from the Sheboygan River, and to
test certain design components to evaluate their full-scale feasibility as a remedial
alternative.
2. Evaluation of sediment removal technologies.
3. Evaluation of sediment control devices, i.e. silt curtains and other measures to prevent
and control resuspension of sediments.
4. In situ armoring of low PCB concentration sediments.
5. A monitoring program designed to assess the effectiveness of the removal, armoring and
biodegradation of sediments.
6. Bench-scale studies of removed sediments performed under laboratory conditions,
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52 Areas of Concern
including PCB extraction, chemical fixation, armoring, supplemental biodegradation,
dewatering and physical characterization.
Pilot Confined Treatment Facility
A Confined Treatment Facility (CTF), built on-site on property owned by Tecumseh, is being
utilized to study the effectiveness of enhanced natural biodegradation for treatment of PCBs in
sediments. Sediments with maximum PCB concentrations ranging from 640 ppm to 4500 ppm
will be utilized in the pilot study. Additionally, certain physical components of the CTF are
being studied to evaluate the feasibility of a full-scale CTF, including a series of permeable
treatment walls (i.e. water treatment system).
Preliminary studies by Dr. John F. Brown, Jr., of General Electric's Research and
Development Center, have shown that PCBs in the river and harbor sediments apparently are
being transformed by at least three processes. According to his letter reporting his findings to
B&B, which is in Appendix J of the RI/ES Report, Dr. Brown reports that "...PCBs in the river
and harbor sediments are being transformed by two types of reductive dechlorination
processes... and one type of oxidative biodegradation (process)."
In addition, bench scale biodegradation tests have been in progress at the University of
Michigan. The results of these tests will be used to develop operating parameters for the CTF.
Data will be forthcoming at their completion in the Fall.
The CTF has a capacity for approximately 1500 cubic yards of sediments, which are being
dredged from the upper portion of the river. The CTF is divided into four treatment cells
which will provide different testing environments in which to study the effectiveness of
degrading PCBs by enhanced natural biodegradation.
Two of the treatment cells will receive enhancements, such as a nutrient mixture, a carbon
source, or a surfactant. Other factors may also be controlled, including oxygen and pH. The
types of enhancements, rate of application, and control of other factors will be determined
through the on-going bench-scale biodegradation studies at the University of Michigan.
Biodegradation will be studied under both anaerobic and aerobic conditions. It is expected that
each of the treatment cells will undergo both anaerobic and aerobic degradation cycles. The
two remaining cells will act as control cells where bacterial activity is not enhanced or
controlled.
The CTF is constructed of structural steel sheet piling, 25 feet in length, driven 15 feet into
the ground. Facility dimensions measure approximately 106 feet long by 135 feet wide by 10
feet high. The 14,000 square foot facility has a double liner in each of the cells and
incorporates a leak detection/leachate removal system in-between the two liners. Overflow
protection is also provided for through the use of piping to collect excess water to the
treatment system. A piping system has been placed into each cell for aeration, drainage, or
addition of enhancements. The water treatment system consists of four permeable treatment
walls which will be used to evaluate four different mediums for filtering the water thereby
removing the PCBs. A backup carbon adsorption system will also be utilized to ensure a
discharge that meets effluent requirements.
Sediment Removal
Extensive sampling and analysis of river sediments have identified specific depositional
areas for removal. Dredging has been accomplished through the use of mechanical equipment.
A crane with a modified clamshell bucket was used, from either a land base or from a barge.
The clamshell bucket was modified whereby the joints were sealed to minimize losses of
dredged material to the water column. Operational controls were utilized to reduce
resuspension - through controlling the speed of the bucket through the water column,
maintaining a smooth movement of the bucket, and not dragging the bucket over the dredged
bottom to smooth it out. In addition, sediment control devices consisting of double-lined
siltation control curtains were placed completely around the sediment area prior to dredging to
contain any resuspended sediments. Sediments were placed into a sealed box which was then
transported by barge or truck to the CTF.
To ensure that all prescribed sediments were removed, removal was conducted in two steps.
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B.L. Eleder 53
In step one, the majority of the sediments were dredged to a depth based on study data.
Suspended particles within the siltation control curtains were then allowed to settle at least 24
hours. The second-step dredging operation removed an additional six-inch layer comprising any
settled solids and allowing for a safety factor. Following another period to allow any sediments
to resettle, the area was probed for any remaining sediments and samples were collected for
PCB analysis. Based on the probing and sampling results, either additional material was
removed or the siltation control curtains were removed.
Sediment Armoring
Armoring confines the sediments in place by covering them with successive layers of
materials in order to minimize their resuspension and to minimize or stop the loss of PCBs to
the water column. This pilot study will also assess the effects of armoring on in-place
biodegradation of PCBs.
Approximately 20,000 square feet of sediments will be armored in place. After placement of
siltation control curtains around the sediment area to be armored, a geotextile material is first
placed over the sediment area. A line of sandbags temporarily holds the fabric in position.
The sediment area is then armored with roadbed material consisting of fine- to coarse-grained
material. A second layer of geotextile is then placed over the roadbed material and gabions
are placed around the edges to permanently hold the fabric in place. The sandbags were
previously removed. A layer of cobbles is placed over the geotextile. Finally, a layer of
roadbed material is spread over the gabions. Once any resuspended solids (from the roadbed
material) had settled, the siltation control curtains were removed.
Monitoring
An extensive monitoring program has been established in order to evaluate the effectiveness
of the activities associated with the removal, armoring and biodegradation of PCB-contaminated
sediments. This program has incorporated sampling and analysis of the water column,
sediments, and fish as follows:
1. Baseline sampling and analysis of the water column for PCBs (filtered and unfiltered),
total suspended solids, volatile suspended solids, turbidity;
2. During removal and armoring activities, weekly water column sampling and analysis for
PCBs, TSS;
3. Daily monitoring of the water column during removal and armoring activities for TSS
and turbidity;
4. Sediment and water sampling for PCB analysis within silt curtained area post-dredging;
5. Long-term sampling of resident fish - PCBs and lipid content;
6. Caged fish studies using fathead minnows and tethered clams are being conducted pre-,
during and post- removal/armoring activities - for PCBs and lipid content;
7. Sediment sampling for congener-specific PCBs to evaluate biodegradation in CTF and
under armoring materials.
V. Initial Results
The analytical results for samples of the water column, sediments, native species of fish,
fathead minnows, and clams have been coming in over the past several months, and will
continue over the next 1 to 2 years. Once the data has been reviewed and compiled, it will be
released. The analytical results thus far indicate that there has been no or minimal measured
impact on the water column due to sediment dredging and armoring activities, based on
analysis for TSS, turbidity, and PCBs. Preliminary conclusions show that the use of double-
lined siltation control curtains and operational controls by the crane operator are successful in
controlling and preventing the loss of resuspended sediments and PCBs into the surrounding
water column during these activities. Samples collected after the completion of two rounds of
dredging have shown that mechanical dredging using a modified clamshell bucket has reduced
PCB concentrations as follows: 4500 ppm to 4.9 ppm; 830 ppm to 2.5 ppm; and 1000 ppm to
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54 Areas of Concern
0.49 ppm.
VL Future ASM Activities/Project Schedule
Future activities include completion of the ASRI Pilot Study and the Feasibility Study. The
following lays out the current draft schedule of activities:
ASRI Activities
* Complete dredging of sediment - July 1990
* Complete armoring of sediments - Fall 1990
* Continued monitoring activities - On-going
* Finalization of operational parameters from bench-scale biodegradation studies for
enhancement of PCB biodegradation in CTF pilot Study - Sept. 1990
* Evaluation/monitoring of pilot CTF & armoring - Through Oct. 1991
* ASRI Report - Late 1991
FS Activities
* ARARS Finalization - Sept. - Dec. 1990
* Determination of clean-up standards
* Draft FS Report- March 1992
* Record of Decision - 4th quarter 1992
References
1. Sheboygan River and Harbor Remedial Investigation/Enhanced Screening Report, May 1990.
Prepared by Blasland & Bouck Engineers.
2. Sheboygan River and Harbor Alternative-Specific Remedial Investigation Work Plan/QAPP,
August 1989. Prepared by Blasland & Bouck Engineers.
3. Sheboygan River and Harbor Superfund Site File. U.S. EPA Region V.
-------
4 POLYCHLORINATED BIPHENYLS (PCBs)
4.1 Aerobic Biodegradation of PCBs
Ronald Unterman
Vice President, R&D
Envirogen, Inc.
Lawrenceville, New Jersey 08648
On the spectrum of chemical targets from easiest to most difficult, the polychlorinated
biphenyls are definitely one of the more challenging for bioremediation. Research programs
over the last 18 years have clearly demonstrated that PCBs can be biologically destroyed by
bacteria and fungi. However a continuing development effort will be required to transition
some very promising laboratory results to commercial cleanup technologies.
Unlike easier targets such as gasoline and simple pesticides, the microbes which degrade
PCBs do not utilize them as a source of carbon and energy. These bacteria break down PCBs
in a cometabolic process whereby the organism grows on one substrate, for example, biphenyl,
and then fortuitously degrades the target substrate, in this case PCBs. There is no energy
derived from the breakdown of PCBs and in some cases this process consumes energy.
Therefore, whereas indigenous microbes are generally sufficient to degrade simple targets
because of their selective growth advantage, aerobic technologies to degrade PCBs will probably
require the introduction of exogenous organisms. Another challenge posed by PCBs is their
insolubility with the result that some PCBs are often not bioavailable. Therefore, some of the
technology efforts currently underway are attempting to develop physical and chemical
pretreatment and cotreatment approaches for increasing the bioavailability of this difficult
substrate.
Generally, one can consider two approaches for a PCB bioremediation system. The first
would be an in situ approach whereby bacteria would be introduced directly to the
contaminated soils or sediments and then incubated under conditions to facilitate the
degradation of the target. Alternatively, one can excavate the soils or sediments and treat
them in a soil slurry bioreactor with added microbes and nutrients. For strictly aerobic
biodegradation of PCBs, the latter is the most promising technology for the near term.
However, in situ approaches are a major goal of this technology and in the short term may be
most applicable for anaerobic sediments.
Discoveries over the last several years have now shown that PCBs can be broken down by
both aerobic and anaerobic microbial systems. This paper will discuss solely aerobic
approaches for the biodegradation of PCBs. However, other papers at this meeting will address
the complementary anaerobic technologies.
It is important to keep in mind that PCBs are a complex family and not a single
chemical target. There are 209 different theoretical PCB congeners from mono- through
decachlorobiphenyl. However, all do not exist in the environment. Generally, the more
chlorine atoms on the molecule, the more recalcitrant it is. However, the position of the
chlorine atoms is also a critical factor in the biodegradability of PCBs. For example, 2,5,2',5'-
tetrachlorobiphenyl is readily degradable by Type II bacterial strains, whereas 3,5,3',5'-
tetrachlorobiphenyl and 2,6,2',6'-tetrachlorobiphenyl are not biodegradable to any extent by any
known bacterial species.
55
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56 Polychlorinated Biphenyls
As produced by Monsanto under the trade name Aroclor, each PCB mixture contained
from 30-60 congeners. Aroclors 1242 and 1248 are more easily biodegradable by aerobic
systems than the higher chlorinated mixtures. These two Aroclors contain mostly di-, tri-,
tetra, and some pentachlorobiphenyl. The congeners in Aroclors 1254 and 1260 contain tetra-,
penta-, hexa-, and heptachlorobiphenyls and pose a much greater challenge to aerobic bacteria.
In the short-term, we believe that a strictly aerobic approach to PCB biodegradation will be
limited to the lower Aroclors (1242 and 1248) and probably at concentrations no higher than
1,000 - 5,000 ppm. For the higher Aroclors (1254 and 1260) and for higher concentrations of
the lower Aroclors, a dual anaerobic/aerobic biotreatment system will probably be required.
The initial studies to degrade PCBs from university and industrial labs throughout the
world were all aerobic and the approach taken was generally similar. Soil from PCB
contaminated sites was used to inoculate minimal media containing either biphenyl or
monochlorobiphenyls as sole source of carbon and energy. From these enrichments, many
bacterial strains have now been isolated which can degrade PCBs. These strains, however vary
dramatically in their capabilities to degrade PCBs. Some can only degrade lower chlorinated
congeners such as mono-, di-, and trichlorobiphenyls, whereas some cultures can degrade PCB
congeners as highly chlorinated as tetra-, penta- and hexachlorobiphenyl. It is the use of these
more active strains that will be the basis for commercial PCB bioremediation technologies.
In addition to the differing capabilities of PCB-degrading bacterial strains in terms of the
chlorine content of the ring, another discovery has shown that at least two different types of
bacteria exist which degrade complementary PCB congeners. This can be illustrated by two of
Envirogen's more active strains, one of which (Type I) readily degrades PCB congeners which
are substituted in both para positions (4,4'-chlorobiphenyl family), whereas the strain (Type II)
has the capability of readily degrading PCB congeners with a 2,5-substitution pattern on one
ring. The congener complementary of these two strains forms the basis for the current
development program for a commercial aerobic PCB biotreatment system.
Biochemical studies of these and other strains have now elucidated the biodegradative
pathways for PCBs. This pathway is similar to other aromatic compounds whereby the initial
attack is by a dioxygenase followed by a dehydrogenase and subsequent ring cleavage by
another dioxygenase. The end product from this initial oxidation is generally the chlorinated
benzoic acids. Other microbes in mixed cultures have the capability of further breaking down
chlorobenzoic acids to carbon dioxide and water. The molecular genetics of PCB-degrading
strains is now under investigation and the genes from various of these organisms have been
isolated from several laboratories, including Envirogen. It is the goal of these studies to
develop superior PCB-degrading strains which will express higher levels of PCB-degradative
enzymes. Additionally, the use of genetic engineering will permit us to uncouple biphenyl
metabolism in these strains from PCB biodegradation, thereby allowing these cultures to be
grown on a common inexpensive carbon source and then utilized as biocatalysts to degrade the
target PCBs.
The initial microbiological, biochemical and genetic studies of PCB-degrading strains have
now led to research and development projects with soils and sediments contaminated with
PCBs. This, of course, is the ultimate goal of development programs and what is needed for
addressing problems such as those in the Great Lakes Basin. Several laboratories are
attempting to develop commercial systems for the biotreatment of PCBs on soils and sediments,
however, it has been critical for biodegradation process-modeling research to demonstrate that
bacterial soil decontamination results are unequivocally due to biological activity. One pitfall
that both scientists and regulators must be concerned with is congener depletion in a
"biodegradation" process that is actually due to physical loss of the PCB and not to true
biological degradation. With Aroclor studies, these processes can easily be distinguished,
because biodegradation results in depletion of specific congeners yielding gas chromatographic
(GC) profiles that are distinctly different from those of the original Aroclor mixtures. Physical
depletion, on the other hand, results in uniform depletion of all congeners (e.g., adsorptive loss)
or depletion of lower-chlorinated congeners due to their higher volatility (e.g., evaporative loss).
The production of PCB metabolites is of course another unequivocal method for demonstrating
the biological basis of PCB depletion.
In order to better evaluate results for open-air, aerated, stirring reactors, a model process
was set up to mock a biologically mediated treatment, but whose conditions were adjusted so
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R. Unterman 57
as to preclude biodegradation (study conducted at GE, CRD, Schenectady, NY). A sample of
soil contaminated with Aroclor 1260 (7800 ppm) was used in this study. Argon (instead of air)
was bubbled through a water/soil slurry in a round-bottom flask. A florosil sample tube was
attached to trap PCBs in the argon off-gas as it exited the vessel. Samples were taken
periodically while mixing to ensure a homogeneous sample. The volume in the vessel was
maintained by adding distilled water after each sampling, and the florosil sample tube was
replaced each time a sample was taken. The soil was mixed and purged with argon at room
temperature for 19 days, after which the vessel was disassembled.
Each soil sample and florosil tube was extracted for GC analysis. The remaining soil and
water were removed and pooled for GC analysis. The entire vessel was extracted several times
to remove any PCBs bound to the vessel. These extracts were also pooled for GC analysis.
Upon disassembly, a tar-like substance was observed sticking to the Teflon stirrer. This was
removed and added to the soil and water fraction for PCB analysis.
Neither oxygen nor bacterial inoculum was introduced in this mock process, yet the
analytical results could be mistaken for biodegradation. Although the time-course analysis
indicated greater than 90% PCB depletion, it was clear from the mass balance calculations
(86% PCB recovered) that the aeration and stirring of the soil resulted in the redistribution of
PCBs from soil to the difficult-to-sample locations in the reactor (i.e., glassware, stirrer, and
coalesced droplets of PCB). The GC profiles also demonstrated that the observed depletion was
not due to a biological process, since all GC peaks were depleted proportionally. Therefore,
experiments that purport to show biodegradation of PCBs by quantifying total GC peak areas
must be carefully evaluated. It is for this reason that nonbiodegradable PCB internal
standards should be included wherever possible. If such standards or dead-cell controls cannot
be included, then one must rely on differential congener depletion (or metabolite production) as
evidence for the biological basis of PCB "biodegradation" processes.
Our studies to date are concentrating on utilizing Envirogen's two best Type I and Type
II PCB-degrading strains to develop a strictly aerobic biotreatment process. Similar studies
done at General Electric CR&D, both in the laboratory and eventually in a direct field
application using one Type II microbe, clearly demonstrated that PCBs on soil can be
biodegraded, however, several limiting factors were identified. In the General Electric
experiments, a PCB-contaminated soil containing 500 ppm of an Aroclor 1248-like mixture
could be degraded to the extent of 50 percent PCB destruction in an actively-mixed system in
1-3 days. In a laboratory modeled in situ approach, this extent of degradation was not
achieved until 100 days of incubation. In field studies with the single Type II strain in an in
situ mode, only 25 percent destruction was achieved in approximately 100 days (i.e., one-half
that observed in the laboratory).
From those General Electric studies, several key parameters were identified as necessary
in order to improve the extent of aerobic degradation from 50 to 90 or 99 percent. These
parameters include bioavailability, temperature, and better utilization of microbial strains.
The issue of bioavailability is critical when developing treatment systems for highly
insoluble substrates such as PCBs. Studies are currently underway to develop physical and
chemical pretreatment approaches to facilitate the desorption of PCBs form soils and sediments,
thereby increasing the rate of uptake by bacterial strains. In the short term, we believe that
reactor-based technologies will show more promising results due to the active mixing in a soil
slurry system as opposed to the diffusion limitations of in situ approaches. The bioavailability
of PCBs can also be affected by co-contaminating substrates such as simple hydrocarbon oils.
Studies at Envirogen have shown that a co-contamination oil at a PCB-contaminated site
dramatically reduces the extent of PCB biodegradation by otherwise competent microorganisms.
From this study, and others done at General Electric, it is apparent that PCBs are sequestered
into the oil phase and are therefore not available for the PCB-degrading bacteria.
In terms of the best utilization of microbial strains, we believe that the use of co-cultures
of Type I and Type II bacteria will result in the most extensive degradation of the broadest
range of congeners. As discussed above, Type I strains preferentially degrade double-para
substituted congeners and Type II strains are better able to degrade congeners with 2,5-
substituted rings. Studies utilizing individual versus co-cultures of Type I and Type II strains
have now clearly shown the advantages of using the two complementary strains in concert.
For example, Envirogen strain ENV 307 can degrade 59 percent of Aroclor 1248 in a standard
-------
58 Polychlorinated Biphenyls
30 ppm, 20-hour assay. Envirogen strain ENV 360 can degrade 58 percent. However, when
both strains are utilized together, greater that 70 percent of the PCB is destroyed in these 20-
hour assays.
Soil from a PCB-contaminated location which contains 290 ppm Aroclor 1248 has now
been shown to be biodegradable down to levels of less than 100 ppm utilizing the two
complementary strains. The challenge now is to develop this technology to get greater levels of
destruction. One approach that one can envision is the development of a genetically-engineered
strain which will encode both Type I and Type II PCB degradative pathways in a single
organism.
Ultimately, the development of dual anaerobic/aerobic biotreatment systems will also allow
us to address higher concentrations and higher chlorinated congeners by initially performing an
anaerobic biotreatment step to first remove chlorine atoms from the biphenyl nucleus, thereby
transforming Aroclors to lower chlorinated mono-, di-, and trichlorobiphenyl products. These
are readily degradable by aerobic bacteria.
In summary, a direct aerobic biodegradation treatment technology is the first, short-term
goal for biotreatment of PCB-contaminated soils and sediments. This will be reactor-based and
address Aroclor 1242 and 1248 problems at concentrations under 5,000 ppm. Technology
advances using anaerobic cultures, genetically-engineered strains, and soil pretreatment steps
will in the future extend the capability of PCB biotreatment systems to higher Aroclors and
higher PCB concentrations.
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J.F. Quensen, SJL Boyd, and J.M. Tiedje 59
4.2 Anaerobic Dechlorination and the Bioremediation of PCBs
J. F. Quensen, S. A. Boyd, and J. M. Tiedje
Department of Crop and Soil Sciences
Michigan State University
East Lansing, MI 48824
Anaerobic reductive dechlorination is a newly recognized environmental fate of
polychlorinated biphenyls (PCBs) that is of potential importance in risk assessment, deciding
remediation strategies for contaminated sites, and in developing treatment systems for the
biodegradation of the more heavily chlorinated PCB mixtures (Aroclors). Dechlorination both
reduces the toxicity of a PCB mixture and makes it more aerobically degradable. Thus a
sequential anaerobic/aerobic treatment system has the potential to degrade the more heavily
chlorinated PCB congeners that are resistant to aerobic degradation.
We here review our research findings on PCB dechlorination and further discuss the
implications for bioremediation of PCB contaminated sediment and soil.
First Evidence for Anaerobic PCB Dechlorination
The PCB congener distribution patterns obtained for core samples from the upper Hudson
River first suggested the anaerobic dechlorination of PCBs (2). Aroclor 1242 was known to be
the primary input to this section of the river, and surface sediments showed a congener profile
similar to Aroclor 1242. Deeper and potentially anaerobic sediments, however, showed a
depletion of the more chlorinated congeners and a corresponding increase in mono- and
dichlorobiphenyls. This suggested that anaerobic bacteria might be dechlorinating the PCBs in
the deeper sediments. Microbially mediated reductive dechlorination of other chlorinated
aromatics had at that time been recently demonstrated (13). Similar differences between
congener profiles for sediments and their most probable PCB inputs have since been observed
at other sites (1,3).
Demonstration of Biologically Mediated PCB Dechlorination
We first demonstrated biologically mediated reductive dechlorination of PCBs by adding
low concentrations of pure PCB isomers to PCB-contaminated Hudson River sediments.
Dechlorination was observed, but attempts to obtain dechlorination in the absence of sediments
were not successful. The high levels of putative dechlorination products in these sediments
made it impractical to study the dechlorination of Aroclors by adding them to the contaminated
sediments.
Methods
To study the dechlorination of Aroclors, we therefore developed a method in which
microorganisms from the contaminated sediments were transferred to non-PCB contaminated
sediments (9,10). First serum bottles or Balch tubes of methanogenic "clean" sediments were
prepared and autoclaved. These were then inoculated with supernatant from a anaerobic
slurry of a contaminated sediment. Control bottles or tubes were then autoclaved. An Aroclor
was then added in a small quantity of acetone while flushing with filter sterilized
nitrogen:carbon dioxide (80:20) and the vessels were sealed with Teflon lined rubber stoppers.
Periodically samples were extracted and analyzed for changes in the PCB congener profile by
capillary gas chromatography with an electron capture detector.
-------
60 Polychlorinated Biphenyls
Dechlorination of Aroclor 1242 by Hudson River Microorganisms
Dechlorination of 700 ppm Aroclor 1242 (on a sediment dry weight basis) by
microorganisms eluted from the Hudson River sediments was readily apparent from a visual
inspection of the chromatograms for live samples taken after 16 weeks of incubation (Figure
4.2.1). There was a marked decrease in the peak heights for later eluting, more heavily
chlorinated congeners and an increase in the early eluting mono and dichlorobiphenyls. Closer
inspection revealed that nearly all of the dechlorination occurred from the meta and para
positions (Figure 4.2.2). Little if any ortho dechlorination occurred. This resulted in the
accumulation of mostly 2-chlorobiphenyl (2-CB), 2,2'-CB and/or 2,6-CB (coeluting isomers), and
2,2',6-CB. The detector response for 2-CB is particularly weak; while the peak height for this
congener is small (bottom panel, Figure 4.2.1) this congener actually represented 63% of all
PCBs recovered from the live samples receiving 700 ppm Aroclor 1242 after 16 weeks of
incubation. Most of the other persistent congeners were chlorinated in both ortho and meta
positions, indicating some preference for dechlorination of the para positions over the meta
ones.
Concentration Dependence
The extent of dechlorination observed was concentration dependent (Figure 4.2.2), being
greatest at 700 ppm. At that concentration the average number of meta and para chlorines
decreased from 1.98 to 0.31 after 16 weeks, but decreased to only 1.19 in the 140 ppm
treatment. There was no apparent dechlorination in the 14 ppm treatment.
There are two possible causes for the observed dependence on concentration. It may be
related in part to bioavailability. Higher concentrations in the sediments should result in
higher solution concentrations (4), and it may be that only PCBs in solution are available for
microbial uptake (8). It is also possible that greater population growth of the dechlorinating
organisms occurred during the experiment at the higher PCB concentrations resulting in more
extensive dechlorination. Two subsequent experiments give credence to this interpretation.
Dechlorination activity has been maintained through eight successive transfers in the presence
of 1000 ppm Aroclor 1242. The activity would have been lost if no growth of the
dechlorinating population occurred. In a second experiment, the dechlorination rates at 500
and 5000 ppm of Aroclor 1254 were compared. Initial dechlorination rates (calculated as Cl
released per week) were similar for the first 8 weeks, but between 8 and 16 weeks the rate at
5000 ppm was 10 times higher than at 500 ppm.
Terminal Products
The high accumulation of the ortho-only substituted products in the above experiment
suggested that they were terminal products. The high total recovery did not indicate that there
was significant degradation of the biphenyl structure under our incubation conditions. We
therefore conducted experiments in which only biphenyl, 2-CB, 2,2'-CB, or 2,6-CB was added to
sediment slurries inoculated with Hudson River microorganisms. There was no evidence for
the dechlorination or degradation of any of these compounds during a one year incubation. We
therefore conclude that these are in fact terminal products under our experimental conditions.
Selection for PCB Dechlorinators
There appears to have been selection for PCB dechlorinating microorganisms at several
PCB contaminated sites. Assays for the presence of PCB dechlorinating microorganisms in
PCB-contaminated sediments generally give positive results within 4 weeks and extensive
dechlorination within 12 to 20 weeks. In contrast, assays for the presence of these organisms
in non-PCB contaminated sediments yield at most very modest activity after incubation periods
of 20 weeks or more. We therefore believe that PCB-dechlorinating microorganisms may be
widely distributed at low levels, but they are much more abundant at PCB-contaminated sites
that are also otherwise favorable for their growth.
There may be two related advantages to microorganisms from the dechlorination of PCBs.
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J.F. Quensen, SA. Boyd, and JM. Tiedje 61
The dechlorination process may be serving as a terminal electron sink in anaerobic sediments.
Terminal electron acceptors are usually the limiting factor for microbial growth in anaerobic
habitats so that any microorganism that could use PCBs for this purpose would be at a
selective advantage in such habitats. The second related advantage is that it is possible that
energy can be gained from the dechlorination step itself. The chlorobenzoate dechlorinating
strain DCB-1 can apparently gain energy from dechlorination (5,6}.
Temperature Dependence
Dechlorination assays using Aroclor 1242 and Hudson River microorganisms were
conducted at 12, 25, 37, 45, and 60°C. The greatest dechlorination rate was at 25°C while
about half as much dechlorination occurred at 12°C (Figure 4.2.3). There was no
dechlorination at temperatures of 37°C or above. Such temperature effects are characteristic of
enzyme catalyzed reactions. It is also noteworthy that significant dechlorination can occur at
normal environmental temperatures.
Dechlorination of Other Aroclors by Hudson River Microorganisms
The dechlorination of more heavily chlorinated Aroclors by the Hudson River
microorganisms was also investigated (Figure 4.2.4). Aroclors 1242, 1248, 1254, and 1260
average approximately 3, 4, 5, and 6 chlorines per biphenyl, respectively. Dechlorination of all
of these Aroclors was observed although the dechlorination rate and extent tended to decrease
with increasing degree of chlorination, particularly for Aroclor 1260 (Table 4.2.1). As before
there was no evidence for dechlorination from the ortho position. For Aroclors 1242, 1248, and
1254 only 2-CB and 2,2'-CB and/or 2,6-CB accumulated to an appreciable extent. The meta
position was apparently more effectively dechlorinated than in our first Aroclor 1242
experiment. In the case of Aroclor 1260 the accumulation of ortho and meta substituted
products, particularly 2,2',5,5'-CB, was noted.
From these and other experiments it appears that there are at least two distinct
dechlorination activities associated with the Hudson River sediments. One preferentially
dechlorinates the meta position while the other preferentially dechlorinates the para position.
When both activities are expressed the only prominent products are the ortho-only substituted
chlorobiphenyls.
Miscellaneous Observations
Several miscellaneous observations regarding PCB dechlorination by Hudson River
microorganisms may be made from our attempts to characterize the dechlorinating
microorganisms. While they can survive intermittent oxygen exposure, anaerobic conditions are
required for dechlorination. Further, dechlorination has always been accompanied by
methanogenesis. BESA (a specific inhibitor of methanogenesis), molybdate (an inhibitor of
sulfate reduction), sulfate, and nitrate all inhibit dechlorination.
The frequency and amount of substrate added may influence PCB dechlorination.
Continuously available yeast extract or acetate inhibited dechlorination. The dechlorinators
may be out competed by other organisms when readily utilizable carbon sources are
continuously available. Small amounts of pyruvate, fed at intervals such that all is consumed
between additions, appear to support a dechlorinating community. Similarly made additions of
glucose, methanol, or acetone appear to stimulate dechlorination (7).
Aroclor Dechlorination by Silver Lake Microorganisms
We have also investigated the dechlorination of Aroclors by Silver Lake microorganisms.
Silver Lake in Massachusetts was contaminated primarily with Aroclor 1260 and some 1254.
Initially the Silver Lake microorganisms dechlorinated Aroclor 1242 at a rate comparable to the
Hudson River microorganisms (Table 4.2.1), but after the first 4 weeks dechlorination ceased,
leaving an average of approximately one chlorine in the meta and/or para position. Closer
inspection revealed the accumulation of ortho and para substituted products. Thus
-------
62 Polychlorinated Biphenyls
dechlorination of Aroclor 1242 by the Silver Lake microorganisms appears to be limited to
removing primarily meta chlorines. Dechlorination ceased when most of the meta chlorines had
been removed.
The Silver Lake microorganisms were more effective than the Hudson River ones at
dechlorinating Aroclor 1260 (Table 4.2.1). Dechlorination was first evident after only 8 weeks
of incubation and continued throughout the course of the experiment. As with Aroclor 1242,
dechlorination appeared to be limited to removing meta chlorines. The most prominent
dechlorination product was 2,2',4,4'-CB.
Dechlorination Patterns
The existence of the different dechlorination patterns or specificities for the Hudson River
and Silver Lake microorganisms implies the existence of different PCB dechlorinating species or
strains. We have found at least one other unique dechlorination activity associated with New
Bedford Harbor, MA sediments.
Implications for Bioremediation
The anaerobic dechlorination of PCBs has three important implications for the
development of a biological treatment system for the destruction of PCBs. First, because of the
preferential removal of meta and/or para chlorines, anaerobic dechlorination alone reduces the
toxicity of a PCB mixture. Second, the process is capable of transforming the more heavily
chlorinated congeners that are so resistant to aerobic biodegradation. And third, if anaerobic
dechlorination is coupled to a subsequent aerobic biodegradation step, then greater
mineralization of the PCBs can be expected than from aerobic treatment alone.
Toxicity Reduction
The most toxic of the PCB congeners are generally considered to be the coplanar
congeners 3,3',4,4'-CB, 3,3',4,4',5-CB, and 3,3',4,4',5,5'-CB. In a coplanar configuration, these
congeners are structurally similar to 2,3,7,8-tetrachlorodibenzodioxin (TCDD) and exhibit similar
toxicity effects. PCB congeners like these but with a single ortho chlorine also have similar
toxicity effects but are much less potent.
We first confirmed the dechlorination of two of these toxic PCB congeners (3,3',4,4'-CB and
2,3,3',4,4J-CB) by adding them to the Aroclor 1242 used in a dechlorination assay with Hudson
River microorganisms. These two congeners were dechlorinated at rates similar to other tetra
and penta chlorobiphenyls in the Aroclor 1242 (Figure 4.2.5).
Both 3,3',4,4'-CB and 2,3,3',4,4'-CB coelute with other congeners in Aroclor 1242. Because
we used an electron capture detector which does not distinguish among coeluting congeners, it
was necessary to amend the Aroclor with each of these congeners in order to be sure they were
being dechlorinated when present in a mixture. More recently Lopshire and Encke in the
Department of Chemistry at Michigan State University have developed a sensitive GC-MS-MS
technique to directly quantify these and other toxic PCB congeners. This new method allowed
us to determine the percent reduction of each of the toxic congeners after a 16 week incubation
of Aroclor 1242 with the Hudson River microorganisms.
The dioxin-like toxicity of compounds has been correlated with their potential to induce
P450 enzymes such as aryl hydrocarbon hydroxylase (AHH) and ethoxy resorufin O-deethylase
(EROD) and the toxicities of various PCB congeners have been estimated based on their
potential to induce these enzymes (11,12). Using such toxicity estimates we calculated an 85%
reduction in toxicity in 16 weeks as a result of the dechlorination of Aroclor 1242 by Hudson
River microorganisms.
EROD induction assays were performed on the PCB extracts from live and autoclaved
treatments to directly determine the toxicity reduction affected by 16 weeks of dechlorination of
Aroclor 1242. A 75% reduction was observed, in good agreement with our calculations.
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J.F. Quensen, S^L Boyd, and J.M. Tiedje 63
Extending the Range of Biodegradable Congeners
The dechlorination process also extends the range of PCB congeners that can be
biologically transformed. The dechlorination of Aroclor 1254 and 1260 is particularly notable.
The aerobic transformation of congeners in 1254 is limited and there is no convincing evidence
for the aerobic transformation of Aroclor 1260, yet both of these Aroclors can be dechlorinated
by suitable anaerobic microorganisms.
A sequential anaerobic-aerobic treatment system should also lead to greater mineralization
of a PCB mixture than would aerobic treatment alone. Many of the more chlorinated
biphenyls that are reportedly aerobically degraded are in fact only transformed, often merely to
hydroxylated chlorobiphenyls. However, an anaerobic pretreatment would remove the chlorines
that limit aerobic mineralization.
