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This is to certify that the
dissertation entitled

MICROBIAL REDUCTIVE DECHLORINATION OF CHLORINATED
ETHENES: ECOLOGY AND APPLICATION

presented by

Benjamin Matthew Grifﬁn

has been accepted towards fulﬁllment
of the requirements for the

Doctoral degree in Department of Microbiology and
Molecular Genetics! Institute of
Environmental Toxicology

 

 

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MICROBIAL REDUCTIVE DECHLORINATION OF CHLORINATED
ETHENES: ECOLOGY AND APPLICATION

By

Benjamin Matthew Grifﬁn

A DISSERTATION

Submitted to
Michigan State University
in partial fulﬁllment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY

Department of Microbiology and Molecular Genetics/ Institute of Environmental
Toxicology

2003

ABSTRACT

MICROBIAL REDUCTIVE DECHLORINATION OF CHLORINATED ETHENES:
ECOLOGY AND APPLICATION

By

Benjamin Matthew Grifﬁn

Chlorinated solvents are ubiquitous groundwater pollutants and pose a signiﬁcant
risk to human health. Although tetrachloroethcne (PCB) and trichloroethene ('l‘CE) can
be completely reduced to ethene, dichloroethenes (DCEs) and vinyl chloride often
accumulate during dechlorination. All characterized PCB and TCE degrading bacteria
produce cis-DCF. as the major DCE isomer. 1 found that some anaerobic enrichment
cultures derived from geographically diverse river sediments reductively dechlorinate
PCB and TC‘E to trans- and cis-dichloroethene simultaneously and in a constant ratio of
3(ﬁ;0.5):1. F urthcr dechlorination of DCES was not observed. Two highly enriched trans
and :53 DCE-producing cultures were screened with several PCR primer pairs speciﬁc to
168 rRNA genes of genera known to contain PCE dechlorinating members. Amplicons
from both cultures were 99% identical in sequence to Dehalococcoides Sp. belonging to
the Pinellas subgroup. One cuiture also contained an amplicon 98% similar to
Dehalobactcr restrictus. These results demonstrate that trans-DCE can be a major,
microbially produced daughter product from chloroethene dechlorination. The microbial
production of trans-DCE is important to monitored natural attenuation and source
tracking of contaminant plumes.

Since it is oﬂen assumed that 'reducuve dechlorination of chlorinated cthenes is a

response to anthropogenic contamination, I surveyed for the presence of dehalogenating
bacteria in pristine environments in Antarctica. Anaerobic enrichment cultures were
established from nine melt-pond sediments on Bratina Island and Ross Island, Antarctica,
and screened for reductive dehalogenation of PCB, TCB, 2-bromophenol (2BP), 2-
chlorophenol (2GP), 3-bromobenzoate (BBBA), and 3-chlorobenozoate (3CBA). ZBP
debromination was most widespread followed by PCB dechlorination. Reduction of TCE
occurred at two sites producing cis-DCB and a mixture of trans and cis-DCB; two sites
degraded 3BBA. No dehalogenation of 2CP or 3CBA was observed. PCR amplicons
were produced using 168 rRNA gene primers for Desulfomonile from DNA from all PCB
dehalogenating samples, and for Desulﬁrromonas and Dehalococcoides from some of the
samples. These results demonstrate the presence of reductive dehalogenating bacteria in
Antarctic melt-pond sediments and suggest that they have a role in pristine, anaerobic
environments.

The use of zero-valent iron, lFe(O), for in situ remediation of chlorinated solvents
is gaining acceptance as a cost-effective remediation technology, but synergies and
inhibitions between the microbial and abiotic dechlorination processes are not well
understood. Hydrogenotrophic chlororespirers could use the cathodic hydrogen produced
by Fe(0) to enhance the rate of PCB to DCB dechlorination. Fe(0) then more rapidly
dechlorinated DCB, providing an overall synergy for the combination of processes. The
level of synergism between Fe(0) and microbial dechlorination ultimately depended on

the dechlorination products and rates of each process.

This dissertation is dedicated to my parents, Norm and Barbara Grifﬁn, for their endless
dedication to my education, success, and happiness.

iv

ACKNOWLEDGMENTS

I would like to thank my guidance committee members, Dr. John Breznak, Dr.
Robert Hausinger, and Dr. Stephen Boyd, for their support during my graduate training. I
would also like to recognize the mentoring I received from Dr. Frank Ldfﬂer, who helped
get me started as a graduate student and continues to offer advice and guidance to this
day. Finally, I would like to thank Dr. James Tiedje, my advisor, for all of his support and
patience. He allowed me to explore many creative paths, which fostered my growth as a

scientist. Thank you all.

TABLE OF CONTENTS

LIST OF FIGURES ...................................................................... ix
LIST OF TABLES .................................................................... xii
CHAPTER 1: INTRODUCTION .................................................. 1
Tetrachloroethene ...................................................................... 2
Trichloroethene ......................................................................... 4
1,1-Dichloroethene ..................................................................... 4
1 ,2-Dichloroethene ..................................................................... 5
Vinyl Chloride .......................................................................... 5
Natural production of halocarbons ................................................... 6
Chloroethene dehalogenating bacteria ............................................... 8

Desulﬁtobacterium... 10
Dehalobacter... l4
Desulfuromonas... l4
Sulfurospirillum... 15
Enterobacter... 15
“Dehalococcoides l5
Objectives ............................................................................... 16
REFERENCES ......................................................................... 19

CHAPTER 2: FREQUENCY OF OCCURRENCE OF
CHLORINATED ETHANES, CHLORINATED ETHENES, AND
BTEX AT NATIONAL PRIORITY LIST SITES: EVALUATING THE

NATIONAL DATASETS ............................................................ 22
ABSTRACT ............................................................................. 23
INTRODUCTION ..................................................................... 24
METHODS .............................................................................. 28
EPA Dataset. .................................................................. 28
ATSDR Dataset. .............................................................. 28
Toxicological Proﬁles, ToxFaqs, and CERCLIS Hazardous
Compound Data. .............................................................. 29
CERCLIS ID comparison of EPA and ATSDR datasets. 29
Pairwise co-occurrence ....................................................... 30
RESULTS ............................................................................... 31
Synonyms in the EPA Dataset. ............................................. 31
ATSDR-HazDat contaminated media ....................................... 31
Comparison of frequency of occurrence data from different sources... 34
Site speciﬁc comparison of EPA and ATSDR data ........................ 34
Pairwise co-occurrence ....................................................... 37
DISCUSSION ........................................................................... 40
REFERENCES ......................................................................... 43

vi

CHAPTER 3: ANAEROBIC MICROBIAL REDUCTIVE
DECHLORINATION OF TETRACHLOROETHENE (PCE) TO

PREDOMINATELY trans-1, 2-DICHLOROETHENE (DCE) ............ 44
ABSTRACT ............................................................................ 45
INTRODUCTION ..................................................................... 46
MATERIALS AND METHODS ..................................................... 48
Chemicals. ..................................................................... 48
Inoculum sources, microcosm preparation, and growth conditions. 48
Physiological studies. ......................................................... 49
DNA-based characterization of trans-DCB producing cultures. . . . . 49
RESULTS ............................................................................... 51
trans-DCB producing cultures. ............................................. 51
Physiological studies .......................................................... 51
DNA-based characterization of trans-DCB producing cultures. . . . . 54
DISCUSSION ........................................................................... 57
REFERENCES ......................................................................... 60

CHAPTER 4: SURVEY OF DEHALOGENATION ACTIVITY AND
PCR DETECTION OF DECHLORINATING GENERA IN

ANTARCTIC MELT-POND SEDIMENTS .................................... 64
ABSTRACT ............................................................................ 65
INTRODUCTION ..................................................................... 67
MATERIALS AND METHODS ..................................................... 70
EXPERIMENTAL DESIGN ......................................................... 70

Sample collection, microcosm preparation, and grth conditions. 71

Analytical methods ............................................................ 73

PCR detection of dehalogenating genera ................................... 74
RESULTS ............................................................................... 76
DISCUSSION ............................................................................ 87
REFERENCES ......................................................................... 92

CHAPTER 5: SYNERGISM OF ZERO-VALENT IRON AND
MICROBIAL DECHLORINATION OF CHLORINATED SOLVENTS 96

ABSTRACT ............................................................................. 97
INTRODUCTION ..................................................................... 99
MATERIALS AND METHODS ..................................................... 102
Chemicals ....................................................................... 102
Iron pretreatment ............................................................... 102
Bacterial cultures and medium preparation ................................ 103
H2 consumption ................................................................ 105
Analytical techniques ......................................................... 105
RESULTS ............................................................................... 107

vii

DISCUSSION ........................................................................... 1 17

REFERENCES ......................................................................... 121
SUMMARY AND CONCLUSIONS .............................................. 124
REFERENCES ......................................................................... 130

APPENDIX 1: F E(0) STIMULATION OF PCB REDUCTIVE

DECHLORINATION ................................................................ 131
ABSTRACT ............................................................................. 132
INTRODUCTION ..................................................................... 133
RESEARCH DESIGN AND METHODS .......................................... 141
Research Design ............................................................... 141
Sediment microcosms ......................................................... 141
Extraction and analysis of PCBs ............................................. 142
Headspace analysis ............................................................ 142
Iron, sulfate and sulﬁde analysis ............................................. 142
RESULTS ............................................................................... 143
Methane ......................................................................... 143
Hydrogen ........................................................................ 143
Volatile fatty acids ............................................................ 143
Acid extractable aqueous F e(II) ............................................. 148
PCB analysis ................................................................... 148
DISCUSSION ........................................................................... 151
REF ERENCBS ......................................................................... 153

APPENDIX II: MICROBIAL DEHALORESPIRATION WITH 1,1,1-
TRICHLOROETHANE ............................................................. 1 54

viii

LIST OF FIGURES

Figure 1.1. One electron transfer, or radical, mechanism as proposed for PCB
dechlorination by Holliger et a1. 1999 ...................................................... 12

Figure 1.2. Dechlorination patterns of Chloroethene dechlorinating bacterial
populations. Arrows represent the transformation of substrates (right side) to
dechlorination products (left side). Metabolic, or energy yielding, dechlorination

steps are solid lines, and cometabolic transformations are dashed lines. ............. 13

Figure 2.1. Reported occurrences of vinyl chloride at National Priority List site by
various data sources. EPA — US Environemtnal Protection Agency; HazDat —

Agency for Toxic Substances and Disease Registry HazDat database, Final +

Deleted sites only; CERCLA Priority List - The Comprehensive Environmental
Response, Compensation, and Liability Act Priority List of Hazardous

Compounds; Tox. Prof.,ToxFaq’s, PHS — Toxicological Proﬁle for Vinyl

Chloride, ToxFAQs, Public Health Statements ........................................... 26

Figure 3.1. Typical time course for TCE dechlorination by the Tahqumenon River
enrichment culture (TQ). Lactate (5mM) was added as electron donor, and 25

umol of TCE was added to the 100 ml bicarbonate-buffered medium.

Chloroethenes were measured by headspace GC-F ID. No vinyl chloride, ethene,
ethane, or methane was detected ............................................................ 53

Figure 3.2. A. Inﬂuence of pH on the ratio of trans-DCB and cis-DCE isomer
production by the Parfume River culture. B. Extent of dechlorination at pH 6.8,

7.5, and 8.2. Lactate (5mM) was added as electron donor, and 25 umol of TCE

was added to 100 ml bicarbonate-buffered medium. Chloroethenes were measured

by headspace GC-FID after 4 months of incubation. No vinyl chloride, ethene,

ethane, or methane was detected ............................................................ 55

Figure 4.1 Methane production in anaerobic complex medium inoculated with
Antarctic melt pond sediment ............................................................... 77

Figure 4.2. 2-Bromophenol consumption in Antarctic microcosm sediments.
Cultures were fed to a ﬁnal concentration of 0.25 mM per feeding. Bar height
reﬂects the cumulative number of feedings aﬁer 8—months .............................. 78

Figure 4.3 3-Bromobenzoate and 3-chlorobenzoate consumption in Antarctic
microcosm sediments and production of benzoate. Cultures were fed an initial
concentration of 0.25 mM of each halobenzoate. Consumption is shown for an 8-
month incubation .............................................................................. 79

Figure 4.4. 2-Bromophenol debromination in 1% transfers from halophenol
microcosms after 8 weeks of incubation with 10 mM lactate as electron donor.

Cultures that consumed the initial 0.25 mM 2-bromophenol were amended with
an additional 0.25 mM feeding .............................................................. 80

Figure 4.5. PCB conversion to TCE in Antarctic melt-pond sediment microcosms
amended with lactate (10 mM) and 20 umol PCE. Boulder Dry Pond received an
additional feeding of PCB after the initial amount was transformed. .................. 82

Figure 4.6. PCR detection of Desulfuromonas in hydrogen + acetate and PCB fed
microcosms after 12 months of incubation ................................................ 83

Figure 4.7. PCR detection of Desulfomonile in hydrogen + acetate and PCB fed
microcosms after 12 months of incubation ................................................ 85

Figure 4.8. PCR detection of Dehalococcoides in hydrogen + acetate and PCB fed
microcosms after 12 months of incubation ................................................ 86

Figure 4.9. Summary of dehalogenation activity in melt-pond sediment
microcosms after 16 months of incubation ................................................ 88

Figure 4.10. l6S-rDNA primers targeting known dechlorinators were tested on
DNA extracted from PCB-amended enrichment cultures 90

Figure 5.1. Dual chambered batch reactor. Each side of the reactor was

approximately 160 ml, and the upper portions of each were connected to allow for

a common headspace. The tops were sealed with butyl rubber stoppers and

aluminum crimp tops ......................................................................... 104

Figure 5.2 Dechlorination of 20 umol of PCB in the presence of 80 mg Fe(0) by
Dehalobacter restrictus, a hydrogen utilizing PCE to cis-DCE dechlorinating
bacterium. In the absence of F e(0) and other electron donors, D. restrictus does

not signiﬁcantly reduce PCE ................................................................ 108

Figure 5.3. Dechlorination of 95 umol PCE in 0.2 ml hexadecane in the presence
of 200 mg of Fe(0) after 10 Mdays of incubation (A) with S. multivorans and (B)
withoutS. multivorans... .. 109

Figure 5.4. Dechlorination of 20 umol PCB in in dual chambered batch reactors
in the presence of S. multivorans (A) with 1000 mg of Fe(0) and (B) without after
14 days of incubation ......................................................................... 110

Figure 5.5. Dual chamber batch reactor: Dechlorination of PCB in the presence of
200 ug F e(0) with (diamonds, replicate A, and squares, replicate B) and without
(circles) Dehalobacter restrictus in the accompanying chamber ....................... 1 l 1

Figure 5.6. Dechlorination of TCE in the presence of Fe(0) with and without D.
restrictus ....................................................................................... 1 l3

Figure 5.7. Abiotic, Fe(0)-mediated dechlorination of PCB (diamond), TCE
(square), and cis-DCE (triangle) as parent compounds .................................. 114

Figure 5.8. Effect of inhibitors on F e(0) mediated dechlorination of 20 ul 1,1,1-
trichloroethane. Inhibitors were cobalamin (Cob), sulﬁde (82-), DTT, cysteine
(Cys), rezazurin (Res), AQDS, and humic acids (humics) .............................. 115

Figure AI.1. Generalized structure of polychlorinated biphenyl (PCB) ............... 133

Figure AI.2. Model for FeSO4 stimulation of PCB dechlorination. 1. Sulfate is an
alternate electron acceptor for growth of the dechlorinating population. 2. This
larger bacterial population is able to dechlorinate PCBs at a faster rate. 3. Sulﬁde
from sulfate reduction is detoxiﬁed by co-precipitation with ferrous iron ............ 135

Figure Al.3. Proposed stimulation of PCB by F e(0). 1. Direct dechlorination of

PCBs by F e(0). 2. Fe(0) reduction of water to form ferrous iron and hydrogen.

3. Cathodic hydrogen is an electron donor for microbial dechlorination of PCBs.

4. Ferrous iron is also a thermodynamically favorable electron donor for abiotic

or microbial reductive dechlorination. 5. Ferrous iron co-precipitates endogenous
sulﬁde ........................................................................................... 139

Figure AI.4. Proposed model for the synergism between F680; and Fe(0)
amended PCB reductive dechlorination .................................................... 140

Figure AI.5. Methane production in sediment microcosms containing PCBs. . . . 144

Figure AI.6. Headspace hydrogen in PCB amended microcosms. Hydrogen
production varied from > 1 ppmV to > 40% (note scales) .............................. 145

Figure AI.7. Volatile fatty acids in supernatant of sediment microcosms after 16
weeks of incubation ........................................................................... 146

Figure AI.8. Microcosm pH after 16 weeks of incubation .............................. 147

Figure AI.9. 0.5 N HCl-extractable aqueous F e(II) from sediment microcosms
after 16 weeks of incubation ................................................................. 149

Figure AI.10. PCB anaylsis of Arochlor 1242 amended sediment microcosms
aﬁer 16 weeks of incubation ................................................................. 150

xi

LIST OF TABLES

Table 1.1. Production of the top industrial organohalogen compounds in the

United States. Data from Haggbltim and Bossert, 2003 ................................. 3
Table 1.2. Chlorinated ethene halorespiring bacteria .................................... 9
Table 1.3. Characteristics of Chloroethene reductive dehalogenases ................... 11

Table 2.1. Synonyms of selected chemicals in the EPA database. The value
adjacent to the CAS Reg. # is the number of sites after elimination of duplicates. .. 32

Table 2.2 Occurrence of chemicals in speciﬁc media at NPL sites according to the
HazDat data set ................................................................................ 33

Table 2.3. Comparison of the frequency of occurrence of chemicals at NPL sites.
CERCLA ‘Rank’ is that chemical’s rank on the 2001 Priority List of Hazardous
Compounds. ‘EPA and ATSDR’ data is the sum of unique and shared sites from

each datasource as determined by site IDs ................................................ 35

Table 2.4. Site speciﬁc comparison of EPA—synonym corrected data and ATSDR
data. ............................................................................................. 36

Table 2.5a. Co-occurrence of selected chemicals at NPL sites using EPA +
ATSDR dataset. Co-occurrence is expressed as percent of total NPL sites .......... 38

Table 2.5b. Co-occurrence of selected chemicals at NPL sites using EPA +
ATSDR dataset. Co-occurrence is expressed as percent of each chemical’s
abundance ...................................................................................... 39

Table 3.1 Summary of microcosms and enrichment cultures that produced
mixtures of trans-DCB and cis-DCE ....................................................... 52

Table 4.1. Melt-pond sediment sampling sites on the McMurdo Ice Shelf and
Cape Evans, Antarctica. Sediment was sampled within 0.5 m of the pond shore

and stored at 5 °C for shipment from Antarctica to Michigan State University ...... 72

Table 4.2. PCR primer sets used for the detection of dechlorinating genera in
sediment microcosms ......................................................................... 75

xii

CHAPTER 1

INTRODUCTION

 

 

Organohalogen compounds are ubiquitous. They are found in innumerous
everyday materials (PVC), clothing (Gor’tex), pharmaceuticals (vancomycin), and even
toothpaste (triclosan). Soil has a signiﬁcant organochlorine content, the ocean is a net
source for diverse volatile and semi-volatile compounds, and the atmosphere is a large
reservoir. The high anthropogenic production of organohalogens to meet industrial and
consumer demands has resulted in the substantial release of these compounds into the
environment (Table 1). The following sections summarize the industrial production, use
and fate of chlorinated ethenes as reported in the Toxicological Proﬁles for each
compound. These sections will be followed by summaries of natural organohalogen
production and the speciﬁc objectives of this dissertation.

Tetrachloroethene [1]

Tetrachloroethene, also known as perchloroethylene (PCB), is used commercially
as a chlorinated hydrocarbon solvent and as an intermediate for chemical syntheses. A
1995 estimate of end-uses for PCB showed: 55% for chemical intermediates, 25% for
metal cleaning and vapor degreasing, 15% for dry cleaning and textile processing, and
5% for other unspeciﬁed uses. PCE is a volatile organic compound, and because of this
property releases are primarily to the atmosphere. PCE is also released to surface water
and land from sewage sludge and other liquid and solid wastes. In these cases the high
vapor pressure and Henry’s law constant for PCE usually results in its rapid volatilization
to the atmosphere. PCB has relatively low water solubility and has medium-to-high

mobility in soil, thus its residence time in surface environments is not expected to be

Table 1.1. Production of the top industrial organohalogen compounds in the United States

Data from Haggbltim and Bossert, 2003 [8].

 

Annual productiona

 

Compound (x106 55) Primary use
1,2-Dichloroethane 8140 Production of vinyl chloride, solvent
Vinyl Chloride 6235 Production of PVC, synthesrs of

chlorrnated solvents
Chloromethane 385 Production of silicones
Chloroform 215 Manufacture of HCFC-22, solvent
1,1,1-Trichloroethane 205 Solvent, dry cleaning
Dichloromethane 160 Solvent, paint remover
Carbon Tetrachloride 140b Solvent (production of CF Cs)
Tetrachloroethene 125 Solvent, dry cleaning
Chloroethane 700 Production of ethylcellulose

 

2‘ Production in 1993 unless noted

bProduction in 1991
c Production in 1990

more than a few days. However, PCE may persist in the atmosphere for several months
or more. The half-life in groundwater depends greatly on the prevailing geochemical
conditions.

Trichloroethene [2]

End-uses for tetrachloroethene (TCE) include: vapor degreasing of fabricated
metal parts, 80%; chemical intermediates, 5%; and miscellaneous uses, 5%; exports
accounted for an additional 10%. The most important use of TCE, vapor degreasing of
metal parts, is closely associated with the automotive and metals industries. Most of the
TCE used in degreasing operations evaporates into the atmosphere. Once in the
atmosphere, TCE is degraded by reaction with hydroxyl radicals; the estimated half-life
for this process is approximately 7 days. TCE’s relatively short half-life indicates that it
is not a persistent atmospheric compound. When deposited in surface waters or on soil
surfaces, TCE volatilizes into the atmosphere, buts its high mobility in soil means
substantial portions percolate to the subsurface. In the subsurface, TCE is only slowly
degraded and may be relatively persistent depending on the aquifer conditions. Based on
available federal and state surveys in 1997, between 9% and 34% of the drinking water
supply sources tested in the United States have TCE contamination.
1,1-Dichloroethene l3]

1,1-Dichloroethene (1,1-DCE) is used as an intermediate for captive organic
chemical synthesis and in the production of polyvinylidene chloride copolymers.
Polyvinylidene chloride copolymers are used to produce ﬂexible ﬁlms for food

packaging (e. g. “Saran wrap”). Atmospheric releases of 1,1-DCE are the greatest source

of ambient 1,1-DCB. As a result of waste disposal, smaller amounts of the chemical are
released to surface water and soil. Most of the 1,1-DCB released to the environment
partitions to air or water, including groundwater.

l,2-Dichloroethene[4]

1,2-DCB is produced as individual cis or trans isomers or as a mixture and is used
primarily as a chemical intermediate in the synthesis of other chlorinated solvents. 1,2-
DCE also been used as a solvent for waxes, resins, acetylcellulose, perfumes, dyes,
lacquers, thermoplastics, fats, and phenols. 1,2-DCB is used in the extraction of rubber,
as a refrigerant, in the manufacture of pharmaceuticals and artiﬁcial pearls, and in the
extraction of oils and fats from ﬁsh and meat. 1,2-DCE was also reported as a low-
temperature extraction solvent for decaffeinated coffee. The trans-DCB isomer is more
widely used in industry than either the cis isomer or the commercial mixture.

Most of the 1,2-DCE released into the environment either enters the atmosphere or
groundwater, where it may be subject to biotic or abiotic degradation processes. The
estimated atmospheric lifetimes for cis and trans 1,2-DCE due to reaction with
photochemically generated hydroxyl radicals are 12 and 5 days, respectively. When
released to surface water, volatilization is expected to be the primary fate process, with a
reported estimated half-life of about 3-6 hours in a model river. 1,2-DCE volatilizes
rapidly from moist soil surfaces and leaches through subsurface soil, to groundwater.
Vinyl chloridelS]

Vinyl chloride is an important industrial chemical because of its wide variety of
end-use products and the low cost of production of its polymer. Polyvinyl chloride (PVC)

is one of the most efﬁcient construction materials available when analyzed on an energy-

equivalent basis. Major end-use products include PVC products, such as automotive parts
and accessories, furniture, packaging materials, pipes, wall coverings, and wire coatings,
as well as vinyl chloride-vinyl acetate copolymer products, such as ﬁlms and resins. End
use estimates for 1992 indicated 98% of vinyl chloride monomer production was for
making PVC and various PVC copolymers; the other 2% was for miscellaneous uses.
Most of the vinyl chloride released into the environment is eventually transported to the
atmosphere with lesser amounts transported to groundwater. In the atmosphere, vinyl
chloride is removed by reaction with photochemically generated hydroxyl radicals with a

half life of 1-2 days.

Natural production of halocarbons

In 2003, Gribble reported over 3700 naturally formed halocarbons had been
identiﬁed [6]. These halogenated compounds arise in nature from biogenic and abiotic
production. Organochlorines are represented in nearly every organic chemical class and
are a major group of natural products [6]. Approximately half of the known naturally
occurring halocarbons are chlorinated; most of the other half is brominated [7]. Iodinated
compounds are infrequent and ﬂuorinated metabolites are extremely rare [8].
Halogenation reactions are performed by a variety of marine organisms including
seaweeds, sponges, corals, tunicates, and bacteria [6]. Biogenic production by terrestrial
organisms is also documented for plants, fungi, lichens, bacteria, insects, frogs, and
humans [6]. The diversity of organohalogen-producing organisms is great, and the
production of certain compounds is sufﬁcient to affect global budgets. For example,

chloroform is produced by several Australian termite species to levels within their

mounds up to 1000 times greater than the ambient concentration [9]. Production from
these termites was calculated to account for up to 15% of the global chloroform
emissions [6].

Several temperate and subtropical marine macroalgae produce PCE at the rate of
0.0026-8.2 ng/g fresh weight/hour. They also generate TCE, usually at greater rates. The
emission of PCB and TCE from algae to the atmosphere is estimated to impact the global
atmospheric chlorine budget [6].

In addition to biogenic production, organohalogens are formed via geothermal
processes (e.g. volcanoes), natural combustion (e.g. forest ﬁres), and early diagenetic
processes in soil [10]. All of these reactions involve radical chemistry. Volcanoes
produce a diversity of volatile halocarbons, including relatively rare ﬂuorinated
compounds. Volcanic production starts from methane, ethene, and ethyne in the presence
of halides and is catalyzed on hot mineral surfaces [10]. Lava gas samples contain a
variety of small aliphatic halocarbons including tetrachloroethene and trichloroethene;
however, volcanic emissions are not estimated to affect global atmospheric budgets [1 l].
Biomass burning also involves high temperatures where, in the presence of halide ions,
primarily methylhalides are formed. Early diagenesis in soils and sediments occurs at
ambient temperatures, is driven by redox-sensitive elements (e.g. iron), and involves the
radical chemistry of organic material in the presence of halides [10]. Halogenation by
early diagenic processes forms volatile organohalogens, haloacetic acids, and total

organic halogen (i. e. halogenated humus).

The implication of natural halogen cycles for environmental clean-up was
recently eloquently expressed:

“T he wealth of reliable and unambiguous evidence that chlororganics
may arise through natural processes, are ubiquitous and are formed in
large quantities focuses attention on the means by which such compounds
containing carbon-chlorine bonds are broken down in the environment.
Bearing in mind that production of organochlorine compounds is
probably an ancient characteristic of the natural world, the
concentrations now present require the existence of natural processes for
effecting their breakdown and ultimate mineralization. Understanding of
these processes will be relevant in the search for effective technologies to
bring about the purposeful degradation of persistent man-made

chlorinated pollutants and the purposeful remediation of contaminated
media or their removal from waste streams. ” — Neil Winterton 2000.

Chloroethene dehalogenating bacteria

Known Chloroethene halorespiring organisms are classiﬁed into four phylogentic
groups: low G+C Gram positive Bacteria, S—subgroup of the Proteobacteria, e—subgroup
of the Proteobacteria, and the “Dehalococcoides” group that is most closely related to the
green nonsulfur bacteria (Table 2). Use of either hydrogen or acetate as electron donors
can be used to broadly categorize Chloroethene halorespirers (Table 2), although some
strains can also use other fermentation products such as formate, pyruvate, and lactate.
Likewise, most strains can also use other inorganic and organic compounds as electron
acceptors. Dehalococcoides and Dehalobacter strains, however, are metabolically
restricted and appear to be obligate halorespiring hydrogenotrophs. Hydrogen-utilizing,

halorespiring bacteria have low hydrogen thresholds and, based on thermodynamic

Table 1.2. Chlorinated ethene halorespiring bacteria (adapted from [15])

 

Dechlorinationa

 

Electron Donors

 

Or anism Stra'n
g l Substrate Product H2 or acetate
Dehalobacter restrictus PER-K23 PCE cis-DCE H2
Dehalobacter restrictus TEA PCE cis-DC E H;
Dehalococcoides etheneogenes 195 PCE VC(Ethene) H2
Dehalococcoides sp. F L2 TCE (PCE) VC(Ethene) H2
Dehalococcoides sp. BAVl DC Es, VC Ethene H2
Desulﬁtobacterium frappieri TC El PC E cis-DCE H2
Desulfitobacterium sp. Y51 PCB cis-DCE H2
Desulﬁtobacterium sp. PCEl PC E TC E H2
Desulﬁtobacterium sp. PCE-S PC E cis-DCE H2
Desul/itobacterium sp. Vietl PC E TCE H2
Desulﬁtromonas chloroethenica 'I'I‘4B PC E cis-DCE Acetate
Desulﬁiromonas michiganensis 881 PC E cis—DCE Acetate
Desul/iiromonas michiganensis BRSl PCE cis-DCE Acetate
Enterobacter sp. MS-l PCE cis-DCE Acetate
Enterobacter agglomerans CDC 4~74 PC E cis-DCE n.r.
Sulﬁtrospirillum halorespirans PC E cis-DCE H2
Sulfurospirillum multivorans PCE-M2 PCE cis-DCE H2

 

3 Compounds in parentheses are only used or produced cometabolically.

n.r. — not reported

calculations, should out compete methanogens, acetogens, and sulﬁdogens in anaerobic

environments [1 2].

Puriﬁed Chloroethene reductive dehalogenases are corrinoid-containing enzymes
(Table 3) and are suggested to catalyze dechlorination through a dissociative one electron
transfer [13](Figure 1). Gene and genome sequencing of Chloroethene-degrading strains
and other reductively dehalogenating bacteria suggest that individual strains may contain
several dehalogenase-like gene sequences, most of which have unknown functions [14].
The following sections describe the genera of Chloroethene dechlorinating

bacteria and their dechlorination patterns (Figure 2).

Desulﬁtobacterium

Desulﬁtobacterium is an important, newly described, dehalogenating genus in the
low G+C Gram-positive Bacteria [15]. Ten of the 12 isolates in this genus are
dechlorinators and can dechlorinate chlorophenolic compounds and/or PCE or TCE.
There are ﬁve Chloroethene dechlorinating strains. Strains PCE-S, TCEl, and Y51
dechlorinate PCB and TCE to cis-DCE. Strains Vietl and PCB] only reduce PCB to
TCE. Desulfitobacterium strains are metabolically versatile, utilizing several electron
donors including: hydrogen, formate, lactate, pyruvate, and butyrate. In addition to
chlorinated compounds, alternate electron acceptors include nitrate, sulﬁte, thiosulfate,
fumarate, and some metals. The genome of the non-chloroethene-degrading strain, D.
hafniense strain DCB-2, was sequenced, and analysis of the genome revealed nine

putative dehalogenase-like sequences [14].

10

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Figure 1.1. One electron transfer, or radical, mechanism as proposed for PCE
dechlorination by Holliger et al. 1999 [13].

C" ,CI

He ﬂ.“ ﬁlm“ .. 9'

cu Cl

.- l»
c-

H *1 c'x—W”

c1 °
C1
or ‘C

“U
CIH/—\ Cl

 

 

Figure 1.2. Dechlorination patterns of Chloroethene dechlorinating bacterial populations.
Arrows represent the transformation of substrates (right side) to dechlorination products
(left side). Metabolic, or energy yielding, dechlorination steps are solid lines, and
cometabolic transformations are dashed lines.

PCE ——>TCE —’ DCEs —*VC —> ETH

>
Some Desulﬁtobacterium

 

 

>
Some Desulfitobacterium, Desulfuromonas,
Sulfurospirillum, Dehalobacter, Clostridium, Enterobacter

 

> ————— >
Dehalococcoides ethenogenes strain 195 Cometabolic

 

> >
Cometabolic Dehalococcoides sp. strain FL2 Cometabolic

 

’
Cometabolic Dehalococcoides sp. strain BAV1

l3

Dehalobacter

Dehalobacter are phylogenetically related to Desulﬁtobacterium as members of
the low G+C Gram positive Bacteria [15]. Two of the three Dehalobacter strains utilize
PCE or TCE as electron acceptors (the other strain, TCAl, dechlorinates 1,1,1-
trichloroethane and 1,1-dichloroethane). Dehalobacter are obligate, hydrogen-oxidizing,
halorespirers. The chloroethene-dechlorinating D. restrictus strains PER-23 and TEA

produce cis-DCE as the dechlorination end product.

Desulfuromonas

The genus Desulfuromonas is part of the S-subdivision of the Proteobacteria [15].
Three strains belonging to D. chloroethenica (l) or D. michiganensis (2) are capable of
PCB dechlorination to cis-DCE with acetate, but not H2, as electron donor. D.
michiganensis strain B81 and strain BRSl also use lactate, pyruvate, succinate, malate,
and fumarate, but not hydrogen, formate, ethanol, propionate, or sulﬁde as electron
donors. In addition to PCB and TCE, these strains also use fumarate, malate, ferric iron,
and sulfur as electron acceptors. Strain BB1 was isolated from pristine sediment from the
Pére Marquette river, Michigan. Strain BRSl was isolated from PCE-contaminated
aquifer material in Oscoda, Michigan. Environmental surveys with PCR primers to
detect Desulfuromonas dechlorinating populations suggest that these organisms may be

widespread in anaerobic environments [16].

14

Sulfurospirillum

Sulfurospirillum is a member of the s-subdivision of the Proteobacteria [17]. Two
species, S. halorespirans, and S. multivorans, dechlorinate PCE and TCE to cis-DCE. S.
mulitvorans, formerly Dehalospirillum multivorans, can use hydrogen in addition to
pyruvate, formate, lactate, ethanol, glycerol and sulﬁde as electron donors. Nitrate
fumarate, arsenate, and selenate are used as electron acceptors; S. multivorans can also
ferment fumarate, and pyruvate [13]. S. multivorans has a constitutively expressed PCE
reductive dehalogenase that produces cis-DCE as the dechlorination product [18, 19].
The PCE dehalogenase has been puriﬁed and characterized; the gene encoding the
enzyme was also cloned and sequenced [18]. The dehalogenase appears to contain a
novel corrinoid cofactor [20]. A strain of S. multivorans without the ability to

dechlorinate was shown to have a mutation in the cofactor biosynthesis pathway [21].

E nterobacter

Enterobacter belongs to the y-subdivision of the Proteobacteria [15].
Enterobacter sp. strain MS-l and E. agglomerans CDC 4-74 dechlorinate PCE to cis-
DCE. These strains are facultative anaerobes, and strain MS-l was shown to use acetate

as an electron donor for dechlorination.

“Dehalococcoides ”
Dehalococcoides forms a deeply branching taxon most closely related to the
green non-sulfur bacteria [15, 22]. The four isolates from this genus are all obligatory

hydrogen-oxidizing halorespirers. Dehaloccoides isolates do not contain peptidoglycan

15

which makes them resistant to high concentrations of cell wall synthesis inhibitors such
as ampicillin and vancomycin. Three strains grow on Chloroethenes; the other, strain
CBDBl, dechlorinates chlorobenzenes and dioxins. D. ethenogenes strain 195
dechlorinates PCE to VC with further cometabolic conversion of VC to ethene. Two
enzymes are responsible for Chloroethene dechlorination. One dechlorinase reduces PCE
to TCE and the other dechlorinates TCE to VC. Dehalococcoides sp. strain F L2 also
dechlorinates TCE to VC, but it only dechlorinates PCE to TCE cometabolically [15].
The recently isolated Dehalococcoides sp. strain BAV1 is the only pure strain capable of
the metabolic conversion of VC to ethene [23]. BAV1 also converts all three DCE
isomers to ethene, but only cometabollically dechlorinates higher chlorinated ethenes.

Several molecular methods have been developed for the detection of
Dehaloccoides[24-26], and PCR-based methods have detected Dehalococcoides in a
variety of contaminated and pristine sediments capable of PCB dechlorination [24, 25].
Dehalococcides related sequences are often detected in 16S rRNA clone libraries ﬁ'om
anaerobic environments including sites not expected to have dechlorinating activity.

The genome of D. ethenogenes strain 195 was sequenced. Strain 195 contains 17
dehalogenase-like sequences [27]. Only the PCB and TCE dechlorinating enzymes,

however, have known functions.

Objectives
I have addressed the following four objectives in this Dissertation:
1. _I_assessed the mortance of chlorinated ethenesirs environmentgl contaminants

by evaluating their occurrence at US EPA Natiorlal Priority List sites (Chather 2).

16

The hazard of contaminants is, in part, determined by their occurrence in the
environment. Assessing chemical occurrence in nature, however, is difﬁcult. The
US EPA National Priority List is an important set of contaminated sites as they
have been designated as the Nation’s most contaminated and are undergoing long-
term remediation. Several publicly available databases contain the frequency of
chemical occurrence at these sites. I describe substantial inconsistencies between
these data sources and evaluate the frequency of occurrence of chlorinated

ethenes and other volatile organic compounds based on the best available data.

I chargcterized enrichment cultures with a novel PCE dechlorination pathwaﬁ that
produces trans-DCB as the mjaior dechlorigrtion product (Chapter 3. All
previously characterized bacteria and enrichment cultures produce cis-DCE as the
sole DCE isomer during dechlorination of PCB and TCE; thus, trans-DCB was

not believed to be a microbial end product of Chloroethene degradation.

I surveyed reductive dehalogenation of chlorinated ethenes and haloaromatic
compounds in pristine environments of Antarctica (Chapter 4L Reductive
dehalogenating bacteria are increasingly being found in pristine environments, but

their roles in nature are largely unknown.

I evaluated the role of halorespiring byterig during Fe(OI-remediation of

chlorinated ethenes (Chapter 5). The importance of halorespiring bacteria for

17

bioremediation is well known; however, microbes may also contribute to
dechlorination in remediation strategies designed to stimulate abiotic, chemical
dechlorination. Synergism between microbial and abiotic dechlorination is shown

to occur in only select conditions.

18

10.

11.