Biotreatment Scenarios
Several biotreatment systems employing a sequential anaerobic/aerobic sequence may be
envisioned. Sediments might be treated ire situ, although some form of containment will likely
be required. Alternatively, sediments could be removed to a containment facility where greater
control over conditions is possible. Sediments may or may not have to be inoculated with
dechlorinating organisms. Soils would likely have to be slurried in a containment facility and
inoculated with appropriate microorganisms to effect dechlorination. The aerobic treatment step
would require some form of mixing or aeration.
Site Assessment
While the details of a sequential anaerobic/aerobic biotreatment system for the destruction
of PCBs have not yet been worked out, it is possible to appraise the likelihood that a
particular site can be treated in this manner. A particular site would have to be evaluated for
the following:
1) The presence of dechlorinating microorganisms.
2) In situ dechlorination and dechlorination patterns.
3) Sediment type.
4) Nutrients / organic carbon.
5) Presence of inhibitors.
6) Bioavailability of the PCBs.
The presence of dechlorinating microorganisms.
If dechlorinators are absent the sediment or soil will have to be inoculated. This might
be accomplished by adding laboratory grown microorganisms or, if in a containment
facility, by simply mixing with a second sediment that does contain dechlorinators.
In situ dechlorination and dechlorination pattern.
In some cases dechlorination may have already proceeded to the point where only aerobic
treatment is needed. In some cases the particular dechlorinators present may have
limited dechlorination capabilities. For example, they may remove only meta chlorines or
have limited activity on some tri- and tetrachlorobiphenyls. Then it may still be desirable
to add a different dechlorinating microorganism. If no dechlorination has occurred, it may
be because dechlorinators are absent or because inhibitors are present. Appropriate
bioassays can distinguish between these possibilities.
Sediment type.
Oxygen diffuses much more slowly through fine grained sediments so that anaerobic
conditions are more likely to develop. Anaerobic conditions are of course necessary for
dechlorination to occur.
Nutrients / organic carbon.
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64 Poly chlorinated Biphenyls
There must be enough substrate initially present in the sediments to deplete oxygen in
the sediments and consume other electron acceptors such as sulfate which inhibit
dechlorination. It is still unclear what substrates are required to support the
dechlorinators. Different dechlorinators may have different requirements.
Presence of inhibitors.
Bioassays may be conducted to determine if dechlorination of PCBs can occur in a given
soil or sediment. We have encountered some sediments which do not support
dechlorination, probably because of high levels of heavy metals.
Bioavailability of the PCBs.
The PCBs can only be dechlorinated and degraded if they are available to the
microorganisms. Bioassays may also be conducted to determine bioavailability.
Research Needs
Areas requiring further research are :
1) Environmental rates / time course
2) Concentration effects
3) Enhancement of activity
4) Propagation / Identification of dechlorinators
5) Factors affecting bioavailability
6) Suitable aerobic microorganisms for an anaerobic / aerobic treatment sequence
Environmental rates / time course.
When the evidence for environmental dechlorination of PCBs was first presented, it was
assumed to have been a slow continuous process taking decades to reach the extent of
dechlorination observed. In laboratory experiments we have achieved the same extent of
dechlorination within a few weeks. It is therefore important to determine whether PCB
dechlorination in situ was a relatively rapid event of short duration (and has since
stopped), or is in fact a continuous process. The answer to this question has important
implications for both predicting the environmental fate of PCBs in anaerobic environments
and in its implications for in situ bioremediation.
Concentration effects.
A pronounced concentration effect was observed in our laboratory experiments; no
detectable dechlorination occurred within 16 weeks at a concentration of only 14 ppm of
Aroclor 1242. Some environmental samples at comparable concentrations, however, do
show evidence of dechlorination. It is important to determine if in situ dechlorination of
low concentrations does occur and at what rate.
Enhancement of activity.
Ways to increase dechlorination rates will be important to developing a treatment system.
Areas include determining best supporting substrates, increasing bioavailability, and
counteracting inhibitory substances. It will be necessary to enhance aerobic activity by
providing oxygen or aerating in some way.
Propagation / identification of dechlorinators.
Before dechlorinating microorganisms can be grown in adequate numbers to use in
inoculating a sediment or soil to be treated, it will be necessary to determine how to
propagate them in the absence of PCBs. Identification and isolation of the dechlorinators
will aide this effort, and also allow whole new sets of experiments aimed at better
understanding the dechlorination process itself.
Factors affecting bioavailabilitv.
It is a common concern in developing biological treatment systems for poorly water soluble
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J.F. Quensen, SA. Boyd, and J.M. Tiedje 65
compounds that rates may be inadequate and target levels not be achieved because of
limited bioavailability of the compounds to the microorganisms. Even among the
relatively few sediments and industrial sludges we have examined so far, the
bioavailability of PCBs apparently varies widely. A better understanding of the reasons
for these differences between sediments and sludges may lead to ways to increase
bioavailability. It should be noted also that the generally longer time scales typical for
anaerobic dechlorination may be an advantage when desorption from sediments is slow.
Suitable aerobic microorganisms for an anaerobic / aerobic treatment sequence.
In the most complete dechlorination we have observed, PCBs substituted at only the ortho
positions accumulate. It is therefore important to obtain strains capable of degrading
these compounds. A serious problem with the aerobic degradation of PCB mixtures is
that the PCBs are actually cometabolized and the pathway must first be induced by the
addition of biphenyl. However, some strains capable of growth on monochlorobiphenyls
are known. The high levels of 2-CB that are produced by the dechlorination process may
induce such strains to cometabolize the remaining congeners.
References
1.
709-712.
Brown, J.F., D.L. Bedard, M.J. Brennan, J.C. Carnahan, H. Feng, and R.E. Wagner
(1987a). Polychlorinated biphenyl dechlorination in aquatic sediments. Science, 236:
2. Brown, J.F., R.E. Wagner, D.L. Bedard, M.J. Brennan, J.C. Carnahan, R.J. May, and
T.J. Tofflemire (1984). PCB transformations in upper Hudson sediments. Northeast.
Environ. Sci., 3: 167-179.
3. Brown, J.F., R.E. Wagner, H. Feng, D.L. Bedard, M.J. Brennan, J.C. Carnahan, and
R.J. May (1987b). Environmental dechlorination of PCBs. Environ. Toxicol. Chem., 6:
579-593.
4. Chiou, C.T., L.J. Peters, and V.H. Freed (1979). A physical concept of soil-water
equilibrium for non-ionic organic compounds. Science, 206: 831-832.
5. Dolfing, J. (1990). Reductive dechlorination of 3-chlorobenzoate is coupled to ATP
production and growth in an anaerobic bacterium, strain DCB-1. Arch. Microbiol., 153:
264-266.
6. Mohn, W.W. and J.M. Tiedje (1990). Strain DCB-1 conserves energy for growth from
reductive dechlorination coupled to formate oxidation. Arch. Microbiol., 153: 267-271.
7. Nies, L. and T.M. Vogel (1990). Effects of organic substrate on dechlorination of Aroclor
1242 in anaerobic sediments. Appl. Environ. Microbiol., 56: 2612-2617.
8. Ogram, A.V., R.E. Jessup, L.T. Ou, and P.S.C. Rao (1985). Effects of sorption on
biological degradation rates of 2,4-Dichlorophenoxyacetic acid in soils. Appl. Environ.
Microbiol., 49: 582-587.
9. 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: 2360-2369.
10. Quensen, J.F., J.M. Tiedje, and S.A. Boyd (1988). Reductive dechlorination of
polychlorinated biphenyls by anaerobic microorganisms from sediments. Science, 242:
752-754.
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66 Poly chlorinated Biphenyls
11. Safe, S. (1987). Determination of 2,3,7,8-TCDD toxic equivalent factors (TEFs): Support
for the use of the in vitro AHH induction assay. Chemosphere, 16: 791-802.
12. Sawyer, T.W. and S. Safe (1982). PCB isomers and congeners: induction of aryl
hydrocarbon hydroxylase and ethoxy resorufin o-deethylase activities in rat hepatoma
cells. Toxicol. Lett., 13: 87-93.
13. Suflita, J.M., A. Horowitz, D.R. Shelton, and J.M. Tiedje (1982). Dehalogenation: A
novel pathway for the anaerobic biodegradation of haloaromatic compounds. Science,
218: 1115-1116.
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J.F. Quensen, S-4. Boyd, and J.M. Tiedje
67
Table 4.2.1.
Maximal observed dechlorination rates (means with standard deviations) of the Aroclors
tested for microorganisms collected from the two sites. Significant differences between
rates (Least Significant Difference test, 0.05 confidence level) are indicated by different
capital letters next to means (From Quensen et al., 1990).
Site
Aroclor
Rate
(ug atoms Cl"
removed/g
sediment week)
Time Period
(weeks)
Percent
m & p Cl
removed
HR
1242
1248
1254
1260s
1260b
0.31ab(0.03)
0.34° (0.01)
0.22C (0.02)
0.00d (0.03)
0.04° (0.005)
0-8
0-8
0-8
0-25
16-24
85C
75*
63d
Od
15e
SL
1242
1260
0.30b (0.02)
0.21C (0.01)
0-4
12-16
46f
19f
a serum bottle experiment, sediments collected 1/88
b serum tube experiment, sediments collected 8/88
c after 12 weeks
d after 25 weeks
° after 50 weeks
f after 16 weeks
-------
68
Poly chlorinated Biphenyls
5001
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Figure 4.2.1. Capillary gas cliromatograrns showing tlie anaerobic dcchlorinntion of 700-ppm
Aroclor 12/12 after 16 weeks of incubation
-------
J.F. Quensen, SLA. Boyd, and J.M. Tiedje
69
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Figure 4.2.2. Decrease in the average number of chlorines by position at three Aroclor 1242
concentrations as a result of dechlorination by Hudson River microorganisms
-------
Average Total Chlorines
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-------
J.F. Quensen, S./L Boyd, and J.M. Tiedje
71
Aroclor Dechlorlnation by
Hudson River Inoculum
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Figure 4.2.4 Decrease in the average number of chlorines for four Aroclors as a result of
dechlorination by Hudson River microorganisms
-------
72
Polychlorinated Biphenyls
o
LI
cu
o
4-
34-34-CB
25-34-CB
23-34-CB
Incubation Time (Weeks)
4-
3-
1 -
O-1
234-34-CB
245-25-CB
235-24-CB
0 5 10
Incubation Time (Weeks)
Figure 4.2.5. Comparison of the dechlorination rates of 3,3',4,4'-CB, 2,3,3',4,4'-CB, and selected
tetra- and penta- CBs present in Aroclor 1242
-------
G-Y. Rhee and B. Bush 73
4.3 Dechlorination and Biodegradation of Chlorinated Biphenyls in
Anaerobic Sediments
G-Yull Rhee and Brian Bush
Wadsworth Center for Laboratories and Research
New York State Department of Health
and
School of Public Health
State University of New York at Albany
Albany, N.Y. 12201-0509
Polychlorinated biphenyls were thought to be highly resistant to biodegradation, especially
in anaerobic environments due to their thermodynamic stability. The results of earlier studies
reinforced this postulation. Although partial degradation of some monochlorobiphenyls was
observed in some cases, breakdown by facultative anaerobes could not be excluded.
Recently, our laboratory reported anaerobic biodegradation of lightly-chlorinated PCB
congeners in the laboratory by mixed cultures obtained by Aroclor 1221 enrichment of Hudson
River sediments. A comparison of congener patterns in the deeper sediment layers of the
Hudson River with those of the PCBs presumed to have been discharged into the river (Aroclor
1242), also led to speculation that PCBs were anaerobically dechlorinated (1). This hypothesis
was later disputed owing to several quantitative problems (2). Quensen et al. (3), however,
have shown unambiguous evidence for anaerobic microbial dechlorination in the laboratory with
Hudson River sediments. With Aroclor 1242, they found the accumulation of lightly
chlorinated, mostly o-substituted congeners as a result of dechlorination of m- and p-chlorines.
However, they failed to find any loss of chlorobiphenyls on a molar basis.
Our investigation of anoxic Hudson River sediments showed no sign of dechlorination,
especially for highly chlorinated congeners. Rather, lightly chlorinated congeners in ambient
sediment PCBs exhibited significant decreases. PCBs in this study consisted mainly of lightly-
chlorinated congeners. Therefore, we undertook an investigation with a mixture with greater
proportions of highly chlorinated congeners and a single hexachlorobiphenyl congener to
determine whether similar degradation also occurred and measure the rate. Six different
concentrations ranging from 100 to 1500 ppm on a sediment dry-weight basis using Aroclor
1242 were used.
PCB-free sediments, contaminated with Aroclor 1242 at 100, 300, 500, 800, 1200, and
1500 ppm on a sediment dry weight basis and enriched with 1000 ppm biphenyl, were made
into slurries by adding 20 ml of a cystine sulfide-reduced synthetic medium. They were
autoclaved and then inoculated with the supernatant of Hudson River sediment slurries (0.5
ml) except for the controls for each concentrations and incubated in triplicate under N2
atmosphere. Experiments with the single congener 2,3,4,2',4',5'-hexachlorobiphenyl was also set
up in the same manner at 300 ppm. The sediment PCBs were extracted and analyzed by GC
(Hewlett-Packard 5840A) with a Ni-63 electron capture detector and 50 m capillary column
(DB-5).
The first analysis Aroclor 1242 after a 3-month incubation clearly demonstrated
dechlorination and their dependence on the total Aroclor 1242 concentrations; the sediments
with the initial concentrations of 300 and 500 ppm showed dramatic changes in congener
patterns with highly significant accumulation of mono-, di- and some tri-chlorobiphenyls and
concomitant decrease of most congeners with three or more chlorines. Out of 39 major peaks
in the gas chromatogram comprising about 98% of the total PCB, 10 (2; 2,2'+2,6; 2,4+2,5; 2,3';
2,4'+2,3; 2,6,2'; 4,4'+2,4,2'; 2,4,3'; 2,4,4') showed significant increases after 3 months.
The accumulation of dechlorination products relative to the control was highest in the
congeners with ortho-substituted chlorines (2-, 2,2'-, 2,6,2-chlorobiphenyls). In absolute
-------
74 Poly chlorinated Biphenyls
concentrations, 2,2'-, 2,4', and 4,4'+ 2,4,2' exhibited the highest increase.
Although the congener profile also changed in sediments with 100 and 800 ppm Aroclor
after 3 months, a t-test showed no statistically discernable difference for most individual
congeners. However, the difference became highly significant at 4.5 and 6 months.
Dechlorination appeared to be inhibited at high concentrations, since at 1200 and 1500 ppm no
change was evident even after 6 months.
Dechlorination at different substitution positions reflected the concentration dependence of
overall dechlorination. m-Chlorine was most readily dechlorinated; an average number of this
chlorine per biphenyl (0.98) decreased about 75% in 6 months in the 300 ppm sediments, p-
Chlorine (0.88) was much slower at about 20%. However, the average number of o-chlorine did
not appear to change.
Despite such extensive dechlorination, no significant decrease in the total molar
concentration of the mixture was found by 6 months. These results indicate that no
biodegradation beyond dechlorination has taken place.
The single congener 2,3,4,2',4',5'-hexachlorobiphenyl (300 ppm) was investigated using
sediments reduced biologically or the same sediments further reduced with cystine sulfide.
These sediments were incubated under C02 atmosphere with and without biphenyl enrichment.
At 3 months, the sediments with cystine sulfide exhibited an extensive dechlorination, yielding
daughter congeners with fewer chlorines. However, the total molar concentration of
chlorobiphenyls showed no significant change. In the sediments which were not reduced
chemically, the parent congener was recovered quantitatively with no dechlorination.
At 9 months, however, all treatments showed dechlorination. The first dechlorination
product was 2,5,2'4'5'-pentachlorobiphenyl, which was further dechlorinated in the next step to
2,4,2',5'- or 2,4,2',4'- + 2,2'4'5'-tetrachlorobiphenyl. These products were then dechlorinated
mostly to 2,2'5'-trichlorobiphenyl, with a small amount of 2,2'4'-trichlorobiphenyl also produced.
They were then converted to 2,2-dichlorobiphenyl. This dechlorination pathway appeared to be
the same for all incubation conditions. In all treatments, the chlorination of products
decreased with time.
In summary, polychlorinated biphenyls (Aroclor 1242) were dechlorinated in anaerobic
sediments by indigenous microbial populations from Hudson River sediments when incubated
with biphenyl enrichment under N2 atmosphere, m- and p-Chlorines were most readily
dechlorinated, but o- were not. The dechlorination rate was concentration-dependent; it was
fastest at a sediment PCB concentration of 300 ppm and slower at lower (100 ppm) and higher
(500 and 800 ppm) concentrations. At 1200 and 1500 ppm, no sign of dechlorination was
observed after 6 months. As a result of dechlorination, mono-, di- and some tri-chlorobiphenyls
increased with concomitant decreases in highly chlorinated congeners. Anaerobic incubation
of the single congener 2,3,4,2',4',5'-hexachlorobiphenyl produced daughter congeners with 2 - 5
chlorines with the degree of chlorination decreasing with time. The relative concentration of
dechlorination products of the hexachlorobiphenyl appeared to vary with incubation conditions.
Total molar concentration of the parent compound and its dechlorination products did not
appear to change at 9 months.
References
1. Brown, J. F., Jr., et al (1987). Science, 236: 709.
2 Brown, M. P., B. Bush, G-Y. Rhee, and L. Shane (1988). Science, 240: 1674.
3. Quensen III, J. F., J. M. Tiedje, and S. A. Boyd (1988). Science, 242: 752.
-------
W.C. Sonzogni and MJI. David 75
4.4 PCB Dechlorination in the Sheboygan River, Wisconsin
William C. Sonzogni and Margaret M. David
Laboratory of Hygiene and Water Chemistry Program
University of Wisconsin
Madison, WI 53706
The Sheboygan River in Wisconsin flows into Lake Michigan at the city of Sheboygan,
located 90 km north of Milwaukee. Due to the high concentrations of PCBs in the river
sediment, the Sheboygan River and Harbor area has received national attention. The main
source of contamination was from a die casting plant located in the Village of Sheboygan Falls.
The contamination source area is about 22 km upstream from the mouth of the river.
Hydraulic fluids containing PCBs were used by the die casting plant from 1959 to 1971
(11). Apparently, a large fire occurred at the plant prior to 1959 that was caused by
combustion of the hydraulic fluids then in use. Fluids containing PCBs were subsequently put
in use because of their fire resistance. Based on interviews and available records, a product
called Pydrol F9 was used between 1959 and 1969 and a product called Chemtrend HF30 was
used between 1970 and 1971. Pydrol F9 contains Aroclor 1248, while Chemtrend HF30
contains mostly Aroclor 1254 with a small percentage of Aroclor 1248. In 1971 the use of
hydraulic fluids containing PCBs ceased.
Material from the plant (oil soaked rags, hoses and other refuse) and soil from around the
plant was used to construct a low dike at the edge of the Sheboygan River. The dike sloped at
a 45 degree angle to the river, so erosion of the diked material into the river occurred
relatively easily (11). Concentrations of PCBs in the soil samples were as high as 120,000
ug/g. The dye casting plant is the only known major source of PCBs to the river, therefore,
the congeners deposited in the sediments were most likely the components of Aroclor 1248 and
1254.
In an article in Science it was reported that biological reductive dechlorination of PCBs was
occurring in Hudson River sediments. There was evidence that anaerobic dechlorination was
also occurring in other aquatic sediments, including Sheboygan River sediments (4). The
Sheboygan evidence was based on observations of chromatograms obtained from the U.S. Army
Corps of Engineers.
As a result of the published reports that degradation could occur and because of new
analytical capabilities to do congener specific PCB analysis, research was begun to examine the
distribution of PCB congeners in the Sheboygan River sediment and to determine whether
anaerobic dechlorination may be occurring. The intent was to determine the congener
distribution in Sheboygan River sediment and assess whether some transformations had
occurred. Results of congener distributions in Sheboygan River sediment relative to
distributions in Aroclors will be summarized below as well as information on the occurrence of
"toxic" congeners. Finally, a summary of evidence so far to degrade PCBs in the laboratory
using bacteria from Sheboygan River sediments will be made.
Total PCB concentrations ranged from 1586 ug/g found downstream from the source to 0.04
ug/g above the source (considered to represent background levels). Although there is
considerable variation in the sediment PCB concentrations, in general, the values decreased
with distance downstream from the source. Highest PCB concentrations were found in areas of
sediment deposition in the river. In the individual cores, the top segment of core (0-15 cm)
and the bottom segment of core (45-60 cm) had relatively low concentrations of PCBs. The
highest concentrations were found in the 15-45 cm segments.
Sediment samples were also analyzed for PCB congeners using high resolution gas
chromatography. Samples containing total PCB concentrations greater than 50 ug/g appeared
to be enriched with the lower chlorinated congeners whereas those with less than 50 ug/g PCBs
were not. Samples containing 50 ug/g or more PCBs had significantly higher concentrations of
-------
76 Poly chlorinated Biphenyls
mono- and di- chlorinated congeners when compared to Aroclors 1248 and 1254 which were
originally introduced into the river. Using a multivariate ANOVA statistical test, samples
containing greater then 50 ug/g were found to be statistically different from Aroclor 1248,
Aroclor 1254 and an equal parts mixture of Aroclors 1248 and 1254 (p values were < 0.05). In
sediment samples containing PCB concentrations less than 50 ug/g, the homolog patterns were
more similar to the patterns of the Aroclor 1248 and 1254 than the more contaminated
sediments.
The most prominent congeners in the sediments with total PCB concentrations greater than
50 ug/g were (IUPAC #) 5/8, 17, 16/32, 47/48, and 28/31. Relative to the original Aroclors,
particularly high concentrations of congeners 5 and 8 were seen (congeners 5 and 8 coelute).
To confirm the presence of congeners 5/8, six of the samples containing high concentrations of
PCBs (342.8 ug/g on average) and high concentrations of congener 5/8 (43.7 percent on average)
were analyzed using an electron impact gas chromatography mass spectrometer. All samples
contained high concentrations of dichlorinated congeners. In samples containing less than 50
ug/g of PCBs, the most prominent congeners were similar, but their weight percents were
generally reduced.
The results from the sediment analyses indicate that the PCB congeners and their
respective weight percentages in sediments with high PCB concentrations are significantly
different from the Aroclors originally introduced into the river. Although physical-chemical
processes such as sediment-water partitioning are important in determining the distribution of
congeners in sediments, it is unlikely that it is the dominant process influencing the
distribution of congeners. Sediment-water partition coefficients generally increase with
molecular weight and thus an enrichment of the higher chlorinated congeners in the sediments,
not the lower chlorinated congeners as observed in the sediments, would be predicted.
Diffusion of congeners out of the sediment and into the water is slow relative to
sedimentation rates and is inversely related to partition coefficients (7,8). Therefore, a
distribution enriched in the higher chlorinated PCBs would be predicted (opposite of what was
observed in this study).
Another possibility to account for the change in congener patterns is abiotic chemical
reactions. PCBs have been shown to undergo abiotic reductive dechlorination in the laboratory;
however, the conditions in the laboratory (high temperatures, excess base, and the presence of
a catalyst) are considerably different from those in the environment (3). In general, it is
thought that there are very few abiotic pathways which completely mineralize organic
contaminants (1).
It is possible, however, that the Aroclors undergo biological dechlorination. Recent work by
Quensen et al.(9), Chen et al. (6), Rhee et al. (10), and Brown et al (4,5) suggest that PCBs
can undergo anaerobic microbial degradation. Several results in this study suggest such a
process.
First, there is a shift in the congener pattern from the higher chlorinated congeners to the
lower chlorinated congeners as observed by Quensen et al. (9) in a laboratory experiment and
as noted by Brown et al. (4) in a field study of Hudson River sediments. This enrichment in
lower chlorinated congeners cannot be accounted for by physical-chemical partitioning
relationships or diffusion processes.
Second, there appears to be a structural selectivity as to which congeners are depleted in
the sediment. Congeners containing chlorines in the ortho position are enriched, whereas
congeners containing chlorines in the meta and para position are depleted. This is consistent
with the results obtained by Quensen et al. (9) and Brown et al. (4) in their anaerobic
microbial dechlorination work.
Third, several congeners are found in abundance that would not be expected based on
physical-chemical partitioning relationships or on the original weight percentages present in the
Aroclor mixtures. In sediment samples with concentrations of PCBs above 50 ug/g, congeners
that were significantly enriched are 5/8, 19, 17, 24/27, 16/32, 26, and 47/48. Congeners that
were significantly depleted are 18, 28/31, 52, 44, 70/76, 66/95, 56/60, 101, 77/110, 132/153/105,
and 138/163. These changes are comparable to Brown et al.'s (5) findings.
Fourth, the concentration of PCBs in the sediment appears to be important. Microbial
degradation is often restricted to areas of high substrate concentrations. For example, toluene,
xylene, and naphthalene are metabolized by bacteria at high concentrations but not at low
-------
W.C. Sonzogni and MM. David 77
concentrations (2). Threshold concentrations exist for many contaminants and are the
minimum concentration of a chemical which is needed to support growth of a microbial
population (2). Below the threshold concentration, additional energy sources must be found to
support growth of the population, since the organism is no longer able to completely mineralize
the substrate. Based on the different chromatographic patterns seen for high and low
concentrations of PCBs in the Sheboygan River, it may be that a threshold concentration exists
for the anaerobic dechlorination of PCBs.
To confirm that microbial processes are actually responsible for degrading PCBs, laboratory
experiments have been conducted similar to those reported by Quensen et al. (9) and Rhee et
al. (Iff). Using bacteria extracted form Sheboygan sediments, degradation was attempted using
growth medium and anaerobic conditions suitable for microbial dechlorination of PCBs.
However, to date no dechlorination has been observed in the experiments. The reasons for the
lack of dechlorination activity is not clear, but it is suspected that the conditions that favor
degradation are very complicated (e.g., may involve very precise Eh conditions and may involve
several different species or strains or organisms) and may be difficult to consistently reproduce
in the laboratory.
Finally, Sheboygan sediments have been analyzed for the presence of non-ortho or coplanar
PCBs. These congeners are believed to be the most toxic (at least in terms of dioxin like toxic
properties), but generally coelute with other congeners using capillary column gas
chromatography. A multidimensional ("heart cutting") gas chromatograph that uses two high
resolution columns in series was used to separate coeluting congeners. Results to date indicate
that several congeners of toxicological interest are found in sediment samples, albeit at low
concentrations. Congeners 118, 105 and 77 were detected in 83 percent of the samples
analyzed, at average concentrations of about 0.25, 0.06, and 0.04 ug/g, respectively. The
average composition of these congeners was 0.13, 0.03 and 0.02 percent, respectively.
Congeners 81, 114, 167, 126 and 169 were also detected in some of the samples, all at
concentrations less than 0.03 ug/g. While the concentrations of these congeners are low
relative to total concentrations of PCBs or to the congener they coelute with, the fact that they
are present may be important toxicologically. Research is ongoing in this area.
This work was supported by grants from the Wisconsin Coastal Management Program and
the Wisconsin Sea Grant Program.
References
1. Alexander, M (1981). Biodegradation of chemicals of environmental concern. Science,
211: 132-138.
2. Alexander, M (1985). Biodegradation of organic chemicals. Environ. Sci. Technol., 18:
106-111.
3. Boyer, S.K., J. McKenna, J. Karliner, and M. Nirsberger (1985). A mild and efficient
process for detoxifying polychlorinated biphenyls. Tetrahedron Letters, 26: 3677-3680.
4. Brown, J.F., D.L Bedard, M.J. Brennan, J.C. Carnahan, H. Feng, and R.E. Wagner
(1987a). Polychlorinated biphenyl dechlorination in aquatic sediments. Science, 236:
709-712.
5. Brown, J.F., R.E. Wagner, H. Feng, D.L. Bedard, M.J. Brennan, J.C. Carnahan, and
R.J. May (1987b). Environmental dechlorination of polychlorinated biphenyls. Environ.
Toxicology and Chemistry, 6: 579-593.
6. Chen, M., C.S. Hong, B. Bush, and G.Y. Rhee (1988). Anaerobic biodegradation of
PCBs by bacteria from Hudson River sediments. Ecotoxicology and Environ. Safety, 16:
95-105.
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78 Polychlorinated Biphenyls
1. DiToro, D.M., J.M. Jeris, and D. Clarcia (1985). Diffusion and partitioning of
hexachlorobiphenyl in sediments. Environ. Sci. Technol., 19: 1169-1172.
8. Fisher, J.B., R.L Petty, and W. Lick (1983). Release of polychlorinated biphenyls from
contaminated lake sediments: flux and apparent diffusivities of four individual PCBs.
Environ. Pollut. Series B., 5: 121-132.
9. Quensen, J.F., J.M. Tiedje, and S.A. Boyd (1988). Reductive dechlorination of
polychlorinated biphenyls by anaerobic microorganisms from sediments. Science, 242:
752-754.
10. Rhee, G.Y., B. Bush, M.P. Brown, M. Kane, and L. Shane (1989). Anaerobic
biodegradation of polychlorinated biphenyls in Hudson River sediments and dredged
sediments in encapsulation. Water Research, 23: 957-964.
11. Wisconsin Department of Natural Resources (1989). Sheboygan River remedial action
plan. Environmental Quality Division, Madison, Wisconsin.
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DJL Abramowicz and M.J. Brennan 79
4.5 Anaerobic and Aerobic Biodegradation of Endogenous PCBs
Daniel A. Abramowicz and Michael J. Brennan
GE Research and Development Center
P.O. Box 8 Schenectady, NY 12301
INTRODUCTION
Environmental reductive dechlorination of PCBs has been widely observed in contaminated
sediments, and has recently been reviewed (1,3). In addition, microbial anaerobic
dechlorination of PCBs in aquatic environments has been confirmed in the laboratory (2,5).
This report will focus on recent findings involving the acceleration of dechlorination in Hudson
River sediments, the dechlorination of endogenous PCB contamination, as well as the sequential
anaerobic/aerobic treatment of contaminated sediments.
The acceleration of anaerobic dechlorination in Hudson River sediments was observed upon
the addition of a complex nutrient mixture, surfactants, or a simple trace metal mixture. The
latter result may indicate that low levels of a trace metal in the sediment may limit the rate
of PCB dechlorination occurring in the environment today. Dechlorination of endogenous PCB-
contamination has been observed in three different soils and sediments. This result indicates
that anaerobic microorganisms have access to PCBs in even "aged" soil environments. In
addition, sequential microbial treatment via anaerobic dechlorination and aerobic biodegradation
has been demonstrated on such endogenous PCB contamination.
RESULTS AND DISCUSSION
Rate Enhancement
The addition of a minimal medium to the sediment slurry resulted in a dramatic increase
in the observed rate of anaerobic dechlorination after 8 weeks (see Figure 4.5.1). The RAMM
minimal medium contained nutrients, trace minerals, and bicarbonate (6). The control was
autoclaved and incubated along with the samples; no change was observed in any of the heat
treated controls during the experiments. The control (Figure 4.5.1A), therefore, represents the
PCB distribution in the original mixture added to the sediment (70% Aroclor 1242, 20% Aroclor
1254, 10% Aroclor 1260). The sample mixed with distilled water in place of the minimal
medium is shown in Figure 4.5. IB. Only slight dechlorination was observed in this
experimental sample after an 8-week incubation. This is contrasted by the significant change
observed in the sample to which RAMM minimal medium has been added (Figure 4.5.1C). The
selective mete-and para- dechlorination observed in this sample is consistent with the
environmental changes observed in the Hudson River (4). This result suggests that a limiting
nutrient present in the RAMM medium may be restricting the rate of dechlorination in Hudson
River sediments. It should be noted that at later timepoints significant dechlorination was also
observed in the sediment to which no nutrients were added. These changes were similar
although less extensive than the sample to which nutrients were added. Therefore, nutrient
addition can decrease the lag time before activity is initiated, as well as increase the extent of
dechlorination observed.
The RAMM medium was subdivided into four different components to further investigate
nutrient stimulation of this PCB dechlorination activity. Individual components were added in
various combinations and concentrations; results are shown in Table 4.5.1. Note that the
addition of the trace metals (Zn+2, Cu+2, Ni+2, Se03"2, B03'3) correlates with nearly a two-fold
increase in the rate of dechlorination of 234-34-chlorobiphenyl (CB). This effect suggests that
one of these trace metals, added at less than 0.02 ppm level, may represent the component
that limits the PCB dechlorination in Hudson River sediments. Other agents that have
-------
80 Poly chlorinated Biphenyls
demonstrated to increase the rate and/or extent of PCB dechlorination in Hudson River
sediments include non-ionic high molecular weight surfactants (e.g. Triton X-705) and the
addition of a complex carbon source (e.g. yeast extract or fluid thioglycollate medium with beef
extract). In addition, PCB dechlorination has been observed over a broad range of
temperatures (5-30°C) and PCB concentrations (20-1500 ppm).
Patterns
It has been observed that minor modifications to the RAMM medium dramatically affect
the observed dechlorination pattern for Hudson River sediments. This effect is demonstrated
by preferential dechlorination for 24- or 25- chlorophenyl PCB congeners in Figure 4.5.2. In
Figure 4.5.2B the addition of the minimal medium (RAMM) to the sediment results in
extensive dechlorination of the mixture, with corresponding large increases in the resultant 2-;
2-2-; 2-3-; 2-4-; and 26-2 chlorobiphenyl peaks. The shaded peak, 2356-245-heptachlorobiphenyl,
is virtually untouched in these systems and serves as an internal reference for comparisons. In
Figure 4.5.2C, the effect of the minimal medium and the reductant cysteine hydrochloride is
shown.
The addition of minimal salts (Figure 4.5.2B) supports the growth of a microbial population
which more readily attacks PCB congeners containing a 25-dichlorophenyl ring than a 24-
dichlorophenyl ring (pattern M). Note that the congeners 25-25-; 25-4; and 25-2-chlorobiphenyl
have all decreased in area while the corresponding 24-24-; 24-4-; and 24-2- chlorobiphenyls
have not decreased. But the addition of reductant (Figure 4.5.2C) now supports a microbial
population which prefers the 24- over the 25-dichlorophenyl groups (pattern Q). It is also
possible to determine conditions which support the growth of both of these microbial
populations (data not shown).
Endogenous PCBs
It is possible that biodegradation studies on soils spiked with PCBs may not provide
accurate kinetic data for similar experiments on endogenous, aged PCB contamination. It has
been observed with South Glens Falls dragstrip soil that aerobic biodegradation rates can be
limited due to bioavailability issues (data not shown). Therefore several different contaminated
soils and sediments were investigated to directly monitor the PCB dechlorination rate of the
endogenous contamination.
Hudson River sediments contaminated >15 years ago have already been extensively
dechlorinated in the environment (4). Such sediments can be even further dechlorinated by the
addition of RAMM nutrients (see Figure 4.5.3B). The dechlorination rate observed is
comparable to that found in spiked samples. Endogenous PCB contamination can also be
dechlorinated in Woods Pond sediments (Aroclor 1260, data not shown).