REFERENCES

Toxicological proﬁle for tetrachloroethylene. 1997, US. Department of Health
and Human Services,Public Health Service, Agency for Toxic Substances and
Disease Registry.

Toxicological proﬁle for trichlorethylene. 1997, US. Department of Health and
Human Services, Public Health Service, Agency for Toxic Substances and
Disease Registry.

Toxicological proﬁle for 1,1-dichloroethene. 1994, US. Department of Health
and Human Services, Public Health Service, Agency for Toxic Substances and
Disease Registry.

Toxicological proﬁle for 1,2-dichloroethene. 1996, US. Department of Health
and Human Services, Public Health Service, Agency for Toxic Substances and
Disease Registry.

Toxicological proﬁle for vinyl chloride. 1997, US. Department of Health and
Human Services, Public Health Service, Agency for Toxic Substances and
Disease Registry.

Gribble, G.W., The diversity of naturally produced organohalogens.
Chemosphere, 2003. 52: 289-297.

Gribble, G.W., The diversity of naturally occurring organobromine compounds.
Chemical Society Reviews, 1999. 28: 335-346.

Haggblém, M.M. and ID. Bossert, Halogenated Organic Compounds - A Global
Perspective, in Dehalogenation: microbial processes and environmental
applications, M.M. Haggbldm and ID. Bossert, Editors. 2003, Kluwer Academic
Publishers: Boston. p. 3-29.

Khalil, M.A.K., et al., The Inﬂuence of Termites on Atmospheric Trace Gases —
CH4, C02, CHCI3, N20, C0, H2, and Light-Hydrocarbons. Journal of
Geophysical Research-Atmospheres, 1990. 95: 3619-3634.

Scholer, HF. and F. Keppler, Abiotic Formation of Organohalogens During
Early Diagenetic Processes, in Natural Production Organohalogen Compounds,
G.J. Gribble, Editor. 2003, Springer-Verlag: Berlin. p. 63-84.

Jordan, A., Volcanic Formation of Halogenated Organic Compounds, in Natural
Production Organohalogen Compounds, G.J. Gribble, Editor. 2003, Springer-
Verlag: Berlin. p. 121-139.

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l3.

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22.

Ldfﬂer, F.E., J .M. Tiedje, and RA. Sanford, Fraction of electrons consumed in
electron acceptor reduction and hydrogen thresholds as indicators of
halorespiratory physiology. Appl. Environ. Microbiol., 1999. 65: 4049-56.

Holliger, C., C. Regeard, and G. Diekart, Dehalogenation by Anaerobic Bacteria,
in Dehalogenation .° microbial processes and environmental applications, M.M.
Haggblbm and ID. Bossert, Editors. 2003, Kluwer Academic Publishers: Boston.
p. 115-157.

Villemur, R., et al., Occurrence of several genes encoding putative reductive

dehalogenases in Desulﬁtobacterium hafniense/frappieri and Dehalococcoides
ethenogenes. Can. J. Microbiol., 2002. 48: 697-706.

Lbfﬂer, F.E., et al., Diversity of dechlorinating bacteria, in Dehalogenation :
microbial processes and environmental applications, M.M. Haggbldm and ID.
Bossert, Editors. 2003, Kluwer Academic Publishers: Boston. p. 53-87.

Ldfﬂer, F .E., et al., 16S rRNA gene-based detection of tetrachloroethene-

dechlorinating Desulfuromonas and Dehalococcoides species. Appl. Environ.
Microbiol., 2000. 66: 1369-74.

Luijten, M.L.G.C., et al., Description ofSulfurospirillum halorespirans sp. nov.,
an anaerobic, tetrachloroethene-respiring bacterium, and transfer of

Dehalospirillum multivorans to the genus Sulfurospirillum as Sulfurospirillum
multivorans comb. nov. Int. J. Syst. Evol. Microbiol., 2003. 53: 787-793.

Neumann, A., G. Wohlfarth, and G. Diekert, Tetrachloroethene dehalogenase
from Dehalospirillum multivorans: cloning, sequencing of the encoding genes,

and expression of the pceA gene in Escherichia coli. J. Bacteriol., 1998. 180:
4140-5.

Eisenbeis, M., K.P. Bauer, and M.H. Scholz, Studies on the dechlorination of
tetrachloroethene to cis-1,2-dichloroethene by Dehalospirillum multivorans in
bioﬁlms. Water Science and Technology, 1997. 36: 191-198.

Neumann, A., et al., Tetrachloroethene reductive dehalogenase of

Dehalospirillum multivorans: substrate specificity of the native enzyme and its
corrinoid cofactor. Archives of Microbiology, 2002. 177: 420-426.

Siebert, A., et al., A non-dechlorinating strain of Dehalospirillum multivorans:
evidence for a key role of the corrinoid cofactor in the synthesis of an active

tetrachloroethene dehalogenase. Archives of Microbiology, 2002. 178: 443-449.

Maymo-Gatell, X., et al., Isolation of a bacterium that reductively dechlorinates
tetrachloroethene to ethene. Science, 1997. 276: 1568-1571.

20

23.

24.

25.

26.

27.

He, .12., et al., Detoxiﬁcation of vinyl chloride to ethene coupled to growth of an
anaerobic bacterium. Nature, 2003. 424: 62-65.

Hendrickson, E.R., et al., Molecular analysis of Dehalococcoides 16S ribosomal
DNA from chloroethene-contaminated sites throughout north America and
Europe. Appl. Environ. Microbiol., 2002. 68: 485-495.

Lbfﬂer, F.E., et al., 1 6S rRNA gene-based detection of tetrachloroethene-

dechlorinating Desulfuromonas and Dehalococcoides species. Appl. Environ.
Microbiol., 2000. 66: 1369-1374.

Yang, Y.R. and J. Zeyer, Specific detection ofDehalococcoides species by
ﬂuorescence in situ hybridization with 16S rRNA-targeted oligonucleotide probes.
Appl. Environ. Microbiol., 2003. 69: 2879-2883.

Villemur, R., et al., Occurrence of several genes encoding putative reductive
dehalogenases in Desulﬁtobacterium haﬁriense/frappieri and Dehalococcoides
ethenogenes. Canadian Journal of Microbiology, 2002. 48: 697-706.

21

CHAPTER 2
FREQUENCY OF OCCURRENCE OF CHLORINATED ETHANES,
CHLORINATED ETHENES, AND BTEX AT NATIONAL PRIORITY LIST

SITES: EVALUATING THE NATIONAL DATASETS

22

ABSTRACT

Substantial inconsistencies between the reported frequencies of occurrence of
common environmental contaminants at National Priority List (NPL) sites were found
among the publicly available databases. Agency for Toxic Substances and Disease
Registry (ATSDR) based reports, including Toxicological Proﬁles and CERCLIS Priority
List of Hazardous Substances, in general reported higher frequency of occurrences for a
given compound than returned from queries of the Environmental Protection Agency
(EPA) NPL database. Duplicates due to chemical name synonyms further complicate
EPA data. This study shows that EPA and ATSDR report both unique and shared
information on the frequency of occurrence at NPL sites, suggesting that a composite of
both datasets provides more comprehensive occurrence data. On average 27% of the sites
reported in the EPA corrected data were not found in the ATSDR data. The number of
unique and common sites that were in either dataset was calculated and used to assess the
reporting of each dataset. The EPA, corrected synonyms, data underestimated the
frequency of occurrence by an average of 51%, and ATSDR underestimated by 13%. A
composite dataset is presented for the 27 chemicals examined in this study. The following
compounds were found at a majority of NPL sites with a Final or Deleted from Final
status: benzene (63%), toluene (69%), trichloroethene (66%), and tetrachloroethene

(59%).

23

INTRODUCTION

The frequency at which a contaminant occurs in the environment is an important
component, along with toxicity and exposure, in determining the hazard of a compound
and, thus, the impetus for studies of its toxicity and remediation. The frequencies of
occurrence of chemicals at National Priority List (NPL) sites are of speciﬁc interest as the
US Environmental Protection Agency (EPA) has designated these as the Nation’s most
seriously contaminated and are planned for long-term remediation. Sites are added and
deleted from the NPL through a multi-step process [1]. As of July 2003, there were 1506
NPL sites with either Final (1234) or Deleted from ﬁnal (272) status [2]. An additional 66
sites are proposed for Final list, and 65 have been removed from the proposed list. NPL
sites are a relatively large, practically important, and strictly deﬁned set, which are ideal
to assess the chemical occurrence of individual compounds and co-occurrence of
contaminant mixtures.

The frequencies of chemical occurrence at NPL sites are used explicitly in
determining the CERCLA Priority List of Hazardous Substances [3]. The Priority List
ranks compounds according to their frequency at NPL sites, exposure, and toxicity. The
Priority List is used to determine which compounds will be subjects of Toxicological
Proﬁles compiled by the Agency for Toxic Substances and Disease Registry (ATSDR).
In addition to providing a comprehensive review of the toxicology of a compound,
Toxicological Proﬁles point out data gaps and research needs on the chemical [4].
Toxicological Proﬁles are summarized for the public in ATSDR’s ToxFAQs summaries

[5] and Public Health Statements [6]. The frequencies of chemical occurrence at NPL

24

sites are stated in of each of these documents and can easily be queried in the ATSDR’s
HazDat and EPA NPL on-line databases [2, 7]. The frequencies of chemical occurrence
at NPL sites are thus available to be used by researchers in publications and grant
proposals to justify budget expenditures and by policy makers, the media, and public to
infer the relative importance environmental contaminants.

Comparisons of the frequencies of occurrence of several chemicals at NPL sites
from these sources show striking inconsistencies, as illustrated for vinyl chloride in
Figure 1. Namely, the ATSDR data sources and CERCLA Priority List data appear to
report substantially higher frequencies of occurrence of chemicals than are retrieved from
the EPA NPL database. For example, a query of the EPA NPL database for vinyl chloride
occurrence returned 17- 31% fewer NPL sites than retrieved from ATSDR or CERCLA
Priority List data sources (Figure 1). Since multiple data sources exist to determine the
frequency of chemical occurrence at NPL sites and there appear to be substantial
inconsistencies among them, our purpose in this study was to compare these sources,
provide some corrections, extract useful information, and foster a more accurate use of
the available data.

We focused on three important classes of contaminants at NPL sites: chlorinated
ethanes (CAs), chlorinated ethenes (CBS), and benzene, toluene, ethylbenzene and xylene
(BTEX). Chlorinated ethanes and ethenes are common industrial solvents used in a wide
range of applications primarily as degreasers, chemical synthesis intermediates, and in
dry cleaning. The BTEX compounds are the most water-soluble fraction of petroleum
and, along with CAs and CBS, are frequent groundwater contaminants. All of these

compounds are considered toxic to varying degrees; vinyl chloride and benzene are

25

Figure 2.1. Reported occurrences of vinyl chloride at National Priority List sites by
various data sources. EPA — US Environemtnal Protection Agency [1]; HazDat - Agency
for Toxic Substances and Disease Registry HazDat database, Final + Deleted sites only
[2]; CERCLA Priority List - The Comprehensive Environmental Response,
Compensation, and Liability Act Priority List of Hazardous Compounds[3]; Tox.
Prof.,ToxFaq’s, PHS — Toxicological Proﬁle for Vinyl Chloride [4], ToxFAQs [5], Public
Health Statements [6].

 

700

600 .

500 -

400 -

300 -

200 -

Final + Deleted NPL Sites

100 .

l

0 -. .- ...--_._._-_.. M.H...-Ar

EPA HazDat CERCLA Tox. Prof.,
Priority List ToxFaq's., PHS

 

 

26

 

known human carcinogens [5]. Although this study is limited to CAs, CBS, and BTEX,

the analyses and insights discussed here can be applied to other NPL contaminants.

27

METHODS
EPA Dataset. Numbers of Final and Deleted NPL sites containing CA, CE, and BTEX
were determined from queries of the EPA superfund website on June 7, 2003 [2]. To
examine synonyms of CAs, CBS, and BTEX compounds, an advanced query was
performed on each chemical name for Final and Deleted NPL sites. For further analyses
using EPA data, in order to ensure that all synonyms for a compound were queried, a
categorical search of volatile organic compounds (VOCs) was performed. The query
returned a text ﬁle with chemical name, NPL status, and CERCLIS ID (a unique site
identiﬁcation number). Using Microsoft Excel the 11,102-record ﬁle was then visually
checked for synonyms of the chemicals of interest and these records were changed to
standard chemical names. Aﬁer consolidating the synonyms, duplicates and proposed,
removed and non-NPL sites were eliminated. To be consistent with the reports from
Toxicological Proﬁles this study is limited to sites with Final and Deleted status, which
are hereaﬁer simply referred as “NPL sites”. The resulting ﬁle containing records for 27
CAs, CBS, and BTEX compounds was termed the “EPA corrected dataset” and used for

the rest of the study.

ATSDR Dataset. ATSDR data were obtained from the HAZDAT database on June 7,
2003 [7]. Chemical name, CERCLIS ID, and NPL status could not be obtained from a
single query. NPL status and CERCLIS IDs were returned from an “ATSDR Site
Activity” query for all Final and Deleted NPL sites. Then each chemical of interested was

queried separately using “contaminant query”. Chemical queries were downloaded to

28

Microsoft Excel and the CERCLIS ID ﬁeld was extracted. Duplicate CERCLIS IDs from
separate ATSDR document types were eliminated. Files containing CERCLIS IDs of
sites containing each chemical were then matched to the NPL status using the previous
Site Activity query data. The resulting ﬁle was termed “ATSDR dataset.”

To examine the occurrence of chemicals in speciﬁc media, CERCLIS ID, CAS
Reg #, medium type were extracted from each chemical record. Records for all chemicals
of interest were combined, and ‘medium’ ﬁelds were generalized to one of the following
categories: air, groundwater, leachate, sediment, sludge, soil, surface water, waste
material and other/not reported. Records indicating mixed media (i.e. soil/sediment) were
included as other/not reported. Duplicates were eliminated. A Microsoft Excel pivot
table of Cas Reg #, media type and CERCLIS ID was used to tabulate the frequency of

occurrence of each chemical in each medium.

Toxicological Proﬁles, ToxFaqs, and CERCLIS Hazardous Compound @La. Site data
from Toxicological Proﬁles and ToxFAQs Summaries were obtained from the online
documents for each compound [4-6]. Occurrence at NPL sites was listed in the Public
Health Statement section. CERCLIS Hazardous Substance Priority List data were found

in the Priority List supplementary material provided by ATSDR [3].

CERCLIS ID comparison of EPA and ATSDR datasets. Records for each chemical in
the EPA corrected dataset and ATSDR dataset were combined into a single Excel ﬁle.
The data were tabulated for each chemical in an Excel pivot table of CERCLIS ID, data

source, and NPL status. Data in this table were used to calculate the number of sites

29

containing a given chemical that were reported in either the EPA corrected or ATSDR
datasets; the composite data were termed the “EPA + ATSDR” dataset and used for co-
occurrence analysis.

ﬁrwise co-occurrence. Eighteen chemicals that were present at more than 300 (>20%)
NPL sites in the EPA + ATSDR dataset were used to examine trends in contaminant co-
occurrence. An Excel Pivot table of CERCLIS IDs, chemical name, and NPL status was
constructed using the EPA + ATSDR dataset. The resulting table showed which of the 18
chemicals was present at each site; this table was used to calculate the co-occurrence for
each pair of compounds. The number of sites that contained a pair of compounds was
expressed both as a percentage of the total number of NPL sites and of the number of

sites at which each chemical was present.

30

RESULTS
Synonyms in the EPA Data_sat. Queries of the EPA database for different chemical
synonyms returned widely varying numbers of sites (Table 1). Tetrachloroethene, for
example, also appeared in the EPA database as PCE, perchloroethene, perchloroethylene,
and tetrachloroethylene. These records were not cross-referenced and each synonym
query returned a different number of NPL sites ranging, for tetrachloroethene, from two
to 222. The sum of the sites returned from each tetrachloroethene synonym query was
544, or 36% of Final + Deleted NPL sites. Comparisons of the site CERCLIS IDs from
each synonym query, however, showed that some single sites listed more than one
synonym. Summing the synonym queries, thus, overestimated the total number of sites
by counting some sites multiple times. By downloading the CERCLIS ID numbers for
each synonym queried, duplicates were removed, and for tetrachloroethene the frequency
of occurrence was 411, or 27% of NPL sites. Similar problems with synonyms were

found for other polychlorinated ethenes and ethanes and the xylene isomers (Table 1).

ATSDR-HazDat contaminated media

HazDat data for chemical occurrence at ﬁnal and deleted NPL sites were analyzed
by contaminated media type (Table 2). In general groundwater was the most frequently
contaminated medium, followed by soil and sediments. Some changes in the rank order
of individual compounds occurred with media type. Toluene, for example, was second
overall in this analysis but was fourth in groundwater and ﬁrst in soil, sediment, and -

waste material.

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33

Comparison of frequency of occurrence data from different sources.

The EPA corrected dataset contained often substantially lower numbers of sites
for the chemicals examined than the ATSDR datasources (Table 3). The Hazdat data for
final + deleted sites were similar to the CERCLIS priority list data. The percentages
show large differences with the EPA corrected dataset reporting trichloroethene at 482, or
32%, of NPL sites where as HAZDAT cites 961, or 64%, of NPL sites.

Site speciﬁc comparison of EPA and ATSDR data.

To examine the nature of the inconsistencies between EPA corrected and ATSDR
frequency of occurrence numbers, CERCLIS IDs were compared for query results from
each data source (Table 4). Although the EPA corrected dataset generally showed lower
occurrence values than the ATSDR dataset; the EPA corrected dataset was not a subset of
the ATSDR dataset. On average 27% of the sites reported in the EPA corrected data was
not found in the ATSDR data.

The number of sites that were at least found in either dataset was calculated and
used to assess the reporting of each dataset. The EPA corrected dataset underestimated
the frequency of occurrence by an average of 50.7%, and ATSDR underestimated by
13.2% (Table 4). Underestimates of the site numbers from the EPA corrected and
ATSDR data varied for each chemical, and thus, the rank order of chemical frequency of
occurrence determined from each data source varied. For example, of the 27 compounds
examined, 14 had a change in rank order from the ATSDR dataset to the EPA + ATSDR

dataset, although none changed more than two positions.

34

 

 

 

 

 

 

 

 

 

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36

Pairwise co-occurrence.

Since the chemicals examined were present at up to 69% (toluene) of NPL sites,
many of these compounds must co-occur at the same site. To examine trends in chemical
co—occurrence, a subset of CAs, CBS, and BTEX compounds containing 18 compounds
that were present at more than 300 (20%) NPL sites were examined in detail. Of the 1506
NPL sites almost 200 did not report any of these chemicals, and seven sites contained all
18 compounds. On average NPL sites contained 8.1 of the chemicals examined. As
expected, the more frequently occurring chemicals co-occurred more oﬁen. Benzene and
toluene co-occurred with each other and in 12 other pairs at more than 50% of NPL sites
(Table 5a). Analysis of the pairwise co-occurrence of chemicals in terms of each
compound’s individual occurrence further shows these compounds are present with
several other compounds (Table 5b). For example, of the 349 NPL sites that contain
1,1,2,2-tetrachloroethane, the following compounds also occurred at greater than 90% of
the sites: toluene (94%), benzene (95%), TCE (95%), PCB (91%), ethylbenzene (90%),
and 1,1-DCA (91%). All the chlorinated solvents examined co-occurred with TCE at
more than 90% of the sites in which they were present; chlorinated solvent co-occurrence

with PCE was greater than 86%.