The endogenous PCBs bound to dragstrip soil were also available for dechlorination via
anaerobic microorganisms (see Figure 4.5.4). In this experiment, 25% by weight Hudson River
sediments were added to the dragstrip soil containing RAMM medium. Again, this microbial
process can successfully attack the endogenous PCB contamination. This result is particularly
encouraging since this same soil demonstrated bioavailability limitations upon aerobic
treatment.
Sequential Anaerobic/Aerobic Treatment
Hudson River sediments that had previously undergone environmental dechlorination were
then treated by aerobic PCB-degrading organisms to demonstrate the effect of this combined
process (see Figure 4.5.5). Figure 4.5.5A represents the original contamination (Aroclor 1242);
Figure 4.5.5B displays a recently obtained sediment sample from the Hudson River that has
been environmentally dechlorinated (>85% mono- and di-CB); Figure 4.5.5C displays the
resulting chromatogram after aerobic treatment. This initial trial demonstrated >70% reduction
in the PCB concentration after one day of aerobic treatment.
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DJL Abramowicz and M.J. Brennan 81
CONCLUSIONS
Hudson River Sediments contain anaerobic microorganisms capable of extensively
dechlorinating PCB mixtures. The observed rate of laboratory dechlorination can be stimulated
by the addition of trace metal mixture or detergents at low concentrations. Different
dechlorination patterns observed under different experimental conditions indicate that these
sediments contain a complex microbial population of different dechlorinating organisms.
Experiments on a variety of PCB contaminated soils have demonstrated that this anaerobic
process will effectively attack endogenous PCB contamination. No significant rate difference
was observable for endogenous or spiked PCB samples. In addition, sequential
anaerobic/aerobic treatment of the PCB contamination present in Hudson River sediments have
resulted in a >70% reduction in total PCB concentrations and a dramatic shift in PCB
distribution to lightly chlorinated material.
REFERENCES
1. Abramowicz, D.A. (1990). In: CRC Critical Reviews in Biotechnology (eds.), G.G. Steward
and I. Russell, CRC Press, Inc., in press.
2. Abramowicz, D.A., M.J. Brennan, H.M. Van Dort and E.L. Gallagher (1990). In: Chemical
and Biochemical Detoxification of Hazardous Waste II (ed.), J. Glaser, Lewis Publishers, in
press.
3. Bedard, D.L. (1990). In: Biotechnology and Biodegradation, (eds.), D. Kaemly, A.
Chakrabarty, G.S. Omenn, Adv. Appl. Biotechnol. Series, Vol. 4, Portfolio Pub. Co., The
Woodlands, TX., pp. 369-388.
4. Brown, J.F., Jr., D.L. Bedard, M.J. Brennan, J.C. Carnahan, H. Feng and R.E. Wagner
(1987). Science, 236: 709-712.
5. Quensen, J.F., Jr., J.M. Tiedje, and S.A. Boyd (1988). Science, 242: 752-754.
6. Shelton, D.R. and J.M. Tiedje (1984). Appl. Environ. Microbiol, 47: 850-857.
-------
82 Polychlorinated Biphenyls
TABLE 1: Effect of RAMM Components on Dechlorination Rate
Relative Dechlorination Rate A B C D
105% +
90 + +
105 + + +
171 + -f +
210 -f ++ +
171 + + ++ +
202 + -f + +
210 + + + +
191 + +
A = Phosphate Salts, Cystein, HCO3 '
B = Nitrogen + Minerals (CaCl2, MgCl2, FeCl2)
C = MnCl2, Mo04 "2, CoCl2
D=Trace Metals (BO3 '3, Zn^2, Cu+2, Ni+2, SeO3 ~2)
Table 4.5.1. Effect of RAMM components on dechlorination rate
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DA. Abramowicz and M.J. Brennan
83
A
iWiiULiu*
B
uv
JUiLLl
C
L
M || II II II II I II, 11 I
Figuro 4.5.1. Acceleration of the reductive dechlorinntion of PCBs upon addition of nutrients (8
week timepoint). A) autoclaved control; B) includes distilled water; C) includes
RAMM minimal medium. All samples contain 500 ppm PCB (70% Aroclor 1242,
20% Aroclor 1254, 10% Aroclor 1260) inoculated with sediments from the Hudson
River.
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84
Poly chlorinated Biphenyls
t r. /\ T /\
r-i ^ ' v -O I »
B
/\ 1
r-t rsi r)
C
Figure 4.5.2. Dechlorination patterns observed under different conditions (18 week timepoint).
A) autoclaved control; B) includes RAMM (pattern M); C) includes RAMM +
cysteine hydrochloride at 1 gm/L (pattern Q)
-------
DA. Abramowicz and M.J. Brennan
85
PQ
U
A
JUv
CQ
U
Figure 4.5.3. Dechlorination of endogenous PCB contamination in Hudson River sediments
with RAMM (18 week timepoint). A) autoclaved control; B) experimental.
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86
Polychlorinated Biphenyls
Jj_J
UU
A
B
A/ .A.JVA
t tl IIJJI III I 1 ill!
Figure 4.5.4. Dechlorination of endogenous PCB contamination in South Glens Falls soil with
25% Hudson River sediment (23 week timepoint). A) autoclaved control; B)
experimental.
-------
DA. Abramowicz and M.J. Brennan
87
LJ_.
\. -J X •_
B
CO
CM
Figure 4.5.5. Sequential Anaerobic/Aerobic treatment of endogenous PCB contamination in
Hudson River sediments. A) Aroclor 1242; B) environmentally dechlorinated
Aroclor 1242; C) B+ aerobic treatment (1 OD cells; 1 day timepoint).
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88 Poly chlorinated Biphenyls
4.6 Remediation Pilot Study in the Sheboygan River, Wisconsin, USA
Dawn S. Foster, P.E.
Blasland & Bouck Engineers, P.C.
Syracuse, NY 13214
The Sheboygan River and Harbor site, located approximately 55 miles north of Milwaukee,
Wisconsin, was placed on the National Priorities List (NPL) in May 1986. The site includes
approximately 14 miles of river and a 100-acre harbor. The river, which flows easterly toward
Lake Michigan, drains 432 square miles of central Wisconsin countryside. The contaminants of
concern include PCBs and various metals.
The Sheboygan River and Harbor Remedial Investigation/Feasibility Study (RI/FS) Program
began in 1986. Blasland & Bouck Engineers, P.C., (Blasland & Bouck) on behalf of Tecumseh
Products Company (one of three identified potentially responsible parties), developed the work
plan and appropriate project plans for the investigation work efforts. Remedial Investigation
field efforts, conducted in a phased approach, began in 1987. The first phase involved
obtaining a number of "key" samples from the river and harbor which were subsequently
analyzed for the hazardous substance list (HSL). Based on results from this first round, the
contaminants of concern were confirmed to include PCBs and eight metals. During the course
of the remedial investigations, approximately 250 sediment cores, three rounds of water column
samples and 20 floodplain soil samples were analyzed for these contaminants.
After a preliminary screening of potentially applicable technologies for remediation of the
site (if deemed necessary), it became apparent that additional information would be necessary
to perform a meaningful comparative analysis of remaining technologies. This was especially
true for those technologies considered to be both promising and innovative.
The results of the preliminary screening, coupled with the EPA's request to remove three
sediment areas with elevated concentrations, prompted a proposal by Tecumseh to conduct an
Alternative Specific Remedial Investigation (ASRI) to provide the information necessary to
conduct a comprehensive feasibility study. The proposed ASRI activities fall into two distinct
categories. The first consists of a pilot study to investigate the feasibility of enhancing natural
biodegradation of PCBs. The process of biodegradation is believed by experts to be already
occurring in the river. The second category includes various bench-scale studies of other
potentially applicable technologies and additional investigative efforts to further supplement the
remedial investigations. More specifically, the primary objectives of the ASRI work efforts,
including the pilot study activities, are as follows:
A. Study the potential for enhancing natural biodegradation
B. Evaluate mechanical dredging in the Sheboygan River
C. Evaluate the effectiveness of in situ capping or "armoring"
D. Monitor the impact of activities on the water column
E. Conduct bench-scale studies of promising and innovative remedial technologies for site
sediments
Each of these objectives is further defined in the text which follows.
Blasland & Bouck designed a pilot scale confined treatment facility (CTF) to study the
effectiveness of using enhanced biodegradation for treatment of contaminated sediments
removed from the river. In addition, it was determined that enhancement of biodegradation
should be investigated with sediments remaining in the river. This was to be accomplished by
capping or "armoring" the sediments, and then monitoring to see if the conditions for natural
biodegradation could be improved.
The pilot scale CTF can accommodate approximately 2500 cubic yards (cy) of sediments
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D. Foster 89
removed from the river, and is constructed of structural steel sheet piling. The 14,000 square-
foot structure is divided into four separate cells, two study cells and two control cells. Each
cell is lined with two high density polyethlyene liners with a leak detection system in between.
Each cell has an independent discharge which exits the cell by flowing through a
permeable treatment wall (PTW). This special design feature provided to study alternative
means for treatment of water discharges, consists of various configurations of sand and
"organic" material. Water from the four cells will be allowed to flow from top to bottom
through the PTWs for cells 1 and 2, and horizontally through an unlined sheet piling wall for
cells 3 and 4. The discharge from each PTW will be comparatively evaluated for treatment
effectiveness. The intent of the alternative water treatment study is to identify an effective
effluent treatment material for possible scale up to a full scale CTF, should this be deemed
necessary. In addition to the special CTF design features mentioned above, an amendment
distribution system is provided in the bottom of each cell to facilitate introduction of materials
for the enhancement of biological activity.
Bench-scale biodegradation studies are currently underway at the University of Michigan,
under the direction of Dr. Timothy Vogel. Dr. Vogel's efforts include work with both anaerobes
and aerobes from the Sheboygan River and other PCB-contaminated sites. Research efforts are
ongoing and will continue into the fall. The results from this work (and that of other
researchers) should provide the information necessary to enhance the process that nature has
already begun.
Sediment removal activities were designed to minimize exposure of the sediments to the
air, in order to preserve the conditions necessary for the indigenous bacteria. As such, the
sediments are mechanically removed utilizing a sealed clamshell. The sediments are then
placed into a sealed 7-cubic yard capacity transport box. Within the river, each sediment area
to be removed is surrounded with a double silt curtain system to prevent the downstream
movement of materials suspended in the water column during the removal process. The silt
curtain system consists of an outer geomembrane curtain and an inner geotextile curtain.
Each curtain is weighted with flexible chain or cable to conform to the configuration of the
river bottom.
Sediments are removed in two "passes". The first pass removes the majority of the
sediment deposit, usually to the hard underlying clay. The area within the curtains is then
allowed to "rest" (remain quiescent) for a minimum of 12 hours prior to conducting a second
pass. The intent of the second pass is to enable removal of the settled fines to the extent
possible, and provide for a "buffer" of underlying material to be removed. After all sediment is
thought to have been removed from each area, the river bed is probed, and post removal
sediment and water column samples are obtained for analysis.
Armoring confines the sediments in place by covering the deposits with successive layers of
materials to minimize resuspension. The same silt curtain system is employed during the
armoring activities. Armoring of the sediments is accomplished as follows:
A. Geotextile material is placed on the sediment deposit, extending beyond the sediment
limits by five feet;
B. A 6-inch layer of run of bank material is placed;
C. Another layer of geotextile is placed;
D. Rock-filled wire cages, called gabions, are placed along the periphery of the sediment
area to anchor the geotextile layers; and
E. A layer of stone is placed for ballast.
To accommodate the monitoring of biological activity under the armoring materials, a
sampling port is provided in a number of the armored sediment areas. This sampling port will
allow for retrieval of sediment samples every six months from underneath the armoring
material. These samples will be subjected to congener specific analyses to assess PCB
biodegradation.
As with any pilot study activity, monitoring of the effects before, during and after the
activity is important; this pilot study is no different in this respect. River monitoring activities
include daily monitoring of the water column, both upstream and downstream of the actual
work area, whenever work is being conducted. Samples are obtained and total suspended
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90 Polychlorinated Biphenyls
solids (TSS) and turbidity determined. In addition to the daily monitoring, weekly water
column samples are obtained and analyzed for PCBs (total and filtered).
Biological monitoring (both in situ caged fish studies and analysis of resident and
migratory species) is also being employed. The in situ fish monitoring consists of caged fish
studies (42-day exposure). Pre-construction monitoring was conducted in September 1989
(before any river activities were initiated) and removal/armoring monitoring took place in
December 1989. The final set of in situ fish studies will be conducted well after all river
activities are complete. This post construction study will be conducted during a similar time
period as the pre-construction study to minimize the effects of water temperature.
Extensive monitoring of resident and migratory species of the Sheboygan River is already
ongoing to develop an adequate data base with which to compare future fish results after the
pilot study and final remedy are completed (should the latter be necessary). The fish selected
for monitoring include:
A. Chinook salmon
B. Steelhead trout
C. Small-mouth bass
D. Sucker species (preferably young-of-the-year)
Collection of these fish occurs throughout the year, as appropriate.
Bench-scale study efforts include gathering additional information on the physical
characteristics of the sediment and applicability of various technologies for remediation.
Further physical characterization includes obtaining supplemental information for the sediments
such as in situ density, particle size, affinity of PCBs for different sized particles, and
Atterberg limits. Other sediment characteristics related to handling prior to treatment or for
design purposes require further definition. These include settleability, dewatering ability,
consolidation, and leachability.
Preliminary technology assessment and determination of applicability to the Sheboygan
River and Harbor sediments will be conducted on a number of technologies. These include
biodegradation (previously mentioned), a number of extraction methods, stabilization/fixation,
and in situ armoring. Many of these studies are already ongoing or are to be conducted in the
near future.
Sediment removal/armoring activities are anticipated to be complete by the end of 1990.
As previously mentioned, it is hoped that biodegradation studies within the CTF will be
initiated by late fall. The other bench-scale treatability studies and the sediment
characterization work will be presented in a final ASRI report to the reviewing agencies. It is
anticipated that this ASRI report will be available by the end of 1991. The final Feasibility
Report for the site will be developed thereafter.
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5 POLYCYCLIC AROMATIC HYDROCARBONS
5.1 The Use of a Mycobacterium sp. in the Remediation of Polycyclic
Aromatic Hydrocarbon Wastes
Carl E. Cerniglia, Ph.D.
Microbiology Division
National Center for Toxicological Research
Food and Drug Administration
Jefferson, Arkansas 72079
Abstract
Recent investigations in my laboratory on tbe biodegradation of PAHs has led to the
isolation of a Mycobacterium sp., which was able to extensively degrade PAHs containing up to
five fused aromatic rings. This microorganism has been shown to mineralize naphthalene
(59.5%), phenanthrene (50.9%), pyrene (63.0%), fluoranthene (89.7%), 1-nitropyrene (12.3%),
dihydropyrene. 18O2 incorporation experiments
dihydrodiol isomers were catalyzed by dioxygenase and monooxygenase enzymes, respectively.
Similar studies with naphthalene indicated that the Mycobacterium initially hydroxylated
naphthalene to form cis- and £rans-l,2-dihydroxy-l,2-dihydronaphthalene in a ratio of 20:1,
respectively. The cis-naphthalene dihydrodiol was further metabolized to ring cleavage products
via the classical meta cleavage pathway. Initial oxidation of 1-nitropyrene occurred in the 4,5-
and 9,10- positions to form cis-4,5- and 9,10-1-nitropyrene dihydrodiols. Fluorenone-1-carboxylic
acid was identified as a predominant ring cleavage product in the degradation of fluoranthene
by the Mycobacterium.
The ultimate usefulness of the Mycobacterium in the bioremediation of PAH contaminated
sediments depends upon its survival and function in diverse ecosystems. The Mycobacterium
survived and mineralized PAHs in sediment and water microcosms. Microcosms inoculated
with the Mycobacterium showed enhanced mineralization, singly and as components in a
mixture, for 2-methylnaphthalene, phenanthrene, pyrene and benzo[a]pyrene. Studies utilizing
pyrene as a sole PAH substrate showed that the Mycobacterium survived in microcosms for six
weeks in both the presence and absence of PAH exposure. The versatility of the
PAH-degrading Mycobacterium and its potential for use in the bioremediation of PAH
contaminated sediments will be discussed.
Introduction
Polycyclic aromatic hydrocarbons (PAHs) are a major class of environmental contaminants
originating from both petrogenic and pyrogenic sources (22,24,25,27,28,34,38,41). Many PAHs
are cytotoxic, mutagenic and carcinogenic to both lower and higher eucaryotic organisms
(13,24,29,33,37) (Figure 5.1.1). Due to their hydrophobic nature, most PAHs in aquatic
ecosystems rapidly become associated with particles and are deposited in sediments. A variety
of processes, including volatilization, sedimentation, chemical oxidation, photo-decomposition,
and microbial degradation are important mechanisms for environmental loss of PAHs (Figure
5.1.2). Microbial degradation of PAHs can have a significant effect on the PAH distribution in
91
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92 Polycyclic Aromatic Hydrocarbons (PAHs)
sediment, especially near the sediment-water interface (2,3,6,31).
There is considerable interest in the use of microorganisms to decontaminated PAH-
polluted environments (42). Successful bioremediation is dependent upon the availability of
microorganisms which possess the catabolic enzymes needed to degrade PAHs. Mono- and
dioxygenases are two groups of enzymes which are important to the microbial catabolism of
PAHs. Dioxygenases incorporate both atoms of the oxygen molecule into the PAH. This
dioxygenase reaction is the major mechanism for the initial oxidative attack on PAHs by
bacteria, which leads to the formation of dihydrodiols that are in the cis- configuration (6).
Enzymatic fission of the aromatic ring is also catalyzed by dioxygenases (Figure 5.1.3). In
contrast to bacteria, fungi oxidize PAHs via a cytochrome P-450 monooxygenase by
incorporating one atom of the oxygen molecule into the PAH and the other into water (7-12).
Chemical pathways and enzymatic mechanisms for the microbial metabolism of PAHs
containing two or three aromatic rings have been well studied (6). However, there are very
few studies on the microbial degradation and detoxification of higher molecular weight PAHs.
Our current knowledge on the microbial degradation of PAHs is summarized below:
1. Biodegradation of lower molecular weight PAHs by a wide variety of microorganisms has
been demonstrated and the biochemical pathways have been investigated (6).
2. There is limited information on the microbial utilization of PAHs containing four or more
aromatic rings; however, cometabolism of high molecular weight PAHs by bacteria has been
demonstrated (1,16,17,19,20,30,32,39,40).
3. Biodegradation of unsubstituted PAHs always involves the incorporation of molecular
oxygen catalyzed by monooxygenase(s) or dioxygenase(s) (6). However, there is also
increasing interest and speculation concerning anaerobic decomposition of PAHs (35,36).
4. Many of the genes coding for bacterial degradation of PAHs are plasmid-associated (5,45).
5. Fungi hydroxylate PAHs as a prelude to detoxification, whereas bacteria oxidize PAHs as a
prelude to ring fission and assimilation (6,7-12).
6. Fungal metabolism of PAHs is highly regio- and stereoselective (8,11).
7. White-rot fungi have the ability to cleave the aromatic rings of PAHs (4).
8. Microbial degradation of PAHs can occur under denitrifying conditions (35,36).
9. Lower molecular weight PAHs, such as naphthalene and phenanthrene, are degraded
rapidly in sediments, whereas higher weight PAHs, such as benz[a]anthracene or
benzo[a]pyrene, are quite resistant to microbial attack (2,15,21).
10. Environmental factors can have a significant effect on PAH biodegradation (43).
11. There are higher biodegradation rates for PAHs in PAH-contaminated sediments than in
pristine sediments (15,18,21).
12. Procaryotic pathways for naphthalene metabolism predominate in sediments from
freshwater and estuarine sediments (18).
Recent investigations in my laboratory on the biodegradation of PAHs has led to the
isolation of a Mycobacterium sp. which is able to extensively degrade PAHs containing up to
five fused aromatic rings (16,19). The ultimate usefulness of the Mycobacterium in the
bioremediation of PAH-contaminated sediments depends upon its survival and function in
diverse ecosystems (17). The versatility of the PAH-degrading Mycobacterium and its potential
for use in the biodegradation of PAH contaminated sediments will be reported.
-------
C.E. Cerniglia 93
Materials and Methods:
Isolation of the polycyclic aromatic hydrocarbon degrading bacterium.
The bacterium was isolated from a 500 ml microcosm containing 20 g of sediment, 180 ml
of estuarine water and 100 )ig of pyrene (16,19). The sediment was obtained from a drainage
pond chronically exposed to petrogenic chemicals. After incubation of the microcosm for 25
days under aerobic conditions, the sediment samples were serially diluted and screened for the
presence of PAH degrading microorganisms (16,19).
The screening medium consisted of mineral salts medium (44) containing (per liter): NaCl,
0.3 g; (NH4)2S04, 0.6 g; KNO3, 0.6 g; KH2PO4, 0.25 g; I^HPO,, 0.75 g; MgSO4 • 7H2O, 0.15 g;
LiCl, 20 ug; CuSO4 • 5H20, 80 ug; ZnS04 • 7H2O, 100 ug; A12(SO4)3 • 1611,0, 100 ug; NiCl •
6H20, 100 ug; CoS04 • 7H2O, 100 ug; KBr, 30 ug; KI, 30 ug; MnCl2 • 4H20, 600 ug; SnCl2 •
2H20, 40 ug; FeSO4 • 7H20, 300 ug; agar, 20 g and distilled H2O, 1000 ml.
The surfaces of the agar plates were sprayed with a 2% (wt/vol) solution of a PAH
dissolved in acetone: hexane (1:1, vol/vol) and dried overnight at 35°C to volatilize the carrier
solvents. This treatment resulted in a visible and uniform surface coat of the PAH on the
agar. Inocula (100 ul) from the 10'1, 10'2, 10"3, and 10"* dilutions of microcosm sediments were
gently spread with sterile glass rods onto the agar surface; the plates were inverted and
incubated for three weeks at 24°C in sealed plastic bags to conserve moisture.
When colonies surrounded by clear zones (Figure 5.1.4) due to polycyclic aromatic
hydrocarbon uptake and utilization were observed (after 2 to 3 weeks), they were subcultured
into fresh mineral salts medium containing 250 ug/1 each of peptone, yeast extract, and soluble
starch and 0.5 ug/ml of a PAH dissolved in dimethylformamide. After three successive
transfers, a bacterium was isolated which was able to degrade pyrene, a PAH containing 4
aromatic rings.
Growth of Organism and Culture Conditions,
The Mycobacterium sp. was grown in 125 ml Erlenmeyer flasks containing 30 ml of basal
salts medium (19) supplemented with 250 ug/ml each of peptone, yeast extract, and soluble
starch and 0.5 ug/ml of pyrene dissolved in dimethylformamide. The cultures were incubated
in the dark at 24°C for 72 h on a rotary shaker operating at 150 rpm. Cells in the
mid-logarithmic phase of growth were harvested by centrifugation at 8000 x g for 20 min at
4°C. The harvested cells were resuspended in sterile 0.1 M £m(hydroxymethyl)aminomethane
buffer (pH 7.5) at a concentration of 3 x 106 cells/ml and used as inoculum for studies of PAH
biodegradation.
Biodegradation experiments.
Biodegradation of PAHs by the Mycobacterium sp. was monitored in a flow-through
microcosm test system (14,23,26). This system enables simultaneous monitoring of
mineralization (complete degradation to C02) and the recovery of volatile metabolites,
nonvolatile metabolites, and residual PAH. Microcosms in this test system consisted of 500 ml
glass mini-tanks containing 100 ml of minimal basal salts medium, 0.92 uCi of 14C-labeled PAH
and 50 ug of unlabeled PAH. The PAHs used and their sources were [l,4,5,8-14C]naphthalene
(5.10 mCi/mmole), Amersham/Searle Corp., Arlington Heights, 111.; [9-14C]phenanthrene (19.3
mCi/mmole), Amersham/Searle; [3-uC]fluoranthene (54.8 mCi/mmole), Chemsyn Science
Laboratories, Lenexa, Kansas; [4-14C]pyrene (30.0 mCi/mmole), Midwest Research Institute,
Kansas City, Mo.; 3-[6-14C]-methylcholanthrene (13.4 mCi/mmole) New England Nuclear Corp.,
Boston, MA. and 6-nitro-[5,6,ll,12-14C]chrysene (57.4 mCi/mmole), Chemsyn Science
Laboratories.
Each microcosm was inoculated with 1.5 x 10* cells/ml, mixed twice weekly, incubated at
24° C for 14 days, and continuously purged with compressed air. The gaseous effluent from
each microcosm was directed through a volatile-organic-trapping column containing 7 cm of
polyurethane foam and 500 mg of Tenax GC (Alltech Associates, Inc., Deerfield, II.) and a 14C02
trapping column (50 ml of monoethanolamine: ethylene glycol, 7:3, vol/vol). Mineralization was
measured at various intervals by adding duplicate 1 ml aliquots from the 14C02 trapping
column to scintillation vials containing 15 ml of a 1:1 mixture of Fluoralloy and methanol
(Beckman Instruments Co., Fullerton, Ca.). Autoclaved inoculated microcosms, and microcosms
-------
94 Polycyclic Aromatic Hydrocarbons (PAHs)
lacking the Mycobacterium sp., were included to detect abiotic PAH degradation.
Results and Discussion
There are four major objectives in my research program concerning PAH biodegradation.
1. To determine the relationships between chemical structure and PAH degradation by
measuring mineralization rates in microcosms, getting good mass balance accountability of
undegraded PAH and of volatile and non-volatile metabolites.
2. To isolate microorganisms from environmental sites chronically exposed to PAHs, which
have the ability to degrade PAHs containing four or more aromatic rings.
3. To elucidate biochemical pathways and reaction mechanisms for PAH degradation in
environmental samples.
4. To determine if PAH-degrading bacteria would be useful in the biological decontamination
and detoxification of PAH-polluted sites.
It is clear from previous investigations that it is relatively easy to isolate microorganisms,
using classical enrichment and plating techniques, which can utilize lower molecular weight
PAHs containing 2 or 3 rings. The focus of research in my laboratory is to isolate
microorganisms which degrade the higher molecular weight PAHs. A summary of our recent
investigations is reported below.
Enrichment of PAH degrading bacterium.
A pyrene-degrading bacterium was isolated by direct enrichment from sediment samples
taken from an oil field near Port Aransas, Texas (Figure 5.1.4). By repeated streaking and
isolation, we obtained an isolate, strain Pyr-1, which was identified as a Mycobacterium sp. on
the basis of the following morphological and biochemical properties (44). It formed
gram-positive, acid-fast rods (1.4 um in length and 0.7 um in width). The 15 biochemical tests,
mole percent G+C analysis of 66% and the characterization of the mycolic acids with a carbon
chain length of C58 to C64 were consistent with the assignment of this organism to the genus
Mycobacterium.
Utilization of PAHS by Mycobacterium.
The Mycobacterium utilized naphthalene, phenanthrene, fluoranthene, pyrene,
3-methylcholanthrene, 1-nitropyrene and 6-nitrochrysene when grown in mineral salts medium
supplemented with low levels of peptone, yeast extract and soluble starch (16). This bacterium
was unable to utilize these PAHS as the sole source of carbon and energy. Pyrene induced
Mycobacterium cultures readily degraded naphthalene (59.5%), phenanthrene (50.9%),
fluoranthene (89.7%), pyrene (63.0%), 1-nitropyrene (12.3%), 3-methylcholanthrene (1.6%), and
6-nitrochrysene (2.0%) to CO2 within 48 h of incubation (Figure 5.1.5). Pathways for the initial
degradation of pyrene, naphthalene, fluoranthene, and 1-nitropyrene are shown in Figures
5.1.6-5.1.9.
The Mycobacterium sp. initially oxidized pyrene to form both pyrene cis- and
£rcms-4,5-dihydrodiols (20). Oxygen-18 incorporation experiments showed that both atoms of the
cis-pyrene dihydrodiol were derived from molecular oxygen where as only one atom of molecular
oxygen was incorporated into the trans-pyrene dihydrodiol (Figure 5.1.6). 4-Phenanthroic acid,
4-hydroxyperinaphthenone, cinnamic acid were identified as ring fission products (20). The
Mycobacterium sp. initially oxidized naphthalene in the 1,2-positions to form
naphthalene-l,2-dihydrodiols. Similar to pyrene oxidation both the naphthalene cis and
£rans-l,2-dihydrodiols were isolated in a ratio of 20:1. The naphthalene cis-l,2-dihydrodiols is
further metabolized to salicylate and catechol by the classical bacterial oxidation of naphthalene
pathway (Figure 5.1.7). The Mycobacterium sp. extensively degrades fluoranthene to C02
(Figure 5.1.8). However, a ring cleavage metabolite was isolated and identified as
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C.E. Cerniglia 95
9-fluorenone-l-carboxylic acid. 1-Nitropyrene is degraded very slowly by the Mycobacterium sp.
and little mineralization occurs which indicates that the nitro-substituent may sterically block
initial enzymatic attack and ring cleavage enzymes since pyrene is rapidly degraded. However,
1-nitropyrene cis-4,5- and 9,10-dihydrodiols were isolated and characterized (Figure 5.1.9).
Microcosm studies to evaluate the PAll-degrading capacity and survival of the
Mycobacterium when added to pristine sediments.
Figure 5.1.10 indicates that 2-methylnaphthalene and phenanthrene were mineralized to
10% and 14%, respectively after 28 days in microcosms containing sediment and water from De
Gray Reservoir, Arkadelphia, Arkansas. De Gray Reservoir is a pristine lake, which receives
relatively little chemical inputs, and has a low-PAH degrading microbial population (15,17,18).
When similar microcosms were inoculated with the Mycobacterium sp. (1.5 x 105 cells/g of moist
sediment), mineralization of 2-methylnaphthalene and phenanthrene increased to 26% and 71%,
respectively. In addition, pyrene and benzo[a]pyrene degradation were observed, whereas
previously we did not see degradation of high-molecular weight PAHs in De Gray Reservoir
sediments lacking the Mycobacterium. Therefore, the Mycobacterium sp. competed with
indigenous microflora and enhanced mineralization of PAHs (17).
Our research indicates that the Mycobacterium sp. isolated from an oil-contaminated
estuarine site is very versatile and can mineralize low and high molecular weight PAHs. The
process is co-oxidation, since low levels of organic nutrients are necessary to initiate growth
and metabolism of the PAHs. The mechanism of oxidation is unique, since the Mycobacterium
has both mono- and dioxygenases to catalyze the initial attack on the PAH.
In conclusion, when one discusses the use of microorganisms in the remediation of
hazardous wastes, such as PAHs, some bioremediation issues that should be addressed are:
(1) A complete understanding of the chemical and ecological characterization of the site.
(2) More data on the fate, metabolism and kinetics of high-molecular weight PAH
biodegradation at the site.
(3) Biochemistry and mechanisms of many of the high-molecular weight PAH degradative
pathways.
(4) What conditions will insure the survival of the biological detoxification system?
(5) How can the biological detoxification system be effectively transported to the site?
(6) Development of procedures for employing immobilized cells to decontaminate PAH
contaminated soils.
(7) Is bioremediation a cost effective means of cleanup of PAH contaminated wastes?
(8) How do you get the PAH degrading microorganisms (large biomass) there and make
them grow and function?
(9) What is the fate of plasmid DNA or recombinant strains in wastewater or sediments?
(10) How do we optimize a PAH degrading microbial system for environmental use?
(11) Basic research on coupling aerobic and anaerobic biodegradation systems.
(12) Research on specific bacteria used at a site, such as salt tolerant or chemical tolerant
bacteria.
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96 Polycyclic Aromatic Hydrocarbons (PAHs)
References
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3. Bauer, J.E., and D.G. Capone (1988). Effects of co-occurring aromatic hydrocarbons on the
degradation of individual polycyclic aromatic hydrocarbons in marine sediment slurries.
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4. Bumpus, J.A. (1989). Biodegradation of polycyclic aromatic hydrocarbons by Phanerochaete
chrysosporium. Appl. Environ. Microbiol., 55: 154-158.
5. Burlage, R.S., S.W. Hooper and G.S. Sayler (1989). The TOL (pWWO) catabolic plasmid.
Appl. Environ. Microbiol., 55: 1323-1328.
6. Cerniglia, C.E. and M.A. Heitkamp (1989). Microbial degradation of polycyclic aromatic
hydrocarbons in the aquatic environment. In: Metabolism of polycyclic aromatic
hydrocarbons in the aquatic environment, U. Varanasi (ed.), CRC Press Inc., Boca Raton
FL.
7. Cerniglia, C.E., W.L. Campbell, J.P. Freeman, and F.E. Evans (1989). Identification of a
novel metabolite in phenanthrene metabolism by the fungus Cunninghamella elegans.
Appl. Environ. Microbiol., 55: 2275-2279.
8. Cerniglia, C.E., W.L. Campbell, P.P. Fu, J.P. Freeman, and F.E. Evans (1990).
Stereoselective fungal metabolism of methylated anthracenes. Appl. Environ. Microbiol 56:
661-668.
9. Cerniglia, C.E., J.P. Freeman, G.L. White, R.F. Heflich, and D.W. Miller (1985). Fungal
metabolism and detoxification of the nitropolycyclic aromatic hydrocarbon 1-nitropyrene.
Appl. Environ. Microbiol., 50: 649-655.
10. Cerniglia, C.E., D.W. Kelly, J.P. Freeman, and D.W. Miller (1986). Microbial metabolism of
pyrene. Chem. Biol. Interact., 57: 203-216.
11. Cerniglia, C.E., D.W. Miller, S.K. Yang, and J.P. Freeman (1984). Effects of fluoro
substituents on the fungal metabolism of 1-fluoronaphthalene. Appl. Environ. Microbiol.,
48: 294-300.
12. Cerniglia, C.E., G.L. White, and R.H. Heflich (1985). Fungal metabolism and detoxification
of polycyclic aromatic hydrocarbons. Arch. Microbiol., 50: 649-655.
13. Dipple, A., R.C. Moschel, and C.A.H. Bigger (1984). Polynuclear aromatic carcinogens. In:
Chemical carcinogens. 2nd ed., C.E. Searle (ed.), American Chemical Society, Washington,
D.C., pp. 41-163.
14. Heitkamp, M.A. and C.E. Cerniglia (1986). Microbial degradation of f-butylphenyl diphenyl
phosphate: A comparative microcosm study among five diverse ecosystems. Toxicity
Assessment, 1: 103-122.
15. Heitkamp, M.A. and C.E. Cerniglia (1987). Effects of chemical structure and exposure on
the microbial degradation of polycyclic aromatic hydrocarbons in freshwater and estuarine
ecosystems. Environ. Toxicol. Chem., 6: 535-546.
-------
CJE. Cemiglia 97
16. Heitkamp, M.A. and C.E. Cerniglia (1988). Mineralization of polycyclic aromatic
hydrocarbons by a bacterium isolated from sediment below an oil field. Appl. Environ.
MicrobioL, 54: 1612-1614.
17. Heitkamp, M.A. and C.E. Cerniglia (1989). Polycyclic aromatic hydrocarbon degradation by
a Mycobacterium sp. in microcosms containing sediment and water from a pristine
ecosystem. Appl. Environ. MicrobioL, 55: 1968-1973.
18. Heitkamp, M.A., J.P. Freeman, and C.E. Cerniglia (1987). Naphthalene biodegradation in
environmental microcosms: estimates of degradation rates and characterization of
metabolites. Appl. Environ. MicrobioL, 53: 129-136.
19. Heitkamp, M.A., W. Franklin and C.E. Cerniglia (1988). Microbial metabolism of polycyclic
aromatic hydrocarbons: Isolation and characterization of a pyrene degrading bacterium.
Appl. Environ. MicrobioL, 54: 2549-2555.
20. Heitkamp, M.A., J.P. Freeman, D.W. Miller and C.E. Cerniglia (1988). Pyrene degradation
by a Mycobacterium sp.: Identification of ring oxidation and ring fission products. Appl.
Environ. MicrobioL, 54: 2556-2565.
21. Herbes, S.E., and L.R. Schwall (1978). Microbial transformation of polycyclic aromatic
hydrocarbons in pristine and petroleum-contaminated sediments. Appl. Environ. MicrobioL,
35: 306-316.
22. Kites, R.A., R.E. Laflamme and J.G. Windsor (1980). Polycyclic aromatic hydrocarbons in
marine/aquatic sediments: Their ubiquity,. In: Petroleum in the marine environment. L.
Petrakis and F.T. Weiss (eds.), Advances in Chemistry Series, American Chemical Society,
Washington, D.C., pp. 289-311.
23. Huckins, J.N., J.D. Petty and M.A. Heitkamp (1984). Modular containers for microcosm
and process model studies on the fate and effects of aquatic contaminants. Chemosphere,
13: 1329-1341.
24. International Agency for Research on Cancer (1983). Polynuclear Aromatic Compounds.
Part 1, Chemical, environmental and experimental data. In: IARC Monographs on the
evaluation of the carcinogenic risk of chemicals to humans, World Health Organization,
Lyon, France, pp. 95-451.
25. Jacob, J., W. Karcher, J.J. Belliardo and P.J. Wagstaffe (1986). Polycyclic aromatic
hydrocarbons of environmental and occupational importance. Fresenius Z. Anal. Chem.,
323: 1-10.
26. Johnson, B.T., M.A. Heitkamp and J.R. Jones (1984). Environmental and chemical factors
influencing the biodegradation of phthalic acid esters in freshwater sediments. Environ.
Pollut. Ser. B, 8: 101-118.
27. Johnson, A.C. and D. Larsen (1985). The distribution of polycyclic aromatic hydrocarbons
in the surficial sediments of Penobscot Bay (Maine, USA) in relation to possible sources
and to other sites worldwide. Mar. Environ. Res., 15: 1-16.
28. Jones, K.C., J.A. Stratford, K.S. Waterhouse, and N.B. Vogt (1989). Organic contaminants
in Welsh soils: polynuclear aromatic hydrocarbons. Environ. Sci. TechnoL, 23: 540-550.
29. Keith, L.H., and W.A. Telliard (1979). Priority pollutants I. a perspective view. Environ.
Sci. TechnoL, 13: 416-423.
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98 Polycyclic Aromatic Hydrocarbons (PAHs)
30. Kelley, I., and C.E. Cerniglia (1990). The metabolism of fluoranthene by a species of
Mycobacterium. J. Ind. Microbiol. (in press).
31. Lewis, D.L., R.E. Hodson and L.F. Freeman (1984). Effects of microbial community
interactions on transformation rates of xenobiotic chemicals. Appl. Environ. Microbiol., 48:
561-565.
32. Mahaffey, W.R., D.T. Gibson and C.E. Cerniglia (1988). Bacterial oxidation of chemical
carcinogens: Formation of polycyclic aromatic acids from benz[a]anthracene. Appl. Environ.
Microbiol., 54: 2415-2423.
33. Martelmans, K.S. Haworth, T. Lawlor, W. Speck, B. Tainer and E. Zeiger (1986).
Salmonella mutagenicity tests II. Results from the testing of 270 chemicals. Environ.
Mutagen., 8 (Suppl. 7): 1-119.
34. Means, J.C., S.G. Ward, J.J. Hassett and W.L. Banwart (1980). Sorption of polynuclear
aromatic hydrocarbons by sediments and soils. Environ. Sci. TechnoL, 14: 1524-1528.
35. Mihelcic, J.R. and R.G. Luthy (1988). Degradation of polycyclic aromatic hydrocarbon
compounds under various redox conditions in soil-water systems. Appl. Environ. Microbiol.,
54: 1182-1187.
36. Mihelcic, J.R. and R.G. Luthy (1988). Microbial degradation of acenaphthene and
naphthalene under denitrification conditions in soil-water systems. Appl. Environ.
Microbiol., 54: 1188-1198.
37. Miller, E.G. and J.A. Miller (1981). Searches for ultimate chemical carcinogens and their
reactions with cellular macromolecules. Cancer, 47: 2327-2345.
38. Morehead, N.R., B.J. Eadie, B. Lake, P.D. Landrum and D. Berner (1986). The sorption of
PAH onto dissolved organic matter in Lake Michigan waters. Chemosphere, 15: 403-412.
39. Mueller, J.G., P.J. Chapman, B.O. Blattmann, and P.H. Pritchard (1990). Isolation and
characterization of a fluoranthene-utilizing strain of Pseudomonas paucimobilis. Appl.
Environ. Microbiol., 56: 1079-1086.
40. Mueller, J.G., P.J. Chapman, P.H. Pritchard (1989). Action of a fluoranthene-utilizing
bacterial community on polycyclic aromatic hydrocarbon components of creosote. Appl.
Environ. Microbiol., 55: 3085-3090.
41. National Academy of Sciences (1983). Polycyclic aromatic hydrocarbons: Evaluation of
sources and effects. National Academy Press, Washington, B.C.
42. Nicholas, R.B. (1987). Biotechnology in hazardous waste disposal: An unfulfilled promise.
ASM News, 53: 138-142.
43. Shiaris, M.P. (1989). Seasonal biotransformation of naphthalene, phenanthrene and
benzo[a]pyrene in surficial estuarine sediments. Appl. Environ. Microbiol., 55: 1391-1399.
44. Skerman, V.B.D. (1967). A guide to the identification of the genera of bacteria, 2nd ed.
The Williams & Wilkins Co., Baltimore.
45. Williams, P.A. (1981). Genetics of biodegradation. In: Microbial degradation of
xenobiotics and recalcitrant compounds, T. Leisinger, R. Hutter, A.M. Cook, and J. Nuesch
(eds.), Academic Press, Inc., New York, pp. 97-130.
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C.E. Cerniglia
99
PAHs
CD
O
C
03
J.J
ctf
O
CD
DC
Solubility
mg/L
31.7
.07
1.3
.002
.003
DA - ONA adducts
CA - chromosomal abberations
Ames - Salmonella typhlmurlum reversion assay
Genotoxicity &
Carcinogenicity
-+- Ames
+ SCE
+Ames
+ UDS
+ SCE
+ CA
+ Carcinogen
+ CA
+ DA
4- Carcinogen
SCE - sister chromated exchange
UDS - unscheduled DNA synthesis
Figure 5.1.1. The structures and chemical and toxicological characteristics of polycyclic aromatic
hydrocarbons.
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100
Polycyclic Aromatic Hydrocarbons (PAHs)
PAH
Volatilization
Photooxidation
Sedimentation
Chemical
Oxidation
Bloaccumulatlon
Initial
Degradation
(Biotransformation)
Detoxification?
CO2
Complete Mineralization
Removal
Figure 5.13. Schematic representation of the environmental fate of polycyclic aromatic
hydrocarbons.
-------
C.E. Cerniglia
101
Aerobic Metabolism
of Aromatic Hydrocarbons
Cis-Dihydrodiol
Catechol
Catcher! 2.3-Oxygenase
Catechol 1 ,2
Oxygenase
Figure 5.1.3. Major pathways of bacterial oxidation of polycyclic aromatic hydrocarbons.
-------
102
Polycyclic Aromatic Hydrocarbons (PAHs)
fcaKSte
Figure 5.1.4. Photograph of Afyco&acterium sp. colonies on MBS agar containing low-levels of
nutrients and coated with pyrene. The clear zones around the bacterial colonies
indicate pyrene utilization.
-------
C.E. Cerniglia
103
120
100
CjcT
o~-
-o 80
o>
£4
1 60
20
0
3 Methyteholantbrene
6-Nltrochrysene
1-Nitropyrone
Phanantiireno
Naphthalene
Mineralized PAHs (%)
Figure 5.1.5. Mineralization of naphthalene, phenanthrene, pyrene, fluoranthene, 1-nitropyrene,
6-nitrochrysene, and 3-methylcholanthrene by the Mycobacterium sp.
-------
104
Polycyclic Aromatic Hydrocarbons (PAHs)
Dioxygenase
1802
H
18,
PyreneX ^
1802\
OH
c/s-4,5-Pyrene-
dihydrodiol
Monooxygenase
Epoxide
Hydrolase
H2O
Pyrene-4,5-Oxide
CO
OH
frans-4,5-Pyrene-
dihydrodiol
Figure 5.1.6. The pathways utilized by the Mycobacterium sp. for the oxidation of pyrene.
-------
C.E. Cerniglia
105
8 1
Naphthalene
Q f Dioxygenase
OH
cis -1,2-Dihydroxy-
1,2-dihydronaphthalene (20)
Salicylic acid
1
o
-
Calechol
I
Ring Cleavage
I
CO2
Naphthalene-1,2-oxide
i
H OH
H
trans -1 ,2-Dihydroxy-1 ,2-dihydro-
naphthalene(l)
Figure 5.1.7. The pathways utilized by the Mycobacterium sp. for the oxidation of naphthalene.
-------
106
Polycyclic Aromatic Hydrocarbons (PAHs)
8 9
5 4
9-Fluorenone-1 -carboxylate
Figure 5.1.8. The pathways utilized by the Mycobacterium sp. for the oxidation of fluoranthene.
NO
NO,
6 5
1 -Nitropyrene
O-Dihydro-9,1 O-
Dihydroxy-1 -Nitropyrene
NO,,
" OH
c/s-4,5-Dihydro-4,5-
Dihydroxy-1 -Nitropyrene
Figure 5.1.9. The pathways utilized by the Mycobacterium sp. for the oxidation of 1-nitropyrene.
-------
C.E. Cerniglia
107
No Mycobacterium
With Mycobacterium
28 Days
Figure 5.1.10.
Mineralization of phenanthrene, 2-methylnaphthalene, pyrene, and
benzol a Ipyrene in microcosms from De Gray Reservoir sediments and water
with and without Mycobacterium inoculation.
-------
108 Polycyclic Aromatic Hydrocarbons (PAHs)
5.2 Fungal Degradation of Hazardous Wastes
John A. Glaser
United States Environmental Protection Agency
Risk Reduction Engineering Laboratory
26 W. Martin Luther King Dr.
Cincinnati, Ohio 45268
Detoxification of hazardous wastes is important as a means to reduce the risk associated
with such waste. The potential of biological processes to detoxify hazardous waste is beginning
to be recognized. The ability to degrade/detoxify organic and inorganic waste constituents
requires two complementary features: microbial competence to degrade target pollutants and
effective contact between the biomass and the pollutants. The competence of an organism is
best understood in terms of the biochemicals (enzymes) that enable the organism to convert the
contaminant chemical to a non-toxic end product. Due to the variety of possible chemicals
forming pollutant mixtures, either a single organism of exceptional competence or a multiplicity
of compatible organisms of complimentary competence is necessary.
The bioavailability or contact between the organism and the pollutant substrate is a
function of mass transfer and the most appropriate reactor conditions that maintain the
activity of the selected degrading organisms. Selection and development of reactor
configurations and operating conditions are necessary to sustain economic operation of the
treatment.
An example of a single organism of high competence is the wood degrading fungus,
Phanerochaete chrysosporium. This fungus possesses great potential to degrade aromatic
components of toxic and hazardous waste, based on the widely recognized ability to degrade
lignin, a persistent biogenic polymer. The degradation of lignin required largely non-specific
enzyme systems to accomplish this remarkable biodegradation. The fungus is also non-
pathogenic to plants and animals permitting possible application of this technology to solve a
variety of environmental contamination problems. The related development of the fungal
biomass as a food supplement for livestock serves to underscore the potential utility and benign
aspects associated with this organism.
The current catalog of pollutants degraded by this fungus ranges from polynuclear aromatic
hydrocarbons, polychlorinated biphenyls, pesticides to dyes. The wide range ability of this
organism to degrade these diverse pollutant classes is a tribute to the activity of the enzymes
systems secreted.
Two areas of liquid and soil treatment have been investigated recently. The liquid
treatment shows promise and is under pilot scale evaluation. However, the application to soil
contamination has shown the most exciting success in the last year.
The ability of the organism to treat under field conditions has recently been evaluated. An
Oshkosh, Wisconsin site was selected for field tfial applications of the white rot fungal
treatment to contaminated soil. The area of application was a former "tank farm" where
above-ground storage tanks contained a wood preservative formulation known as "Woodlife".
The composition of this product was predominantly mineral spirits (high boiling pentanes and
hexanes) and 5% pentachlorophenol. Extensive screening for pentachlorophenol in the tank
farm identified concentrations of 1 to 4435 mg/kg to depths of 30 cm. Within the confines of a
protective berm for the tank farm, the field trial study was laid out according to a specific
treatment design. After thorough mixing of the soil, nine plot borders were installed in a
three-by-three configuration. The plot borders were constructed deep. Plot borders were
worked into the soil surface and filled to a depth of 25 cm with soil outside the border.
Approximately 370 kg (dry weight) of soil were added to each plot. Two different fungi (P.
chrysosporium and P. sodida) were selected as candidate treatment species. In each case, the
-------
JJL Closer 109
fungi were added to the contaminated area through the use of inoculated wood chips with the
appropriate fungal species. The treatability trial began in early August 1989 and continues to
the end of September 1989. The pentachlorophenol concentration was depleted by 82% and
85% respectively, for the two fungal species after 46 days of treatment.
The investigation of field utility of this organism will continue to be pursued. The scope of
the fungal treatment is not limited to the currently selected series of pollutants under study.
Additional pollutant classes such as PCBs, pesticides and herbicides will be explored under
field conditions to determine the general utility of this organism. Development of the best
reactor configuration for field use and maintenance of the organism's biodegrading activities is
underway.
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110 Polycyclic Aromatic Hydrocarbons (PAHs)
5.3 Recent Studies on the Microbial Degradation of PAHs and Their
Relevance to Bioremediation
James G. Mueller1, Peter J. Chapman2,
Suzanne E. Lantz1 and P. Hap Pritchard2
(presented by John E. Rogers3)
Southern BioProducts, Inc., Gulf Breeze, Florida1,
U.S. EPA Environmental Research Laboratories at
Gulf Breeze, Florida2 and Athens, Georgia3
Polycyclic aromatic hydrocarbons (PAHs) are an ubiquitous class of chemicals whose
presence in the environment can be attributed to a number of natural and anthropogenic
sources (13). While the majority of these chemicals are innocuous, several, especially the
higher-molecular-weight (HMW) PAHs, have been shown to exhibit adverse health effects. A
number of areas contaminated with this class of chemical (i.e., coal gasification sites, petroleum
refineries, creosote vvaste sites) often contain sufficient amounts of HMW PAH carcinogens and
other toxic chemicds to pose a significant threat to environmental and human health.
Biological degradation represents a major route through which PAHs and many other
organic chemicals are removed from contaminated environments. By and large, lower-
molecular-weighc PAHs containing 2 or 3 rings are readily degraded biologically (1,5), and the
catabolic pathways for the degradation of these compounds by certain organisms have been
established (2,3,4). Conversely, HMW PAHs are less readily biodegraded and do persist in
contaminated environments. Consequently much less is known of the microbiology and
biochemistry of their degradation. This dearth of information is of particular concern since
HMW PAHs represent the greatest risk to public and environmental health.
Because HMW PAHs are less amenable to microbial attack, their removal from
contaminated environments has proven to be especially difficult for bioremediation technologies.
However, for bioremediation to be considered as an acceptable remedial action alternative for
these types of wastes, biotreatment processes must prove to be capable of destroying these
chemicals in a reliable, timely and predictable manner. To this end, efforts were undertaken to
isolate microorganisms capable of degrading HMW PAHs. These studies resulted in the
discovery of the first axenic bacterial cultures which utilized HMW PAHs as sole sources of
carbon and energy for growth (6,7). Moreover, complete mineralization of a number of these
compounds has been demonstrated (8).
Making use of this new source of novel biocatalysts, a multi-phasic biological treatment
strategy has been developed which effectively integrates physical separation technology
(membrane extraction) with microbial degradation processes. Recent bench-scale studies have
evaluated the effectiveness of a tri-phasic treatment approach (Figure 5.3.1) for remediation of
creosote-contaminated soil and sediment present at the American Creosote Works Superfund
site at Pensacola, Florida: soil washing (phase 1), membrane extraction/pollutant fractionation
(phase 2) and biodegradation (phase 3). A bi-phasic approach comprising membrane extraction
followed by biodegradation of concentrated organics was also evaluated. Performance data from
these studies clearly demonstrate the superiority of the multi-phasic biotreatment strategy over
conventional biotreatment approaches such as land-farming, slurry-phase and in situ
bioremediation (9,10,11,12).
-------
J. Mueller
REFERENCES
1. Bossert, I. and R. Bartha (1986). Structure-biodegradability relationships of polycyclic
aromatic hydrocarbons in soil. Bull. Environ. Contamin. Toxicol., 37: 490-495.
2. Cerniglia, C.E. (1984). Microbial metabolism of polycyclic aromatic hydrocarbons. Adv.
Appl. Microbiol., 30: 31-71.
3. Cerniglia, C.E. and M.A. Heitkamp (1989). Microbial degradation of polycyclic aromatic
hydrocarbons (PAH) in the aquatic environment. In: Metabolism of PAH in the Aquatic
Environment, U. Varanasi (ed.), CRC Press, Boca Raton, Fl. pp. 41-68
4. Gibson, D.T. and V. Subramanian (1984). Microbial degradation of aromatic hydrocarbons.
In: Microbial Degradation of Organic Compounds, D. Gibson (ed.), Marcel Dekker, Inc.
New York, pp. 181-252.
5. McGinnis, G.D. , H. Borajani, L.K. McFarland, D.F. Pope, D.A. Strobel and J.E. Mathews
(1988). Characterization and laboratory soil treatability studies for creosote and
pentachlorophenol sludges and contaminated soil. EPA 600/2-88/055.
6. Mueller, J.G., P.J. Chapman and P.H. Pritchard (1989). Action of a fluoranthene-utilizing
bacterial community on polycyclic aromatic hydrocarbon components of creosote. Appl.
Environ. Microbiol., 55: 3085-3090.
7. Mueller, J.G., P.J. Chapman, Beat 0. Blattmann and P.H. Pritchard (1990). Isolation and
characterization of a fluoranthene-utilizing strain of Pseudomonas paucimobilis. Appl.
Environ. Microbiol., 56: 1079-1086.
8. Mueller, J.G., P.J. Chapman, S.E. Lantz, B.C. Blattmann, and P.H. Pritchard (1990).
Mineralization of fluoranthene by Pseudomonas paucimobilis strain EPA505 and
identification of biotransformation products. Appl. Environ. Microbiol., (submitted).
9. Mueller, J.G., S.E. Lantz, B.C. Blattmann and P.J. Chapman (1990). Bench-scale
evaluation of alternative biological treatment processes for the remediation of creosote
contaminated materials: solid-phase bioremediation. Environ. Sci. Technol., (submitted).
10. Mueller, J.G., S.E. Lantz, B.O. Blattmann and P.J. Chapman (1990). Bench-scale evaluation
of alternative biological treatment processes for the remediation of creosote-contaminated
materials: slurry-phase bioremediation. Environ. Sci. Technol., (submitted).
11. Mueller, J.G., D.P. Middaugh, S.E. Lantz, and P.J. Chapman (1990). Biodegradation of
creosote and PCB in contaminated groundwater: chemical and biological assessment. Appl.
Environ. Microbiol. (submitted).
12. Middaugh, D.E., J.G. Mueller, R.L. Thomas, S.E. Lantz, M.J. Hemmer, G.T. Brooks and
P.J. Chapman (1990). Detoxification of creosote-contaminated groundwater by
ultrafiltration: chemical and biological assessment. Arch. Environ. Contam. Toxicol.
(submitted).
13. National Academy of Sciences (1983). Polycyclic aromatic hydrocarbons: Evaluation of
sources and effects. National Academy Press, Washington, D.C.
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112
Polycyclic Aromatic Hydrocarbons (PAHs)
CONVENTIONAL
SOIL
WASHING
SBP
MEMBRANE
EXTRACTION
SBP / EPA
BIOREACTOR
Figure 5.3.1. Tri-phasic treatment approach
-------
H.J. van Veen 113
5.4 Biological Remediation of Contaminated Sediments in the
Netherlands
H.J. van Veen and G.J. Annokkee
T.N.O., Apeldoorn
The Netherlands
INTRODUCTION
In the Netherlands, contaminated sediments are manifest as an environmental problem in
a dual way:
- As contaminated aquatic soil with the corresponding environmental impact
- As a dredged-sludge problem because many watercourses in the Netherlands must be
dredged for nautical and for water management reasons.
The dredged-sludge problem is currently dominating. This means that remediation of
contaminated sediments in the Netherlands refers to dredged-sludge remediation in particular.
Until a few years ago, all dredged sludges were increasingly being processed in order to
improve their quality. In this respect, a number of harbors had been remediated that were
seriously contaminated with PAH, oil, and metals, specifically. Their remediation included the
dredging and processing of the sludge by means of classification and dewatering into a fraction
for beneficial use and into a concentrate to be disposed.
TNO is one of the institutes that carries out research to improve the remediation
technology. This research into the processing of dredged sludge takes place within a program
in the order of approximately $2 million (U.S.) for 1990. Also participating in the research
project are government, trade, and private companies.
The outlines of the program are as follows:
- Optimization of environmental dredging. The purpose is to remove contaminated
sediments selectively.
Classification of dredged sludges into fractions with different contaminant
concentrations.
Biological, chemical, and physical treatment of the sludge aimed at immobilization of
the contamination.
This paper gives a survey of the current state of full-scale aquatic soil remediation in the
Netherlands and the development of biological remediation of dredged sludges.
CURRENT STATE OF CONTAMINATED SEDIMENT REMEDIATION IN PRACTICE
Since 1985, technology has been applied to restrict the quantitative volume of contaminated
dredged sludge to be disposed of. The process applied, consists of a combination of two
techniques: hydrocyclones and dewatering. In this way a relatively clean fraction is separated
from the dredged sludge while the residual fraction is reduced in volume as much as possible.
-------
114 Polycyclic Aromatic Hydrocarbons (PAHs)
Hydrocyclones
Particle classification is carried out by hydrocyclones (Figure 5.4.1). A hydrocyclone has
one inlet, two outlets, the vortex finder, and the apex nozzle. The outlet flows are called
overflow and underflow. The fluid feed enters the cyclone tangentially bringing about a
downward flow that circulates near the wall of the cyclone. The flow reverses near the apex
into an upstream in the center of the cyclone and leaves the cyclone by way of the vortex
finder.
When a heavy particle enters the feed, the downward flow moves this particle by
centrifugal force to the wall of the cyclone. The particle leaves the cyclone through the apex.
A less heavy particle does not have enough time to reach the wall of the cyclone; thus, it
leaves the cyclone together with the larger part of the water in the overflow. In this way, a
hydrocyclone classifies dredged sludge into heavy sand particles, on the one hand, and into
fines and organic material, on the other hand.
Fines and organic material have a high contaminant content compared with sand, on
account of the differences in sorption properties. This means that hydrocyclones separate a
relatively clean sand fraction from the slime fraction in which a concentration of contaminants
is found.
The effect of hydrocyclones is characterized by two aspects: the distribution of the dry
matter [E^]1 and the distribution of the contaminant [EJ2.
The applicability of hydrocyclones for the treatment of contaminated dredged sludge was
recognized as early as 1983. The technique has been applied in a number of dredging
operations, but does not always offer a solution, in particular, not for dredged sediments with a
high content of very small particles and high organic matter content (peat). Figure 5.4.2 shows
a number of results obtained in hydrocyclone experiments with dredged sludge from various
sites, as well as with various contaminants. The effect of hydrocyclones is more favorable as
the data point is closer to the origin of the diagram. From the figure it appears that
hydrocyclones often give good results, but not always.
Dewatering
There is various dewatering equipment. Three apparatuses qualify for the dewatering of
dredged sludges and of the slime fraction of dredged sludges: the belt press, the filter press,
and the decanter. In general, it can be said that the filter press results in the highest dry-
matter content, whereas the decanter results in the lowest dry-matter content. The use of
flocculants is, in most cases, necessary for dewatering. When a belt press and filter press are
applied, flocculants bring about a good filterability; in the case of a decanter, flocculants help
in reaching a clear decantate. All three apparatuses mentioned are applicable to practical
dredged sludge treatment.
The purpose of dewatering is to reach a volume reduction of the sludge or slime fraction
produced by hydrocyclones. Figure 5.4.3 shows the effect of dewatering on volume, starting
from a slime fraction with a dry-matter content of 5% after using hydrocyclones. The figure
shows that as the dry-matter content increases, a considerable volume reduction is reached in
the first instance. However, at higher dry-matter contents (approximately 40%), the volume
decreases less strongly at increased dry-matter contents.
Since dewatering is aimed at volume reduction, it appears from this figure that further
dewatering becomes less cost effective. Dewatering costs increase strongly as a higher dry-
matter contents are reached.
separation efficiency for the dry matter; this is the percentage of the dry matter that leaves the
hydrocyclone as underflow (sand fraction)
E, = separation efficiency for the contaminants; this is the percentage of the contaminants that leave the
hydrocyclone with the underflow
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H.J. van Veen
Cases
Table 5.4.1 gives the results of a number of practical cases of the treatment of dredged
sludges by hydrocycloning/dewatering. It appears that, on a practical scale, favorable results
have been reached by using hydrocyclones and dewatering.
In a number of remediation cases that are currently being carried out, the following
bottlenecks have been ascertained:
- It turns out that the information obtained in the preliminary investigation strongly
differs from the real situation. For instance, the composition of the soil strongly
deviates from the composition expected on the basis of the preliminary investigation.
This makes it difficult for the contractor in charge of the remediation to meet the
results described in his quotation.
- After using hydrocyclones, the sand fraction, in some cases, still shows a high PAH
concentration because the PAHs are not adsorbed to the slime fraction, but are present
as some kind of tar particle that can hardly be separated from the sand.
Future research will pay considerable attention to these two bottlenecks.
BIOLOGICAL REMEDIATION
General
Dutch research into biological remediation is particularly focused on the biodegradation of
oils and PAHs because these organic micropollutants occur most frequently. TNO has carried
out laboratory-scale exploratory research into the biological remediation of the dredged sludges
contaminated with mineral oils and PAHs (Table 5.4.2). This research has shown that effective
biological cleaning is possible for a number of dredged sludges. Spontaneous degradations have
been found in these dredged sludges, if the conditions for these sludges are biologically
favorable (as in the case in a bioreactor). From a biological point of view, such a degradation
often goes by quickly.
Present research is done along two lines:
1. Development of biological remediation techniques up to a practical scale. This concerns
the development of designs for the biodegradation process that link up with the
dredging process.
2. Broadening of the fundamental knowledge pertaining to the degradation of PAHs and
other substances such as chlorinated hydrocarbons.
At present, Dutch research emphasizes the former line.
For the practical application of biological remediation TNO has three treatment ways in
view:
1. Large scale, extensive treatment in aeration basins
2. Intensive treatment in bioreactors
3. Landfarming
These three ways of treatment, together, form a complete process for remediation contaminated
dredged sediments. The differences in dredged sediments refer to: granular composition,
distribution of the contaminants among the particle size fractions, contaminant content, and
-------
116
degradation rate.
Polycyclic Aromatic Hydrocarbons (PAHs)
COMPLETE DIAGRAM FOR THE BIOLOGICAL REMEDIATION
OF CONTAMINATED DREDGED SEDIMENTS
Pretreatment by classification
I
Hydrocyclones
Sand fraction (underflow)
usually slightly contaminated
Contaminated Dredged Sediments
I
I
No Classification
(complete dredged sediment)
I
I
I
Bioreactor (intensive)
Slime fraction (overflow)
usually strongly contaminated
a.
b.
Bioreactor (intensive)
Landfarming (extensive)
Aeration basin (extensive)
This diagram is based on the following major arguments:
* No Classification
Due to its physical behavior, the non-separated (i.e., "complete") sediment can be treated in
a bioreactor only. It often contains too many fine particles for landfarming, in other
words, its porosity is too small. For an aeration basin there are too many coarse particles
which can hardly be brought into suspension.
* Pretreatment by classification
- The most important reason for classification is that it results in two fractions which can
be treated separately very well; whereas, this is not true for the original dredged
sludge.
- Depending on contaminant content, the (usually) slightly contaminated fraction can have
a direct beneficial use or can be remediated by way of bioreactors or landfarming. The
advantage is that part of the dredged sediment can be treated in a relatively short
time. Landfarming demands a coarse particle size due to the high porosity needed.
The sand fraction has these characteristics.
The slime fraction can be treated as a liquid; consequently, it can be remediated as
waste water. Therefore, an aeration basin is a large-scale possibility.
Apart from the above aspects pertaining to the composition of the dredged sediment which
determines the treatment method to be applied, factors of a pragmatic character and local
conditions play an important part in the treatment method to be chosen, such as:
* Available space.
—» If there is sufficient space available, an aeration basin can be considered. Such an
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H.J. van Veen
aeration basin demands a large surface area; depending on the quantity of dredged
sediment to be treated, this may approximate some tens of thousands m2. If there is
not sufficient space available, the bioreactor offers a possibility (up to a maximum of
some hundreds m2).
* Available time.
-» If the remediation should take place within a short period of time (and if the
degradation rate is sufficiently high), bioreactor treatment is the appropriate method.
-> If the remediation may take a relatively long period of time (from months up to one or
two years), there are possibilities for large-scale, extensive methods (landfarming,
aeration basin).
* Quantity of dredged sediment to be remediated.
-> If the remediation involves a relatively small quantity of dredged sediment, a bioreactor
can be used.
-> If the remediation involves a considerable quantity, it is necessary to apply a large
scale method.
The biological treatment methods (bioreactor, aeration basin, and landfarming) mentioned
above are all subject to investigation.
Intensive Versus Extensive Treatment Methods
Practical biodegradation offers a choice between intensive and extensive methods.
Intensive implementation methods
An intensive implementation method is aimed at:
operating a process as intensively as possible (with much exertion)
thus realizing conditions as optimum and verifiable as possible
resulting in as short as possible a treatment period.
These implementation methods refer to process-type treatment methods (e.g. bioreactor)
Extensive implementation methods
These implementation methods are meant to:
operate remediation methods with relatively slight exertion (extensive)
- usually implying less optimum and verifiable conditions.
These implementation methods refer to large-scale, more or less batchwise treatments, like
biodegradation by means of landfarming and/or treatment as a slurry in an aeration basin.
Whether an intensive or an extensive way of implementation is chosen for remedial
operations, it is determined by a number of choice criteria. Below, by means of some
important choice criteria more detailed grounds are given as to why, in some cases, the
application of large-scale, extensive ways of implementation can be an alternative for intensive
(process-type) methods.
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118 Polycyclic Aromatic Hydrocarbons (PAHs)
* Period of time
The period in which a certain quantity of dredged sediment can be cleaned depends on the
capacity of the remediation method applied. For extensive methods the relevant quantity of
dredged sediment to be cleaned is treated in one batch during a long time. Intensive methods
of implementation may involve short treatment times, but remediation plants have a relatively
limited capacity. In fact, practice will show that the remedial operation for a reasonably sized
quantity of dredged sediment also takes a long time by way of intensive methods. This means
that the application of intensive implementation methods does not necessarily have much
advantage over extensive methods, as far as period of time is concerned.
* Costs
In view of the simplicity of extensive implementation techniques and the small exertion
needed, it is expected that these techniques will cost less than intensive implementation
techniques. An important condition in this respect is that the costs of building a facility in
which the large-scale remediation can take place (e.g. depot or basin) cannot be fully taken into
account in the remediation costs, because:
the depot/basin has to be built anyway to store the dredged sediment; in that case, the
depot/basin will not be built just for the remedial operation
The depot/basin can be used various times to remediate dredged sediments from
different sites.
With respect to the construction of an aeration basin, in practice it is possible for such a
basin to be part of the harbor that is screened off from the rest. In that case, the costs are
expected to be considerably lower than for a new basin.
* Space needed
Much space is needed for extensive implementation methods, in contrast to intensive
methods. If this space is available, extensive implementation methods are a reasonable
alternative.
* Implementation in dredging operations
In the Netherlands the remediation of dredged sludge is carried out by dredging companies
and contractors. Past experience has demonstrated that implementation of new technology
leads to great problems within companies. Extensive implementation methods are more
compatible with the factory management.
Figure 5.4.4 gives the results of laboratory experiments carried out with respect to the
biodegradation of PAHs in a Rotterdam harbor sediment (i.e., 'Geulhaven' sediment). The
following three treatments were carried out:
1. Treatment of the original sample in a bioreactor. The laboratory bioreactor is a
rotating drum with baffles, with a contents of approximately 10 liters.
2. After hydrocyclones, the slime fraction of the 'Geulhaven' sample was treated in an
aeration column; the material was aerated four times a day for one hour, with a total
quantity of 25 m3 air/m3 suspension per day. This is approximately 10% of the air
quantity fed into a biological sewage treatment.
3. Treatment of the sand fraction in a laboratory landfarm. A 20 cm thick layer of sand
fraction was put into aim2 tray. About once a month the sand was mixed with a
hand shovel. To prevent the sand from drying out, it was moistened every week
resulting in a dry matter content of approximately 80-90%.
-------
H.J. van Veen 119
From Figure 5.4.4, it is quite evident that intensive bioreactor treatment goes by quickly.
After a longer period, however, high degradation percentages are also reached by way of
extensive methods. The scale for these laboratory experiments is 10 m3.
TNO's Concept for Extensive Bioremediation of Dredged Sludge
Based on results, some of which are mentioned in this paper, a plan has been developed
for the extensive treatment of dredged sludges. This plan comprises the separation of the
dredged sludge by hydrocyclones, after which, the sand is treated in a landfarm, and slime in
an aeration basin.