37

 

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39

DISCUSSION

The frequencies of chemical occurrence at NPL sites are useful in prioritizing
research on contaminant toxicity, environmental fate, and remediation strategies. Several
data sources exist that either provide or allow the extraction of frequency of occurrence
data at NPL sites; the values returned, however, vary by data source (Table 2). Queries
of the EPA NPL database typically returned substantially lower numbers of NPL sites
containing a given chemical than the ATSDR documents or HazDat database. The nature
of these discrepancies was explored here.

The reported frequencies of chemical occurrence between EPA and ATSDR data
sources are related to the purpose for which each agency collects contaminated site data.
The EPA evaluates potential NPL sites to determine if the level of contamination
warrants further investigation and cleanup. The ATSDR, however, looks at each site from
a public health risk standpoint.

The EPA NPL contaminant data are extracted from the Hazardous Ranking
System (HRS) Documentation Record (Personal Communication). The HRS is used to
evaluate which sites are placed on the NPL and is based on Preliminary Assessments and
Site Investigations. These early evaluations are designed to screen which sites require
additional investigation and are not detailed risk assessments. In this system, monitoring
ceases after a ceiling score is met to justify proposing a site for the NPL. “The level of
effort devoted to scoring a site is governed by two competing requirements: (1) to
accurately determine the relative threat posed by the site, and (2) to efﬁciently use EPA’s

limited data collection and analysis resources” [2].

40

While the EPA HRS is not a detailed risk assessment, ATSDR does perform
Health Assessments on NPL sites, in addition to several other public health activities and
documentations. Health Assessments are mandated to be completed within 2 years of a
site being listed as Proposed for the NPL. Health Assessments can draw from a variety
of sources for contaminant data including data from EPA documents. Data from Health
Assessments, Health consultations and all other ATSDR activities, NPL or not, are
abstracted in to the HazDat database. Information on the frequency of chemical
occurrence from HazDat is then used in Toxicological Profiles and the CERCLA priority
list of Hazardous compounds. Thus, to ensure an efﬁcient use of resources, EPA provides
limited, early screens of potential NPL sites, and ATSDR later performs in depth Health
Assessments.

In addition to HRS, EPA collects additional site information, including
monitoring data, in later stages of the NPL clean-up process. Apparently, this later, more
thorough data is not included in the EPA NPL database but may be included in ATSDR
Health Assessments.

The different timing and purposes for collecting site contaminant data help
explain why EPA returns fewer contaminated sites contaminated with certain chemicals
than ATSDR. Our analyses, however, suggest that the smaller EPA dataset also contains
information not included in the ATSDR database; that is, EPA is not a subset of ATSDR
(Table 3). Assuming the data are equally valid, EPA and ATSDR both provide unique
chemical occurrence data. Thus, a composite dataset is more comprehensive and

informative. The composite dataset produced here reported an average of 13% more

41

contaminant sites than HazDat alone, and for some chemicals the difference was much
higher (i.e. pentachloroethane 79%).

If users are interested in chemicals other than CAs, CEs, or BTEX that we report
here, we recommend using the CERCLA Priority List data to obtain the frequency of
occurrence of a compound at NPL sites. The CERCLA Priority List data is contained in a
peer reviewed document and is published every two years. Unfortunately, the NPL
frequency data is not available online and must be obtained from ATSDR directly.
Toxicological Profiles are also peer-reviewed documents, but since the publication dates
vary for each compound — and thus the number of NPL sites - comparisons of NPL
frequency data among chemicals is not as convenient as the CERCLA Priority List data.
Unlike the CERCLA Priority List NPL frequency data, however, Toxicological Proﬁles
and their summary documents, ToxFaq’s and Public Health Statements, are accessible
on-line. If more extensive analyses are required, then we suggest using the HazDat
database, or a composite of a corrected-EPA NPL and HazDat data, to extract the needed
information, as the EPA NPL data alone only captures 73% of the available contaminant

data and is complicated by chemical synonyms.

42

REFERENCES

U.S._Environmental_Protection_Agency, Superfund: Clean up process.
2003, http://www.epa.gov/superfund/action/process/sfproces.htm.

U.S.__Environmental_Protection_Agency, National Priorities List:
Advanced Query. 2003,
http://www.epa.gov/superfund/sites/query/advquery.htm.

Agency_for_Toxic_Substances_and_Disease_Registry, 2001 CERCLA
Priority List of Hazardous Substances. 2003,
http://www.atsdr.cdc.gov/clist.html.

Agency_for_Toxic_Substances_and_Disease_Registry, Toxicological
Proﬁle Information Sheet. 2003, http://www.atsdr.cdc.gov/toxproZ.html.

A gency_for_Toxic_S ubstances_and_Disease_Re gistry, T oxF A Qs.
2003.http://www.atsdr.cdc.gov/toxfaq.html.

Agency_for_Toxic_Substances_and_Disease_Registry, Public Health
Statements. 2003, http://www.atsdr.cdc.gov/phshome.html.

Agency_for_Toxic_Substances_and_Disease_Registry, HazDat Database.
2003, http://www.atsdr.cdc.gov/hazdat.html.

43

CHAPTER 3
ANAEROBIC MICROBIAL REDUCTIVE DECHLORINATION OF
TETRACHLOROETHEN E (PCE) TO PREDOMINATELY trans-l, 2-

DICHLOROETHENE (DCE)

44

ABSTRACT

Although complete microbial reduction of tetrachloroethene (PCE) and
trichloroethene (TCE) to ethene and chloride is known, dichloroethenes (DCEs) and vinyl
chloride often accumulate during anaerobic dechlorination. All characterized PCB and
TCE dechlorinating anaerobic bacteria and their puriﬁed dehalogenases produce cis-DCE
as the major DC E isomer. Certain anaerobic enrichment cultures, however, were shown
to produce more trans-DCB than cis-DC E. These cultures reductively dechlorinated PCE
and TCE to trans-DCB and cis-DCE simultaneously and in a ratio of 3(i0.5):1 that was
stable through multiple serial transfers, with a variety of electron donors, and in both
methanogenic and nonmethanogenic cultures. The dechlorination rates were always slow,
2.5 umol liter"l day", relative to cis-DCE producing enrichments. Dehalococcoides
populations are implicated in this reduction because two highly enriched cultures
contained 168 rRNA gene sequences 99% identical to Dehalococcoides sp. strain FL2,
and both cultures were resistant to ampicillin, a characteristic of Dehalococcoides. PCR
primers specific for Dehalococcoides tceA reductase gene did not produce an amplicon in
either culture. These results implicate Dehalococcoides populations in the novel
production of trans-DCB as the major daughter product from Chloroethene
dechlorination. This study shows trans-DCB can be the major end product from PCB and
TCE microbial dechlorination, thus a high fraction of trans-DCB at a site is not

necessarily from source contamination.

45

INTRODUCTION

Chlorinated ethenes are significant groundwater contaminants and are present in
aquifers as parent compounds and daughter products. Of the 1506 current or former US.
Environmental Protection Agency National Priority List Sites, tetrachloroethene
(perchloroethylene, PCE) is reported at 62% of the sites; trichloroethene (TCE) at 67%,
1,1-dichloroethene (1,1-DCB) at 46%, trans-1,2-dichloroethene (trans-DCB) at 44%, cis-
1,2-dichloroethene (cis-DCE) at 23%, and vinyl chloride (VC) at 44% (Chapter 2). Under
anaerobic conditions Chloroethenes are subject to reductive dechlorination
(hydrogenolysis) resulting in the stepwise conversion of PCB to TCE, DCE isomers, VC,
and ethene. Since all chlorinated ethenes are toxic, complete conversion to ethene is
critical for site remediation.

Several anaerobic bacterial populations have been isolated that use PCE or TCE
as electron acceptors for growth through a process termed chloridogenesis or
(de)halorespiration [1]. These isolates belong to several genera including Dehalobacter
[2, 3], Desulfuromonas [4, 5], Desulﬁtobacterium [6-8], Dehalococcoides [9],
Clostridium [10], Enterobacter [l 1], Dehalospirillum [12] and Sulfospirillum [13, 14].
Except for a few Desulﬁtobacterium strains that only convert PCE to TCE and two
Dehalococcoides isolates that dechlorinate PCE or TCE to VC and ethene [15], all other
isolates dechlorinate PCE to predominately cis-DCE as the end product. Chloroethene
reductive dehalogenases have been characterized from Dehalospirillum multivorans [16],
Desulﬁtobacterium sp. strain PCE-S [l7], Desulﬁtobacterium sp. strain Y5] [l 8],
Dehalococcoides ethenogenes [l9], and Dehalobacter restrictus [20]. The reductively

dechlorinating enzyme systems contain corrinoid cofactors in addition to iron-sulfur

46

clusters and catalyze dechlorination through a radical mechanism [21]. A recent
computational study examining the stability and transformation rates of radical
intermediates explains the favored cis-DCE production by B12 catalyzed reactions [22].
The occurrence of trans-DCE at chloroethene-contaminated sites can either be
from source material or from biological reduction of PCB or TCE parent compounds.
Although cis-DCE is the major isomer produced by the known reductive mechanisms,
trans-DCB can accumulate and persist in microcosms or aquifers because cis-DCE is
often more readily degraded than trans-DCB. Other mechanisms may contribute to
trans-DCB occurrence and may be important for natural attenuation monitoring and
point-source tracking. Because trans-DCB is typically more resistant to degradation than
cis-DCE, further understanding of trans-DCB formation is important for chlorinated
ethene contaminated site characterization and for choice of remediation strategies. In this
study, I describe certain microcosms and cultures that reduce PCB and TCE to trans—DCE
and cis-DCE in a ratio of 3: l. A preliminary report of this work was presented at the 2000

International Society for Environmental Biotechnology [23].

47

MATERIALS AND METHODS

Chemicals. The following analytical grade chlorinated compounds were used in this
study: PCE, TCE, trans-DCB, cis-DCE, all of which were obtained from Supelco,
Belfonte, Pennsylvannia; and VC, which was obtained from F luka Chemical Corp.,
Ronkonkoma, NY. Other chemicals used in the study were obtained from Sigma-Aldrich
(Milwaukee, WI).
Inoculum sources, microcosm preparation, and growth conditions. Five microcosms
that produced a mixture of trans-DCB and cis-DCE were derived as part of a larger set
of sediment microcosms initiated with PCE as potential electron acceptor and lactate as
electron donor. Microcosms derived from the Tahquamenon River in Michigan, the
Parfume River in Vietnam, and Chitwan National Park, Nepal were established as
previously described [24]. A mixture of agricultural and forest soils, and river sediment
samples collected in Michigan were pre-incubated under nitrate reducing conditions and
used to establish PCE dechlorinating cultures. Approximately 10 g of the nitrate depleted
slurry was added to 100 m1 of basal salts medium in 160 ml serum bottles. Lactate (5
mM) was added as electron donor and 20 umol PCE was added as potential electron
acceptor. Pine River (St. Louis, Michigan) sediments were mixed with RAMM medium
(21) under anaerobic conditions. Seven milliliters of the sediment slurry (containing 2
gdw of sediment material), lactate (10 mM), and 10 umol of PCE were added to each
vial. Three microcosms from each site were autoclaved for 60 min. at 121 °C on three
consecutive days and served as sterile controls.

Chloroethenes were analyzed in headspace samples performed at time zero and

periodically thereafter. All cultures were monitored for Chloroethene transformation and

48

electron donor consumption as described previously and replenished as needed [24-27].
Transfers were 1% (vol/vol) into fresh basal salts medium. Culture bottles were incubated
in an inverted, stationary position at 22—25° C unless otherwise noted.

Physiological studies. To test for spore forming ability of the dechlorinating populations,
cultures were pasteurized by inoculating fresh medium and immediately heating the
bottles to 80°C in a water bath. The temperature was measured in parallel in 100 ml of
medium in a 160 ml serum bottle ﬁtted with a mercury thermometer through the septum.
Timing began at 77°C, and the cultures were incubated for 10 min. The bottles were then
cooled in a water bath to room temperature, before vitamins and Chloroethenes were
amended.

The Tahquamenon River culture was used to test the effect of electron donor
addition on trans-DCE and cis-DCE production. TCE (20 umoles) and 200 Dmol
hydrogen plus 1 mM acetate, or 5 mM of either formate, acetate, succinate, pyruvate,
lactate, or glycerol, were added to freshly inoculated medium. To examine the effect of
temperature on growth and dechlorination, TCE and lactate fed cultures were transferred
to fresh medium and incubated at 15, 25, or 30 °C. To examine the effect of cell wall
synthesis inhibitors on dechlorination, ampicillin was added from a sterile stock solution
to a final concentration of 400 mg/l in freshly inoculated TCE and lactate fed cultures.
The effect of pH on dechlorination was determined in basal salts medium amended with
10 mM TES (N-tris[hydroxymethyl]methyl-2-aminoethane sulfonic acid) and adjusted to
a pH of 6.8, 7.5, or 8.2.

DNA-based characterization of trans-DCB producing cultures. Nucleic acids were

extracted from the Tahquamenon and Parfume River cultures after they were grown on

49

TCE and lactate. The 100-ml cultures were forced through sterile, 0.22 pm membrane
ﬁlters. The membranes were cut with sterile razor blades and added directly to the tubes
provided in the MoBio UltraClean Soil Kit (Solana Beach, CA). DNA was extracted
following the manufacturer’s instructions.

To screen for the presence of certain genera known to contain Chloroethene
degrading members, PCR ampliﬁcation was performed with the following primers
speciﬁc for the 16S rRNA genes of the following genera: Dehalobacter [28],
Desulfuromonas [29], Desulﬁtobacterium (2 sets, [30] ), Dehalococcoides [31]. The
TceA reductase genes similar to that of Dehalococcoides ethenogenes strain 195 were
screened for using previously described primers [32]. PCR reactions were typically 20ul
or 50 [41 as described in each reference.

The 168 rRNA genes of Dehalococcoides were recovered using primers from
Hendrickson et. al [31]. PCR products were cloned using TOPO TA (Invitrogen) kit and
inserts from at least three positive clones were sequenced from each culture by Michigan

State University’s Genomic Technology Support Facility.

50

RESULTS
trans-DCB producing cultures. Five enrichment cultures that reduced PCE to primarily
trans-DCB were selected from a larger set [24, 26], most of which produced cis-DCE
and/or ethene. These ﬁve cultures, plus a sixth enriched on 1,2—dichloropropane produced
trans-DCB and cis-DCE in a consistent ratio of 3(:i:0.5):1 (Table 1). In addition to three
inocula from Michigan sources, trans-DCB producing cultures were also established in
sediments from Nepal and Vietnam rivers. The cultures were at different levels of
enrichment, from methanogenic sediment microcosms to sediment-free,
nonmethanogenic cultures, with two transferred over 25 times. The ratio of DCE isomers

was stable through all transfers and no further reduction of DCEs to VC was observed.

Physiological studies. Trans—DCB and cis-DCE formation was studied in more detail in
cultures from the Tahquamenon and Parfume Rivers because these cultures were
nonmethanogenic and had been transferred more than 30 and 25 times, respectively.
Time courses for the dechlorination of TCE showed the simultaneous production of both
trans- and cis-DCE (Figure 1). The lag time before the onset of dechlorination varied
from culture to culture but was always greater than 7 days and could be up to several
weeks. The lactate fed Tahquamenon River culture reduced TCE at a rate of 0.25
umol/day. The ratio of trans-DCB to cis-DCE was constant over the period of active
dechlorination and trans-DCE accounted for up to 75% of the total Chloroethenes in the
culture.

The Tahquamenon culture was screened for dechlorination activity with various

electron donors. The average trans- to cis-DCE ratio in measurements from cultures of all

51

electron donors was 3.13 (SD 0.24, n = 22 measurements) when grown with the

Table 3.1 Summary of microcosms and enrichment cultures that produced mixtures of
trans-DCB and cis-DCE.

 

 

Source Enriched a . . trans:cis
on Transfers Sediment Methanogene51s DCEb

Tahquamenon River,

PCE >30 No No 3.1
MI (TQ)
Parfume River,

PCE >25 No No 2.9
Vietnam (Parf)
Red Cedar River,

. . 1,2-Dc 12 No No 2.5

Michigan
Pine River, Michigan PC E 6 No Yes 3.4
Chitwan National

PCE 1 Yes Yes 3
Park, Nepal
Mixed MI soil &

PCE 1 Yes Yes 3.5

sediment Inocula

 

“ transfers were typically 1%
b ratio is the average of ratios for all data points above detection limit
° 1,2-dichloropropane

52

Figure 3.1. Typical time course for TCE dechlorination by the Tahqumenon River
enrichment culture (TQ). Lactate (5mM) was added as electron donor, and 25 umol of
TCE was added to the 100 ml bicarbonate-buffered medium. Chloroethenes were
measured by headspace GC-FID. No vinyl chloride, ethene, ethane, or methane was
detected.

 

N“
01°

20‘

15‘ _I— trans- DCE

-O— cis- DCE
+ TCE

.A
O

 

 

 

Chloroethenes (umollbottle)
O

O

10 20 3O 40 50
Time (d)

53

following electron donors (mean ratio, SD, n measurements): 200 umol hydrogen + 1
mM acetate (3.33,0.05,4), 5 mM forrnate (3.14, 021,4), 5 mM acetate (3.21,-,1), 5 mM
succinate (2.93, 006,3), 5 mM pyruvate (2.92, 0.41, 2), 5 mM lactate (3.16, 0.31, 5), 5
mM glycerol (3.25,0.28,3). No dechlorination was supported with 5 mM propionate or 5
mM benzoate.

No dechlorination was observed in the Tahquamenon or Parfume River cultures at
12 or 30 °C after more than 4 months of incubation. Pasteurized inocula from stationary
cultures were no longer able to dechlorinate. The Tahquamenon River culture also did not
dechlorinate at pH 8.2, and only one replicate at pH 6.8 was active. The Parfume River
culture had a linear decreasing trend in cis-DCE production with decreasing pH (trans-
:cis-DCE ratio = -0.16*pH + 3.79, R2 = 0.9994, Figure 2) although the extent of
dechlorination was greatest at pH 7.5. The difference between 6.8 and 8.2 was
signiﬁcant at p = 0.027 (t-test). Dechlorination of TCE to trans-DCE and cis-DCE

occurred in lactate fed Tahquamenon and Parfume River cultures treated with ampicillin.

DNA-based characterization of trans-DCE producing cultures. Genomic DNA from
the Tahquamenon River and Parfume River cultures was screened using PCR primers
targeting 16S rRNA genes of genera known to contain dechlorinating populations.
Amplicons from primers targeting Desulfuromonas, Desulﬁtobacterium, or
Desulfomonile were not detected. In both cultures, however, target-sized amplicons were
produced using the Dehalococcoides-targeted primers pairs tested. Sequences from the

1377 base pair PCR products from both cultures were 99% identical to

54

(5mM) was added as electron donor, and 25 pmol of TCE was added to 100 ml
bicarbonate-buffered medium. Chloroethenes were measured by headspace GC-FID after
4 months of incubation. No vinyl chloride, ethene, ethane, or methane was detected.

A.

Total DCE Produced (umol)

trans:cis DCE ratio

 

2.9 -
2.7 -
2.5 -
2.3 -

16.0
14.0 -
12.0 -
10.0 ]

2.1

 

 

 

I j I I I I I

6.6 6.8 7.0 7.2 7.4 7.6 7.8 8.0 8.2 8.4

pH

 

8.0 -
6.0 W
4.0 7
2.0 1

 

 

 

+ ,.

 

 

 

 

 

 

 

0.0

6.8

7.5 8.2

55

Dehalocococcoides sp. strain F L2, which belongs to the Pinellas subgroup [31].
Additionally, Dehalobacter-targeted primers produced a band of the expected size from
the Parfume River culture DNA. Sequence from the 800 base pair amplicon was 98%
identical to the 168 rRNA gene sequence of Dehalobacter restrictus. No ampliﬁcation of
the D. ethenogenes tceA reductase gene was detected using either the Tahquamenon

River and Parfume River cultures.

56

DISCUSSION

Chlorinated ethenes are frequent groundwater contaminants that are proving
amenable to remediation by microbial reductive dechlorination [33, 34]. Problems arise,
however, if toxic or persistent intermediates such as 1,1-DCE, trans-DCE, cis-DCE, or
VC accumulate. Over the last several years our laboratory has established several
hundred microcosms for various Chloroethene dechlorination studies. In several of these
cultures I observed the formation of trans-DCE in excess of cis-DCE. This is interesting
because to date, all characterized PCB and TCE dechlorinating bacterial isolates and
puriﬁed enzymes produce cis-DCE as the major DCE isomer. The predominately trans-
DCE producing cultures represent only a small fraction of the total number of cultures I
have studied, consistent with the more infrequent production of large amounts of trans-
DCE in nature. In most cases this trans-DCE producing physiology was observed in
microcosms from sites that also performed other dechlorination activities [15,24]. For
example, a sediment-free nonmethanogenic culture, derived from Red Cedar River
sediment (Michigan) that was originally enriched on 1,2-dichloropropane, also converted
PCE to predominateldy trans-DCE (Table 1)[26].

The ratio of DCE isomers was remarkably stable, with the trans-DCE and cis-DCE
ratio remaining 3:1 during serial transfer and with different electron donors. Only a slight
decrease in the ratio was observed with increasing pH in the Parfume River culture. The
ratio of 3:1 trans to cis isomers is interesting because there is no precedence in microbial
systems for this product distribution, and it is not explained by current understanding of
dehalogenase catalyzed reduction. The consistency of the ratio, however, is suggestive of

a yet unknown underlying biochemical mechanism. Abiotic Zn(0) catalyzed

57

hydrogenolysis of TCE produces trans-DCE and cis-DCE in a 2.5:] ratio [35]. There
was no signiﬁcant source of zinc in the medium used in this biological system.

The production of trans-DCE in a 3:1 ratio with cis-DCE as described in this study is
distinct from the accumulation of trans-DCE by Dehalococcoides ethenogenes strain
195. Strain 195 produced cis-DCE (and 1,1-DCE) prior to and faster than trans-DCE, but
cis-DCE was subsequently consumed at a faster rate [3 6]. T rans—DCE only accumulated
to a low percentage of the total Chloroethene mass. In my cultures trans-DCE and cis-
DCE were produced simultaneously and in a constant 3:1 ratio with no further reduction
to VC. This comparison suggests that, in addition to source contamination with trans-
DCE, there are at least two biological TCE reduction pathways, which may lead to the
accumulation of trans-DCE in nature.

Dehalococcoides populations may be involved in the primarily trans-DCE production
in the Tahquamenon and Parfume River cultures since their 16S rRNA gene sequences
were readily recovered in these highly enriched cultures. This suggestion is supported by
the fact that dechlorination of TCE also occurred in the presence of ampicillin. Resistance
to cell wall synthesis inhibitors such as ampicillin is a characteristic trait of
Dehalococcoides isolates and is consistent with their lack of peptidoglycan, [9, 15, 29,
37]. The dechlorination rates in all trans-DCE producing cultures were slow, but similar
to those observed in other Dehalococcoides cultures. Similar lag periods before the onset
of dechlorination were also observed in Dehalococcoides cultures that grew using VC
reduction [15]. My hypothesis is that speciﬁc populations have the unique catalytic
capability to produce the primarily trans-DCE distribution. Three years of effort to

isolate such strains from these highly enriched cultures has not been successful due to the

58

slow and unreliable grth of the cultures. In nature trans-DCE producing populations
are likely out to be competed by the faster growing cis-DCE producing populations in
most cases. Although some contaminated ﬁeld sites, like the Key West, FL Navel Air
Facility, show a transzcis-DCE ratio of 2 to 3:1 [38].

Both the Tahquamenon River and Parfume River cultures contained 16S rRNA gene
sequences most closely related to Dehalococcoides sp. strain F L2. The four known
Dehalococcoides isolates are closely related by 16S rRNA sequence identity and
dechlorinate various chlorinated substrates. D. ethenogenes strain 195 and strain FL2
both reduce TCE to VC and can cometabolically produce ethene. Unlike strain 195,
strain F L2 only reduces PCE to TCE through a co—metabolic reaction. Dehalococcoides
sp., strain BAV1, performs the critical and ﬁnal Chloroethene dechlorination steps by
reducing DCE isomers and VC to ethene. Dehalococcoides sp. strain CBDBl
dechlorinates chlorinated benzenes [37] and dioxins [39]. From the from sequence
similarities of known dehalogenase genes and the genome sequence of D. ethenogenes, it
appears strain 195 contains up to 17 dehalogenases-like open reading frames [40]
suggesting a diversity in the dehalogenation potential of strain 195 and other
Dehalococcoides populations that remains largely unexplored. Because D. ethenogenes
tceA reductase gene was not detected using speciﬁc PCR primer pairs in either the
Tahquamenon or Parfume River cultures, a different dehalogenase is likely involved in
TCE reduction in the trans-DCE producing cultures. In addition, since the
Dehalococcoides populations are so closely related, it seems unlikely that trans-DCE

producing populations could be distinguished by 168 rRNA-based methods alone.

59

10.

11.

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63

CHAPTER 4

SURVEY OF DEHALOGENATION ACTIVITY AND PCR DETECTION OF

DECHLORINATING GENERA IN ANTARCTIC MELT-POND SEDIMENTS

64

ABSTRACT

Sediments from Antarctic melt ponds on Ross Island and on the McMurdo Ice
Shelf near Bratina Island were used to test for the presence of microbial reductive
dehalogenation and screen for the occurrence of dechlorinating bacterial genera using
PCR detection techniques. Anaerobic enrichment cultures were established with lactate,
acetate, or hydrogen as potential electron donors and tetrachloroethene (PCE),
trichloroethene (TCE), 2-bromophenol (2BP), 2-chlorophenol (2CP), 3-bromobenzoate
(3BBA), or 3-chlorobenozoate (3CBA) as potential electron acceptors. Two sediments
debrominated 3BBA in primary enrichments, and no site dechlorinated 3CBA. The
3BBA degrading culture was nonmethanogenic and accumulated benzoate. 2CP was not
transformed in any sediment, but seven sites debrominated 2BP to phenol within 2
months. Phenol was not readily degraded and accumulated in all debrominating cultures.
The 2BP debrominating activity was transferable and was not inhibited by 2-
bromoethanesulfonate. A most probable number estimate of 2BP degraders at one site
revealed less than 103-104 cultivatable debrominators per gram of sediment (wet weight).
Five microcosms dechlorinated PCE to TCE with acetate, or hydrogen plus acetate as
electron donors. Dechlorination was slow, occurring over 2 to 4 months, and initially did
not proceed past TCE in these cultures or in primary enrichments with TCE plus lactate.
After prolonged incubation, cis-DCE and a mixture of trans-DCE and cis-DCE
accumulated in certain enrichments. Substantial methane production accompanied PCE
dechlorination at two sites; whereas, the other dechlorinating site was nonmethanogenic.

PCE dechlorinating activity continued upon multiple feedings and was observed in 1%

65

serial transfers. PCE amended microcosms were screened with PCR primer pairs
targeting 16S rRNA genes of genera known to contain dehalogenating populations.
Desulfomonile targeted primers produced amplicons from all active cultures. These
results demonstrate the presence of reductive dehalogenating bacteria in Antarctic melt-

pond sediments and show their potential for a role in pristine, anaerobic environments.

66

INTRODUCTION

Halogenated organic compounds are oﬁen mistakenly equated only with their
anthropogenic production and environmental pollution. Substantial evidence, however,
clearly documents the diversity and abundance of naturally occurring halocarbons, with
over 3700 compounds currently identiﬁed [1-7]. Natural halocarbons may contain any
combination of halogen atoms (Cl, Br, I, F) and may be from mono- to per-halogenated.
Mixtures of halogen atom constituents further expand the potential permutations
available for halocarbon chemical diversity. Biogenic production of halocarbons can be
performed by a variety of organisms including marine plants, sponges, terrestrial plants,
fungi, bacteria, insects, marine animals, higher animals and even humans [2]. Geogenic
production of halocarbons is also well documented in volcanoes [8], in addition to the
abiotic production of halocarbons in soil [9]. The biogenic production of some
halocarbons, such as monohalomethanes, is large enough to inﬂuence global budgets
[10]. On a smaller scale, the amount and diversity of halocarbons produced in a single
organism can be extraordinary. For example “limu kohu” (Asparagopsis taxiformis), an
edible seaweed, contains more than 50 organobromine compounds comprising “80% of
its weight” [3]. Once produced, naturally occurring halocarbons are turned over as part of
larger halogen cycles [1, 3, ll, 12]. Thus, in addition to halocarbon production, central to
halogen cycles, are degradation mechanisms.

Microbes have shown a remarkable ability to dehalogenate a variety of
halogenated compounds [13], and substantial research on microbial dehalogenation has
been conducted in pursuit of pollution abatement technologies. Microbial dehalogenation

of halocarbons can serve three primary eco-physiological roles in anaerobic environments

67

[14]. (1) Halocarbons can be dehalogenated to allow for the subsequent oxidation of the
carbon skeleton for carbon and energy. This is suggested particularly for halo-aromatics,
which contain substantial oxidative energy in their carbon ring structures [15, 16]. (2)
Halocarbons can be co-metabolically dehalogenated by reduced co-factors involved in
other anaerobic processes such as acetogenesis and methanogenesis [17]. Co-metabolic
dehalogenation, however, is almost exclusively associated with aliphatic halocarbons
[17]. (3) Halocarbon dehalogenation can serve as an electron sink for oxidative
metabolism [18]. Some organisms can conserve energy for growth released from
dehalogenation reactions through a process termed (de)halorespiration. Due to the
exergonic nature of reductive dehalogenation, it is a competitive respiratory process
capable of out competing sulﬁdogenesis, methanogenesis, and acetogenesis in anaerobic
environments [19, 20]. Many bacterial strains have now been isolated that grow by
halorespiration, and they belong to only a few phylogenetic lineages: low G + C Gram
positive bacteria, 6- and e-Proteobacteria, and the “Dehalococcoides ” group related to
the Green Non-Sulfur Bacteria [13]. These strains are metabolically diverse including
fermentors, acetogens, facultative anaerobes, and even strictly hydrogenotrophic
halorespirers. The distribution of reductively dehalogenating organisms in nature is only
recently being evaluated and was greatly aided by the development of PCR based
detection methods [21-23]. Consistent with the evidence of naturally occurring
halocarbons, dehalogenating bacteria have been enriched, isolated, and detected using
PCR methods from several non-contaminated sites.

Antarctica is an appropriate location for the study of natural halogen cycles.

Biogenic production of brominated and iodated C l and C2 alkanes has been reported in

68

Antarctic melt pond ecosystems [24]. These ponds are located on the McMurdo Ice Shelf
in the Ross Sea, are dominated by thick photosynthetic microbial mats, and have been
studied extensively for photosynthetic processes. Cyanobacteria and Nostoc are taxa
known to have halocarbon producing populations [2] and are present in these ponds,
although their role in halocarbon production has not been conﬁrmed [24]. Recently,
studies of the anaerobic sediments below the photosynthetic mats in McMurdo Ice Shelf
ponds examined competition between sulﬁdogenesis and methanogenesis in controlling
terminal carbon ﬂow [25]. These studies have been extended to include the effects of
pond geochemistry, temperature and freeze-thaw cycling on anaerobic processes [26].

We hypothesize that biogenic halocarbons produced in the microbial mats in melt
ponds could serve as competitive microbial electron acceptors in the underlying
anaerobic sediment and hence support dehalogenating populations. This study examines
the potential for dehalogenation of model halocarbons by sediments of nine anaerobic
melt ponds located on the McMurdo Ice Shelf near Bratina Island and on Ross Island,
Antarctica. Sediment microcosms were screened for the occurrence of several bacterial
genera known to contain dehalogenating populations using PCR-based detection

methods.

69

MATERIALS AND METHODS

Experimental design

The experimental design was focused to limit the permutations of enrichment
conditions. Model compounds were chosen that represent known biogenic or geogenic
compounds and whose degradation is well studied in mesophilic anaerobic systems.
Both halogenated aliphatics and aromatics were tested. Chloroethenes are produced by a
variety of macroalgae [2] and are present in volcanic gases [8]; they are well-established
substrates for halorespirers [27]. Both PCB and TCE were used as Chloroethene
substrates because until recently [28] all Chloroethene respiring isolates could use one or
both of these compounds as starting materials. PCE was tested with both hydrogen +
acetate and acetate as electron donors. For TCE and the aromatic halocarbons tested,
lactate was chosen as electron donor. 2-Halophenols and 3-halobenzoates were chosen as
model aromatic halocarbons. Only one position was chosen for each class but both
brominated and chlorinated compounds were tested, and to further reduce experimental
permutations the chlorinated and brominated compounds were included simultaneously.
Medium composition, pH, and incubation temperature were the same as previously used
for mesophilic enrichments. incubations were conducted at 20-23 °C, to favor more rapid
activity. Antarctic isolates often are psychro-tolerant, not psychrophylic, and have growth

optima near 20 °C.

70

Sample collection, microcosm preparation, and growth conditions

Samples were collected in the austral summer (January 2001) during the NSF
Training Course in Antarctic Biology (Table 1). The following ponds were sampled near
Bratina Island on the McMurdo Ice Shelf (unofﬁcial names): Orange Pond, Foam Pond,
Russel Pond, and Skua Pond. The following ponds were sampled at Cape Evans on Ross
Island: Rock Pond, Skua Pond, Boulder Pond, and Boulder Dry Pond. Pond geochemistry
was not determined; however, physicochemical characterization of McMurdo Ice Shelf
ponds is well documented in the literature [26]. Sediment was sampled within 0.5 m of
the pond shore into 50 ml Corning tubes (plastic) and stored at 5° C while in Antarctica.
The samples were shipped to Michigan State University from Christchurch, New Zealand
on ice. Sediments were stored at 5° C and microcosms were prepared within 1 month.

Methane production was screened for in 60 ml serum bottles using 30 m1 of a
complex medium modiﬁed from Cote 1994 [29]. No mercaptoethanesulfonic acid was
added, and 5 mM sodium acetate was included as electron donor. Duplicate cultures were
measured every 2 to 4 days for methane formation.

Dehalogenation was screened for in microcosms prepared with 100 ml of
anaerobic bicarbonate buffered mineral medium (pH 7.3) in 160 ml serum bottles [30].
The medium was reduced with 0.4 mM sodium sulﬁde. Lactate (10 mM), acetate (10
mM), or 10 ml hydrogen + 2 mM acetate, were added as electron donors and carbon
sources. PCB (20 umol), TCE (20 umol), 2-bromophenol + 2-chlorophenol (250 [1M
each), or 3-bromobenzoate + 3-chlorobenzoate (500 uM each) were added as potential

electron acceptors. Microcosms without halogenated compounds and uninoculated

71

Table 4.1. Melt-pond sediment sampling sites on the McMurdo Ice Shelf and Cape
Evans, Antarctica. Sediment was sampled within 0.5 m of the pond shore and stored at 5
°C for shipment from Antarctica to Michigan State University.

 

 

Name Location Island Description
Orange Pond McMurdo Ice Shelf Bratina Moist/wet mud
Foam Pond McMurdo Ice Shelf Bratina Very wet mud
Russel Pond McMurdo Ice Shelf Bratina Wet mud

Skua Pond McMurdo Ice Shelf Bratina Wet mud

Rock Pond Cape Evans Ross Mud + mat chunks
North Pond Cape Evans Ross Rocks

Skua Pond Cape Evans Ross Rocks

Boulder Pond Cape Evans Ross Rocks

Boulder Dry Pond Cape Evans Ross Dry (Eganic matter

 

72

cultures served as controls. Cultures were incubated statically in the dark at 20-22° C and
were measured periodically for halocarbon transformation for 16 months.

Foam Pond sediments were chosen to estimate the 2-bromophenol debrominator
population size by Most Probable Number (MPN). Five replicates at ﬁve serial dilutions
were used. Medium was bicarbonate buffered mineral medium with lactate (5 mM) and
acetate (2 mM) as carbon and electron donors and supplemented with 100 mg/L yeast
extract and 0.4 mM sodium sulﬁde. 2-Bromophenol (125 uM) was added as electron
acceptor. Foam Pond sediment (l g wet weight) was added to medium for the lowest
dilution. Phenol, 2-bromophenol, and methane were measured after 2 and 16 months.
The 95% conﬁdence intervals of the debrominating and methanogenic population sizes

were estimated using an Microsoﬂ Excel worksheet [31].

Analytical methods

Chlorinated ethenes and methane were measured in headspace samples at 25°C by
direct GC injection. For chlorinated ethenes, 200 pl samples were analyzed on an HP
model 6850 GC equipped with a Megabore model DB-624 column (30 m by 0.543 mm; J
&W Scientiﬁc), a ﬂame ionization detector and nitrogen carrier gas. The temperature
program was 50°C for 2.5 min, increased at 50°C/min to 200°C, and held at that
temperature for 2 min. For ethene, ethane, and methane, 200 pl headspace samples were
injected onto a stainless steel Porapak Q1 column and detected with a ﬂame ionization
detector with nitrogen carrier gas. The oven temperature was isocratic at 60°C.

Standard curves for volatile compounds were prepared by adding a known
amount of the compound to serum bottles with the same liquid-to-headspace ratio as that

for the cultures being analyzed, and analyzed at the same temperature as the samples.

73

A portion (1 ml) of the microcosm supernatant was withdrawn with a syringe,
ﬁltered, and used for analysis of aromatic compounds. Halophenols and halobenzoates
were analyzed by HPLC with a Hibar RP-18 (10 um) column and a ﬂow rate of 1.5
ml/min with an eluent of HzO-CH3CN-H3PO4 (66:33zl). Phenol, 2-bromophenol, and 2-
chlorophenol were analyzed by UV detection at 218 nm. Benzoate, 3-bromobenzoate,

and 3-chlorobenzoate were analyzed by UV detection at 230 nm.

PCR detection of dehalogenating genera

Nucleic acids were extracted from the microcosms amended with PCB and
hydrogen + acetate. A portion (5 ml) of the dilute sediment slurry was withdrawn from
the microcosms, and pelleted by centrifugation. The sediment pellet was added directly to
the tubes provided in the MoBio UltraClean Soil Kit (Solana Beach, CA). DNA was
extracted following the manufacturer’s instructions with bead beating.

The 16S rDNA primer pairs in Table 2 were used to screen for the presence of
following genera known to contain Chloroethene degrading members: Dehalobacter [32],
Desulfuromonas [33], Desulﬁtobacterium [23], Desulfomonile [34], and Dehalococcoides
[3 5]. PCR reactions were either 20 ul or 50 it] and were performed as described for each
reference. A non-degenerate primer pair speciﬁc for the tceA reductase gene of
Dehalococcoides ethenogenes strain 195 was also used to screen for the presence of this

known dehalogenase [36].

74

Table 4.2. PCR primer sets used for the detection of dechlorinating genera in sediment

microcosms.