Landfarmins of the sand fraction
Landfarming is a technique that is frequently applied in practical (terrestrial) soil
remediation. Much experience has been gained with respect to the degradation of mineral oils
and PAHs in particular. Briefly, the contaminated soil is put down in layers (20 - 50 cm
thick) in a field that is especially equipped for this purpose. Care is usually taken for:
Manuring
Good water balance (draining or watering)
Oxygen supply by means of tillage (plowing, harrowing, working with a rotary cultivator)
Increasing the porosity-increasing means (such as peat and bark) for a better oxygen and
water balance.
Sometimes,
Inoculation with special cultivation or activated sludge takes place, as well as
Temperature increase by means of leading steam or hot water through pipes, or
constructing a covering of transparent plastic foil (greenhouse).
Treatment times depend on contaminant content and vary from six months to two years.
Aeration basin for the treatment of the slime fraction
The size of the basin is determined by the quantity of suspension to be treated and the
treatment period. The quantity of suspension (overflow of the hydrocyclone) depends on the
quantity of dredged sediment to be cleaned; usually a minimum of 1,000 m3 suspension per site
is assumed. From research it can be deduced that the treatment period for this extensive
method will be at least some months. This means that basins of some thousands to some tens
of thousands m3 are needed. In this respect one should think of:
(temporary) depots that are dug or surrounded by earthen dikes
screened off part of the relevant site.
For the aeration of basins up to a size of ten thousand m3, the distribution of air within
the basin is an important aspect. In this respect, a comparison is made with an aeration basin
of a sewage treatment plant, where as short as possible a treatment time is strived after. This
means:
a. Installation of aeration elements across the whole surface of the aeration space, and
b. A sufficient mixing of the waste water (turbulence).
-------
120 Polycyclic Aromatic Hydrocarbons (PAHs)
These two conditions are not considered feasible for aeration basins that have to treat the
slime fraction, in view of the size of such basins. Taking care that there is sufficient
turbulence and oxygen for the aeration basin is considered non-realistic. The TNO plan
considers the installation of a large-scale treatment depot for the slime fraction, with
intermittent aeration. This aeration is realized by moving a pontoon with aerators and mixers
slowly to and fro across the length of the basin.
FINAL REMARKS
Further research into biological remediation will incorporate the following:
- Together with trade and industry, further auxiliary research into the scale-up of the
techniques presented in this paper
- Further fundamental research into biodegradation. This aspect will be considered by
both TNO and universities.
In the Netherlands, the introduction of treatment technology for contaminated sediments in
dredging operations has started only recently. The introduction of relatively simple techniques,
such as hydrocyclonage, already appears to cause many problems. These problems are among
other things the result of :
An inadequate preliminary survey of the site to be dredged; in this way remediation
plans are based on incomplete information which later turns out to be incorrect.
The dredging companies underestimating the degree of complexity of the remediation
technology.
The research results being scaled up too quickly to a practical scale, researchers
underestimate the implementation problems.
In the Netherlands, there is still a relatively large antitreatment lobby; it consists of
representatives of government and companies who do not consider the treatment of dredged
sludge worthwhile and want to dump everything. This lobby has intensified as a result of the
introductory problems. Therefore, it is of utmost importance to start with simple technology for
the development of remediation technology for contaminated sediments; only at a later stage is
a more sophisticated technology desirable. This is one of the most important reasons why we
are convinced of the feasibility of the TNO concept for biological remediation of contaminated
sediments.
-------
H.J. van Veen
121
Table 5.4.1. Results of Practical Hydrocyclone Applications
Project
Process (*)
Capacity (m3/h)
Separation
Contaminants
diameter (micron)
Eto
E,
Concentration
in sand (mg/kg)
Barendrecht
1985
1
20
20
metals
oils
50%
metals: ± 15%
Zn: 169
Cu: 28
Cd: 1.8
Roozendaal
1986
2
18
50-60
metals,
oils
20%
metals: 1-5%
oils: ± 0.5%
Zn: 63
Cu: 18
Cd: 0.2
oils: 93
Nijerkerk
1986
3
300
50-60
PAH
70%
PAH: 5-10%
PAH: 1-2.9
Dordrecht
1988
4
300
50-60
PAH,
metals
60%
PAH: ± 5%
oils: ± 10%
PAH: 0.38
Zn: 150
Cu: 38
Cd: 0.9
Volume reduction by
hydrocyclones/
dewatering
75
50
1. Test installation consisting of a storage basin, a preseparator (CBC = Circulation Bed Classifier), a
buffer basin and hydrocyclones.
2. Installation consisting of a hydrocyclone and a sieve belt press.
3. Installation consisting of a sieve, three hydrocyclones, a sediment tank and a sludge depot.
Flocculants have been dosed in the delivery pipe to the depot for a quick first sedimentation, thus
making a quick water drainage possible.
4. See 3. The sludge depot has been replaced by a flat-bottom craft in which the fine fraction has
settled.
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122
Polycyclic Aromatic Hydrocarbons (PAHs)
Table 5.4.2. Results of Biodegradation for Various Sediment Samples
PAH content mg/kg' Geulhaven
Scheveningen Dordrecht
Dodewaard
Amstel/Drecht
Original material
After 7 days
After 30 days
After 60 days
Oil Content mg/kg
Original material
After 7 days
After 30 days
Cleaning efficiency (%
PAH after 60 days
Oil after 7 days
Oil after 30 days
212
49
22
16
12000
1040
>)
92
91
351
237
188
175
2580
1027
50
60
817
217
232
145
2826
768
82
73
156
150
125
379
357
20
6
372
320
333
275
1372
1045
26
24
* 16 PAH (EPA)
-------
H.J. van Veen
123
feed
overflow
(slime fraction)
underflow
(sand fraction)
Figure 5.4.1. Hydrocyclone
-------
124
Polycyclic Aromatic Hydrocarbons (PAHs)
Pirt of contaminant
in underflow (Ex) %
Wialhaven
Geulhaven
Naarden
Anstel-Oretht-
kanaai
6 Kaflipen
Mm IRoermond)
8 Eenskanaal
9 Naordzetkanaal
10 Singtlgracht
11 Zaan
12 Arnhen
13 Schcveningen
Dordrecht
15 Dodevaard
Pirt of dry matter
in underflow (Ed.ra.)
Figure 5.4.2. Hydrocyclone results.
-------
H.J. van Veen
125
1000
volume
m
500
Based on:
- 1 m1 slime fraction
• 1m. content 5X
50
dry matter content (%)
100
Figure 5.4.3. Volume reduction by dewatering.
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126
Poly cyclic Aromatic Hydrocarbons (PAHs)
PAH conctntrjllon "H, 116 of EPA)
blor«act»r •critloi
lorlgtotl biiiln
sample) Ultat)
200
150.
100.
SO.
52* ,
500
(00
300
200
100
(und)
SO
X of orlglnil conte«tr«t)on
I
10
so.
\
\
sind friction to
landfir*
•rljlnal
b blor«actor
sttae (rittlon to
hassln
(0
120
tine Idifi)
200
Figure 5.4.4. Intensive versus extensive treatment (Geulhaven Rotterdam).
-------
6 METALS
6.1 Bacterial Leaching of Metals from Various Matrices Found in
Sediments, Removing Inorganics from Sediment-Associated Waters
Using Bioaccumulation and/or BIOFEX Beads
Paulette Altringer and Shane Giddings
U.S. Department of the Interior
Bureau of Mines
Salt Lake City Research Center
Biotechnology Group
729 Arapeen Drive
Salt Lake City, UT 84108
INTRODUCTION
The Bureau of Mines Salt Lake City Research Center (SLRC) has been conducting
biotechnological research for waste remediation over the past 5 years. Bacteria and
immobilized biomass are being used to remove heavy metals and toxic process chemicals from
solution; and bacteria are being applied to remediate mining and milling tailings and
sediments. Biotechnology is being used by itself and in conjunction with chemical treatments
to "polish" solutions to the stringent requirements imposed by environmental legislation.
Biotechnology may reduce contaminants to lower concentrations than those achievable using
chemical treatment, and may provide on-the-shelf technology for environmental problems
untreatable with conventional physical and chemical technology today.
The SLRC has considerable expertise in treating liquid hazardous wastes. Arsenic,
cadmium, cyanide, lead, mercury, selenium, and other commonly encountered toxic metals and
process chemicals have been removed from a wide variety of wastewaters using both
conventional and newly emerging technologies. Conventional techniques utilized include
chemical precipitation, ion exchange, and solvent extraction. Newly emerging technologies
under investigation at the SLRC include biosorption using viable biological materials such as
live bacteria and algae, and immobilized biomass. These latter techniques are currently being
addressed by ongoing projects, and have been particularly effective in treating liquid wastes
containing dilute concentrations of toxic metals.
Research is being expanded to include bioleaching of inorganic contaminants from sediments
and tailings using bacteria. This approach has potential to clean up one of the largest
contamination problems in the United States. Tailings are contaminating a large portion of the
waterways in the West, especially in Montana. We are currently investigating several
biotreatment techniques for the mixed tailings-soils along these arsenic contaminated
waterways. Bacteria have been identified which aid in leaching arsenic directly from
contaminated soils. Sediments are contaminated with man-made waste throughout the country,
and especially in the Great Lakes Region. The nature of these low-level, high-volume wastes
makes most processing options extremely expensive. Bacterial leaching in situ or on heap pads
may provide an answer to this wide-spread problem.
127
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128 Metals
ONGOING SLRC BIOTECHNICAL RESEARCH PROJECTS
Biohydrometallurgical Decontamination of Mining and Milling Waters
The biohydrometallurgy project develops new biotechnical techniques for decontaminating
mining and milling waters containing heavy metal ions and toxic process chemicals. This
includes using live bacteria to remove heavy metals, such as arsenic, cadmium, cobalt,
chromium, lead, selenium, and zinc from solution, as well as destroying cyanide in solution.
Chemical techniques are also investigated, but to a lesser degree, to enhance results.
Successful development of the biotechnical techniques will provide alternative remediation
technology to those available today for liquid wastes including those that must be treated
during heap leach closure.
Immobilized Extractant Technology for Wastewater Treatment
The immobilized extractant technology project is investigating procedures to immobilize
biological materials in polysulfone beads. The beads have excellent handling characteristics
and have been utilized to extract toxic metal ions such as cadmium, lead, and mercury from
wastewaters. Successful application of this technology will provide an innovative method for
removing and recovering heavy metals from a wide variety of mining and mineral processing
wastewaters.
Hazardous Wastes on Federal Lands
The federal lands project investigates remediation of hazardous wastes specifically under the
jurisdiction of the United States government. Sites under investigation include (1) the inactive
Midnite Mine on the Bureau of Indian Affair's Spokane Reservation contaminated with
uranium and radium, (2) the Olson-Neihart tailings just outside of Heber, UT, generated by a
lead-zinc operation, now under the jurisdiction of the Bureau of Reclamation, and (3) recent
involvement with the U.S. Forest Service on permitting gold operations on Forest Service lands
including closure technology. Development of this technology could help alleviate the wide-
spread hazardous waste problems on federal lands which the Federal Government must
remediate.
Technical Consultation and Support
This project provides technical consultation and support to the Bureau and other
cooperating agencies in assessment of techniques to decontaminate Superfund and metal mining
waste sites. This includes reviewing the technical credibility of recommendations relating to
Superfund Sites, such as those included in Environmental Impact Statements (EIS), Remedial
Investigations (RI), Feasibility Studies (FS), and Records of Decision (ROD). These manuscripts
are reviewed at the request of the Bureau's Washington Office.
Cooperative Efforts with Other Agencies and the Private Sector
Cooperative demonstrations of SLRC developed procedures for wastewater and solids
remediation are being conducted with various private and public agencies.
• Memorandums of Agreement (MOAs), for cooperative work, are being signed with the
mining industry for bacterial cyanide destruction in mine tailings ponds and in spent ores.
• A blanket MOA is in place between the Bureau of Mines and the U.S. Forest Service for
remediation of acid mine drainage waters. Part of the SLRC effort is using immobilized
biomass to remove heavy metals from waste waters. Another effort has begun for the
SLRC to review of EIS, RI, and FS documents for permitting gold mining and milling
operations including closure technology.
-------
P. Altringer and S. Giddings 129
• A blanket MOA has been in effect with the Bureaus of Mines and Indian Affairs (BIA), and
a specific MOA has been in effect between the SLRC and the BIA in Spokane, WA, for the
past 3 years for remediation of Midnite Mine including treatment of 500 million gal of
impounded water to meet NPDES limits, checking for the reactivity of rock impounded on
site, and reviewing draft closure technology.
• An MOA is in place with the Colorado Department of Natural Resources for using
immobilized biomass to remove heavy metals from waste waters.
• An MOA has been completed with the Bureau of Reclamation out of Sacramento, CA, for
bacterial removal of selenium from agricultural drainage waters.
• An MOA is nearing completion with the Bureau of Reclamation out of Provo, UT, for
determining chemical and biological oxidation of Olson-Neihart mine tailings (once proposed
to be proposed for the NPL list) during drying and relocating in a new impoundment
Heber, UT.
• Over the past 4 years we have reviewed treatment alternatives in various EIS, RI, FS, and
RODs at the request of our Washington Office under an MOA between the Bureau of Mines
and USEPA.
BIOREMEDIATION OF LIQUID WASTE
Background
The bioaccumulation of metals is the reverse reaction of bioleaching; instead of mobilizing
metals from minerals, microorganisms remove soluble metal ions from contaminated water.
Bioaccumulation has received considerable attention in the scientific community. Several
important conclusions are evident:
• Bioaccumulation is effective, often superior to conventional metal removal systems such as
solvent extraction or ion exchange.
• Bioaccumulation processes can be applied to a wide variety of metals. In this report,
cadmium, cobalt, nickel, zinc, and uranium will serve as prominent examples, although
other metals are known to be susceptible to bioaccumulation and will be mentioned where
appropriate.
• A variety of biological mechanisms play a role in these processes. In some cases, the
bioaccumulation of a metal involves the active uptake of the metal into the cell; in other
situations, passive adsorption of the metal to the cell wall may occur; still another
mechanisms deals with the complexing of metals by specific metabolic binding wastes such
as H2S. There are approximately a dozen recognized metal accumulating mechanisms.
• Metal removal processes can be devised to make use of live and dead cells. In many cases,
the use of bioaccumulation using non-viable cells is as effective or more effective than using
living cells.
• Bioaccumulation has a recognized, established use in the mining industry; Schist Lake,
Manitoba, and various other sites, make use of microbial systems to remove metal
pollutants (generated from mining activities) from surface waters.
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130 Metals
Specific Examples Cited in the Literature
Cadmium
Reports of environmental cadmium contamination of waters has dramatically increased over
the last quarter of a century. The Marathon Battery site, located near Buffalo, NY is a
notorious example. This heavy metal presents a variety of environmental problems; it is
among the more toxic elements, and is readily accumulated by many living organisms,
including man. John Poldoski, EPA ERL, studied the possibility of cadmium accumulation by
the microorganism Daphnia magna in the late 1970s (46); however, much of the cadmium
bioaccumulation research originated in England. In 1982, researchers at Oxford University
reported a novel approach for cadmium removal (38). These scientists selectively altered the
enzymatic pathway of a bacterial strain belonging to the enteric family. This strain of
Citrobacter was engineered to produce a cellular phosphatase at levels well above its normal
range. This enzyme cleaves organic phosphate, yielding HP042~ which then binds soluble
cadmium to the cell membrane as insoluble cadmium phosphate (39,42). The SLRC performed
some exploratory research on Marathon Battery Superfund sediments in 1986 (4). Research
focused on using such an enzyme system produced by Pseudomonas aeruginosa to remove heavy
metal ions from contaminated ground water.
The same enzyme-metal binding pathway has been reportedly used by the same bacteria for
lead (1,2,3), strontium (40), and uranium (41).
Other cadmium removal mechanisms exist. Two species of yeast accumulate cadmium
through a biosorption mechanism: Aureobadisium pullans (43) and Saccharomyces cerevisiae
(44). This biosorption mechanism is explored below in the section on cobalt and nickel
recovery.
Cobalt and Nickel
Cobalt has many important uses in industry and is considered a strategic metal. Cobalt is
frequently found in nickel-bearing deposits. To quote Brierley, "Advances in biomining
technology may make it possible to recover not only some of the nickel (worth $60 billion at
1982 prices) but also some of the cobalt ... the emphasis is not so much on rapid reaction rates
as it is on lower capital investment, greater recovery of metal, and reduced environmental
damage" (13). The need for increased bioaccumulation research is clear; as Brierley points out,
biological accumulation of metals may be relatively cheap and effective.
Investigation of cobalt bioaccumulation is not new. In 1954, Parker and O'Brien studied
the bioaccumulation of cobalt by Saccharomyces cerevisiae, common brewer's yeast. They found
a cobalt resistant strain that would accumulate the metal ion at uptakes of 10 pet of the dry
weight of the organism. The cobalt accumulating ability of S. cerevisiae was verified in 1977
by Norris and Kelly (44) as was mentioned in the previous section on cadmium.
Kuyucak and Volesky (28) reported a recovery system that used microalgae to remove cobalt
from solution. Their process is complex, but works extremely well. The mechanism,
biosorption, works as follows: The cell wall of many microorganisms is porous to allow uptake
of organic nutrients and trace minerals. Metal ions enter cells through the same porous
channels, where they can bond to a variety of anionic cellular components such as sulfhydryl
groups, phosphate groups, amino acids, or polysaccharides. These anionic cellular components
act as electrostatic magnets for a number of metal cations including nickel, lead, zinc,
chromium, copper, iron (28), uranium, thorium (48), germanium (17) and gold (29). Biosorption
is a phenomenon that has been linked to many bacteria and yeasts, as well as algae.
The rate of metal recovery using biosorbants compares favorably to ion exchange systems,
often working faster, and removing more metal from solution while costing less than the
conventional technology. As Brierley also notes, three metal removal systems were in effect in
1982 that made use of such biosorption mechanisms; at Schist Lake in Manitoba, at the New
Lead Belt in Missouri, and at the Grants Uranium District in New Mexico the mining industry
has used the accumulating ability of various microorganisms to remove a wide variety of
soluble, toxic metal cations (13).
Kuyucak and Voleski (28) also point out that the sorbed metals could be recovered through
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P. Altringer and S. Giddings 131
chemical stripping, allowing the metals to be sold to help offset costs. Their process also
illustrates that living and dead algal biomass could be reused, without losing its effective metal
binding properties. This supports results from the immobilized biomass bead research being
conducted at the SLRC today. This new technology may make Brierley's comments a reality.
Zinc
Zinc is a trace nutrient for most living systems; however, it has been demonstrated that a
number of microbial fungi can accumulate the metal. White and Gadd (50) explored the
uptake and intracellular accumulation of zinc by S. cerevisiae. They found that this useful
microorganism accumulates the metal in a two-step, energy dependent reaction.
The mechanism behind the accumulation of metals by S. cerevisiae has been studied; the
phenomenon may be related to the presence of an intracellular binding protein such as
metallothionine. These proteins are high in sulfhydryl groups and have significant metal
binding properties. Bacteria, yeasts, and molds are all capable of forming metallothionines in
response to the presence of heavy metals. The specific zinc binding metallothionine of S.
cerevisiae has been documented by researchers at the University of Utah (51).
A high affinity zinc accumulation system was demonstrated using the yeast Candida utilis
(20). The yeast cells were grown under abnormally low zinc conditions; however, when the cells
were exposed to higher levels of the trace element, they hyperaccumulated the metal - almost
10 times the normal level - up to 1 pet of the dry weight biomass.
Researchers at the Bureau of Mines, SLRC, have developed an effective bioaccumulating
system that removes zinc (and most other divalent metal cations) from aqueous solutions
(12,30). Their research has explored the sorptive powers of a number of biomass sources, both
living and dead. The procedure which they have developed works well and has been effectively
tested on the heavy metal tainted waters from near Leadville, Colorado. This work followed
the success of Darnall and others (18) who effectively removed zinc, uranium, barium, gold, and
other metals from solution using immobilized microalgae.
Uranium
Uranium has been mentioned throughout this report; much of the conventional uses of
bioaccumulation have focused on the recovery of this metal. Uranium, found principally as
hexavalent uranium (U+s present as UO22t), is among the easiest of metals to remove by
biosorption (18,49). Tsezos evaluated the uranium accumulating ability of a variety of molds
and bacteria; he applied the same technology to the removal of thorium and radium.
As mentioned previously, Macaskie and Dean (40) used the same system to remove
cadmium and uranium from waters with the uranium precipitating as cell-bound
uranylphosphate.
A third uranium removal system demonstrates a different way that microorganisms remove
metal cations from aqueous systems: by the production of metabolic wastes. Near Ambrosia
Lake, New Mexico, uranium mine discharge water is percolated through soil. The water
contains elevated molybdenum, selenate, and sulfate levels as well. The removal of the
minerals from the water is due to the presence of soil bacteria, most notably Clostridium and
Desulfovibrio. These bacteria metabolize the sulfate and selenate to hydrogen sulfide and
elemental selenium, respectively (27). The hydrogen sulfide reduces the uranium to insoluble
uranium dioxide and binds the molybdenum as insoluble MoS2. In this case, the treatment is
effective in reducing the level of all of the minerals below required levels.
The importance of this project is that conventional treatment technology was ineffective in
removing the metals from solution; however, applied microbiology did the trick. It can be
postulated, based on this model, that other soluble metal cations can be removed from solution
in a similar manner. Cadmium, cobalt, lead, mercury, nickel, tin, zinc, and other metals are
subject to the binding power of hydrogen sulfide, precipitating as insoluble metal-sulfides. With
a careful design, it would be possible to remove these metals from solution and recover the
metals for future refinement.
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132 Metals
Examples of SLRC Biotechnical Removal of Organics from Contaminated Waters
Bacterial Cyanide Destruction and Selenium Removal From Precious Metals Solutions and
Tailings
Ponded cyanide solutions in the environment pose a serious threat to migratory birds and
are responsible for several bird kills each year. An additional threat is posed by selenium,
which is also present in some of these solutions. While selenium is a necessary nutrient in
trace amounts, high concentrations cause death and deformities in wild fowl. The Bureau is
conducting research to reduce costs of cyanide destruction and selenium removal. Presentations
at both the TMS and SME Annual 1990 Meetings describing the Bureau's research on cyanide
and selenium removal from tailing waters generated considerable interest (5,8,34,35,36,37).
More problems exist with selenium than was originally thought. In a number of precious
metal operations, both cyanide and selenium are potential problems and operators need
information on how to deal with them. As an alternative to expensive chemical destruction of
the cyanide, the Bureau has cultured and isolated cyanide-destroying bacteria from toxic,
pH 10.5 precious metals tailings pond water containing 280 ppm CN. The cyanide-destroying
bacteria are Pseudomonas pseudoalcaligen.es and Pseudomonas diminuta. The bacteria have
been oxidizing 85 to 95 pet of the cyanide from two cyanide solutions obtained from different
industrial operations for over a year in a continuous system. Results from treating one water
are shown in Figure 6.1.1. In addition to degrading the cyanide, the bacteria also remove
other contaminants. Most of the iron, lead, nickel, and zinc are removed; however, copper,
selenium, and silver are not. Selenium is the major remaining contaminant. Selenium can be
chemically precipitated from solution using copious amounts of ferrous sulfate, upwards of 600
times the stoichiometric amount, to approach the drinking water standard of 10 ppb. Once
again, bacterial treatment was investigated, but no selenium-reducing bacteria were found in
this toxic precious metals solution. Luckily, earlier SLRC research involved the selenium-
contaminated waters of the Kesterson Reservoir, located in the San Joaquin Valley of California
(9,32,33). Of all the selenium-reducing bacteria isolated, Pseudomonas alcaligenes reduce
selenium fastest under anoxic conditions. After the cyanide is destroyed, these bacteria reduced
the selenate to selenite and then to elemental selenium which precipitates from solution as a
red amorphous mass. These promising results may provide effective technology for application
during heap leach closure for precious metals operations. This technique might also have
application to remediation of plating waste sediments and associated waters.
Arsenic Removal Using Anaerobic Bacteria
SLRC researchers isolated anaerobic bacteria that reduce arsenic and precipitate it from
solution. Continuous and batch tests are ongoing to optimize parameters and devise an
operational treatment system. Arsenic removal of 23 pet has been achieved in the continuous
system, and upwards of 70 pet of the arsenic has been removed in batch tests (6).
Cadmium Removal Using Aerobic Bacteria
Bacteria that reduce cadmium from solution in the presence of nickel and cobalt were
isolated from sediments in the Marathon Battery Superfund Site.
Metal Contaminant Removal Using BIOFIX Beads
The SLRC has developed a material which utilizes immobilized biomass for removing metal
contaminants from a wide variety of mining and industrial wastewaters (22). The original
objectives of this work were to produce a material compatible with conventional equipment and
procedures, produce a reusable and easily regenerated material, and recover the sorbed metal
ions. These objectives have been met and have resulted in polymeric beads designated as BIO-
FIX beads (11,21,30). These beads are being awarded an R&D 100 award in 1990 as one of
the 100 most valuable domestic inventions by Research and Development Magazine.
The beads, which are spherical in shape and somewhat similar in appearance to ion
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P. Altringer and S. Giddings 133
exchange resins, can be produced in a variety of sizes depending on the targeted application.
BIO-FIX beads are produced from readily available raw materials including high-density
polysulfone pellets, an organic solvent, and dried, thermally-killed biomass obtained from
microorganisms and aquatic flora.
The fabrication procedure consists of dissolving polysulfone in the organic solvent, blending
dried, minus 100-mesh biomass into the polymer solution, and spraying fine droplets of the
mixture into water. Spherical beads are immediately formed, and are ready for use after
curing in water for 12 to 16 hours. The cured beads are porous, resistent to attrition, and
stable in strong acid and base solutions.
Types of biomass immobilized in the beads include algae, common duckweed, peat moss,
and other materials. Sphagnum peat moss has been the most effective material utilized thus
far, and has the added advantage of being abundant and inexpensive.
An attractive feature of the beads is that their effectiveness in sorbing the metal
contaminants most frequently encountered in mining and industrial wastewaters: cadmium,
mercury, lead, etc. Although sorption of these metal ions is a characteristic of the individual
biomass used, some materials, particularly peat moss and certain algae, will remove most of
these metals from wastewaters. Evidence of the effectiveness of the beads for removing metal
ions from dilute wastewaters is shown in Figure 6.1.2. Cadmium, copper, manganese, and lead
were removed from various waters, and in each case the resulting effluent met National
Drinking Water Standards. These tests were conducted in fixed-bed columns and stirred tanks,
and biomass types utilized included peat moss and 2 species of algae. Contact times were 5 to
10 minutes in each test.
An important feature of BIO-FIX beads is that sorbed metals are readily eluted of sorbed
metals using dilute mineral acids. Since only a small volume of acid is required for elution
and regeneration, significant concentration of the metal values is possible. As an example, acid
mine drainage water containing 10.5 parts per million zinc and 4.3 parts per million
manganese was processed in a 3-column fixed-bed circuit. Over a period of several loading-
elution cycles, the effluent consistently met all discharge standards, and elution with 20 g/L
sulfuric acid produced an eluate containing about 100 times as much zinc and manganese as
the original wastewater. Subsequent tests indicated that this eluate could be further
concentrated using conventional hydrometallurgical techniques for eventual recovery of the
metal values.
Although most of the work with BIO-FIX beads has involved conventional processing
equipment, recent laboratory and field tests have indicated the potential for use of the beads in
passive systems having low maintenance and labor requirements. One promising technique
consists of enclosing the beads in porous bags fabricated from polypropylene, placing the bags
in a natural trench or constructed trough, and allowing wastewater to flow through the bags by
gravity. Periodically, the beads most fully loaded with metal ions would be collected and
replaced with fresh beads. The loaded beads would then be regenerated on site or transported
to a central location and regenerated. This type of system may be especially useful for treating
small seeps where conventional technologies are often difficult and expensive to apply. Tests
have indicated that beads enclosed in porous bags exhibit the same loading and elution
characteristics as beads utilized in other equipment.
The development of BIO-FIX beads has resulted in a material well-suited for removing
metal contaminants from mining and mineral processing wastewaters. The beads are
fabricated from easily obtained raw materials, accommodate a wide variety of biomass, have
excellent handling characteristics in conventional equipment, and demonstrate long-term
chemical and physical stability. In addition, the beads readily sorb metal contaminants from
dilute solutions, selectively sorb toxic and heavy metals over calcium and magnesium, exhibit
good sorption and elution kinetics, and are readily eluted and regenerated.
BIOREMEDIATION OF SOLID WASTE
Background
One of the oldest, most studied methods of removing metals from rocks and soils is
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134 Metals
leaching. There are two principle types of mineral leaching: chemical and biological, however,
the boundary between the two is often obscure. Microbes (bacteria and molds) are able to
mobilize metals from rocks and sediments through a variety of ways, but the best studied
examples link metal leaching to the production of metabolic waste acids: nitric, sulfuric, or a
wide variety of organic.
The familiar example of bioleaching is found in the copper industry. The soil bacteria
Thiobacillus ferrooxidans oxidizes sulfide minerals to sulfuric acid, in the process metallic ions
are liberated. These metals can subsequently be concentrated from heaps of rock ore and
collected for refinement. It has been estimated that 20 pet of all copper mined in the U.S. is
recovered from bioleaching operations. Thiobacillus mediated leaching is also important in the
recovery of domestic uranium and other non-ferrous metals. A more detailed study of sulfide
oxidizing bacteria will be examined later in the report.
This example has served as a model to illustrate the potential for other bioleaching
operations. Many minerals are subject to biological mobilization, however, the possibilities for
bioleaching are barely being recognized. Most of the actual applications of bioleaching have
taken place to enhance mineral recovery from low yield ores. These ores are frequently
unamenable to any other treatment.
Information on the use of bioleaching to remove metal pollutants from soils and sediments
is scarce, however, the possibility for this type of application is bright. The same treatment
systems that are now used to recover uranium for the mining industry would work equally well
to remove metallic wastes. The reason is simple: Biotechnology is flexible. The appetite of
leaching bacteria is fairly non-specific. The same bacteria that mobilize copper from sulfides
also mobilize zinc, lead, and other metals in the process. Microorganisms have diverse
appetites; they are capable of deriving energy from the degradation of carbonates, phosphates,
sulfides, oxides and other minerals - liberating metals in the process.
As seen in the mining industry, practical applications of bioleaching are relatively
inexpensive and fairly easy to maintain. In situ leaching of metal pollutants may be possible
with the simple addition of a microbial nutrient source. In other cases, where metal toxicity
would be expected to disrupt a biological setup, direct contact between the bacteria or mold and
the metal would not be required. A simple, two-step operation may be possible: First, the
production of microbial-generated metabolic acids; two, application of the leach solution to
remove the metals from tainted sediments.
Specific Examples in the Literature
Sulfides
This example is detailed to show the wide range of metals that can be mobilized through
bioleaching. As stated in the introduction, a large portion of the domestic copper market is
filled by the recovery of bioleached copper. The mechanics of this leaching are well studied
(14). Many bacterial groups are capable of degrading sulfide sources including Thiobacillus,
Sulfolobus, Thermothrix, and Leptospirillium. These bacteria derive energy from the oxidation
of sulfide minerals such as covellite (CuS), chalcopyrite (CuFeSz), and pyrite (FeS2). Some
members of these groups find uses for the metallic component of the mineral as well:
Thiobacillus is known to get electrons from the oxidation of ferrous iron, Sulfolobus uses
molybdenum as a metabolic electron sink, reducing Mo6* to a lower valence (13).
The usual metabolic waste product from sulfide degradation is sulfuric acid. A wide variety
of metals, other than copper, can be liberated by this process. Bioleaching has been applied in
Canada since 1971 to recover uranium from ores that contain minute traces of the metal (23).
Several projects have been conducted in the U.S. and Canada since that time (31). Precious
metals, gold, silver, and platinum-group metals are being subjected to bioleaching prior to
cyanidation. The metals are often found as discreet metal-sulfides; this pre-treatment makes
the metals more responsive to recovery than can be achieved through only chemical means (16).
Sphalerite (ZnS), galena (PbS), cinnebar (HgS), and pentlandite [(Fe,Ni)9S8] yield soluble
zinc, lead, mercury, and nickel when subjected to bioleaching. Cobalt is frequently recovered in
trace amounts from nickel ores (13).
These examples show that bioleaching is being applied to enhance metal recovery from
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P. Altringer and S. Giddings 135
mining operations. It would not, however, be difficult to modify these systems to recover metal
pollutants from sediments. Figure 6.1.3 shows a conceptual configuration for bioleaching
sediments based on industrial leaching of copper ores.
Carbonates and Phosphates
The ability of microbes to attack carbonate and phosphate minerals is an important part of
the natural cycling of these elements; the specific details of these cycles can be found in texts
on geochemistry or microbial ecology. These microorganisms are able to solubilize insoluble
phosphate and carbonate deposits through the production of acidic metabolic end-products that
lower pH and attack the metal-anion bonds.
Calcareous minerals (limestones) can contain elevated concentrations of certain metals. The
shellfish and coral that originally made up these sediments are capable of concentrating metals
many thousand times over surrounding levels. When these marine animals died, the metals
they had accumulated were trapped in a calcium carbonate matrix. Acidic metabolites,
produced by bacteria and molds dissolve the carbonate material, eventually yielding carbon
dioxide and a residual cationic metal (Ca2+). The trace metals which were trapped in the
sediments are liberated in the process. As an example, selenium is found in abundance within
shale deposits of the San Joaquin Valley of California. This selenium is being mobilized from
the exposed shale as the carbonaceous minerals are subjected to biological (and natural)
weathering. This mobilization has been linked to problems associated with high selenium
levels in that area.
The cycle for phosphorus is similar to the cycle for carbonate. Calcium phosphate is
vulnerable to a variety of organic acids produced by bacteria, formic, oxalic, and citric acids
being notable. Ferric and other metal phosphates are subject to the metal-liberating power of
hydrogen sulfide, a common bacterial metabolite. Thus, rocks or sediments that contain metal
phosphates would be susceptible to bioleaching.
Silicates
Little information is available on the leaching of silicate ores, although a good example has
been provided from the Soviet Union (26). Spodumene, LiAlSi206, has been subjected to the
solubilizing power of biologically generated organic acids; the process liberating the lithium and
aluminum. Other authors have investigated the application of biotechnology to remove silicate
from low grade aluminum ores (24). The mechanism(s) behind leaching of silicates are not
known, however, this illustrates the usefulness and wide range of applied microbiology.
Oxides
Research on the reduction of metal-oxides shows the applicability of microbial systems.
Manganese is an important non-ferrous metal that is being recovered through the use of
microbial geochemical agents. Many bacteria are capable of reducing MnO2 if they are
provided with an oxidizable nutrient source. There are several ways that MnO2 can be
reduced.
First, the metal-oxide can serve as a terminal electron acceptor for respiratory enzymes,
replacing oxygen. The model for this example is:
RH2 + MnO2 —> Mn(OH)z + R
Enzyme preparations that accomplish this reduction were isolated by Bautista and
Alexander in 1972 (10). A large number of prominent bacterial groups have demonstrated this
ability, including species of Bacillus, Clostridium, Micrococcus, and Pseudomonas. Several
molds have also shown this enzymatic capability.
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136 Metals
Second, as with the carbonates, phosphates and silicates, biological metabolic waste acids
are effective in leaching manganese from oxide materials. This reaction is:
H*
Mn4* > Mn2*
(Mn02)
The hydrogen ion being provided by bacteria and other microbes. Paponetti (45) reported on
the recovery of manganese using citric acid produced by the mold Aspergillus niger. Gupta and
Ehrlich (25) used a mixed microbial culture to remove manganese from silver ores, prior to
cyanidation, because the manganese interferes with the silver recovery process. [Ehrlich used a
similar experiment to demonstrate the possibility of bioleaching nickel and cobalt from sea
nodules (19)1
Researchers at the Bureau of Mines Reno Research Center have studied the reduction of
manganese using a species of Thiobacillus. Information on this research will be presented in
this session. This bacteria is able to solubilize manganese from oxides by a pathway that is
similar to it's other metal liberating systems. Thiobacillus thiooxidans oxidize sulfur
compounds in sediments and ores, producing sulfuric acid. The acid effects the reduction of
manganese from the insoluble 4+ state to the soluble 2+ state. There is a curious highlight to
this biological mediated leaching - the bacteria were able to solubilize more manganese than
could be achieved by use of non-bacterial generated sulfuric acid. It was concluded that the
difference may be that the bacteria are able to liberate more manganese because they are
additionally using the Mn4* ion as a terminal electron acceptor.
Exploratory SLRC Biotechnical Research on Great Lakes Sediments
Preliminary research into the possibility of removing heavy metal contaminants from Great
Lakes sediments through biotechnology is encouraging. Experiments being conducted at the
SLRC are based on the following premise: that many bacteria produce organic and inorganic
acids as a byproduct of metabolism. These acids can be successfully used to leach metals from
minerals and sediment compounds.
This experiment has a precedent. In a previous test, manganese, cobalt, cadmium, and lead
were successfully leached from simulated Great Lakes sediments using the system described
above. Simulated Great Lakes sediments were used due to the lack of actual sediments and
were created by crushing sea nodules. The sea nodules are rich in insoluble manganese and
cobalt compounds; to this artificial sediment, insoluble cadmium and lead salts were added.
The organic acids were produced by a species of Klebsiella, a member of the enteric family.
The Grand Calumet/Indiana Harbor (GC/IH) sediments were used as a model for these
tests. From the GC/IH site, a wide variety of bacteria were cultivated that could possibly be
used to produce acids which would leach the metals from the sediments. The GC/IH site is
rich in organic waste (sewage, oils, and aromatic compounds) and attempts are underway to
determine if bioleaching can be accomplished using these on-site bacterial feed compounds.
This would alleviate the cost of adding a "bulk" carbon compound nutrient, such as sugar.
The Saginaw River (SR) and Buffalo River (BR) sediments have received little attention as yet,
however, it is felt that any remediation system that is devised for the GC/IH site would apply
equally well to these other two sites.
Identification of Bacteria From the Grand Calumet/Indiana Harbor. Saginaw River, and Buffalo
River Sediments
Research is being conducted at the SLRC under an MOA with EPA on beneficiation of
GC/IH, SR, and BR sediments for decontamination. Small samples of these sediments were
obtained by the Biotechnology Group for exploratory studies. A population study was begun on
the sediments from the various sites to determine the types of bacteria present. (This study
made no attempt to account for molds or viruses in the sediment slurry.) There were two
reasons for checking the bacterial types found at the sites. First, we needed to establish if the
bacteria found in the sediments were capable of producing acid metabolites that could leach the
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P. Altringer and S. Giddings 137
heavy metals. Second, we needed to determine if any potential biological health hazard exists
in working with the sediments. This analysis was performed when it was discovered that a
large portion of the organic sediment at the GC/IH site is sewage.
The population study of the Grand Calumet sediments revealed an extremely diverse
collection of microorganisms. Aerobic, facultative, and anaerobic bacteria were all identified;
many of the bacteria encountered are typical water and sediment types including several
species of Pseudomonas (P. aeruginosa, P. fluorescens, and P. putida), a Bacillus (B. subtilis),
Acinetobacter anitrus, and a sulfate reducing bacteria (probably Desulfovibrio). Clostridium
sporogenes, and anaerobic sediment bacteria, was also identified. Tests revealed approximately
5xl010 bacteria/mL in the sediments sample we received (the sediment sample was 40-pct solid).
Total coliform and fecal coliform tests were run to determine the level of bacteria
introduced to the GC/IH site through sewage run-off; the numbers were 3x10* coliform/mL and
1.8x10* fecal coliform/mL respectively. These last two numbers show the seriousness of the
biological contamination of the GC/IH site; the National Drinking Water Standard allows for
only 1 coliform per 100 mL of water. The coliforms that were identified included E. coli,
Enterobacter cloacae, Citrobacter freundii, Klebsiella pneumonia, Salmonella enteritis and
Streptococcus faecalis. Staphylococcus aureus as well as a Haemophilus sp. were also identified.
These last four bacteria are potential pathogens. Other bacteria suspected as present, but not
confirmed include Methylococcus, Aeromonas, and Selenomonas. These tests were run at the
SLRC and verified, independently, by the Utah State Health Laboratory. Identification on
GH/IH bacteria is continuing.
The Saginaw River (SR) and Buffalo River (BR) sites were similar to the GC/IH site in
many respects. The total cell counts were similar and all three sites contained many of the
same water and sediment bacteria including Bacillus and Pseudomonas species. One important
difference, however, the coliform tests for the SR and BR sediments showed that these waters
are below the NDWS guidelines for coliform bacteria.
From the results of the identification work, two conclusions were reached. First, that a low
level health hazard existed which could be remedied through safe-handling techniques
(minimization of contact and washing). Second, that acid producing bacteria, such as Bacillus,
were endemic to the sediments.
Bacteria in De-oiled Sediments
The first step in beneficiation research being conducted at the SLRC is de-oiling the
sediments. Bacteria were also cultured from GC/IH sediments which had been de-oiled with
(1) a double methanol wash, followed by (2) repulping in equal volumes of methanol and
refiltering, followed by (3) drying at 105° C, followed by (4) soxhlet extraction with 1,1,1
trichloroethane, followed by (5) drying at 105° C. The process of de-oiling the GC/IH sediments
appears to have eliminated the vast majority of bacteria that were found in the sediments.
Two bacteria, Bacillus subtilis (a spore forming bacteria) and a strain of Pseudomonas, probably
P. aeruginosa, were isolated from a sample of de-oiled sediments however the bacteria were not
very numerous.
Inorganic Leaching Capability of Indigenous Bacteria
The objective of the exploratory studies currently being conducted is to determine if the
bacteria will use the organics present in the sediments as a nutrient, or if they need a
supplemental nutrient. A batch experiment was set up using GC/IH sediments; 18 samples
were placed in flasks with an equivalent amount of water (to make sampling easier). Two
sterile controls were produced by autoclaving those flasks to kill the microorganisms present.
Two other flasks were sterilized and were then inoculated with Enterobacter and Bacillus
species to determine how well these single species would leach metals. Enterobacter was
chosen because it is closely related to the bacteria from the synthetic soil experiment which
was described above. Bacillus was chosen because it is a good acid producer and it is tolerant
of heavy metals. The remaining flasks were not autoclaved. Six of the flasks contained only
the diluted GC/IH sediments to determine if the organics present in the sediments could serve
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138 Metals
as a source of bacterially generated acids. To the remaining nine flasks, sugar (dextrose) was
added in different concentrations. Results are pending chemical analysis.
CONCLUSION
We think that bioremediation of inorganics in sediments shows potential. Successful
development of the biotechnical techniques may provide on-the-shelf technology for
environmental problems untreatable with conventional technology today.
ACKNOWLEDGEMENT
The authors wish to express appreciation to G. Semerad for her assistance on the bacterial
identification studies.
REFERENCES
1. Aickin, R.M. and AC.R. Dean (1977). Lead Accumulation by Microorganisms. Microbios
Letters, 5: 129-33.
2. Aickin, R.M. and AC.R. Dean (1979a). Lead Accumulation by Pseudomonas fluorescens and
by a Citrobacter sp. Microbios Letters, 9: 55-66.
3. Aickin, R.M. and AC.R. Dean (1979b). Electron Microscope Studies of the Uptake of Lead
by a Citrobacter sp. Microbios Letters, 9: 7-15.
4. Altringer, P.B. Biohydrometallurgical Methods for Metals Removal. Presented
to USEPA, NYEPA, and EBASCO, New York, NY, Nov. 13, 1986.
5. Altringer, P.B. Biological Cyanide Destruction and Selenium Removal From Precious Metals
Solutions. To be presented at the 1990 Intermountain AIME/SME Minerals Conference,
Vail, CO, Aug. 9-10, 1990a.
6. Altringer, P.B. and B.E. Dinsdale. Biological Arsenic Removal From Mining
and Milling Waters by Anaerobic Sulfate Reducing Bacteria. To be presented and published
in the 1991 SME Annual Meeting and Exhibit, Environmental Management Symposium,
Water Quality Concerns in the Mining Industry Session, Denver, CO, Feb. 25-28, 1991.
7. Altringer, P.B., R.H. Lien, and K.R. Gardner. Determining Mechanisms of
Anoxic Bacterial Selenium Removal. Interagency Agreement No. 9-AA-20-08500 between
U.S. Department of Interior, Bureau of Reclamation and Bureau of Mines, April, 1990b, p.
19.
8. Altringer, P.B., R.H. Lien, and K.R. Gardner. Biological and Chemical Selenium Removal
From Precious Metals Solutions. To be presented and published at the 1990 SME
GOLDTech 4 "North American Practices," Technical Session on "Advances in Gold and
Silver Processing", Reno, NV, Sept. 10-12, 1990b.
9. Altringer, P.B., D.M. Larsen, and K.R. Gardner (1989). Bench-Scale Process Development of
Selenium Removal From Wastewater Using Facultative Bacteria. In: Biohydrometallurgy, J.
Salley, R. G. L. McCready, and P. L. Wichlacz (eds.), CANMET SP89-10, pp. 643-658.
10. Bautista, E.M. and M. Alexander (1972). Reduction of Inorganic Compounds by Soil
Microorganisms. Proc. Soil Sci. Soc. Amer., 36: 918-20.
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P. Altringer and S. Giddings 139
11. Bennett, P. G., and T. H. Jeffers. Passive Biological Treatment of Acid Mine Waters. To
be presented at the TMS-AIME Annual Meeting, New Orleans, LA, Feb. 17-21, 1991.
12. Bennett, P.G., and T.H. Jeffers (1990). Removal of Metal Contaminants From a Waste
Stream Using BIO-FIX Beads Containing Sphagnum Moss. In: Proceedings, Western
Regional Symposium on Mining and Mineral Processing Wastes, F. M. Doyle (ed.), Chapter
35, pp. 279-286.
13. Brierley, C. (1982). Microbiological Mining. Sci. Am., 247: 44-54.
14. Brierley, C.L. (1978). Bacterial Leaching. CRC Critical Reviews in Microbiology, 6: 207-
62.
15. Brierley, C.L. (1982). Microbiological Mining. Sci. Am., 247: 44-54.
16. Bruynesteyn, A. (1988). Biotechnology for Gold Ores: The State of the Art. In:
Proceedings International Gold Conference 88, Perth, Western Australia, Oct. 28 - Nov. 1,
1988.
17. Chmielowski, J. and B. Klapcinska (1986). Bioaccumulation of Germanium by Pseudomonas
putida in the Presence of Two Selected Substrates. Appl. Environ. Microbiol., 51: 1099-
1103.
18. Darnall, D.W., B. Greene, M. Hosea, R.A. McPherson, M. Henzl, and M.D. Alexander (1986).
Recovery of Heavy Metals by Immobilized Algae. In: Trace Metal Removal From Aqueous
Solution: Proceedings Symposium Royal Society Chemistry, Annual Chemical Congress, R.
Thompson (ed.).
19. Ehrlich, H.L., S.H. Yang, and J.D. Mainwaring, Jr. (1973). Z. Allg. Mikrobiol., 13: 39-48.
20. Faila, M.L. and B.D. Weinberg (1977). Cyclic Accumulation of Zinc by Candida
utilis During Growth in Batch Culture. J. Gen. Microbio., 99: 85-97.
21. Ferguson, C.R. and T.H. Jeffers. Biosorption of Metal Contaminants From Acidic Mine
Waters. To be presented at the SME Annual Meeting, Denver, CO, Feb. 25-28, 1991.
22. Ferguson, C.R., M.R. Peterson, and T.H. Jeffers (1989). Removal of Metal Contaminants
From Waste Waters Using Biomass Immobilized in Polysulfone Beads. In: Biotechnology
in Minerals and Metal Processing, B.J. Scheiner, F.M. Doyle, and S.K. Kawatra (eds.), pp.
193-199.
23. Gow, W.A., H.H. McCreedy, G.M. Ritcey, V.M. Mcnamara, V.F. Harrison, and G. H. Lucas
(1971). In: Recovery of Uranium, International Atomic Commission, Vienna, pp. 195-206.
24. Groudev, S.N. and V. Groudeva (1988). Microbial Removal of Silicon from Mineral Raw
Materials. In: Biohydrometallurgy Proceedings International Symposium Warwick, Norris,
P. R., and D. P. Kelly (eds.). Univ. of Warwick.
25. Gupta, A. and H.L. Ehrlich (1988). J. Biotechnol., 39: 137-42.
26. Karavaiko, G.I., and S.N. Groudev (eds.), (1985). Biotechnology of Metals, Moscow.
27. Kauffman, J.W., W.C. Laughlin, and R.A. Baldwin (1986). Microbiological Treatment of
Uranium Mine Waters. Environ. Sci. Technol, 20: 243-48.
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140 Metals
28. Kuyucak, N. and B. Volesky. Recovery of Cobalt by a New Biosorbent. In: Proceedings of
the Third Annual General Meeting of Biomet., R. McCready (ed.) Toronto, Canada, pp. 111-
27.
29. Kuyucak, N. and B. Volesky (1986). Recovery of Gold by Biosorption. Proceedings of the
Third Annual General Meeting of Biomet., R. McCready (ed.), Toronto, Canada, pp. 171-72.
30. Jeffers, T.H., C.R. Ferguson, and B.C. Seidel (1990). Biosorption of Metal
Contaminants using Immobilized Biomass. In: Biohydrometallurgy 1989 Proceedings, J.
Salley, R.G.L. McCready, and P.L. Wichlacz (eds), CANMET, pp. 317-327.
31. Lakshmanan, V.I. (1986). Industrial Views and Applications: Advantages and Limitations
of Biotechnology. In: Workshop on Biotechnology for the Mining Industries. Biotech.
Bioeng. Symp. 16, Ehrlich, H.L. and D.S. Holmes (eds.).
32. Larsen, D.M., K.R. Gardner, and P.B. Altringer (1989). Biologically Assisted
Control of Selenium in Process Waste Waters. In: Biotechnology in Minerals and Metals
Processing, B.J. Scheiner and P.M. Doyle (eds.), Ch. 22, pp. 177-185.
33. Larsen, D.M., K.R. Gardner, and P.B. Altringer (1987). A Biohydrometallurgical Approach
to Selenium Removal. In: American Water Resources Association, R.F. Dvorsky (ed.),
Technical Publication TPS-87-4, Bethesda, MD, pp. 419-426.
34. Lien, R.H., and P.B. Altringer. Biological and Chemical Cyanide Destruction
in Heap Leachates and Tailings. To be presented and published in the 1991 SME Annual
Meeting and Exhibit, Environmental Management Symposium, Water Quality Concerns in
the Mining Industry Session, Denver, CO, Feb. 25-28, 1991.
35. Lien, R.H., B.E. Dinsdale, and P.B. Altringer. Biological and Chemical
Cyanide Destruction From Precious Metals Solutions. To be presented and published in the
1990 SME GOLDTech 4 "North American Practices," Symposium on "Advances in Gold and
Silver processing", Reno, NV, Sept. 10-12, 1990c.
36. Lien, R.H., B.E. Dinsdale, K.R. Gardner, and P.B. Altringer (1990a). Chemical and
Biological Cyanide Destruction and Selenium Removal From Precious Metals Tailings Pond
Water. In: EPD 90, D. R. Gaskell (ed.), AIME-TMS, pp. 323-339.
37. Lien, R.H., B.E. Dinsdale, K.R. Gardner, and P.B. Altringer. Chemical and Biological
Cyanide Destruction and Selenium Removal From Precious Metals Tailings Pond Water.
Presented at the AIME-SME Annual Meeting, Salt Lake City, UT, Feb. 26 - Mar. 1, 1990b,
to be published in Proceedings.
38. Macaskie, L.E. and A.C.R. Dean (1982). Cadmium Accumulation by Microorganisms. Enu.
Tech. Letters, 3: 49-56.
39. Macaskie, L.E. and A.C.R. Dean (1984). Cadmium Accumulation by a Citrobacter sp. J.
Gen. Microbiol., 130: 53-62.
40. Macaskie, L.E. and A.C.R. Dean. (1985a). Uranium Accumulation by Immobilized Cells of a
Citrobacter sp. Biotech. Letters, 7: 457-62.
41. Macaskie, L.E. and A.C.R. Dean (1985b). Strontium Accumulation by Immobilized Cells of
a Citrobacter sp. Biotech. Letters, 7: 627-30.
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P. Altringer and S. Giddings 141
42. Macaskie, L. E., A. C. R. Dean, A. K. Cheetham, R. J. B. Jakeman, and A. J. Skarnulis
(1987). Cadmium Accumulation by a Citrobacter sp: The Chemical Nature of the
Accumulated Metal Precipitate and Its Location on the Bacterial Cells. J. Gen. Microbiol.,
133: 539-44.
43. Mowell, J.L., and G. M. Gadd (1984). Cadmium Uptake by Aureobasdisium pullans. J.
Gen. Microbiol., 130: 279-84.
44. Norris, P.R., and D.P. Kelly (1978). Accumulation of Cadmium and Cobalt by
Saccharomyces cerevisiae. J. Gen. Microbiol., 99: 317-24.
45. Paponetti, B.L., C. Abbruzzese, A. Marbini, and M.Y. Duarte (1989). Manganese Recovery
from MnO2 Ores by Aspergillus niger: Role of Metabolic Intermediate. Biotechnology in
Minerals & Metals Processing, B.J. Scheiner and F.M. Doyle (eds.), Las Vegas SME
Proceedings, pp. 33-37.
46. Poldoski, J.E. (1979). Cadmium Bioaccumulation Assays. Their relationship to
various ionic equilibria in Lake Superior water. Environ. Sci. and Tech., 13: 701-706.
47. Trujillo, E.M., T.H. Jeffers, C.R. Ferguson, and H.Q. Stevenson. Biosorption of Metal Ions
on Immobilized Biomass Beads. To be presented at the AIChE 1990 Summer National
Meeting, San Diego, CA, August 19-22, 1990.
48. Tsezos, M. and B. Volesky (1981). Biosorption of Uranium and Thorium. Biotechnol.
Bioeng., 23: 583-604.
49. Tsezos, M. (1984). The Selective Extraction of Metals From Solution by Microorganisms - A
Brief Overview. Can. Met. Quart., 24: 141-44.
50. White, C. and G.M. Gadd (1987). The Uptake and Cellular Distribution of Zinc in
Saccharomyces cerevisiae. J. Gen. Microbiol., 133: 727-37.
51. Winge, D.R., K.B. Nielson, W.R. Gray, and D.H. Hamer (1985). Yeast Metallothiones. J.
Biol. Chem., 260: 14454-70.
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Metals
Figure 1. - CN Removal In Single-Pass
3-Column Trickling Reactor
CN concentration, ppm
300
200
100
87%
75%
93%
90%
88%
91%
91%
87%
83%
84%
68%
8
90%
90%
96%
91%
83%
II
I
1
89%
ll.
1
1 9 16 26 49 96 107 126 155 217 253
7 14 21 37 57 105 111 147 182 245
Time, days
H Feed B Col. 3 Effluent
Figure 6.1.1. CN removal in single-pass 3-column trickling reactor.
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P. Altringer and S. Giddings 143
Figure 2. - Metal Sorption Using BIO-FIX Beads
Metal Concentrations, mg/L
Waste Treated National Drinking Metal
Water Effluent Water Standard Removal, pet
Cadmium 0.060 0.001 0.01 98
Copper 2.0 0.023 1.0 99
Manganese 4.7 0.018 0.05 99
Lead 0.059 0.002 0.05 97
Figure 6.1.2. Metal sorption using BIO-FIX beads.
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144 Metals
Figure 3. - Conceptual Configuration For
Bioleaching Sediments
Waste rock/Low grade ore
(Contaminated sediments)
Sulfuric acid/
Ferric salts
t
Thiobacillus
Sulfolobus
Acidified H,O
Oxidized ores
_ _
I (Sediments) I
Sulfur Ferrous salts Soluble metal
Collecting pond
i
Concentrated metals Acid leach solution Organic/microbial recovery
Figure 6.1.3. Conceptual configuration for bioleaching sediments.
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H. Edenborn 145
6.2 Biological Treatment of Metal-Contaminated Water
Hank Edenborn
Supervisory Research Biologist
U.S. Bureau of Mines
Pittsburgh Research Center
P.O. Box 18070
Pittsburgh, PA 15236
Acid coal mine drainage
Acid mine drainage (AMD) is a common water pollution problem on active and abandoned
coal mine sites in the eastern United States. AMD forms when surface mining brings
unweathered pyrite-containing rocks to the surface or when deep mining allows oxygen to
contact buried pyritic strata. In the absence of neutralizing compounds, the drainage that
results can be extremely acidic and contaminated with dissolved iron, manganese, and sulfate.
Drainages with pH < 3.0 and concentrations of sulfate greater than 1,000 mg/L, iron greater
than 50 mg/L, and manganese greater than 10 mg/L are common. Where water flows through
alkaline materials (such as limestone) before surfacing, the drainage is less acidic and
occasionally circumneutral, but it can still contain high concentrations of sulfate and metals.
Current water quality standards in the United States require that mine discharges have a
pH between 6 and 9, total iron concentration less than 3.0 mg/L, and manganese less than 2.0
mg/L. At thousands of active and inactive mine sites, drainage does not meet these standards
and is being treated before discharge by the mining company. At thousands of other sites that
were abandoned prior to the enactment of water pollution laws or were operated by companies
that have gone bankrupt, untreated AMD is polluting receiving water systems.
The standard mine drainage treatment system involves the addition of alkaline chemicals to
the water, which raises the pH and causes metals to precipitate in a settling pond. These
systems are expensive, often costing tens or hundreds of thousands of dollars per year for
chemicals, operation, maintenance, and disposal of the metal-laden sludge. Because the
drainage on many sites will likely be contaminated for decades, there is financial incentive to
find alternative water treatment systems.
The constructed wetland concept has its roots in observations of natural Sphagnum peat
wetlands that received acid mine drainage and, instead of being adversely affected, appeared to
clean the polluted water. These observations instigated the idea that wetland systems might
be used for the intentional treatment of mine drainage. Because the discharge of AMD into a
natural wetland is prohibited by several laws, it has been necessary to construct wetlands that
act solely as water treatment systems.
Initially, most wetland research and construction efforts mimicked the original observations
by using Sphagnum moss and peat. Despite promising lab results, virtually all field tests of
Sphagnum-dominated constructed wetlands failed to provide sufficient water treatment for more
than several months. Sphagnum proved quite sensitive to the stresses associated with
transplanting, abrupt changes in water chemistry, excessive or insufficient water depth, and
excessive accumulation of iron. At most sites, the moss died within the first growing season.
Today, almost all wetlands constructed to treat AMD are planted with Typha latifolia,
common cattails. Typha is readily available to most sites, transplants well, and has proved
tolerant of a wide range of water conditions. Occasionally, Scirpus spp. (bulrushes) and
Equisetum spp. (horsetails) are also planted, but even these wetlands are generally dominated
by cattails after the first few years of system operation.
Most constructed wetlands include 15-45 cm of an organic substrate in which the emergent
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146 Metals
plants can root. Topsoil, rotten animal manure, spoiled hay, and compost have been used. In
western Pennsylvania, mushroom compost, a waste product of mushroom farming, has become a
widely used organic substrate. At sites with acidic drainage, 8-16 cm of crushed limestone is
often spread underneath the organic substrate to provide some neutralization.
Constructed wetland systems usually consist of a series of shallow pits or cells. This design
makes flow control much easier than with a single, large wetland. The cells are filled with
substrate, planted with Typha, and flooded with mine drainage. In most systems, water depth
is 5-15 cm above the substrate and flow is quite slow. Hay bales and logs are sometimes used
as barriers to enhance serpentine flow pattern, prevent channelization, and increase the contact
of AMD with the wetland substrate and vegetation.
AMD is now being treated biologically in constructed wetlands at over 300 mine sites in the
bituminous coal region of the eastern United States. In general, the processes at work in these
systems are aerobic. The oxidation of ferrous iron to ferric iron and the subsequent
precipitation of iron oxyhydroxide floe, for example, are dominant processes:
Fe2+ + 0.2502 + 1.5H20 --> FeOOH (solid) + 2H* (1)
Ferrous iron tends to autooxidize in aerated solutions at pH values greater than 6, while in
more acidic water naturally- occurring bacteria catalyze the reaction. Although iron oxidation
and hydrolysis processes are effective at removing much of the iron from the AMD, these
processes do nothing to help raise the pH of the water or lower the acidity. In fact, the pH of
water can be lowered by these reactions (equation 1). Many constructed wetlands with
circumneutral pH and iron-contaminated inflow water actually produce water with a lower pH.
Ironically, bacterial processes capable of increasing the pH and alkalinity of AMD entering
constructed wetlands are already found there, but current wetland designs do not take
advantage of them. Probably the most useful of these processes for treating AMD is bacterial
sulfate reduction, a naturally-occurring reaction that proceeds in many environments in the
absence of oxygen and in the presence of suitable organic substrates and sulfate.
Sulfate-reducing bacteria use organic carbon and sulfate in the process of anaerobic respiration:
2CH2O + SO42 —> H2S + 2HCO3- (2)
The reaction has promise in the treatment of acid- and metal- contaminated mine waters
because the by-products of the reaction, hydrogen sulfide and bicarbonate, can precipitate many
metals and raise the pH of the water, respectively.
Sulfate reduction rates have been measured in the sediments of marine and freshwater
environments. Sulfate reduction rates often vary over several orders of magnitude at any given
location due to the heterogeneous nature of sediments. Measured rates range from
approximately 0.4 to 3000 nmol cm"3 day"1. Oxygen, low temperatures, low concentrations of
organic matter and sulfate, and low pH all tend to limit sulfate reduction rates. Recent work
at the Bureau of Mines has established that sulfate reduction does occur in constructed
wetlands and can play a significant role in the treatment of AMD. Water quality data from
several constructed wetlands demonstrating the influence of both aerobic and anaerobic
treatment processes will be shown.
Metal mine drainage
Recently, the U.S. Bureau of Mines has begun to exploit the bacterial sulfate reduction
process studied in wetlands for the treatment of mine waters contaminated with metals other
than iron and manganese. Many heavy metals, such as Cd, Cu, Pb, Hg, Ni, Ag, and Zn, can be
precipitated as insoluble sulfides in the presence of sufficient hydrogen sulfide. Although little
evidence has been accumulated to date, it seems unlikely that wetland systems will be a
satisfactory way to treat these metals due to the likelihood of their bioaccumulation in plants
and animals. Research efforts have therefore been directed towards the development of
contained sulfate reduction systems consisting of barrels or tanks with sufficient organic matter
to enhance anaerobic bacterial activity. Laboratory experiments have been performed and
pilot-scale studies are currently underway at several locations, including the U.S. Bureau of
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H. Edenborn 147
Mines research mine in Pittsburgh, PA, and at a zinc smelter Superfund site near Palmerton,
PA. The results of this work will be discussed and the potential use of wetland and bacterial
sulfate reduction systems in the bioremediation of contaminated sediments will be addressed.
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148 Metals
6.3 Bioleaching of Ores
Elizabeth G. Baglin
U. S. Bureau of Mines
Reno Research Center
1605 Evans Avenue
Reno, Nevada 89512-2295
INTRODUCTION
Metal Biosolubilization in Nature
The most important biogeochemical roles that microbes play in nature have to do with the
transformation and cycling of elements, such as carbon, oxygen, nitrogen, sulfur, and
phosphorus. But metals such as iron, manganese, calcium, potassium, mercury, selenium, and
zinc are also transformed in nature. The various chemical reactions in these cycles are
beneficial, and often essential, to make the minerals available to indigenous flora in soluble
form for their metabolism.
But natural, or uncontrolled, biosolubilization can create environmental problems. For
instance, oxidation of pyritic minerals by native microorganisms, such as Thiobacillus
ferrooxidans or Thiobacillus thiooxidans can lead to serious water pollution problems in coal
mining regions. Thiobacilli are chemolithoautorophs, which obtain their energy by oxidizing
reduced iron and sulfur moieties and their carbon from CO2 in the air. Sulfuric acid produced
by the action of the bacteria on sulfide minerals present in the coal is responsible for
solubilization of metal ions which contaminate the mine waters. Acid mine drainage is also a
problem in metal mines in the west, especially lead and zinc districts, where the sulfidic ores
are attacked by similar microorganisms.
Metal Bioleaching from Ores by Thiobacillus Bacteria
Naturally occurring Thiobacillus bacteria play a significant role in leaching of copper from
heaps and dumps of low-grade ore (1). The microbes can oxidize reduced copper sulfide
minerals by a direct mechanism to produce soluble cupric sulfate, which is concentrated and
recovered as copper metal. But an even more important mechanism is the indirect oxidation of
copper sulfides by ferric iron formed by direct attack of the bacteria on iron pyrite which is
also present in the ore:
DIRECT LEACHING (direct attack of mineral by microbes)
4 FeS2 + 15 02 + 2 H2O bactl!ria ) 2 Fe2(SO4)3 + 2 H2S04
INDIRECT LEACHING (attack by biologically generated ferric iron)
CuFeS2 + 2 Fe2(S04)3 -> CuS04 + 5 FeS04 + 2 S
The leachant is regenerated by further biooxidation:
BACTERIAL LEACHANT REGENERATION
2 S + 3 O2 + 2 H20 bactena > 2 H2SO4
4 FeS04 + O2 + 2 H2SO4 baeteria > 2 Fe2(SO4)3 + 2 H20
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E.G. Baglin 149
By similar mechanisms, direct and indirect, Thiobacillus ferrooxidans is known to aid in the
removal of uranium, zinc, cobalt, nickel, cadmium and other metals from sulfidic ores.
Metal Bioleaching from Ores by Heterotrophic Microorganisms
Heterotrophic microorganisms, those which require organic carbon for their growth and
energy needs, can also solubilize metals from rocks and minerals. Heterotrophs have been
shown to be capable of extracting nickel and aluminum from silicate ores, waste products and
clays (2,3) and bacterially mediated, reductive dissolution of iron- and manganese-bearing ores
has also been demonstrated (4-6). Heterotrophs are also known to be capable of breaking down
silicate minerals, the principal component of most rocks, and can utilize and remove
phosphorus from phosphate ores.
Like the chemolithoautotrophs, heterotrophic microorganisms also employ direct and indirect
leaching mechanisms. Direct leaching utilizes reductive solubilization of metal species as a
form of respiration. Or the microorganisms can indirectly produce an acid, base, or ligand the
solubilized metals.
Bacterial Pretreatment of Ores by Thiobacillus Bacteria
Biooxidation is increasingly being implemented as a pretreatment step in the processing of
refractory precious metal ores (7). For instance, gold which is locked inside iron sulfide
minerals often cannot be extracted by conventional cyanidation. But, ores of this type can be
pretreated with sulfur oxidizing microorganisms, most commonly Thiobacillus ferrooxidans,
which break open the pyrite matrix and allow access to the gold by cyanide during subsequent
processing. A 1500 ton per day biooxidation plant has recently been built in Central Nevada,
and smaller plants have been operated worldwide.
Bacterial pretreatment has also been extensively investigated as a means of removing
pyritic sulfur from coal.
BIOLEACHING TO EXTRACT METALS FROM ORE MATERIALS
The Bureau of Mines has developed a research group at the Reno Research Center which
has been conducting biohydrometallurgical research for the past several years. The focus of
this work has been the biosolubilization of manganese from low-grade domestic ores by
heterotrophic microorganisms, and the biooxidative pretreatment of a sulfide concentrate
containing platinum-group metals using Thiobacillus ferrooxidans bacteria. The research has
taken primarily an applications approach, with the more basic work on mechanisms and
physiology of the microbes being undertaken via contract research at the Idaho National
Engineering Laboratory.
Bioleaching of Manganese Ores
Because manganese is a low-value commodity, this work is directed at low-cost mining and
processing technology - open pit mining, and heap leaching. The microorganisms utilized are
native to the ores or are introduced via the nutrient used to feed them (molasses). In the
laboratory, bioleaching of manganese ores is being investigated on three different scales:
(1) shake-flask tests - conducted as screening experiments to quickly obtain information on
the ability of finely ground ore to be leached under various conditions. These experiments
can be kept sterile, if desired, to test for chemical leaching effects or to run controls.
(2) column tests - utilize information obtained during flask tests. The columns allow for
some control of the biological system and employ larger-sized material. Nutrient medium is
recirculated through the ore bed and is replaced when depleted.
(3) open, non-sterile simulated heaps - allows for contamination of the bioleach system by
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150 Metals
air-borne microorganisms, especially mold spores, which are present in the environment
surrounding the heap. These experiments simulate conditions likely to be incurred in a
real-world heap leaching situation.
Shake-flask tests:
Several nutrient media have been evaluated and the best results have been obtained by
using molasses to feed the microorganisms. The experiments are conducted by slurrying
ground ore with diluted (5 wt pet) molasses and sampling the flasks weekly so that the
manganese content of the solutions can be monitored. The results of shake-flask screening
tests using food-grade molasses are shown in Table 6.3.1. Several oxide ores were found to be
readily amenable to bioleaching, with extractions of more than 95 pet being attained in 4
weeks. But only 27 pet of the manganese was extracted from the sulfidic Black Cloud ore.
This ore would best be treated by sulfide oxidizing bacteria such as Thiobacillus ferrooxidans.
The antimicrobial agent sodium azide was added to some experiments to eliminate biological
activity and to allow determination of the extent of chemical leaching by the molasses. These
control tests showed that chemical leaching by the 5 pet molasses solution varied from 2 to 13
pet for the ores tested.