 

Target

Dehalobacter
Desulfomonile
Desulfuromonas
Desulﬁtobacterium
Desulﬁt. Str. PCP!
Dehalococcoides
Dehalococcoides
Dehalococcoides
Dehalococcoides
Dehalococcoides
Dehalococcoides
Dehalococcoides
tceA reductase
Universal

Pimer Pair
deb179f- deb1007r
DTIl78f— DTIl001r
881 f0 - 831 ro
Del - De2
PCPIG — PCP4D
(l) FnDHCl — RoDHC692
(2) FnDHCI — RnDHClZlZ
(3) FnDHCl — RDDHCI377
(4) FnDHC385 - RoDHC806
(5) FnDHC587 — RDDHC1090
(61 FnDHC774 - RDDHC 1212
(7) FnDHC946 - RDDHCIZIZ

7971' — 2490r
8F - I492R

75

Reference

Schlbtelbure et al. unnub.
Bunee et al. unnub.

Lbfﬂer et al. AEM 2000
Lanthier et al. FEMS Microb.
Lanthier et a1. FEMS Microb.
Hendrickson et al. AEM 2002
Hendrickson et al. AEM 2002
Hendrickson et al. AEM 2002
Hendrickson et al. AEM 2002
Hendrickson et al. AEM 2002
Hendrickson et al. AEM 2002
Hendrickson et al. AEM 2002
Maenuson et al. AEM 2000
Hendrickson et al. AEM 2002

RESULTS

Nine melt pond sediments representing a range of sediment types were collected
from the McMurdo Ice Shelf near Bratina Island and at Cape Evans on Ross Island
(Table 1). Methane production was rapid in Boulder Dry Pond, Skua Pond (Bratina
Island), and Orange Pond, with methane exceeding 7% headspace gas in 18 days (Figure
4.1). Methane production in Foam Pond, Boulder Pond, Rock Pond, and Russell Pond
began after a l2-day lag period. No methane was detected in North Pond sediments.
These results suggest there was a viable anaerobic community capable of growth at 20-
23° C.

2-Bromophenol was degraded by seven of the eight sediments tested (Figure 4.2).
North Pond was the only sediment that did not debrominate the initial feeding of 2BP.
This site showed low activity and was the single site in which methanogenesis was not
detected in complex medium. In contrast, no sediment transformed 2-chlor0phenol.

The initial feeding of 3-bromobenzoate was only degraded by one sediment, Rock
Pond (Figure 4.3). Debromination occurred in the absence of signiﬁcant methanogenesis.
Some 3BBA debromination occurred in Orange Pond and Boulder Dry Pond but only
upon extended incubation. 3BBA debromination was not as rapid or as widely distributed
among the sediments as 2-bromophenol. Like for the 2-chlorophenol enrichments, no
dechlorination of 3CBA was observed.

Some debromination occurred in all transfers of the active 2BP microcosms
(Figure 4.4). Rock Pond had the most active populations. BES did not inhibit

debromination in these secondary Foam Pond and Orange Pond enrichments.

76

Figure 4.1 Methane production in anaerobic complex medium inoculated with Antarctic

melt pond sediment.

 

120000

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— Boulder Dry Pond
— Negative Control

5

77

10
Time (d)

15

 

20

Figure 4.2. 2-Bromophenol consumption in Antarctic microcosm sediments. Cultures
were fed to a ﬁnal concentration of 0.25 mM per feeding. Bar height reﬂects the
cumulative number of feedings after 8-months.

1400

 

1200 1
I 2-Br-Phe no]
1000 -
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000 .
400 a

200 a

 

     

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Orange Foam Pond Russel Skua Pond Rock Pond North Pond Boulder Boulder
Pond Pond Pond Dry Pond

Halophenol Consumed! Phenol Produced (micromoles)

78

 

Figure 4.3 3-Bromobenzoate and 3-chlorobenzoate consumption in Antarctic microcosm
sediments and production of benzoate. Cultures were fed an initial concentration of 0.25
mM of each halobenzoate. Consumption is shown for an 8-month incubation.

 

 

Foam Pond

Orange
Pond

Russel

Pond

79

   
     

 
   

300

E

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E 250 . ﬂ 3-Br-Benmate

3" 3:4

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Skua Pond Rock Pond North Pond Boulder

Pond

Boulder
Dry Pond

 

Figure 4.4. 2-Bromophenol debromination in 1% transfers from halophenol microcosms
after 8 weeks of incubation with 10 mM lactate as electron donor. Cultures that
consumed the initial 0.25 mM 2-bromophenol were amended with an additional 0.25 mM
feeding.

 

300

     

I 2-Br-Phenol
Cl Phenol

N

u:

9
I

     
         

ophenol Consumed/Phenol Produced (micromoles)
G
e

 

 

WW
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80

The ﬁrst sampling of the MPN tubes of Foam Pond only showed debromination
in the lowest dilution corresponding to a population size of 7.2-78.9 cells/g wet weight
(95% CL). After 16 months, however, higher dilutions showed activity, and the
population size was estimated at 103 — 104 cells/g wet weight. At both time points all
replicates at all dilutions showed methane production, suggesting the methanogen
population size was greater than the detection limit of the dilution series, 1'. e. 1.39 x 105
cells/g wet weight.

Five sediments dechlorinated PCE to TCE with acetate, or hydrogen plus acetate
as electron donors (Figure 4.5). Dechlorination was slow, occurring over 2 to 4 months,
and initially did not proceed past TCE in these cultures or in primary enrichments with
TCE plus lactate. After prolonged incubation (>10 months), cis-DCE and a mixture of
trans-DCE and cis-DCE accumulated in certain enrichments. PCE dechlorinating cultures
ﬁrst converted PCE to TCE, before further reduction to DCEs. Only Orange Pond
dechlorinated TCE when it was added as the starting substrate. Again, Orange Pond
produced trans-DCE and cis-DCE in a ratio of approximately 3: 1. PCE dechlorinating
activity continued upon multiple feedings and was observed in 1% transfers of Boulder
Dry Pond.

The hydrogen and acetate fed PCE microcosms were screened for genera known
to contain dechlorinating populations using PCR techniques. DNA from all cultures were
ampliﬁed using 16S rRNA universal primers. Dehalobacter and Desulﬁtobacterium were
not detected in any culture. Orange Pond and Boulder Dry Pond produced amplicons of

expected size using Desulfuromonas targeted primers (Figure 4.6). Boulder Dry Pond and

81

Figure 4.5. PCB conversion to TCE in Antarctic melt-pond sediment microcosms
amended with lactate (10 mM) and 20 umol PCE. Boulder Dry Pond received an
additional feeding of PCB after the initial amount was transformed.

 

 

 

 

 

.2220 N
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Rock Pond also produced unspeciﬁc bands with a larger than expected size.
Desulfomonile targeted primers produced amplicons of expected size from six sediment
microcosms (Figure 4.7), but not from Skua Pond (Ross Island), North Pond, and Russell
Pond. These three ponds also did not dechlorinate PCE or produce methane in the
microcosms from which DNA was extracted.

DNA from three microcosm sediments from Bratina island: Orange Pond, Foam
Pond, and Skua Pond, produced amplicons of the expected size from two different primer
pairs targeting Dehalococcoides 16S rRNA genes (Figure 4.8). Additionally Boulder Dry
Pond (Ross Island) produced a band of expected size from one of those primer sets. The
other ﬁve Dehalococcoides target primer sets did not produce ampliﬁcation products.
Speciﬁc primers targeting the Dehalococcoides ethenogenes strain 195 TceA reductase

also did no amplify DNA from any culture.

84

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DISCUSSION

Antarctica is a promising location to study natural halogen cycles due to its
geographic isolation and relatively limited human impact. We hypothesized that melt
ponds like those on the McMurdo Ice Shelf would harbor dehalogenating microbial
populations since these ponds are known to produce biogenic halocarbons [24]. The
abundant phototrophs in the mats could both produce biogenic halocarbons and create
conditions for an active anaerobic community in the underlying sediment [25, 26].

Although the microbiology of the melt ponds scattered on the McMurdo Ice Shelf
near Bratina Island have been examined for several decades, the small ponds sampled on
Cape Evans have not been reported previously. Three of the Cape Evans ponds, Rock
Pond, Boulder Pond, and Boulder Dry Pond, were small- about 2 m in diameter- and
formed in the shadows of large boulders. Boulder Dry pond, in fact, did not contain
standing water and was noticeable only by a darker color against the otherwise light, dry
soil in that area. The Boulder Dry Pond sample was dense with organic matter and when
rehydrated upon inoculation of the microcosms, turned a deep green color suggesting the
site contained substantial microbial mat material. This dry pond, presumable aerobic at
the time of sampling, was the most rapid to produce methane and was active in several of
the dehalogenating enrichments.

We tested six model compounds to screen for dehalogenation activity (Figure
4.9). Although neither of the chlorinated compounds were dehalogenated, 2-bromophenol
was debrominated in seven of the eight sediments and one site showed debromination of

3-BBA. Aromatics are typically not considered subject to cometabolic dechlorination

[17].

87

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3BBA debromination occurred in the absence of methanogenesis in Rock Pond, and the
two ZBP secondary enrichments tested were not inhibited by BES, a common inhibitor of
methanogenesis and used to test for cometabolic activity of methanogens in
dehalogenating cultures.

Our MPN study showed a 2BP debrominating population of 103-104 cells/ g wet
wt. after 16 months of incubation. All replicates in all MPN dilutions, even those that did
not debrominate, showed methane production after 2 months, further suggesting no
obligatory link between methanogenesis and debromination in these sediments.

The chlorinated ethenes tested, PCB and TCE, have been extensively studied for
bioremediation, and thus much is known about Chloroethene dehalogenation at
mesophilic temperatures. In our enrichments dechlorination was slow and primarily
produced TCE from PCE. This single-step dechlorination is reported as a common co-
metabolic reaction catalyzed by reduced co-factors of acetogens and methanogens [l7].
Desulfomonile tiedjei strain DCBI is also known to metabolically dechlorinate PCE to
TCE when its 3CBA dehalogenase is induced [37]. All PCE dechlorinating microcosms
also yielded PCR products of the expected size with Desulfomonile primer pairs,
suggesting that this may be the debrominating population. However, all these
microcosms also produced PCR products from either two of seven Dehalococcoides
primers or the Desulfuromonas primers suggesting other candidates for the activity
(Figure 4.10). The seven Dehalococcoides primer sets target different regions of the 16S
rRNA gene [22]. The two primer sets that ampliﬁed were from the internal part of the
168 rRNA, positions 385 to 806 and 587 to 1090 (Dehalococcoides positions). The

phylogenetic diversity of the Dehalococcoides group and its neighbors is largely

89

 

 

 

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90

unknown. The Dehalococcoides group and related members of the green non-sulfur
bacteria are commonly recovered in anaerobic environments. Only four Dehalococcoides
isolates are known and they are all closely related obligate hydrogenotrophic
halorespirers. Three are Chloroethene respirers and one is a chlorobenzene respirer. It is
possible that the 168 rRNA genes detected here are only partially related to the known
Dehalococcoides group. Recently, several l6S rRNA library clones from Antarctic
marine sediments were obtained that form a clade most closely related to the
Dehalococcoides group, further suggesting that close relatives of Dehalococcoides may
extend to the Antarctic region [3 8].

These results demonstrate the presence of reductive dehalogenating activity in
Antarctic melt-pond sediments and show that dehalogenators also populate Antarctic
pond sediments. Dehalogenating populations have apparently survived dispersal to the
Antarctic. Whether the dispersal events were ancient or more recent, once there, some
populations grew locally as supported by a population size of 103-104 debrominators/g
sediment. However, the diversity of dehalogenating populations was much less than in
many US and European sediments as judged by the range of substrates transformed and
the very slow enrichment. This is consistent with the more depauperate state of most
species in Antarctica and reﬂects both limitations in dispersal and in colonization.
Nonetheless, ﬁnding dehalogenating populations in a remote, pristine environment is
signiﬁcant to understanding their historical ecological role. We could not conclude which
dehalogenators were responsible for the dehalogenation in Antarctic ponds, but
Desulfomonile, and either novel Dehalococcoides or Desulfuromonas are candidates

since their 168 rRNA genes were detected only in the active dechlorinating microcosms.

91

10.

ll.

12.

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Publishers: Boston. p. 3-29.

Gribble, G.W., The diversity of naturally produced organohalogens.
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Gribble, G.J., The natural production of organobromine compounds.
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Gribble, G.W., The diversity of naturally occurring organobromine compounds.
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Gribble, G.W., Naturally occurring organohalogen compounds. Accounts of
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Gribble, G.W., The natural production of chlorinated compounds. Environmental
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Gribble, G.W., Naturally-Occurring Organohalogen Compounds - a Survey.
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Jordan, A., Volcanic Formation of Halogenated Organic Compounds, in Natural
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Scholer, HF. and F. Keppler, Abiotic Formation of Organohalogens During
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Harper, DB. and J .T.G. Hamilton, The Global Cycle of the Naturally-Occurring
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Gberg, G., The natural chlorine cycle -ﬁtting the scattered pieces. Applied
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Oberg, G., The Biogeochemistry of Chlorine in Soil, in Natural Production
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Haggblom, M.M. and ID. Bossert, Microbial Ecology of Dehalogenation, in
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Gibson, J. and CS Harwood, Metabolic diversity in aromatic compound

utilization by anaerobic microbes. Annual Review of Microbiology, 2002. 56:
345-369.

Héiggblbm, M.M. and LY. Young, Anaerobic degradation of 3-halobenzoates by
a denitrijying bacterium. Archives of Microbiology, 1999. 171: 230-236.

Holliger, C ., C. Regeard, and G. Diekart, Dehalogenation by Anaerobic Bacteria,
in Dehalogenation : microbial processes and environmental applications, M.M.
Haggblom and ID. Bossert, Editors. 2003, Kluwer Academic Publishers: Boston.
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Holliger, C. and W. Schumacher, Reductive Dehalogenation as a Respiratory
Process. Antonie Van Leeuwenhoek International Journal of General and
Molecular Microbiology, 1994. 66: 239-246.

Dolﬁng, J ., Thermodynamic Considerations for Dehalogenation, in
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Haggblbm and ID. Bossert, Editors. 2003, Kluwer Academic Publishers: Boston.
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Lofﬂe, F.E., J .M. Tiedje, and RA. Sanford, Fraction of electrons consumed in
electron acceptor reduction and hydrogen thresholds as indicators of
halorespiratory physiology. Appl. Environ. Microbiol., 1999. 65: 4049-56.

Lofﬂe, F.E., et al., 16S rRNA gene-based detection of tetrachloroethene-

dechlorinating Desulfuromonas and Dehalococcoides species. Appl. Environ.
Microbiol., 2000. 66: 1369-74.

Hendrickson, E.R., et al., Molecular analysis of Dehalococcoides 16S ribosomal

DNA from chloroethene-contaminated sites throughout North America and
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Lanthier, M., et al., Geographic distribution of Desulﬁtobacterium frappieri PCP-
] and Desulﬁtobacterium spp. in soilsfrom the province of Quebec, Canada.
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Mountfort, D.O., et al., Partitioning eﬂects during terminal carbon and electron

flow in sediments of a low-salinity meltwater pond near Bratina Island, McMurdo
Ice Shelf Antarctica. Appl. Environ. Microbiol., 1999. 65: 5493-5499.

Mountfort, D.O., et al., Inﬂuences of pond geochemistry, temperature, and freeze-
thaw on terminal anaerobic processes occurring in sediments of six ponds of the
McMurdo Ice Shelf near Bratina Island, Antarctica. Appl. Environ. Microbiol.,
2003. 69: 583-592.

Damborsky, J ., Tetrachloroethene-dehalogenating bacteria. Folia Microbiol,
1999. 44: 247-62.

He, J .Z., et al., Detoxiﬁcation of vinyl chloride to ethene coupled to growth of an
anaerobic bacterium. Nature, 2003. 424: 62-65.

Cote, RJ. and R.L. Ghema, Nutrition and media, in Methods for General and
Molecular Bacteriology, P. Gerhardt, Editor. 1994, American Society for
Microbiology: Washington DC. p. 176.

Lofﬂe, F .E., et al., Complete reductive dechlorination of 1,2-dichloropropane by
anaerobic bacteria. Appl. Environ. Microbiol., 1997. 63: 2870-2875.

Briones, A.M., Jr. and W. Reichardt, Estimating microbial population counts by
'most probable number’ using Microsoft Excel. J Microbiol Methods, 1999. 35:
157-61.

Schlotelburg, C., Unpublished. 2001.

Lofﬂer, F.E., et al., 1 6S rRNA gene-based detection of tetrachloroethene-
dechlorinating Desulfuromonas and Dehalococcoides species. Appl. Environ.
Microbiol., 2000. 66: 1369-1374.

Bunge, M., Unpublished. 2000.

Hendrickson, E.R., et al., Molecular analysis ofDehalococcoides 16S ribosomal
DNA from chloroethene-contaminated sites throughout north America and
Europe. Appl. Environ. Microbiol., 2002. 68: 485-495.

Magnuson, J .K., et al., Trichloroethene reductive dehalogenase from

Dehalococcoides ethenogenes: sequence of tceA and substrate range
characterization. Appl. Environ. Microbiol., 2000. 66: 5141-7.

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Cole, J .R., 8.2. F athepure, and J .M. Tiedje, T etrachloroethene and 3-
chlorobenzoate dechlorination activities are co- induced in Desulfomonile tiedjei
DCB-1. Biodegradation, 1995.6: 167-72.

Bowman, J .P., et al., Prokaryotic metabolic activity and community structure in
Antarctic continental shelf sediments. Appl. Environ. Microbiol., 2003. 69: 2448-
2462.

95

CHAPTER 5

SYNERGISM OF ZERO-VALENT IRON AND MICROBIAL

DECHLORINATION OF CHLORINATED SOLVENTS

96

ABSTRACT

The use of zero-valent iron, F e(0), for in situ remediation of chlorinated solvents
is gaining acceptance as a cost-effective remediation technology, but the interaction of
microbial and abiotic dechlorination processes is not well understood. To better
understand the role of dechlorinating microbes during Fe(O)-mediated remediation we
examined the immediate electron donors to microbial dechlorination and assessed
synergism between microbial and abiotic processes. Fe(O) served as the sole electron
donor for two hydrogen-oxidizing, tetrachloroethene (PCE) dechlorinating bacteria:
Dehalobacter restrictus and Sulfurospirillum multivorans. Desulfuromonas
michiganensis, which cannot use hydrogen, did not dechlorinate PCE with Fe(O) as the
sole electron donor. None of the strains tested dechlorinated PCE with ferrous iron as an
electron donor. Cathodic hydrogen was proven to be the immediate electron donor for
PCE dechlorination by D. restrictus and S. multivorans in a dual-chambered batch reactor
in which the bacteria were physically separated from the elemental iron while allowing
cathodic H2 and Chloroethenes to diffuse between the two chambers through a common
headspace. We obtained no evidence in support of an immediate electron donor to
dechlorination other than cathodic hydrogen. The microbes used reducing equivalents
from cathodic hydrogen for reductive dechlorination that were otherwise wasted in a
strictly abiotic system. The level of synergism between Fe(0)-mediated and microbial
dechlorination ultimately depended on the nature of the dechlorination pathways and
rates of each process. There was a high level of synergism between abiotic and microbial

dechlorination for Chloroethenes, because microbial dechlorination resulted in rapid

97

dechlorination of PCE to cis-DCE, which was dechlorinated by F e(0) catalyzed reactions
faster than PCE. The synergistic effect was decreased if a lesser-chlorinated ethene,
trichloroethene, was added as the parent compound. In contrast, no synergism was found
for 1,1,l-trichloroethane (TCA) dechlorination as both F e(0) and Dehalobacter sp. strain
TCAl catalyzed reactions that produced primarily 1,1-dichloroethane. In this case,
dechlorination by Fe(O) was substantially faster than microbial dechlorination by
Dehalobacter sp. strain TCAl such that there was no biotic contribution to
dechlorination. F e(O) reactivity was inhibited by several common medium amendments.
These results suggest the level of synergism of microbial and F e(0) mediated reductive
dechlorination depends on the speciﬁc dechlorinating microbes, the complementarities of

the dechlorination rates and pathways, and geochemical, or medium, conditions.