Factory molasses, which is the residual in food-grade molasses production, was also
evaluated. This by-product is sold for animal feed for approximately $0.05/lb and could provide
a low cost nutrient source for bioleaching. Shake-flask bioleaching tests were conducted using
factory molasses in the same manner as the tests with food-grade molasses. Results for
bioleaching Three Kids ore with 5 pet factory molasses are shown in Figure 6.3.1. The results
were similar to those obtained using food-grade molasses - 97 pet of the manganese was
extracted in 7 weeks.
When slurries of unsterilized ore and molasses medium were inoculated with bacteria from
previous experiments, only 70 pet of the manganese was extracted in the first 5 weeks, and
some precipitation of the manganese occurred during continued leaching. Precipitation occurs
when the organic carbon in the medium has been depleted and the pH rises. It appears that
the inoculum probably contained microorganisms which competed with the manganese
solubilizing microbes for the carbon in the medium.
The aerobic dissimilation of carbohydrates, such as glucose, involves a series of enzymatic
changes, which may be divided into two parts: an initial breakdown to pyruvic acid and the
subsequent oxidation of pyruvate to C02 and H2O, via the tricarboxylic acid, or Krebs, cycle.
Other pathways for pyruvate utilization also occur, depending on the microorganism and the
conditions. During the Krebs cycle, pyruvate is converted to citric acid, which undergoes a
series of enzymatic oxidations, decarboxylations, and transformations, forming (in succession)
isocitric, oc-ketoglutaric, succinic, fumaric, malic, and oxaloacetic acids. The net result of this
cyclic process is the complete oxidation of one molecule of acetate to CO2 and H2O with each
turn of the cycle.
To evaluate whether Krebs cycle acids and other organic acids which can be products of
respiration or fermentation pathways are capable of solubilizing manganese from its ores, a
series of shake-flask tests was performed in which Three Kids ore was leached abiotically with
a number of organic acids. The tests were conducted under the same conditions as bioleach
experiments and the carbon content of each leach liquor was 4 g/L. Results are tabulated in
Table 6.3.2. The acids which extracted the most manganese were L-malic (91 pet), a-
ketoglutaric (91 pet), and citric (84 pet). Leaching rate was fastest with citric acid, the first
acid formed during the Krebs cycle. These experiments suggest that an indirect leaching
mechanism, i.e., leaching by organic acids produced during carbohydrate metabolism, may be
responsible for the removal of manganese from Three Kids ore. Examination of solution
potential and pH data indicate that leaching is not the result of acid generation, but probably
involves reduction of the higher oxides of manganese to soluble manganous ions by the organic
compounds, possibly accompanied by chelation of the dissolved metal species.
These tests do not rule out a direct leaching mechanism. Other researchers have shown
that both indirect and direct leaching mechanisms are operative in manganese bioleaching
systems.
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E.G. Baglin 151
Column and heap bioleaching tests:
Column and simulated heap leaching experiments are being conducted by recycling medium
through beds of Three Kids ore. Size of the tests has varied from 400 grams of ore in 5-cm
i.d. columns and 2 L of medium up to 34 kg of ore in a 32 cm i.d. open cylindrical tank and
20 L of medium. Occasional replenishment of the medium is required to keep the bioleaching
going. A typical column bioleaching curve is shown in Figure 6.3.2. Similar curves are
generated during heap bioleaching. Manganese dissolution increases until the medium is spent
and if the test is allowed to proceed without replenishment of the nutrient, the pH of the
system increases and manganese precipitates from solution. As long as sufficient nutrient is
present, the pH remains on the acid side of neutral (5.5 to 7) and manganese stays in solution.
We have operated some columns for as long as a year and manganese has continued to leach
from the ore.
Best results to date have been attained using molasses to bioleach minus-1/4 in. ore in 5
cm columns (Table 6.3.3). For instance, 70 pet manganese extraction was obtained in 29 weeks
in a test using 3 pet molasses, and 30 pet extraction has been obtained after 6 weeks in an
ongoing column which is being leached with 5 pet molasses. A control test, in which sodium
azide was added to the medium, showed only 5 pet extraction in 6 weeks and visible evidence
of growth was negligible compared to active bioleach columns. The data also show that minus-
3/4 in. ore leaches much more slowly than the 1/4 in. material, which is not unexpected. We
are not sure whether the ore particles are being wetted through to the core.
We have isolated various microorganisms from bioleach solutions and identified them by
fatty acid analysis and by using a standard biochemical multitest system. Because molasses
caramelizes at sterilization temperatures, the microbes were cultured on tryptic soy agar (TSA)
plates. This technique may not be definitive, because it is possible that the TSA selects for
microorganisms that do not dominate in the molasses-based leach medium. At this point, we
have not determined which species are responsible for the leaching. We also do not know
whether the manganese solubilizing microorganisms come from the molasses, or from the ore
itself. To answer these questions, we recently hired a person to undertake the microbiological
aspects of this project. We hope to have some results in the next few months.
Biooxidation of Platinum and Gold Ores
As stated earlier, sulfide ores are often refractory, i.e. resistant to conventional cyanidation,
because the sulfide minerals encapsulate the precious metals and prevent access of the leachant
to the insides of the mineral particles. To overcome this problem, refractory gold ores are
treated by pressure oxidation or by roasting prior to cyanidation. Pressure oxidation results in
high capital costs, and roasting produces SO2 which requires off-gas treatment. Biooxidation is
another pretreatment option which has been gaining increased attention in recent years. In
fact, a 1500 tpd biooxidation plant came on-line earlier this year near Austin, NV. Smaller
plants are in operation worldwide, and the technology is being marketed by a number of
companies.
The Bureau of Mines has been investigating biooxidation as a possible environmentally
acceptable alternative to smelting for pretreating a sulfide platinum-group metal concentrate
from the Stillwater Complex in Montana. Smelting is a high-temperature processing step
which, like roasting, evolves SO2, and requires strict environmental controls. The Stillwater
Complex holds the only PGM deposit in the United States and this research is being conducted
as part of the Bureau's strategic minerals program. A mineralogical description of the
concentrate is shown in Table 6.3.4.
Biooxidation is being conducted at 30° C in stirred, batch reactors up to 5 liters in size.
Normally a 10 pet pulp density is maintained. Thiobacillus ferrooxidans bacteria are fed with
a simplified mineral salts nutrient medium composed of three salts dissolved in water
[(NH4)2SO4, KH2P04, MgSO4]. Ferrous sulfate is sometimes added to the medium to give the
bacteria a ready supply of energy and to support rapid growth during the early stages of the
process. We have been trying to wean the bacteria from ferrous iron by gradually decreasing
the amount added to the medium. The intent is to force the microbes to obtain their energy
by oxidizing the sulfides present in the concentrate. The laboratory reactors are aerated with a
mixture of air containing 5% CO2. The dissolved oxygen content and oxygen transfer rates are
-------
152 Metals
important factors in biooxidation reactors, and Thiobacillus ferrooxidans obtains its cellular
carbon from the CO2. We have found that operating the reactors in a draw and fill mode, i.e.,
periodic settling, decantation, and replacement of the medium, enhances sulfide oxidation.
Table 6.3.5 shows the effects of medium replacement and use of aeration. Both factors
enhanced the amount of sulfide oxidation achieved.
Using draw and fill operation, we have been able to oxidize up to 94 pet of the sulfide in
the Stillwater concentrate over a period of 5 weeks. Biooxidation destroys the pentlandite,
pyrite, and chalcopyrite minerals in the concentrate, and leaches the nickel and some of the
copper. The PGM remain in the residue, primarily as sulfide minerals, even though most of
the sulfur has been removed. Mineralogical characterization with the scanning electron
microscope showed the following relative preponderance of PGM minerals in the biooxidized
residue:
PdS > (Pt,Pd,Ni)S » PdTe = PtTe2 > PtFe = PtS.
Chemical Leaching of Biooxidation Residue
After biological pretreatment, the solids are treated chemically to extract the precious
metals. We have investigated several leachants: (1) oxidative chloride (aqua regia and H202-
HC1), the traditional method for extracting PGM from minerals; (2) thiourea, a known gold
extractant which operates in the acid pH range; and (3) cyanidation, which is conducted at
high pH.
Highest extractions to date have been obtained with cyanidation at 80° C. Results are
shown in Table 6.3.6. Palladium, rhodium and gold extractions are quite reasonable, but the
highest platinum extraction obtained so far is only 34 pet. SEM examination showed that the
palladium-bearing minerals in the residue had decreased considerably, but the platinum-
bearing minerals still remained:
(Pt,Pd,Ni)S > PtS » PdS » PtTe2 > PtFe.
The poor leaching of platinum may be a kinetic problem, or it may be the result of
electrochemical (galvanic) effects. Selective metal leaching by galvanic effects is based on the
fact that physical contact between dissimilar metal sulfides immersed in dilute sulfuric
acid/ferric sulfate solution will create a galvanic cell. The sulfide mineral with the highest rest
potential will become cathodically protected (passivated), while ones with lower rest potentials
(anode) will be leached. The Stillwater concentrate is a multimetal sulfide mixture and the
rest potentials of the principal base metal sulfide minerals fall into the following order:
pyrite > chalcopyrite > pentlandite > pyrrhotite.
Pyrrhotite should leach first, pyrite last. Destruction of the pentlandite should liberate the
contained palladium, and we do see good palladium removal during cyanidation. Because
platinum sulfides are very insoluble and stable, their rest potentials are expected to be higher
than that of pyrite. As a result, they should be even harder to leach. It may be necessary to
have almost complete oxidation of the sulfide minerals before high platinum extractions are
possible.
We have recently set up a prototype continuous stirred tank bioreactor system to see if we
can improve sulfide oxidation, in hopes of improving the recovery, especially platinum in the
second stage chemical leach. If that doesn't work, we will be looking at different sulfide
oxidizing microorganisms such as the thermophile Sulfolobus to see if more efficient
biooxidation can be attained.
SUMMARY
Natural biosolubilization of metals can be beneficial, in that it provides minerals to
indigenous flora for their metabolic needs, and it can create environmental problems, such as
acid mine drainage. Controlled biosolubilization of metals from ores, biooxidation and
-------
E.G. Baglin 153
bioleaching, are showing increased importance as alternatives to conventional chemical oriented
procedures for recovering metals from ores.
The bioleaching research group at Bureau of Mines Reno Research Center has been
investigating biological treatment of ores and mineral concentrates for the past several years.
Emphasis has focussed on the biosolubilization of manganese from its ores using heterotrophic
microorganisms and the biooxidation of a sulfidic platinum-group metal concentrate with the
acidophilic chemolithotroph Thiobacillus ferrooxidans. The research has shown that biological
treatment can be used to extract metals from ores and that biological treatment can also be
used to make minerals more amenable to subsequent chemical treatment to remove metal
values.
Sediments and ores are both rock-based substances. The fact that metals can be removed
from ores by biological action indicates that there is good potential that biological extraction of
metals from sediments can be successful.
REFERENCES
1. Hiskey, J. B. and R. Bhappu (1987). Role of Oxygen in Dump Leaching. In: Proc. of
Internatl. Sympos. on the Impact of Oxygen on the Productivity of Non-Ferrous Metallurgical
Processing, G. Kachaniwsky and C. Newman (eds.), Pergamon Press, pp. 165*182.
2. Bosecker, K (1989). Bioleaching of Valuable Metals From Silicate Ores and Silicate Waste
Products. In: Biohydrometallurgy, J. Salley, R. G. L. McCready, and P. L. Wichlaez
(eds.), Canmet SP89-10, pp. 15-24.
3. Groudev, S. N., and V. I. Groudeva (1986). Biological Leaching of Alumina from Clays. In:
Workshop on Biotechnology for the Mining, Metal-Refining, and Fossil Fuel Processing
Industries, H. L. Ehrlich and D.S. Holmes (eds.), John Wiley And Sons, N. Y., pp. 91-99.
4. Ehrlich, H. (1980). Bacterial Leaching of Manganese Ores, Biochemistry of Ancient and
Modern Environments, Springer-Verlag, Berlin, pp. 609-614.
5. Yopps, D. L., E. G Noble, E. G. Baglin, and J. A. Eisele (1989). Bioleaching of Manganese
Ores (Poster). Biohydrometallurgy 1989, Jackson Hole, WY, Aug. 13 - 18, 1989.
6. Holden, P. J., and J. C. Madgwick (1983). Mixed Culture Bacterial Leaching of Manganese
Dioxide. Proc. Australas. Inst. Min. Metall., 286: 61-63.
7. Hutchins, S. R., J. A. Brierley, and C. L. Brierley, Microbial Pretreatment of Refractory
Sulfide and Carbonaceous Ores, 116th Annual AIME Meeting, Feb 23 - 27, 1987, SME
Preprint 87-143, 1987, 16 pp.
-------
154 Metals
Table 6.3.1. Shake-Flask Bioleaching of Manganese Ores.
Source Mineralogy Mn Content, Mn Extraction, pet
Three Kids,
Nevada
Silver Cliff,
Colorado
Algoma-Zeno,
Minnesota
Black Cloud,
Colorado
Mn oxide
quartz
feldspar
Mn oxide
quartz
potassium feldspar
Mn oxide
Fe203
quartz
Mn carbonate
Fe-Pb-Zn sulfides
pet
15.6
3.7
15.5
0.5
Bioleach
79
97
98
27
Azide Control
12
13
13
4
Conditions: 2 g minus 48-mesh ore, 100 mL of 5 wt-pct food-grade molasses medium,
4 weeks, ambient temperature, 200 RPM
Table 6.3.2. Abiotic Leaching of Three Kids Ore with Organic Acids.
Acid Mn Extraction, pet
L-Malic 91
ot-Ketoglutaric 91
Citric 84
Formic 65
Lactic 44
Oxalic 11
Succinic 6
Fumaric 4
Acetic 4
Conditions: 2 g minus 48-mesh ore, 100 mL of solution containing 4 g/L organic carbon, 2 weeks,
ambient temperature, 200 RPM
-------
E.G. Baglin
155
Table 6.3.3. Column and Heap Bioleaching of Three Kids Ore.
Weight
Ore, kg
0.4
0.5
0.5
0.5
9
34
36
Medium:
Ore
Size,
in
-1/4
-1/4
-1/4
-1/4
-3/4
-3/4
-3/4
1GA ..
3GFM
5FGM
Column
i.d., cm
5
5
5
5
12.7
32
heap
...1 wt-pct
....3 wt-pct
....5 wt-pct
Medium
Type
1GA
3FGM
5FGM
5FGM/
azide
3FGM
3FGM
1GA
glucose, (NH4)2S04
food grade molasses
food grade molasses
Table 6.3.4. Stillwater Ore Minerals.
BASE METALS ovrite, chalcom
Medium
Vol, L
2
5
5
5
20
20
20
'rite, pentk
Mn Extn, Leach Time,
pet wks
28 52
70 29
30 6
5 6
2 5
1 15
2 37
indite
PLATINOID MINERALS
PGM sulfides
Pt-Fe alloy
solid solution in pentlandite (Pd)
GANGUE Al, Ca, Fe, Mg silicates
-------
156
Table 6.3.5. Bio-oxidation of Stillwater Flotation Concentrate.
Metals
Medium
Replace-
ment
No
No
Yes
Yes
Air/CO2
Sparge
No
Yes
No
Yes
1
ND
5.6
4.0
5.2
Sulfide
2
ND
4.2
3.5
2.3
in Solids, pet
Week
3
6.1
1.6
ND
ND
4 5
1.4 1.1
ND 3.8
0.9 0.3
Total Sulfide
Oxidized,
pet.
0
83
38
94
Conditions: 300 g concentrate, 6.1 pet sulfide, 3 L ATCC medium 64, 300 mL inoculum
of 72-hr T. ferrooxidans A-6 culture.
Table 6.3.6. Cyanidation of Bioleached and As-Received Stillwater Concentrate.
Extraction, pet
Pt Pd Rh Au
pet sulfide oxidized
during bioleaching
94
79
70
60
0 (as received)
35
24
20
22
16
76
73
74
76
64
90
92
79
91
43
99
98
98
..
97
Conditions: 30 g concentrate in 600 mL of 1 pet or 2 pet CN solution, 80° C, 23 h.
Head Analysis: 10 oz/t Pt, 33 oz/t Pd, 0.3 oz/t Rh, 0.5 oz/t Au, 6.1 % S.
-------
I
«
CO
02
S3
03
cn
CT;
5'
ro*
E3
o
3'
crq
25
al
CO
o
-i
JO
en
•O
o
Mn EXTRACTION.
03
CO
S I i i
2-
en —
en
-------
158
Metals
1 I ' I '
4 6 6 10 12 14
WEEKS OF LEACHING
16
Figure 6.3.2. Column bioleaching of Three Kids ore, 3 pet. food-grade molasses.
-------
A.E. Torma and PJL Pryfogle 159
6.4 Mechanisms of Bacterial Metals Removal From Solids
Arpad E. Torma and Peter A. Pryfogle
Center for Biological Processing Technology
INEL, EG&G Idaho, Inc.
Idaho Falls, Idaho 83415
Abstract
The Great Lakes area sediments are contaminated with varying amounts of heavy metals
and polychlorinated organic matter. With respect to the bioremediation of metallic contents of
these sediments, it was shown that a number of microorganisms exist which can effectively
solubilize heavy metals. The basic reaction mechanisms of bioleaching processes were discussed
and the effects of semiconductor character of the sulfide substrate explained. A special
emphasis was made to comment on INEL's bioremediation capability.
Introduction to Sediment Environments
Sediments are commonly defined as solid material that has settled down from a state of
suspension in a liquid (1). The sediments of marine origin are divided into three main classes
(1,2): detrital material derived from the erosion of the continents, biogenic material that is
formed by biological productivity, and autogenic material that is formed in situ.
Detrital material consists mostly of alumino-silicates. Biogenic components are produced
from plankton tissues, algal mats, and other microbial organisms (4) in the surface waters of
lakes, estuaries and seas, and contain calcites as well as organic matter. Autogenic materials
consist of mineral phases (sulfides, phosphates, and carbonates). Biogenic sediments are most
active in the upper part of the sediments and it is very rare to find biological processes below
a depth of half a meter (3). When a phosphate-rich wastewater is introduced into very hard
lake water (containing high calcium concentration) the following reactions may occur (5):
[1] 5 Ca+2 + OH' + 3 PO/3 — > [Ca5(OH)(P04)3} hydroxyapatite
Near the surface where C02 concentration is relatively high, calcium carbonate is formed:
[2] Ca+2 + 2 HC03- - > (CaC03) + C02 + H2O
When the pH is locally raised by photosynthetic reaction it yields;
[3] Ca+2 + 2 HCO3- + hv — > CH20 + {CaCO3} + 02
A decrease in pH can result in the production of insoluble humic-acid-base metal sediments (6).
Introduction of acidic mine tailings and industrial drainages into rivers and lakes results in
considerable heavy metal contamination and formation of toxic inorganic sediments. Biological
activity is responsible for the formation of sulfide-bearing heavy metal sediments. For example,
there are a number of sulfate reducing bacteria (7), which perform an anaerobic respiration by
using sulfate as final electron acceptor and by oxidizing organic compounds (8); for example:
[4] 2 CH3-CHOH-COONa + H2S04 ^^.>
(lactate) H2S + 2 CH3-COONa + 2 C02 + H2O
The resulted H2S will react with the heavy metal ions in solution:
-------
160 Metals
[5] M2* + H2S .—> {MS} + 2 H*
where M2+ is Fe2*, Mn2+, As3*, Zn2+, Cd2+, etc. The metal sulfide precipitate {MS} will be
contained in the sulfide autogenic sediment. The sulfate reduction reactions will especially be
predominant during the winter season when the surface of the lake is covered by ice and snow
and infiltration of oxygen into the lake water is considerably limited. The sediments may vary
from completely oxic, where the supply of organic carbon is less than the supply of oxygen, to
wholly anoxic, where the supply of organic carbon is much greater than the supply of oxygen.
The changing chemistry of the sediments is reflected in changes in redox potential and the pH.
The redox system was used for the classification of the chemical sediments (2,9) already in the
1940s and 1950s.
Great Lake Sediments
The sediments from the Great Lakes areas (Saginaw Bay, Michigan; Sheboygan Harbor,
Wisconsin; Grand Calumet River, Indiana; Ashtabula River, Ohio; and Buffalo River, New York)
are known to be highly toxic (70). The remedial action plans include 42 areas of the Great
Lakes where the sediments contain varying amounts of heavy metals (arsenic, cadmium,
chromium, copper, iron, lead, manganese mercury, nickel, silver, and zinc) as well as organic
matter (polychlorinated biphenyls, polyaromatic hydrocarbons, oils, greases, and cyanides).
Thus, the sediment environment of these areas is a very complex material, and it is likely that
there is more variability in the sedimentary system than uniformity. Therefore, it can be
anticipated that a simple approach for the remediation of all Great Lakes areas may not be
feasible and methods based upon site specific information must be worked out. In this context,
the bioremediation of contaminated sediments presents an alternative to the chemical and
physical remediation possibilities. This paper will report on the background information on
possible bioremediation of inorganic (metallic) contents of the Great Lakes Sediments and point
out INEL's capabilities.
Microorganisms
It is known and well documented that microorganisms are playing an important role in the
formation and solubilization of mineral deposits since geological time (11, 12). Since the metal
contents of the Great Lakes area sediments are occurring probably in forms of sulfides, oxides,
silicates and carbonates, it is likely that they can be extracted especially by the iron and sulfur
oxidizing thiobacilli, which are know as the leaching microorganisms. The most frequently
studied bacterium is called Thiobacillus ferrooxidans (13). It oxidizes ferrous iron and
reduced-valence inorganic sulfur compounds (14):
[6] 2 FeS04 + H2SO4 + 0.5 02 ^£™.> Fe2(S04)3 + H2O
[7] MS + 2 02 J2!*£»L> MSO4
where M is a bivalent heavy metal. The metal sulfides, MS, are generally insoluble in the
acidic nutrient media, while their corresponding sulfates are soluble. Hence is the dissolution
process. Other leaching bacteria are (15,16) Thiobacillus thiooxidans which oxidizes elemental
sulfur and thiosulfate but not metal sulfides; Leptospirillium ferrooxidans which oxidize ferrous
iron and pyrite; Thiobacillus organoporus oxidizes elemental sulfur and a number of metal
sulfides (PbS, Bi2S3, Sb2S3, ZnS) thermophilic bacteria which are active between 45 and 85°C
(Sulfobacillus thermosulfiodooxidans, Sulfolobus acidocaldarius, Sulfolobus brierley, Sulfolobus
solfataricus) can oxidize metal sulfides, ferrous iron and elemental sulfur. Some of the
thermophilic species require the presence of minute amounts of yeast extracts. In addition to
the above lithotrophic bacteria, there are a number of heterotrophs present in the naturally
occurring heap and dump leach media (17). However their specific contribution to the
solubilization of metal sulfides is not well understood and documented.
-------
A.E. Torma and PA. Pryfogle 161
Leaching Mechanisms
Metals can be extracted from insoluble minerals (sulfides, oxides, carbonates, etc.) directly
by metabolic action of microorganisms or indirectly by the product of their metabolism (18).
The biological reactions relevant for removal of toxic metal contents from sediments are
primarily oxidation reactions of metal sulfides, such as of iron pyrite in which sulfur is in the
-2 oxidation state (19), that will be oxidized to ferric sulfate in which sulfur is in the +8
valence form:
[8] 2FeS2 + 7.5 O2 + H2O JStt!s.> Fe2(S04)3 + H2SO4
Ferric iron is also an oxidizing agent which contributes to the dissolution of metal sulfides, for
example, CuS according to:
[9] CuS + Fe»(SO«)s — > CuS04 + 2 FeS04 + S°
In the presence of bacteria, ferrous iron and elemental sulfur liberated in reaction [9] will be
oxidized to:
[10] 2 FeS04 + H2S04 + 0.5 02 -^^-> Fe2(S04)3 + H2O
and,
[11] S + 1.5 O2 + H2O J=2=1L> H2SO4
The sulfuric acid produced in the metabolic reactions [8 & 11] may also react with the
oxide and carbonate metallic constituents of sediments to yield further metal solubilization:
[12] MO + H2SO4 — > MS04 + H20
[13] MC03 + H2SO4 — > MS04 + H20 + C02
where M is a bivalent heavy metal. Reactions [8,10, and 11] represent the direct leaching
mechanisms of bacterial action. The energy available from these oxidation reactions will be
captured by the microorganisms to cover their energetic needs. Reactions [9,12, and 13]
represent the indirect mode of bacterial leaching activity, where the metabolites (ferric sulfate
and sulfuric acid) react with the insoluble oxide and carbonate inclusions of the sediments.
The literature of bioleaching is vast and is described in a large number of review articles
(13,15,16,18,20-23) and symposium proceedings (11,12,16,24-30). The information available from
these sources is relevant to the treatment of the Great Lakes sediments. The review of the
above information is not the purpose of the present article. However, the authors thought to
be important to include here these references for the benefit of those scientists who would like
to familiarize themselves with the potential possibilities of this emerging technology.
Effects of Semiconductor Character of Metal Sulfides
It was observed by many investigators that in some cases bioleaching of particular MS is
easy, and the same type of MS from a different location is very difficult. For example for
chalcopyrite leaching, it was suggested that the composition Cu(II)Fe(H)S2 is easy to leach, but
when it is in the form Cu(I)Fe(III)S2 crystallographic modification then its biooxidation is very
difficult and slow. Other investigators (31) reported that chalcopyrite from different locations
vary especially in minor and trace amounts of many elements which are present in
isomorphous substitution. For example, silver, gold, platinum, lead, cobalt, nickel, manganese,
tin, and zinc replace copper or iron, while arsenic, selenium and tellurium replace sulfur. In
addition, chalcopyrite, as many of the MS, is a typical semiconductor material, which exhibits
an energy gap of about 0.6 eV and a resistivity of about 10"3 ohmm (32). The naturally
occurring product is never perfect. It contains crystal defects (vacancies) or interstitial
-------
162
Metals
impurities that lead to the formation of extrinsic n-type (having an excess of negative charges)
or p-type (having an excess of positive charge or holes) semiconductor. The valence energy
band (lower energy band) is completely filled with electrons and these are highly bound, while
the conduction band (higher energy band) is empty or only partially filled with electrons which
are largely bound and free to move (33). Between these two bands is the energy gap, which is
often called the forbidden zone of energy. The type of conductivity (n- or p- type) is
determined by the energy levels of the Fermi electrons (Ef) (34) that can be assessed from the
Hall effect. If Ef is located close to or within the conduction band, the semiconductor is
designated as n-type, and in the case when the Fermi electron is located close or within the
valence band, then the semiconductor is called p-type. The Fermi electron energy level
represents the amount of thermodynamic work that has to be provided to the sulfide substrates
(by the redox leach system of bacteria) in order to remove an electron (oxidize) the solid sulfide
mineral. Therefore, the redox potential of the bacterial leach system must be higher than that
represented by Ef of solid sulfide (chalcopyrite) in order for oxidation to take place (35).
On the basis of electron structure of the semiconductor chalcopyrite, it is likely that the
n-type CuFeS-j will be easier to be oxidized by the microorganisms than the p-type ore, since
the electrons in the conduction zone are mobile and loosely bound. This is the reason why
chalcopyrite from different localities has varying leachabilities.
INEL's Bioleaching Capabilities
In the past seven years INEL has been intensively involved in diverse bioleaching activities.
The success in its activities rely on a multidisciplinary approach. Scientists with basic
educational backgrounds in microbiology, biochemistry, molecular biology, genetics, chemistry,
and engineering work within the disciplines of biotechnology, metallurgy and chemical
engineering and collaborate to solve bioleaching problems. In the biohydrometallurgical section
recently the following main topical areas have been investigated:
a) Mechanistic aspects of biocorrosion of copper with exopolymers from Pseudomonas
atlantica indicated that copper was oxidized and a metal film was eroded as measured
by FTIR/ATR coupled with XPS/AES (36).
b) Kinetics of biological cobaltite solubilization (37).
c) Identification of sulfur and iron oxidizing enzymes from Thiobacillus ferrooxidans.
d) Characterization of plasmids from T. ferrooxidans for metal specificity or heavy metal
tolerance.
e) Carbon-fixation efficiency.
f) Biosorption studies (protein and exopolymer isolation, characterization, thermodynamic
measurements).
Part of the above mentioned research has been supported by the U.S. Bureau of Mines and
has been directed at determining the mechanisms as well as the rate and extent of biologically
assisted mineral leaching and recovery from low grade ores.
The INEL laboratory has the following selected specialized equipment for conducting
biotechnological research:
Atomic Absorption Spectrometer
Mobile Riot-Scale Bioreactors
Laminar Flow Hoods
Scanning UV Spectrophotometer
GCs and HPLCs
Image Analysis System
Ion Chromatography
Ultracentrifuge
Environmental Incubator-Shakers
Electrophoretic Gel Sequencing Equipment
Fluorescent Phase Contrast Microscopes
Anaerobic Chamber
Walk-in Environmental Chamber
Bioreactors (column, airlift, RBCs)
-------
AJE. Torma and PA. Pryfogle 163
Conclusion
The bioremediation of toxic heavy metal contents of Great Lakes sediments with iron and
sulfur oxidizing microorganisms is feasible. The bio-mediated extraction processes involve the
direct and indirect leaching mechanisms of bacterial action. INEL's biotechnological
laboratories are well equipped with all analytical tools to cope with the complexity of
bioremediation problems.
Acknowledgement
This work was supported by the U.S. Department of Energy under Contract No.
DE-AC07-76IDO01570.
References
1. Malcolm, S.J. and S.O. Stanley (1982). The Sediment Environment. In: Sediment
Microbiology, D.B. Nedwell and C.M. Brown (eds.), Academic Press, New York, pp. 1-14.
2. Krumbein, W.C. and R.M. Garrels (1952). Origin and Classification of Chemical Sediments
in Terms of pH and Oxidation-Reduction Potentials. Journal of Geology, 60: 1-33.
3. Hallberg, R.O. (1980). In-situ Experimentation with Anaerobic Sediments: Some
Biogeochemical Applications. In: Biogeochemistry of Ancient and Modern Environments,
P.A. Trudinger, M.R. Walter, and B.J. Ralph (eds.), Australian Academy of Science,
Canberra, Australia, pp. 145-155.
4. Philp, R.P., M. Calvin, S. Brown, and E. Yang (1978). Organic Geochemical Studies on
Gerogen Precursors in Recently-Deposited Algal Mats and Oozes. Chemical Geology, 22:
207-231.
5. Manshan, S.E. (1979). Environmental Chemistry, Willard Grant Press, Boston,
Massachusetts, pp. 115-137.
6. Gamble, D.S. and M. Schnitzer (1973). The Chemistry of Fulvic Acid and Its Reactions
with Metal Ions. In: Trace Metals and Organic Interactions in Natural Waters, P.C.
Singer (ed.), Ann Arbor Science Publishers, Inc., Ann Arbor, Michigan, pp. 265-302.
7. Peck, H.D., Jr. (1984). Physiological Diversity of the Sulfate Reducing Bacteria. In:
Microbial Chemoautotrophy, W.R. Strohl and O.K. Tuovinen (eds.), Ohio State University
Press, pp. 309-335.
8. Fischer, U. (1988). Sulfur in Biotechnology. In: Biotechnology, H.J. Rehm and G. Reed
(eds.), VCH Verlagsgesellschaft, Weinheim, Federal Republic of Germany, volume 6B, pp.
463-496.
9. ZoBell, C.E. (1946). Studies on Redox Potential of Marine Sediments. Bulletin of the
American Association of Petroleum Geologists, 30: 477-513.
10. Personal communication with Paulette Altringer, (1990). U.S. Bureau of Mines, Salt Lake
City Research Center, Salt Lake City, Utah.
11. Ehrlich, H.L. (1981). Geomicrobiology, Marcel Dekker, Inc., New York, pp. 1-393.
12. Krumbein, W.E. (1983). Microbial Geochemistry, Blackwell Scientific Publications, Oxford,
England, pp. 1-330.
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164 Metals
13. Ludgren, D.G. and W. Dean (1979). Biogeochemistry of Iron. In: Biogeochemical Cycling
of Mineral Forming Elements, P.A. Trudinger and D.J. Swaine (eds.), Elsevier, Amsterdam,
pp. 211-251.
14. Silver, M. (1978). Metabolic Mechanisms of Iron Oxidizing Thiobacilli. In: Metallurgical
Applications of Bacterial Leaching and Related Microbiological Phenomena, L.E. Murr, A.E.
Torma, and J.A. Brierley (eds.), Academic Press, New York, pp. 3-17.
15. Torma, A.E. (1988). Leaching of Metals. In: Biotechnology, H.J. Rehm and G. Reed (eds.),
VCH Verlagsgesellschaft, Weinheim, Federal Republic of Germany, pp. 367-399.
16. Karavaiko, G.I. (1985). Microbiological Processes for the Leaching of Metals from Ores, A.E.
Torma (ed.), United Nations Environment Pro., Moscow, USSR, pp. 1-69.
17. Wichlacz, P.L. and R.F. Unz (1982). Microbiology of Coal Mine Drainage Treatment,"
Proceedings 64th CIC Coal Symposium, pp. 199-208.
18. Lundgren, D.G. and M. Silver (1980). Ore Leaching by Bacteria. Annual Review of
Microbiology, 34: 263-283.
19. Dugan, P.R. (1989). Microbial Conversion of Sulfur and Their Potential for Bioprocessing
Fossil Fuels. In: Processing of Fossil Fuels Workshop, P.E. Bayer (ed.), U.S. Department of
Energy, Washington, D.C., pp. 2-40.
20. Ralph, B.J. (1986). Geomicrobiolgy and the New Technology. Developments in Industrial
Microbiology, 26: 23-59.
21. Brierley, C.L. (1978). Bacterial Leaching. CRC Critical Reviews in Microbiology, November,
pp. 207-262.
22. Smith, A.J. and D.S. Hoare (1977). Specialist Phototrophs, Lithotrophs, and Methylotrophs
Unity among a Diversity of Procaryotes? Bacteriological Reviews, 41: 419-448.
23. Ehrlich, H.L. (1986). What Types of Microorganisms are Effective in Bioleaching,
Bioaccumulation of Metals, Ore Beneficiation and Desulfurization of Fossil Fuels.
Biotechnology and Bioengineering Symposium, number 16, pp. 127-137.
24. Swartz, W. (ed.), (1977). Conference Bacterial Leaching, Verlag Chemie, Weinheim, pp.
1-270.
25. Murr, L.E., A.E. Torma, and J.A. Brierley (1978). Metallurgical Application of Bacterial
Leaching and Related Microbiological Phenomena, Academic Press, New York, pp. 1-526.
26. Trudinger, P.A., M.R. Walter, and B.J. Ralph (1980). Biochemistry of Ancient and Modern
Environments, Australian Academy of Science, Canberra, Australia, pp. 1-723.