98

INTRODUCTION

Chlorinated solvents are abundant groundwater pollutants. Remediation strategies
to replace ineffective pump and treat technologies are the subject of substantial research
efforts. The use of zero-valent iron, F 6(0), in permeable, reactive barriers is proving to
be a cost-effective remediation strategy for chlorinated solvents [1]. Although the precise
catalytic mechanisms of Fe(0)—stimulated dechlorination are not fully elucidated, the
general reactions involved are well known [1]. The zero valent iron oxidation half
reaction producing F e(II) is quite favorable (-447 mV) [2]. The reducing power of this
couple is sufﬁcient to drive reductive dechlorination of chlorinated solvents by either
hydrogenolysis or elimination pathways [3]. In addition to supporting dechlorination, the
reducing power of F e(0) oxidation can reduce water to hydrogen gas, called “cathodic
hydrogen,” through a process termed “anaerobic corrosion”. Thus, in addition to directly
catalyzing the dechlorination of chlorinated solvents, F e(0) oxidation can produce an
important electron donor, H2, for anaerobic microbes.

A variety of chlororespiring bacteria are capable of growth from the energy
released during the reductive dechlorination of chloroethenes [4, 5]. Remediation
strategies employing these organisms are promising due to their high dechlorination rates.
Halorespiration in Chloroethene-contaminated aquifers is often limited by deﬁcient
electron donor. Biostimulation technologies involve the addition of an exogenous
electron donor, or nutrients, to the aquifer stimulate dehalogenation [6]. Hydrogen is a
favorable electron donor for dechlorination [7, 8] and is even required by some
dehalogenating bacteria as an electron source [9, 10]. The introduction of hydrogen into

ground water at sustained levels to support dechlorinating populations is a major

99

challenge to bioremediation. Since the anaerobic corrosion of Fe(O) produces substantial
cathodic hydrogen, implementation of F e(O) permeable reactive barriers could
biostimulate microbial dechlorination.

Cathodic hydrogen can serve as an electron source for several hydrogenotrophic
anaerobic respiratory processes including: sulﬁdogenesis [11], autotrophic denitriﬁcation
[12], acetogenesis [l 1], and methanogenesis [13]. Analysis of microbial communities
associated with ﬁeld implementation of Fe(O) reactive barriers show stimulation of these
populations down-gradient of the treatment zone [14]. Potential positive effects of
microbes on Fe(O) remediation include removal of hydrogen and the oxidation of iron
from the reactive metal surface [1].

Few studies have examined the direct role of microbes capable of reductive
dechlorination during the implementation of Fe(O)-stimulated remediation. Mixed
cultures showed enhanced dechlorination rates with microbial F e(0) systems for both
chloroform [15] and TCE [16]. A Dehalococcoides containing mixed culture, however,
was suggested to negatively inﬂuence the TCE dechlorination product distribution by
stimulating production of vinyl chloride (VC), which is a known human carcinogen [16].
A study with a pure culture of Desulﬁtobacterium sp. Y51, showed enhanced
dechlorination of PCE in the presence of F e(0) [17]. In addition to the presumed
utilization of cathodic hydrogen, this study suggested that the combined microbial-Fe(O)
system produced favorable redox and pH conditions for microbial dechlorination[l].

In addition to cathodic hydrogen, anaerobic corrosion produces Fe(II), which can
serve as electron donor for certain anaerobic respiratory processes. Several studies have

shown ferrous iron oxidation coupled to nitrate reduction [18-24]. Since halorespiration is

100

similar energetically to nitrate reduction to nitrite [7, 25], thermodynamically,
halorespiring organisms could use F e(II), or even Fe(O) itself, as the immediate electron
donor for reductive dechlorination in addition to, or in place of, cathodic hydrogen.

The purpose of this study was to determine the immediate electron donors to
microbial reductive dechlorination following F e(0) amendment and examine the

synergism of abiotic and microbial dechlorination processes.

101

MATERIALS AND METHODS
Chemicals. The following chlorinated compounds were used in this study:
tetrachloroethene was obtained from Aldrich Chem. Co., Milwaukee, Wis .
Trichloroethene, trans-1,2-dichloroethene, cis-1,2-dichloroethene, vinyl chloride, 1,1,1-
trichloroethane, 1,1-dichloroethane, Chloroethane, 1,2-dichloroethane, ethane, and ethene

were all obtained from Supelco Co. All other chemicals used in this study were obtained

from Aldrich Chem. Co.

Iron pretreatment. The elemental iron used was 100-mesh cast iron powder with an
average particle size of 4.95 um (Aldrich Chem. Co.) and was acid washed to remove any
accumulated corrosion on starting material. To acid wash the iron, distilled water (91.4
ml) was added to a 160 ml serum bottle sealed with a rubber stopper and aluminum crimp
seal. The bottle was evacuated to —28 mm Hg for 5 — 10 min. before ﬂushing with argon.
The evacuation-ﬂush procedure was repeated three times before the rubber stopper was
removed under a stream of argon. Concentrated HCl (11.6 N, 8.6 ml) was added to the
anoxic water to give a ﬁnal HCl concentration of approximately 1 N. The HCl solution
was ﬂushed for 5 min. with argon before the addition of 50 g of iron powder. The
mixture was bubbled for 20 min with argon before sealing with a butyl rubber stopper
and capping. The mixture was allowed to sit for 60-120 min. at room temperature. The
iron suspension supernatant was then decanted under a stream of argon after brief
centrifugation at low speed. The iron was washed ﬁve times with argon-sparged distilled

water. After washing, the acid-treated iron was dried at 100 ° C under a stream of argon

102

for 4-6 h. The treated iron was stored under argon and used within 24 — 48 h of

preparation.

Bacterial cultures and medium preparation. Pure cultures of the following organisms
were used: Sulfurospirillum multivorans (formerly Dehalospirillum multivorans ) [26],
Dehalobacter restrictus strain PerK23 [27], Desulfuromonas michiganis strain BB1 [28],
Dehalobacter sp. strain TCAl [29].

Anaerobic, bicarbonate buffered minimal medium (30 mM, pH 7.3) was prepared
as previously described [30, 31] without the addition of reductant, rezazurin or a
secondary buffer. Wolin’s vitamins and 0.5 mM acetate were added via syringe aﬁer
autoclaving. Iron was added to the medium in small amounts from a stock suspension or
larger amounts were added directly to the medium under a stream of N2:C02 (20%).
Dehalogenating microbes required a reductant, dithiothreitol or sulﬁde, in treatments not
containing iron.

Batch experiments were performed in 160 ml serum bottles with either 80 or 100
ml liquid volume. Bottles were incubated horizontally in the dark on a rotary shaker
operated at 100 rpm. Chlorinated solvents were added neat using Hamilton syringe with
chaney adaptors, except for speciﬁed treatments in which PCE (95 umol) was added in
0.2 ml of a hexadecane carrier phase.

lDual chamber batch reactors resembling two fused serum bottles were purchased
from Lillie Glassblowers (Smyrna, GA; Figure 5.11). Iron powder (1.0 or 0.2 g) was

weighed in to one chamber of a nitrogen-ﬂushed reactor. The reactors were sealed with

103

butyl rubber stoppers and aluminum crimp caps and autoclaved at 121 °C for 25 min.
Minimal medium was prepared anaerobically, 50 ml dispensed into two l60-ml serum
Figure 5.1. Dual chambered batch reactor. Each side of the reactor was approximately

160 ml, and the upper portions of each were connected to allow for a common headspace.
The tops were sealed with butyl rubber stoppers and aluminum crimp tops.

 

 

 

 

 

104

bottles and sterilized. The medium was transferred, 50 ml into each chamber, by ﬂushing
with 80:20 N2:CO2 gas. Wolin’s vitamins were added to each chamber. Dithiothreitol
(0.2 mM) and 1 ml of a S. multivorans or D. restrictus stock was added to the chamber
that did not contain iron. PCE (1 ul, neat) was added to each reactor chamber to initiate

the experiment. Chloroethene headspace analysis was performed periodically.

H2 consumption. Electron donor, Fe(O), limited batch reactors with 22 mg of Fe(O) and
95 umol PCE in 0.2 ml hexadecane were incubated with and without S. multivorans.
Headspace analysis of hydrogen concentrations were determined periodically during a

10-month incubation.

Analytical techniques. Chlorinated ethenes and methane were measured in headspace
samples at 25°C by direct GC injection. For chlorinated ethenes, 200 pl samples were
analyzed on an HP model 6850 GC equipped with a Megabore model DB-624 column
(30 m by 0.543 mm; .1 &W Scientiﬁc), a ﬂame ionization detector and nitrogen carrier
gas. The temperature program was 50°C for 3.5 min, increased at 50°C/min to 200°C,
and held at that temperature for 2 min. For ethene, ethane, and methane analyses, 200 pl
headspace samples were injected onto a stainless steel Porapak Q column and detected
with a ﬂame ionization detector with nitrogen carrier gas. The oven temperature was
isocratic at 60°C. Volatile standards were prepared by adding a known amount of the
compound to serum bottles with the same liquid-to-headspace ratio and temperature as
that for the cultures being analyzed.

Hydrogen was measured on a GC equipped with reduction gas detector (Trace

Analytical, Menlo Park, Calif.) N2 was used as the carrier gas at a ﬂow rate of

105

approximately 30 ml/min. Dilutions of commercial hydrogen standards were used to
construct a standard curve. Under these conditions the detection limit for H2 with a 1 ml
injection loop was less than 0.5 ppmv. For all headspace measurements, gas-tight 250-ul
glass syringes (Hamilton, Reno, Nev.) with gas-tight Teﬂon valves and Luer Lock

adapters were used.

106

RESULTS

Fe(O) could serve as the electron donor for dechlorination by Dehalobacter
restrictus, and Sulfurospirillum multivorans, two hydrogenotrophic dechlorinators. D.
restrictus rapidly converted 20 umol PCE to cis-DCE in the presence of a low amount of
F e(0) (80 mg, 1.43 umol) (Figure 5.2). Dechlorination began before the ﬁrst 4-h time
point, transiently producing TCE before rapid accumulation of cis-DCE. Likewise, S.
multivorans converted 95 umols of PCE to primarily DCE within 10 days when amended
with 200 mg Fe(O) (Figure 5.3). In this system only trace formation of dechlorination
products was detected in the purely abiotic treatment. D. michiganensis, a dechlorinator
unable to use H2, did not dechlorinate PCE in the presence of F e(0) during a 6 month
incubation period. Neither S. multivorans nor D. michiganensis could dechlorinate PCE
with 10 mM FeSOI as electron donor during a 6 months incubation period.

A dual chambered batch reactor with a common headspace was used to ascertain
the use of cathodic hydrogen by halorespiring bacteria. S. multivorans was incubated in
the dual chambered batch reactor and, after 14 days, converted almost all PCE to
primarily cis-DCE with smaller amounts of ethene, ethane, and VC (Figure 5.4). In
controls without Fe(O) only small fractions of cis-DCE and TCE were formed, with the
remainder being recovered as PCE. D. restrictus cultures added to dual chambered batch
reactors, completely dechlorinated PCE within 11 days, whereas, there was no loss in

controls without D. restrictus (Figure 5.5).

107

Having shown that dechlorinating bacteria use cathodic hydrogen, in order to
assess the ability of microbes to inﬂuence hydrogen levels in the system, we incubated
electron donor-limited, Fe(O) batch cultures for 10 months with PCE and with or without
Figure 5.2 Dechlorination of 20 umol of PCB in the presence of 80 mg Fe(O) by
Dehalobacter restrictus, a hydrogen utilizing PCE to cis-DCE dechlorinating bacterium.

In the absence of F e(0) and other electron donors, D. restrictus does not signiﬁcantly
reduce PCE.

30

 

- o- PCE: Fe(O)
only

+cDCE: Fe(O) +
D. restricuts

+TCE: Fe(O) +
D. restrictus

-O— PCE: F e(0) +
D. restrictus

 

Chloroethenes (umollbottle)

 

 

 

Days

108

Figure 5.3. Dechlorination of 95 umol PCE in 0.2 ml hexadecane in the presence of 200
mg of Fe(O) aﬁer 10 days of incubation (A) with S. multivorans and (B) without S.
multivorans

A
l 80-

umollbottle
S 8

N
O
1

 

 

 

PCE TCE DCE vc ETH

1oo -
so -
so -
4o -

umollbottle

20J

 

PCE TCE DCE VC ETH

109

Figure 5.4. Dechlorination of 20 umol PCE in in dual chambered batch reactors in the
presence of S. multivorans (A) with 1000 mg of F e(0) and (B) without after 14 days of
incubation.

umollbottle
_L .5 N
O (II 0

on
I

 

 

 

PCE TCE DCE VC ETH Total

20F
15’

10'

umollbottle

 

 

PCE TCE DCE VC ETH Total

110

Figure 5.5. Dual chamber batch reactor: Dechlorination of PCE in the presence of 200 ug
Fe(O) with (diamonds, replicate A, and squares, replicate B) and without (circles)

Dehalobacter restrictus in the accompanying chamber

PCE (umol/bottle)

30

 

I-t N N
0| 0 UI
I I .

I—I
c
1

UI
I

 

 

0

 

Days

111

 

12

+ Fe(0) + D.
restrictus; Rep A
+ Fe(0) + D.
restrictus; Rep B
'0'- Fe(0) only

S. multivorans. The duplicate bottles of Fe(O) maintained headspace concentrations of 3
and 4% H2. S. multivorans amended duplicates, however, decreased the hydrogen
headspace concentrations several orders of magnitude lower to 1.5 ppmv and 0.5 ppmv.

To examine the synergistic interactions of microbial and abiotic dechlorination of
lesser chlorinated ethenes, F e(0) and TCE were incubated with and without D. restrictus.
In this experiment, the TCE loss rates in microbial and abiotic cultures were almost
identical (Figure 5.6).

In order to determine the trend in abiotic Fe(O) dechlorination rate with degree of
chlorination, the rate of PCE, TCE and cis-DCE dechlorination by F e(0) was compared
over a 10 day period (Figure 5.7). The dechlorination increased with the decreasing
degree of chlorination. Cis-DCE was rapidly removed, whereas, PCE was hardly lost.
After 10 days, dechlorination slowed for both cis-DCE and TCE.

I also evaluated how the saturated chlorinated solvent, 1,1,1-trichloroethane (TCA),
was metabolized by Fe(O) versus its microbial dechlorination by Dehalobacter sp. strain
TCAl. TCA was rapidly transformed by Fe(O). Dechlorination was nearly complete
within 2 days, and 1,1-DCA was the primary product, and only disappeared slowly. No
dechlorination was observed during the experiment in either treatment with strain TCA].
Strain TCAl typically displays a lag upon 1% transfer into fresh medium. However, in
order to explain the inhibition of Fe(O) mediated dechlorination in the presence of TCAl
inoculum, we tested several possible medium components for their effect on Fe(0)-
stimulated dechlorination. All amendments slowed TCA disappearance in the presence of
Fe(O) (Figure 5.8). Cysteine and DTI‘ were the strongest inhibitors of dechlorination,

inhibiting more than 80% the extent of dechlorination at 140 h. Humic acids also

112

Figure 5.6. Dechlorination of TCE in the presence of Fe(O) with and without D.
restrictus.

 

25

   
   

+TCE: Fe(0) + D.
restrictus

+ TCE: Fe(0) Only

N
G

 

 

 

E
‘515
g
G
E
3
m10—
o
[—
5_
07 l I I
0 5 10 15

113

Figure 5.7. Abiotic, Fe(0)-mediated dechlorination of PCB (diamond), TCE (square), and
cis-DCE (triangle) as parent compounds.

 

  
   
  

y = -0.38x + 20
R2 = 0.97

 

10 - y=-l.3x+26
R2=O.99

A

y = -2.8x + 34
R2 = 0.99

Chloroethenes (umoles/bottle)

U!
I

 

 

 

G
d
.l

G
N
A
0‘
no

10 12 14

114

Figure 5.8. Effect of inhibitors on Fe(O) mediated dechlorination of 20 ul 1,1,1-
trichloroethane. Inhibitors were cobalamin (Cob), sulﬁde (S2-), DTT, cysteine (Cys),
rezazurin (Res), AQDS, and humic acids (humics).

 

 

 

 

 

 

 

 

100 +Fe(0) only

80 q +Fe(0) + Cob
m -III-Fe(0) + 52-
E 60 . -)<-Fe(0) + DTI‘
5 *Fe(0) +Cys
a: 40 - +Fe(0) + Res

-l—Fe(0) + AQDS
20 - —Fe(0) + Humics
O
O 50 100 l 50

Hours

115

inhibited dechlorination more than 50% within this time period. Other medium
amendments also slowed F e(0) dechlorination, but each dechlorinated TCA completely

within 140 h.

116

DISCUSSION
To better understand the role of microbes during Fe(O) remediation, we performed
laboratory studies to determine the direct electron donors to pure cultures of halorespiring
organisms and to assess the interaction of microbial and abiotic processes.

Both hydrogentrophic bacteria tested were able to dechlorinate PCE with Fe(O)
added as the sole electron donor. To determine if cathodic hydrogen was the electron
donor, we conducted experiments in a dual chambered batch reactor that allowed only
volatile compounds to exchange between the two chambers. Similar approaches have
been used historically to prove the use of cathodic hydrogen as an electron donor for
methanogenesis and more recently for autotrophic denitriﬁcation [13, 32]. In single
chamber and dual chamber reactors, cis-DCE, a microbial, not abiotic product, was
rapidly produced from PCE proving cathodic hydrogen supported dechlorination.

To examine the possible role of direct electron donors other than cathodic
hydrogen, we tested a PCE dechlorinating bacterium, D. michiganensis strain BB1,
which can not use hydrogen for dechlorination. PCE was not dechlorinated by strain BB1
in the presence of F e(0) or F e804 in incubations lasting 6 months. Similar results with
FeSO4 as electron donor were found with S. multivorans. These experiments with strain
88] indicate that F e(0) or F e(II) do not serve as direct electron donors for PCE
dechlorination for this organism.

To explore the implication of cathodic hydrogen consumption by dechlorinating
microbes, we performed long term incubations of PCB in Fe(O) limited PCE batch
reactors with and without S. multivorans. F e(0) — only treatments showed 3-4% hydrogen

in the headspace versus approximately 1 ppmv hydrogen in the S. multivorans amended

117

reactors. These results illustrate the large amount of hydrogen equivalents produced with
relatively small amounts of elemental iron and also the ability of dechlorinating microbes
to consume these large amounts of hydrogen to levels approaching the threshold for
energy metabolism. Microbial consumption of cathodic hydrogen has been suggested to
maintain the reactivity of the iron surface and help minimize clogging of aquifer pore
spaces with hydrogen gas [1]. The experiments performed here demonstrate another
beneﬁt of a combined microbial-Fe(O) system by showing that halorespiring bacteria
make the system more efﬁcient by channeling electrons from Fe(O) to dechlorination that
would otherwise be wasted as cathodic hydrogen in a purely abiotic scheme.