27. Rossi, G. and A.E. Torma (1983). Recent Progress in Biohydrometallurgy, Associazione
Mineraria Sarda, Iglesias, Italy, pp. 1-752.
28. Strohl, W.R. and O.K. Tuovinen (1986). Microbial Chemoautotrophy, Ohio State University
Press, Columbus, Ohio, pp. 1-351.
29. Lawrence, R.W., R.M.R. Branion and H.G. Ebner (1986). Fundamental and Applied
Biohydrometallurgy, Elsevier, Amsterdam, pp. 1-501.
30. Norris, P.R. and D.P. Kelly (1988). Biohydrometallurgy, Science and Technology Letters,
New Surrey, England, pp. 1-578.
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A.E. Torma and PA. Pryfogle 165
31. Habashi, F. (1978). Chalcopyrite Its Chemistry and Metallurgy, McGraw-Hill, New York,
pp. 1-165.
32. Crundwell, F.K. (1988). The Influence of the Electronic Structure of Solid on the
Dissolution and Leaching of Semiconducting Sulphide Minerals. Hydrometallurgy, 21:
155-180.
33. Shuey, R.T. (1975). Semiconducting Ore Minerals, Elsevier, Amsterdam, pp. 26-318.
34. Solymar, L. and D. Walsh (1988). Lectures on the Electrical Properties of Materials, Oxford
University Press, Oxford, England, pp. 7-327.
35. Choi, W.K., Z.F. Wang, and A.E. Torma (1990). Electrochemical Aspects of a Refractory
Gold Ore Leaching by Thiobacillus ferrooxidans, Reprint No. 90-159, Society for Mining,
Metallurgy, and Exploration, Inc., Littleton, Colorado, pp. 1-8.
36. Gianotto, A.K., P.L. Wichlacz, J.G. Jolley, M.R. Hankins, G.G. Geesey, and R.B. Wright
(1989). The Biocorrosion of Copper by Biopolymers as Examined In Situ, In Real Time
FT-IR/ATR in Conjunction with Pre and Post XPS/AES. In: Biotechnology in Minerals and
Metal Processing, B.J. Scheiner, F.M. Doyle, and S.K. Kawatra (eds.), Society of Mining
Engineers, Littleton, Colorado, pp. 45-51.
37. U.S. Bureau of Mines (1989). Biologically Assisted Minerals Processing. In: Strategic and
Critical Materials Program Annual Report - 1989, Idaho National Engineering Laboratory,
Department of Energy, pp. 1.1-1.12.
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166 Metals
6.6 Linking Biological and Hydrogeochemical Mechanisms of
Sediment Leaching
Robert H. Lambeth
Mining Engineer
and
Barbara C. Williams
Research Civil Engineer
Spokane Research Center
U. S. Bureau of Mines
Spokane, Washington
Introduction
Leaching and fixation of inorganic materials in fluvial and lacustrine sediments result from
hydrogeochemical and biochemical (microbial) processes. The interdependence of such
mechanisms is recognized as critical. Cross-disciplinary research is difficult because of
differences in vocabulary, conceptual models, and methods of experimental design. If
interdisciplinary teams of scientists are to interpret the mechanisms operating in a given
system, experiments must be designed so that the parameters required by scientists from
diverse fields, such as microbiology, hydrology, and hydrogeochemistry, are measured. Ideally,
teams would then interpret findings from new, hybrid perspectives as well as from their
discipline-specific perspectives.
In order to investigate methods of bioremediation of contaminated fluvial and lacustrine
sediments, it is useful for scientists to have at their disposal methods to rapidly simulate the
results of a hypothetical remediation procedure. A valuable class of analytical and predictive
tools is computer models. Calibration and sufficient verification of computer models makes it
possible to predict qualitatively contaminant fate at new sites on the basis of knowledge gained
at sites that have already been studied.
Computer models exist that use data on sediment and pore water chemistry to predict
mineral solubilities, that is, to determine whether inorganic materials are fixated in situ in a
solid matrix in an innocuous form. Other computer models consider data on biological
composition. Some computer models use the hydraulic parameters of aquifers to predict spatial
and temporal distributions of inorganic contaminants in ground water, while others predict
contaminant transport in open estuaries. Generally, those computer models appropriate to
address the interdisciplinary perspectives described above would involve consideration of the
combined kinetic and equilibrium effects of chemical dissolution, biochemical dissolution,
chemical fixation, biochemical fixation, and the dispersion characteristics of moving ground and
surface water. No single "utopian" computer model exists that represents all these processes,
nor is there all the underlying data required to run such a model if it did exist. The
development of a completely linked model would require a large, multidisciplinary team, a
lengthy time frame, and a massive budget. At this time, it is only reasonable to use existing
models separately, with experienced scientists acting as the links, or interface, among the
models.
Background
Staff from the U.S. Bureau of Mines' Spokane Research Center studied sulfide oxidation,
metal leaching, and metal transport in the unsaturated and saturated zones of fine-grained
tailings and a downgradient aquifer. Water quality data were interpreted using an equilibrium
geochemistry computer model, verifying the hypothesis that certain minerals (inorganic chemical
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R.H. Lambeth and B.C Williams 167
compounds) would precipitate, others would dissolve, and yet others would tend to exert
solubility controls on concentrations of various constituents within certain pH ranges. An
additional finding was that an organic layer at the base of the waste impoundment may be
causing the attenuation of metals transported from the tailings.
Key questions remaining in the interpretation of sulfide oxidation at this site include:
What microbial species are present in the tailings? What biochemical effects would these
populations have upon the system? Are different microbes present in the organic layer than
are present elsewhere? Are the microbes that thrive in the variably saturated (air and water)
zone aerobic or anaerobic? Does microbial activity vary seasonally? How do microbes catalyze
(increase the kinetics, or speed of) oxidation reactions occurring within the tailings? Are any
microbes acting to fix metals further downgradient where metal concentrations diminish? For
example, the oxidation of ferrous to ferric iron can be catalyzed by the bacterium Thiobacillus
ferrooxidans. This catalysis may increase the reaction rate by as much as five to six orders of
magnitude. Hydrogeochemists know of the implications of this reaction but generally do not
quantify it. Other possible microbial mechanisms are rarely considered in interpretations of
the chemistry of a site. Cooperative work with microbiologists at this site will be directed to
addressing more such processes.
Part I: Field and Laboratory Data Requirements
In order to determine the interdependence of hydrogeochemical, hydrological, and
microbiological mechanisms in the bioremediation of contaminated fluvial and lacustrine
sediments, data should be collected to address the following questions:
1. What types and amounts of inorganic species can be leached from a matrix of specific
composition by pore water of specific composition and known chemical parameters? At
what rates?
2. What types and amounts can be leached by biota catalysis? At what rates?
3. What types and amounts can be chemically precipitated? At what rates?
4. What biomediated fixation processes are occurring? At what rates and in what
quantities?
5. How are dispersion, dilution, and sorption properties quantified in an aquifer or
estuary?
6. What are the hydraulic properties of an aquifer or estuary?
Questions 1 and 4 are complicated by the necessity of using kinetic (reaction rate) as well
as equilibrium (maximum reaction extent and direction) considerations.
The data required to characterize these interdependent processes are as follows:
Hydrogeochemistry
The field data required for input into most hydrogeochemistry models are relatively
standard and straightforward. Measurements are made for pH, Eh (redox potential),
temperature, and electrical conductivity. Additional measurements are usually made for
alkalinity and reactive dissolved gases such as O2, NH3, S02, and CO2, which are dependent
upon site-specific requirements. Most of these field measurements (with the exception of
temperature, alkalinity, and conductivity) are made with potentiometric electrodes. Another
type of potentiometric electrode, the ion-specific electrode, can measure a limited variety of
mono- and divalent ions, but usually these element concentrations are determined in the
laboratory. Potentiometric methods are usually only reliable and linear within specific ranges
and are subject *o numerous interference problems, particularly in solutions with significant
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168 Metals
levels of dissolved solids. Consequently, in certain instances, other methods, such as titration,
must be used, even in the field. Ion speciation analysis is often also performed in the field if
the element in question changes oxidation states rapidly. Many models utilize an Fe(II)/Fe(III)
ratio, and Fe(II) can be determined readily in the field by colorimetry.
Water samples are analyzed in the laboratory for dissolved elemental or ionic constituents
with instruments. The available methods and their variants are innumerable, but the majority
of analyses of inorganic materials are performed using spectral emission or absorption. There
is no "ideal analytical procedure" for any given substance; the procedure developed will be
controlled by site-specific conditions. Spectral interferences will vary from site to site, as will
concentrations and other factors. All these factors must be evaluated when determining what
analytical procedure to use. The analyses menu will also be site specific; it must be
determined by a general analytical scan, a mineralogic analyses, and requirements of the biota.
The latter illustrates the interrelationship of considerations of the biosphere and the
hydrosphere. A substance which is of no thermodynamic or kinetic consequence may be of
extreme biochemical importance as an energy source or as a substrate or biocide, and its
concentration must be determined.
Mineralogic analyses (solid and dissolved phases) must also be performed. In order to
simulate a leaching from the solid phase, the solid phase must be identified and quantified.
Again there is no "ideal method" for a given substance, and site-specific influences will
determine the suite of techniques selected. Reflective or transmissive optical methods can be
used to identify many inorganic particles, but identifying many amorphous substances will
require more expensive and sophisticated techniques, such as x-ray scanning or ion-
microprobe/SEM analysis. Identification of organometallic complexes can be even more
perplexing, and infrared scanning, nuclear magnetic resonance, chromatography, or mass
spectrometric methods, among others, are often necessary. In certain situations, such as with
solid-solution minerals, the composition of the mineral as well as its identity must be
determined.
Biochemistry
The data requirements for characterizing microbiological processes include the rates at
which microorganisms catalyze reactions. These rates are a function of temperature, initial
population density, and the availability and concentration of energy sources and substrates such
as oxygen, nitrogen, sulfur, carbon, and phosphorus compounds. The concentrations and
identities of the mineral or substance to be catalyzed must be known, and the levels of any
biocides (or rate inhibitors) should be determined. Unfortunately, the identities of inhibitors for
many microbes are not known. The reaction rate will also be a function of mineral surface
area; therefore, a particle size distribution analysis must be performed. This information will
also be required to determine purely chemical reactions.
Perhaps most important is information about the microbes themselves. Important classes of
microorganisms include those free-floating in the aqueous phase and those attached to the
surfaces of the solid phases. The identities, sizes, and population densities of organisms
associated with all phases must be determined. This means that for the solid phase, each
organism for each mineral of interest must be identified, because the type and density of
microbes vary with mineralogy. An added complication is that the activities of the microbe
population must also be determined. A micr&be may be present, but it will not participate in a
reaction unless it is activated. Pulsing cycles are likely to occur if a microbial population
grows exponentially, exhausts an energy source, and dies off. When new nutrients are supplied
by a water-transport process, the population may be reestablished. For microbes that
incorporate liberated inorganic material into their structure, the attrition rate must be known.
If these organisms die, a contaminant could be much more occult to extract from the system.
Attrition could be the end of the normal life cycle or be caused by biocides; also, the
microorganism could be consumed by a more complex life-form that accumulates the
contaminant.
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RJI. Lambeth and B.C Williams 169
Hydrology
The hydrologic setting of a bioremediation effort may be in a groundwater aquifer or an
open estuary. The hydrogeologic parameters necessary to perform flux calculations in
groundwater are relatively well understood, and the calculations are reasonably well developed
and proven. Collection of adequate data to support such models, however, can be quite
expensive. The geometry and internal structure of the sediment mass of interest must be
established. Flow direction and rate are, to a great extent, determined by the stratigraphy of
the sediment and are determined by surface and bathymetric mapping, geophysical surveying,
and well installation (for potentiometric head measurements). Hydraulic head distribution
within each individual stratigraphic unit in the sediment will indicate the natural flow
direction and flux of water under static conditions. Vertical and horizontal coefficients of
hydraulic conductivity for each stratigraphic unit can be determined through aquifer tests in
the field or by permeameter testing of properly prepared undisturbed samples. Horizontal
hydraulic conductivity tends to be much greater than vertical hydraulic conductivity in
sedimentary units because of the tendency of mineral grains to interlock in a horizontal
direction during sedimentation. All porous-media models, analytic or numeric, use Darcy's Law,
Q = - K A (dh/dl), as the primary governing equation.
Contaminant flux in lakes and estuaries is dominated by fluid mechanics. Mixing and
dispersion are fueled by temperature or salinity density differences, deltaic processes, and tidal
movement.
Part II: Computer Model Requirements
There are two types of computer models: analytic and numeric. Analytic models are
essentially exact calculations and are not easily adapted to spatial and temporal variability.
Numeric models are based on successive iterations of controlling formulas and adapt well to
real-world conditions where time and space are varied during simulations. A detailed
discussion of model construction is beyond the scope of this presentation, and the reader is
referred to the bibliography.
A number of models exist that contain various subsets of the components described above.
MINTEQ (U.S. Environmental Protection Agency) and WATEQ and PHREEQE (U.S. Geological
Survey) compare water composition and chemistry to thermodynamic equilibrium data bases to
forecast the tendency of certain minerals to precipitate or dissolve. BALANCE (U.S. Geological
Survey) makes mass balance calculations of changes in water chemistry from single or dual
sources.
Additional models are EQ6 (Lawrence Livermore National Laboratories), which uses a
limited reaction rate data base to predict mineral precipitation; CHEMTRN and TRANQL,
which link equilibrium chemistry with mass balance calculations; and FASTCHEM (Electric
Power Research Institute), which links equilibrium chemistry and mass balance calculations
with advective transport, sorption, and dispersion. FOWL (Electric Power Research Institute),
predicts leachant concentrations from fly ash impoundments, but it is based on an empirical,
not a thermodynamic, data base. RATAP (CANMET) utilizes very limited biochemical,
equilibrium, and kinetics data bases to predict the dissolution rates of pyrite, pyrrhotite,
chalcopyrite, and sphalerite in mine tailings ponds. Obviously, there is a dearth of models that
incorporate biochemistry. A wide variety of ground water flow models is available; one of the
most commonly used is MODFLOW3D (U.S. Geological Survey), but more sophisticated finite-
element models are now available. Most of the models mentioned need fine tuning and field
verification under widely variable conditions, but there is a serious shortage of trained
personnel to do so.
The modular numeric model is the type that might be developed to link hydrogeochemical,
biochemical, and hydrologic models as modules into one large, complex program. In each
iteration, most variables would be held constant while solving for others, but eventually, every
change of every variable in each module would affect many other variable values. Such
changes might be small or large, inversely proportional or directly proportional, linear or
exponential. Such a model would be necessarily constantly testing for limiting factors, such as
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170 Metals
depletion of a substrate or the level of a biocidal constituent. The model would have to search
data bases constantly to fill variables, such as thermodynamic constants, and it would have to
monitor surprising items, such as the decrease in aquifer hydraulic conductivity induced by the
filling of voids in the event of a burgeoning microbe population. No such model exists, but
several groups (e.g., Battelle National Laboratories and the U.S. Geological Survey) are
developing smaller programs that could become modules in such a model. A fully linked model
that begins with a steady state, combines the compositional and biologic data of aquifer
material, pore water, and injection fluid with hydrogeologic data to predict the growth and
leaching characteristics of the microbes and the leaching fluid and the composition of the fluid
at any point in time and space may be beyond the time, budget, and personnel availability of
any industry. There is a severe lack of the most basic research; the data bases necessary to
supply equations with constants are usually incomplete, inaccurate, unproven, or nonexistent.
Even if such a model could be developed, its output would be only as good as the site-specific
input data, and user costs of collecting this data may be prohibitive. Users with partial input
data sets might be tempted to use a model with default values or use published data that are
not necessarily transferable. It is then impossible to determine the defensibility of the output.
Model users frequently fall into this trap.
Beyond field verification and improvements to existing models of all varieties, the greatest
need for future research is in the area of developing new chemical and biochemical data bases
and improving existing ones. Thermodynamic equilibrium data bases have been reasonably
well developed, but they need to be supplemented with additional minerals, and the data for
many minerals is suspect. Kinetic data bases for oxidation/reduction and precipitation
reactions are extremely limited, and the kinetics are poorly understood. Biochemical data bases
are essentially nonexistent, and these will probably be the most difficult and expensive to
develop. The inventory of microorganisms involved in the dissolution and fixation of minerals
and dissolved contaminants is extremely incomplete, and little is known of the nature of the
fixation mechanisms. Nor do we understand completely population growth rates in the
presence of various substrates, energy sources, and biocides. Obviously, there are many
unanswered questions. What will happen to theoretical reaction rates when a myriad of
microorganic and chemical mechanisms compete for the same reaction? How should this
phenomenon be quantified? At what population density do stearic effects induce nonlinearity in
biochemical reactions, and how is this modeled? What organisms incorporate inorganic
contaminants in their cell structure and in what amounts?
These shortcomings must be overcome before any serious efforts at developing a fully
integrated equilibrium/kinetic-biochemical-hydrogeologic ground water model can begin.
Conclusions
Adequate information is not available to link biochemical and hydrochemical mechanisms
realistically in a single model. Future research must be concentrated toward developing new
biochemical data bases and better equilibrium/ kinetic data bases, determining how chemical
and biochemical mechanisms interact, and determining whether it is feasible to incorporate all
components into a dynamic transport model. The resulting program would be modular, require
a very powerful computer, and be very expensive and time consuming to develop. The only
defensible short-term approach is to interpret the results of several existing biochemical and
hydrogeochemical models jointly, with experienced scientists acting as the link, or interface,
among the models.
References
1. Association of Ground Water Scientists and Engineers (1987). Proceedings, Solving
Ground Water Problems with Models, Conference and Exposition: Vol. 1, Denver, CO,
February 10-12, 1987.
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R.H. Lambeth and B.C Williams 171
2. Ball, J. W., and D. K. Nordstrom (1987). WATEQ4F-A Personal Computer Fortran
Translation of the Geochemical Model WATEQ2 with Revised Data Base. U.S.
Geological Survey Open-File Report 87-50.
3. Canada Centre For Mineral And Energy Technologies (1988). Adaptation of the
Reactive Acid Tailings Assessment Program (RATAP) to Base Metal Tailings. Contract
#15SQ-2344-7-9208.
4. Criscenti, L. J., M. L. Kemner, R. L. Erikson, C. J. Hostetler, J. R. Morrey, and J. S.
Fruchter (1989). The FASTCHEM Workstation for Pre- and Postprocessing Functions.
Battelle Pacific Northwest Laboratories for the Electric Power Research Institute. EPRI
EA-5870.
5. Davis, J. A., and K. F. Hayes (eds.) (1986). Geochemical Processes at Mineral Surfaces.
190th Meeting of the American Chemical Society. Chicago, IL, September 8-13, 1985.
ACS Symposium Series 323. American Chemical Society, Washington, B.C.
6. Hem, J. D. (1985). Study and Interpretation of the Chemical Characteristics of Natural
Waters. U.S. Geological Survey Water-Supply Paper 2254.
7. Hostetler, C. J., R. L. Erikson, J. S. Fruchter, and C. T. Kincaid (1989). FASTCHEM
Package: Vol. 1-5. Battelle Pacific Northwest Laboratories for the Electric Power
Research Institute. EPRI EA-5870.
8. Krupka, K. A., R. L. Erikson, S. V. Mattigod, J. A. Schramke, and C. E. Cowan (1988).
Thermochemical Database Used by the FASTCHEM Package. Battelle Pacific Northwest
Laboratories for the Electric Power Research Institute. EPRI EA-5870.
9. McDonald, M. C., and A. W. Harbaugh (1984). A Modular Three-Dimensional Finite
Difference Ground-Water Flow Model. U.S. Geological Survey.
10. National Research Council (1990). Ground Water Models: Scientific and Regulatory
Applications. National Academy Press, Washington, D.C.
11. Parkhurst, D. L., L. N. Plummer, and D. C. Thorstenson (1982). Balance - A Computer
Program for Calculating Mass Transfer for Geochemical Reactions in Ground Water.
U.S. Geological Survey Water-Resources Investigations 82-14.
12. Patterson, J. W., and R. Passino (eds.), (1987). Metals Speciation Separation and
Recovery. Proceedings, International Symposium on Metals Speciation, Separation, and
Recovery. Chicago, IL, July 27-Aug. 1, 1986. Lewis Publishers, Inc.
13. Schwab, A. P., R. L. Schmidt, D. C. Girvin, and J. E. Rogers (1984). Chemical
Attenuation Rates, Coefficients, and Constants in Leacheate Migration. Vol. 1: A
Critical Review. Battelle Pacific Northwest Laboratories for the Electric Power Research
Institute. EPRI EA-5870.
14. Singer, P. C., and W. Stumm (1970). Acid Mine Drainage: The Rate-Determining Step.
Science, 167: 1121-1123.
15. Sumners, K. V., S. A. Gherini, M. M. Lang, M. J. Ungs, and K J. Wilkinson (1989).
MYGRT Code Version 2.0: An IBM Code for Simulating Migration of Organic and
Inorganic Chemicals in Groundwater. Tetra Tech, Inc. for the Electric Power Research
Institute. EPRI EN-6531.
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172 Metals
This page is provided for your notes:
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Appendix I - PROGRAM
Biological Remediation of Contaminated Sediments
with Special Emphasis on the Great Lakes
July 17 - 19, 1990
at the Inn on Maritime Bay
Manitowoc, Wisconsin
Tuesday, July 17
7:30 - 8:30 A.M. Registration; Coffee and Doughnuts
8:30 - 8:45 A.M. Introductory Remarks
8:45 - 12:00 Noon SESSION I. Overview of 5 Primary Areas of Concern
Session Moderator: Dave Cowgill
8:45 Indiana Harbor AOC, Robert Bunner
9:15 Field Brook-Ashtabula River Superfund Site and AOC, Pete Sanders
9:45 Buffalo River AOC, John McMahon
10:15 Break
10:35 Sheboygan River Superfund Site and AOC, Bonnie Eleder
11:05 Saginaw River AOC, Greg Goudy
11:35 Roundtable Discussion.
12:00 Noon GROUP LUNCHEON, at the Inn
1-30 - 5:30 P.M. SESSION II. Laboratory and Field Studies: Biological Degradation of
PCB's
Session Moderator: John E. Rogers
1:30 Dechlorination of Arochlors by Anaerobic Bacteria in Sediments, John F.
Quensen III
2:15 Aerobic Biodegradation of PCB's, Ronald Unterman
3:00 Break
3:30 Anaerobic Biotransformation of PCB's in Sediments, G-Yull Rhee
4:15 Remediation Pilot Study in the Sheboygan River Wisconsin, USA,
Dawn Foster
5:00 Roundtable Discussion
173
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174 Appendix I: Program
Wednesday, July 18
8:00 - 8:30 A.M. Coffee and Doughnuts
8:30 - 12:00 Noon SESSION II (continued). Laboratory and Field Studies: Biological
Degradation of PCB's and PAH's
8:30 Sequential Anaerobic - Aerobic Biodegradation of PCB's, Daniel Abramowicz
9:15 PCB Dechlorination of the Sheboygan River, William Sonzogni
10:00 Break
10:30 PAH Contamination of Hamilton Harbor, Tom Murphy
11:15 Roundtable Discussion
12:00 Noon LUNCH, on your own
1:30 - 5:30 P.M. SESSION IH. Biological Treatment of Metal Species
Session Moderator: Paulette Altringer
1:30 Bacterial Leaching of Metals from Various Matrices found in Sediments,
Removing Inorganics from Sediment-Associated Waters Using Bioaccumulation
and/or BIOFIX Beads, Paulette Altringer
2:15 Use of Wetlands and Anaerobic Bacteria to Remove Metals From Acid Mine
Drainage, and Bactericides to Deactivate Leaching Reactions, Hank Edenborn
2:55 Bioleaching of Manganese, Platinum, and Gold Ores, Betty Baglin
3:15 Break
3:45 Mechanisms of Bacterial Metals Removal From Solids, Arpad Torma and Pete
Pryfogal
4:25 Linking Biological and Hydrogeochemical Mechanisms of Sediment Leaching; I.
Field and Laboratory Data Requirements, and II. Computer Model Requirements,
Bob Lambeth
5:05 Roundtable Discussion
6:30 P.M. GROUP DINNER at the Inn
Address: EPA's Research Program on Biological Remediation of Alaskan Beaches
following the Valdez Oil Spill, John E. Rogers
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Appendix I: Program 175
Thursday, July 19
8:00 - 8:30 A.M. Coffee and Doughnuts
8:30 - 12:00 Noon SESSION IV. Laboratory and Field Studies: Biological Degradation of
PAH Compounds
Session Moderator: Chad T. JafVert
8:30 The Use of Mycobacterium species in the Remediation of PAH Waste, Carl
Cerniglia
9:15 State-of-the-Art Sediment Remediation in The Netherlands. Biological
Remediation of PAH compounds, H. J. van Veen
10:00 Break
10:20 Recent Studies on the Microbial Degradation of PAH's and Their Relevance to
Bioremediation, Dr. James Mueller (presented by John Rogers)
11:05 Fungal Degradation of PAH's, John Glaser
11:50 Roundtable Discussion
12:20 P.M. ADJOURN
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176 Appendix I: Program
This page is provided for your notes:
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Appendix II - List of Attendees
Workshop on
Biological Remediation of Contaminated Sediments
with Special Emphasis on
the Great Lakes
Inn on Maritime Bay
Manitowoc, Wisconsin
July 17-19, 1990
Daniel Abramowicz
General Electric
Research and Development Center
Bldg. K-l Room 3B19
P.O. Box 8
Schenectady, N.Y. 12301-0008
Carl Cerniglia
Dir. Microbiology Division
Natl. Center for Toxicological Research
NCTR Drive
Jefferson, Arkansas 72079
Paulette Altringer
U.S. Bureau of Mines
729 Arapeen Drive
Salt Lake City, Utah 84108
Daniel Averett
U.S. Army Corps of Engineers
Waterways Experiment Station
3909 Halls Ferry Rd.
P.O. Box 631
Vicksburg, MS 39181-0631
Scott Cornelius
Michigan Dept. of Natural Resources
Knapp Center
P.O. Box 30028
Lansing, MI 48909
Dave Cowgill
U.S. EPA
Great Lakes National Program Office
230 S. Dearborn St., 5GL-TUB-10
Chicago, IL 60604
Betty Baglin
U.S. Bureau of Mines
Reno Research Center
1605 Evans Avenue
Reno, Nevada 89512-2295
Tim Doelger
Wisconsin Dept. of Natural Resources
1125 N. Military Ave.
P.O. Box 10448
Green Bay, WI 54307
David Bowman
U.S. Army Corps of Engineers
Planning Division - Environ. Analysis
P.O. Box 1027
Detroit, Michigan 48231
Hank Edenborn
U.S. Bureau of Mines
Pittsburgh Research Center
P.O. Box 18070
Pittsburgh, PA 15236
Robert (Skip) Bunner
Indiana Dept. of Environmental Mgmt.
105 S. Meridian St.
Indianapolis, IN 46225
Bonnie Eleder
U.S. EPA
Office of Superfund
230 S. Dearborn 5H5-11
Chicago, IL 60604
177
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178
Appendix II: List of Attendees
Clell Ford
Oakridge National Laboratory
Environmental Sciences Division
P.O. Box 2008, M.S. 6351
Oak Ridge, TN 37831-6351
Vicky Harris
Wisconsin Dept. of Natural Resources
1125 N. Military Ave.
P.O. Box 10448
Green Bay, WI 54307
Dawn Foster
Blasland & Bouck Eng. P.C.
6723 Towpath Rd.
Box 66
Syracuse, N.Y. 13214
Jan Heath
Technology Applications Inc.
U.S. EPA Athens ERL
College Station Rd.
Athens, GA 30613
Rick Fox
U.S. EPA - GLNPO
230 S. Dearborn St., 5GL-TUB-10
Chicago, IL 60604
Jonathan Herrmann
U.S. EPA - Risk Reduction Eng. Lab.
26 W. Martin Luther King Dr.
Cincinnati, Ohio 45268
Steve Garbaciak
U.S. Army Corps of Engineers
Chicago District CENCC-ED-HE
111 N. Canal St. Suite 600
Chicago, IL 60606-7206
Carol Holden
Wisconsin Dept. of Natural Resources
1125 N. Military Ave.
P.O. Box 10448
Green Bay, WI 54307
Mary Garren
U.S. EPA Region 1
JFK Federal Building; MC-HRCAN3
Boston, MA 02203
John Glaser
U.S. EPA
Risk Reduction Engineering Lab.
26 W. Martin Luther King Dr.
Cincinnati, OH 45268
Paul Horvatin
U.S. EPA
Great Lakes National Program Office
230 S. Dearborn St, 5GL
Chicago, IL 60604
Chad Jafvert
U.S. EPA Athens ERL
College Station Rd.
Athens, GA 30613
Michelle Glenn
U.S. EPA Region 4
345 Courtland St. N.E.
Atlanta, GA 30365
Greg Goudy
Michigan Dept. of Natural Resources
Div. of Water - Surface Water Qual. Div.
P.O. Box 30028
Lansing, Michigan 48909
Tom Janisch
Wisconsin Dept. of Natural Resources
Bureau of Water Resources Mgt.
Box 7921
Madison, WI 53707
Dan Kaemmerer
Wise. Dept. of Natural Resources
P.O. Box 12436
Milwaukee, WI 53212
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Appendix II: List of Attendees
179
Cindy Koperski
Wisconsin Dept. of Natural Resources
Bureau of Water Resources
101 S. Webster
P.O. Box 7921
Madison, WI 53703
Bob Lambeth
U.S. Bureau of Mines
Spokane Research Center
E. 315 Montgomery Ave.
Spokane, WA 99207
Ronald Lewis
U.S. EPA
Risk Reduction Lab.
SITE Demonstration Technology Division
26 W. M.L. King Drive
Cincinnati, OH 45268
Lee Liebenstein
Wisconsin Dept. of Natural Resources
Bureau of Water Resources
101 S. Webster
P.O. Box 7921
Madison, WI 53703
Terry Lohr
Wisconsin Dept. of Natural Resources
Bureau of Water Resources Mgmt.
101 S. Webster, GEF2
P.O. Box 7921
Madison, WI 53703
Steve Luzkow
Michigan Dept. of Natural Resources
Knapps Center
P.O. Box 30028
Lansing, MI 48909
John McMahon
N.Y. State Dept. of Env. Conservation
600 Delaware Ave.
Buffalo, N.Y. 14202
Tom Murphy
Canada Center for Inland Water
867 Lakeshore Rd.
Burlington, Ontario, Canada L7R 4A6
Mary Beth Novy
U.S. EPA Region V
230 S. Dearborn 5HS-11
Chicago, IL 60604
Dave O'Malley
Wisconsin Dept. of Natural Resources
Bureau of Water Resources
101 S. Webster
P.O. Box 7921
Madison, WI 53703
Ian Orchard
Environmental Protection, Environ. Canada
25 St. Clair Ave. East; 7th Floor
7th Floor
Toronto, Ontario
Canada M4T IM2
David Pfeifer
U.S. EPA Region V
230 S. Dearborn
Chicago, IL 60604
Pete Pryfogal
Idaho National Engineering Laboratory
P.O. Box 1625-MS 2203
Idaho Falls, Idaho 83415
Ron Martin
Wisconsin Dept. of Natural Resources
Bureau of Water Resources
101 S. Webster
P.O. Box 7921
Madison, WI 53703
John Quensen
Dept. of Crop & Soil Science
Michigan State University
E. Lansing, MI 48824
G-Yull Rhee
N.Y. State Dept. of Health
Wadsworth Laboratory
Albany, N.Y. 12201-0509
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180
Appendix II: List of Attendees
John Rogers
U.S. EPA Athens ERL
College Station Road
Athens, GA 30613
Philippe Ross
INKS
607 E. Peabody Dr.
Champaign, IL 61820-6970
Pete Sanders
U.S. EPA
Mailcode 5HS-11
230 S. Dearborn St.
Chicago, IL 60604
Robin Schmidt
Wisconsin Dept. of Natural Resources
P.O. Box 7921
Madison, WI 53707
William Schmidt
U.S. Bureau of Mines
2401 E. Street, N.W.
Washington D.C. 20241
Griff Sherbin
Environment Canada
Great Lakes Action Plan Cleanup Fund
25 St. Clair Avenue East
Toronto, Ontario
Canada M4T IM2
Steve Skavroneck
Milwaukee Metropolitan Sewerage District
260 W. Seeboth
Milwaukee, WI 53021-3049
Frank Snitz
U.S. Army Corps of Engineers
Detroit District, CENCE-PD-EA
Box 1027
Detroit, MI 48231-1027
William Sonzogni
Lab. of Hygiene
465 Henry Mall
University of Wisconsin
Madison, WI 53706
Linda Talbot
Wisconsin Dept. of Natural Resources
Bureau of Water Resources
101 S. Webster
P.O. Box 7921
Madison, WI 53703
Dennis Timberlake
U.S. EPA - Risk Reduction Eng. Lab
26 W. Martin Luther King Dr.
Cincinnati, Ohio 45268
Arpad Torma
Idaho National Engineering Laboratory
P.O. Box 1625
IRC Mailstop 2203
Idaho Falls, ID 83402
Marc Tuchman
Water Quality Branch (5WQS)
U.S. EPA Region V
230 S. Dearborn St.
Chicago, IL 60604
Mark Tusler
Warzyne Engineering
P.O. Box 5385
Madison, WI 53705
Ronald Unterman
Envirogen, Inc.
3371 Route 1
Suite 203
Lawrenceville, N.J. 08648
Terese Van Donsel
U.S. EPA Region V
Mailcode 5HS-11
230 S. Dearborn St.
Chicago, IL 60604
Pat Van Hoof
U.S. EPA
Environmental Research Lab. - Athens
College Station Road
Athens, GA 30613
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Appendix II: List of Attendees
H.J. van Veen
Dept. of Environ. Technology
T.N.O.
P.O. Box 342
7300 AH Apeldoorm
Laan van Westenenk 501
The Netherlands
181
Rick Vining
WA State Dept.of Ecology
Sediment Management
Mailstop PV-11
Olympia, WA 98504
Chris Waggoner
Michigan Dept. of Natural Resources
P.O. Box 30028
Lansing, MI 48909
Steve Westenbroek
Wisconsin Dept. of Natural Resources
Bureau of Water Resources
101 S. Webster
P.O. Box 7921
Madison, WI 53703
Steve Yaksich
U.S. Army Corps of Engineers
Water Quality Section
Buffalo District, CENCB-ED-HQ
17776 Niagara St.
Buffalo, NY 14207-1339
Mike Zarull
National Water Research Institute
Environment Canada
Lakes Research Branch
CCIW, P.O. Box 5050
Burlington, Ontario, Canada L7R4A6
U S GOVERNMENT PRINTING OFFICE 1994—547-024
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