Although we show synergism of microbial and abiotic dechlorination of
chlorinated solvents, some limits to the positive interactions are also noted. The
synergism was pronounced when the biological and Fe(O) catalysts had complementary
reaction optima. In the case of the chloroethenes, the PCB to TCE conversion by the
bacteria was much faster than by F 6(0) but the conversion of the lesser chlorinated
ethenes is faster when catalyzed by Fe(O), than the bacteria. Hence there is a synergistic
effect to overall conversion. Reactivity trends with increasing chloroethene halogenation
have been controversial in the literature, however. Early studies suggested that lesser
chlorinated ethenes were dechlorinated more slowly than higher chlorinated ethenes [33].
A subsequent study has clearly demonstrated the reverse trend [3], and differences from
earlier reports were attributed to the persnickety nature of iron mediated dechlorination
and to variations in experimental conditions. Our results support the second report by

showing Fe(O) increasing reactivity with decreasing degree of substrate chlorination.

118

Dechlorination of chlorinated ethenes may represent a unique opportunity for
synergism between microbial and Fe(0) dechlorination, because each catalyst has the
opposite reactivity trend with degree of chlorination of the substrate. This occurs because
microbes and Fe(O) dechlorinate chloroethenes via fundamentally different mechanisms.
Fe(O) reacts primarily through elimination reactions producing chloroacetylenes, which
are further dechlorinated and reduced to ethene and ethane [3]. Halorespiring bacteria,
however, reductively dechlorinate chloroethenes through hydrogenolysis [34].

In contrast to the chloroethenes, the dechlorination of TCA, another common
environmental contaminant, did not show synergism with the two catalysts. Both F e(O)
and strain TCAI catalyzed one step of dechlorination to 1,1-DCA, and the Fe(O) reaction
was so fast that the conversion was complete before the dechlorinating strain completed
its lag phase.

Fe(O) is not reactive towards all organohalogens [1], and other chlorinated
solvents not explored in this study may have the opposite effect to that observed for TCA.
1,2-Didchloroethane and 1,2-dichloropropane, for example, are two compounds that do
not readily react with Fe(O), but for which microbial dechlorination is documented [5].
The role of Fe(O) in the remediation of these compounds would be as a biostimulant that
produces an electron donor, cathodic hydrogen, and creates or maintains low redox
conditions. The Fe(O) would not participate directly in dechlorination.

In general, dechlorinating strains that grew with little or no lag and that could
dechlorinate at a high rate (ca. 100 umol*l"*d") were able to complement abiotic Fe(O)
dechlorination. In our studies these strains were D. restrictus, and D. multivorans, both of

which are PC E to cis-DCE dechlorinators. Dechlorinating strains that dechlorinate more

119

 

slowly, have substantial lag periods, or are sensitive to pH changes, should not be
expected to signiﬁcantly contribute to dechlorination in the presence of Fe(O). The
beneﬁts of the synergism of microbial and F e(0)-mediated dechlorination are determined
by favorable combinations of chlorinated substrate, dehalogenating bacterial strains, and

the geochemical environment.

120

10.

11.

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Arnold, WA. and AL. Roberts, Pathways and kinetics of chlorinated ethylene
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Holliger, C., G. Wohlfarth, and G. Diekert, Reductive dechlorination in the

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Lofﬂer, F.E., et al., Diversity of dechlorinating bacteria, in Dehalogenation :
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Lendvay, J .M., et al., Bioreactive barriers: A comparison of bioaugmentation and
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Dolﬁng, J., Thermodynamic Considerations for Dehalogenation, in
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Dolﬁng, J. and J .E.M. Beurskens, The microbial logic and environmental

signiﬁcance of reductive dehalogenation. Advances in Microbial Ecology, Vol
14, 1995. 14: 143-206.

MaymoGateIl, X., et al., Isolation of a bacterium that reductively dechlorinates
tetrachloroethene to ethene. Science, 1997. 276: 1568-1571.

He, J .Z., et al., Detoxiﬁcation of vinyl chloride to ethene coupled to growth of an
anaerobic bacterium. Nature, 2003. 424: 62-65.

Rajagopal, BS. and J. Legal], Utilization of Cathodic Hydrogen by Hydrogen-
Oxidizing Bacteria. Applied Microbiology and Biotechnology, 1989. 31: 406-412.

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Daniels, L., et al., Bacterial Methanogenesis and Growth from C02 with
Elemental Iran as the Sole Source of Electrons. Science, 1987. 237: 509-511.

Gu, B.H., et al., Microbiological characteristics in a zero-valent iron reactive
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Weathers, L.J., G.F. Parkin, and RI. Alvarez, Utilization of cathodic hydrogen as

electron donor for chloroform cometabolism by a mixed, methanogenic culture.
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Li, T. and J. Farrell, Reductive dechlorination of trichloroethene and carbon

tetrachloride using iron and palladized-iron cathodes. Environmental Science
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Lee, T., et al., Efﬁcient dechlorination of tetrachloroethylene in soil slurry by
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Heising, S., et al., Chlorobiumferrooxidans sp. nov., a phototrophic green sulfur

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Hauck, S., et al., Ferrous iron oxidation by denitrijying bacteria in proﬁmdal
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Heising, S. and B. Schink, Phototrophic oxidation of ferrous iron by a
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Benz, M., A. Brune, and B. Schink, Anaerobic and aerobic oxidation of ferrous

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Straub, K.L., M. Benz, and B. Schink, Iron metabolism in anoxic environments at
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Scholzmuramatsu, H., et al., Isolation and Characterization of Dehalospirillum
Multivorans Gen-Nov, Sp-Nov, a Tetrachloroethene- Utilizing, Strictly Anaerobic
Bacterium. Archives of Microbiology, 1995. 163: 48-56.

Holliger, C., et al., Dehalobacter restrictus gen. nov. and sp. nov., a strictly
anaerobic bacterium that reductively dechlorinates tetra- and trichloroethene in
an anaerobic respiration. Archives of Microbiology, 1998. 169: 313-321.

Sung, Y., et al., Characterization of two tetrachloroethene-reducing, acetate-
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Sun, B., et al., Microbial Dehalorespiration with 1,1,1-Trichloroethane. Science,
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Lofﬂer, F.E., R.A. Sanford, and J .M. Tiedje, Initial characterization of a
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Lofﬂer, F .E., et al., Complete reductive dechlorination of 1,2-dichloropropane by
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Till, B.A., L.J. Weathers, and P.J.J. Alvarez, F e(0)-supported autotrophic
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Johnson, T.L., M .M. Scherer, and PG. Tratnyek, Kinetics of halogenated organic
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Holliger, C., C. Regeard, and G. Diekart, Dehalogenation by Anaerobic Bacteria,
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123

SUMMARY AND CONCLUSIONS

124

In this Dissertation I ﬁrst assessed the inconsistencies in data on the frequency of
occurrence of chlorinated ethenes and related compounds in our environment. Next, I
described a novel dechlorination pathway in which PCB and TCE are converted to a
mixture of trans-DCE and cis-DCE in a ratio of 3:1 by highly enriched cultures. I showed
evidence for dehalogenation activity in the pristine environment of Antarctic melt-pond
sediments. Finally, 1 characterized the role of halorespiring organisms in the engineered
remediation of chlorinated ethenes using Fe(O) as an electron donor. This research
contributes to our current understanding of microbial reductive dechlorination and points
to new research to be done in the ﬁeld.

The analysis of the frequency of pollutant occurrence at NPL sites data clariﬁes
inconsistencies regarding the extent of contamination at the Nation’s most contaminated
sites and the pervasiveness of certain chemicals at the sties. l was particularly troubled by
the inconsistencies in these widely used databases. Differences in these numbers are
important as policy makers, the public, and scientists use them to gauge the importance of
environmental contaminants. I show that the EPA and ATSDR databases contain both
unique and shared information and that a composite dataset best describes the frequency
of chemical occurrence at NPL sites. Chlorinated solvents, especially chlorinated ethenes,
are present at a majority of these sites and are frequently found in mixtures with other
chlorinated solvents or BTEX compounds. Certain chlorinated solvents co-occur at >90%
of the sites in which they are present. The prevalence of chloroethenes at National
Priority List sites underscores the importance of developing remediation strategies and

understanding the environmental chemodynamics of chloroethenes. While I only

125

analyzed chlorinated ethanes, chlorinated ethenes, and BTEX compounds, the same
procedure could be used to obtain more accurate data for other classes of environmental
contaminants. My ﬁndings in Chapter 2 have been communicated to both the EPA and

ATSDR in order to improve the quality of the publicly available databases.

Evaluation of monitoring data, modeling of aquifer chemistry, and choosing
among remediation strategies relies on understanding the transformation processes of the
chemicals of interest. In Chapter 3, I describe a novel microbial dechlorination
transformation in enrichment cultures in which trans-DCE is produced as the major
product from PCB and TCE reduction. The production of trans-DCE in a 3:1 ratio with
cis-DCE is not explained by current understanding of dehalogenase reactions. Catalysis
is likely performed by a novel chloroethene dehalogenase, as the known tceA reductase of
Dehalococcoides ethenogenes strain 195 was not detected in trans-DCE producing
cultures. 1 show that Dehalococcoides populations perform trans-DCE production, but
three years of effort to isolate the dechlorinating populations were unsuccessful. In
general, isolation of Dehalococcoides species has been notoriously elusive. Continued
research into the physiology of the cultures in hand will not only aid future isolation
attempts on these and other cultures but also in the engineered use of Dehalococcoides
for remediation.

One consequence of the research on reductive dehalogenation for pollution
abatement is that many dehalogenating organisms are enriched or isolated from
environments not contaminated with the compound of interest or even at pristine sites.

Several of the enrichments in Chapter 3 that produce trans-DCE were from “pristine”

126

river sediments. Dehalococcoides, like those detected in the trans-DCE producing
cultures (Chapter 3), are obligate halorespiring organisms; how they survive in pristine
environments is still unknown. Chapter 4 was a ﬁrst step in addressing the role of
dehalogenating organisms in natural systems. I hypothesized there was a halogen cycle in
pristine melt ponds in Antarctica containing photosynthetic mats that produce
halocarbons [l], and thus I looked for halorespirers and dehalogenation activity. 1
established an array of enrichments for dehalogenation activity and detected activity for
PCE conversion to TCE. In some cases dechlorination continued to cis-DCE or a mixture
of trans-DCE and cis-DCE in a ratio of 3:1, as observed in Chapter 3. In addition to
chloroethenes, I reported the debromination of both 2-bromophenol and 3-bromobenzoate
by Antarctic melt pond sediments. The brominated aromatics were dehalogenated but
chlorinated aromatics were not. I estimated the population size of 2-BP debrominators to
be 103-104 cells/g sediment suggesting these organisms had grown in this environment
since their original dispersal to the region. Three genera known to contain mesophilic
chloroethene dechlorinators were detected in enrichment cultures that transformed PCE.
Linking identity to function is a major challenge for microbial ecology, and could not be
done in this study. Nonetheless, it is interesting to know that the distribution of these
novel dechlorinating genera extends to Antarctic regions. These are promising results and
justify the further study of halogen cycles in Antarctic melt ponds and the ecological role
of dehalogenation in nature.

Whereas Chapter 4 looked for a new role for dehalogenating bacteria in a pristine
environments, Chapter 5 explores a new role for halorespirers in an engineered system.

Fe(O) remediation technology is promising as a cost effective strategy since, aﬂer

127

installation, it is a passive process. F e(0) remediation is currently being used to clean up
aquifers. An inevitable consequence of adding Fe(O) to groundwater is the substantial
production of hydrogen, an important energy currency in anaerobic systems. The effects
of cathodic hydrogen on microbes in general in these systems are currently under
investigation and even less is known about halorespiring organisms’ roles in F e(0)
remediation. In Chapter 5 I used carefully designed laboratory studies to tease apart the
immediate electron donors to dechlorination and show deﬁnitively that cathodic
hydrogen supports dehalogenation. I further show the synergistic effects of combined
microbial and abiotic dechlorination, but also set limits to this favorable interaction. I
propose that chloroethenes, especially PCE, may provide a unique opportunity for
synergistic effects between microbial and abiotic dechlorination that would not be
realized with other chlorinated solvents.

Future research on microbial dechlorination should begin to focus on other
persistent halocarbons. Chlorinated C2-C4 compounds are frequent environmental
contaminants and relatively little is known about their microbial transformations
compared to the chlorinated ethenes [2]. Dechlorination of chlorinated ethanes are even
slightly more energetically favorable than reductive dechlorination of the ethenes [3]. To
date the only chlorinated-ethane halorespiring-bacteria are: Dehalococcoides ethenogenes
strain 195 [4] and Dehalococcoides sp. strain BAV1[5], which both convert 1,2-
dichloroethane to ethene, and Dehalobater sp. strain TCA 1 , which dechlorinates 1,1,1-
TCA to Chloroethane[6].

Bioremediation is a successful, if not preferred, remediation strategy for

chlorinated ethene contaminated aquifers. Biostimulation and bioaugmentation are being

128

 

performed at hundreds of sites in the United States. Although bioremediation is a proven
remediation strategy for certain chlorinated solvents, no such ﬁeld-ready solution is
available for other major classes of environmental pollutants such as PCBs, DDT, and
dioxins. Future research should apply the lessons learned from chlorinated ethenes to
these other more challenging compounds. In the study reported in the Appendix I tried to
stimulate the microbial dechlorination of PCBs using F e(0); my results reinforce the
difﬁculty of working with low solubility compounds in sediment systems.

Additional work must be done to examine the role of halorespiring organisms in
the dehalogenation of natural halocarbons in order to aid in the cataloging of halocarbon
‘sinks’ for the construction of global budgets. Such budgets are needed to accurately
assess global change and the effects of halocarbon targeted policies and legislation. The
further study of natural halogen cycles undoubtedly would result in further insights into
the physiology and ecology of halorespiring organisms. Knowledge of how microbes turn
over halogenated compounds could be capitalized for the purposeful destruction of

anthropogenic pollution.

129

REFERENCES

Schall, C ., et al., Biogenic brominated and iodinated organic compounds in ponds
on the McMurdo Ice Shelf Antarctica. Antarctic Science, 1996. 8(1): p. 45-48.

De Wildeman, S. and W. Verstraete, The quest for microbial reductive
dechlorination of C-2 to C-4 chloroalkanes is warranted. Applied Microbiology
and Biotechnology, 2003. 61(2): p. 94-102.

Dolﬁng, J ., Thermodynamic Considerations for Dehalogenation, in
Dehalogenation .' microbial processes and environmental applications, M.M.
Haggblom and ID. Bossert, Editors. 2003, Kluwer Academic Publishers: Boston.
p. 89-114.

Maymo, g.X., T. Anguish, and SH. Zinder, Reductive dechlorination of
chlorinated ethenes and 1,2-dichloroethane by "Dehalococcoides ethenogenes ”

195. Applied and Environmental Microbiology, 1999. 65(7): 3108-3113.

He, J .Z., et al., Detoxification of vinyl chloride to ethene coupled to growth of an
anaerobic bacterium. Nature, 2003. 424(6944): p. 62-65.

Sun, 8., et al., Microbial Dehalorespiration with 1,1,1-Trichloroethane. Science,
2002. 298(5595): p. 1023-1025.

130

APPENDIX 1

F E(0) STIMULATION OF PCB REDUCTIVE DECHLORINATION

131

ABSTRACT

Fe(O) abiotically dechlorinates a wide range of chlorinated compounds and is a
proven remediation approach for chlorinated solvent contaminated groundwater.
Increasingly the synergism of microbial and Fe(O) mediated dechlorination is being
recognized. This study tested the hypothesis that microbial dechlorination of PCBs would
be enhanced by Fe(O) amendment. The experimental results reported here clearly
demonstrated that Fe(O) added to sediment microcosms, in fact, inhibited microbial
dechlorination, presumably by increasing the pH beyond the suitable range for
dehalogenating microbes. No abiotic Fe(O) PCB dechlorination was detected either, and it
is suggested that the added iron was corroded at a rate faster than dechlorination.
Analyses of the sediment microcosms for all treatments indicate that upon extended
incubations these systems are extremely electron donor limited. We suggest that slow
release hydrogen compounds may beneﬁt microbial PCB dechlorination, in addition to

new, modiﬁed-Fe(O) amendments.

132

 

INTRODUCTION

Polychlorinated biphenyls (PCBs) are a class of contaminants that are globally
distributed and are responsible for a variety of toxicological effects in humans and
ecosystems. PCBs were manufactured and widely used in industry due to their favorable
physical and chemical properties including low vapor
pressures, low water solubility, excellent dielectric

properties, stability to oxidation, ﬂame resistance, and CI CI

relative inertness [1]. Many of these same properties,

Figure ALI. Generalized
and the now evident toxicity of PCBs, have proven structure of polychlorinated

biphenyl (PCB)
problematic, as massive amounts of PCBs have been
released in to the environment. This environmental burden is estimated at several
hundred thousand kilograms of the estimated 600 million kilograms of PCBs produced
worldwide. Structurally, the carbon backbone of polychlorinated biphenyls (PCBs)
consists of two benzene rings connected by a carbon-carbon bond at the C-1 carbon. A
combination of up to 10 chlorines may occupy the remaining positions; thus, these
permutations lead to 209 possible compounds, or congeners (Figure AI.1). Commercial

mixtures of PCBs vary in their congener composition; the chemical and physical

properties of each congener are also different.

The capacity for intrinsic biodegradation of PCBs appears to be widespread,
occurring in both contaminated and pristine sites [2]. The extent of this dechlorinating
capability varies considerably among sites and is dependent on many environmental

factors including temperature, pH, and availability of carbons sources, electron donors

133

 

(hydrogen), and competing electron acceptors [1]. These factors along with the
bioavailability of the PCBs and the composition of the active microbial community
determine the rate, extent, and route of PCB dechlorination. Despite the widespread
potential of microbial reductive dechlorination, degradation of PCBs at contaminated
sites is often slow and incomplete [3].

Efforts to stimulate PCB dechlorination through the addition of exogenous
electron donors have had some success. Since dechlorination is a reductive process, a
source of electrons is needed to complete the reaction. An obvious approach to stimulate
reductive dechlorination then is the addition of supplemental electron donors. Alder et al.
[4] found that addition of a mixture of fatty acids stimulated PCB dechlorination in
sediment slurries with low carbon contents but not sediments with high carbon.
Sediment-free enrichment cultures were stimulated to varying degrees by the addition of
pyruvate and malate [1]. Hydrogen was also found to alter the pathway and products of
the dechlorination of 2,3,4-trichlorobiphenyl (-CB) by Hudson River Sediment
microorganisms [5]. Incubations with a headspace of H2/CO2 resulted in the
dechlorination of 234-CB to 24-CB and 23-CB, followed by conversion to 2-CB. Under
an N2 or N2/CO2 headspace, 234-CB was only converted to 23-CB. These studies suggest
that populations of dechlorinating and non-dechlorinating bacteria compete for and utilize
exogenous electron donors, thus determine the resulting dechlorination activity [1].
Although not completely understood, the addition of exogenous electron donors, like
hydrogen, has shown some promise in stimulating speciﬁc PCB dechlorinating

populations.

134

Until recently, the only consistently effective means of stimulating the reductive
dechlorination of PCBs involved the addition of speciﬁc PCB congeners or analogs.
These compounds “primed” the process, allowing for more extensive dechlorination of
the endogenous PCBs. Addition of more PCBs, or other similarly toxic compounds,
however, has not been accepted favorably by environmental regulators. Zwiemik et al.
[6] recently demonstrated F eS04 amendments stimulate extensive dechlorination of
PCBs. FeSOII is an inexpensive and innocuous compound and was found to enhance PCB
dechlorination through two mechanisms (Figure .. donor

2
A12). First, they showed FeSOI acts as an alternate "3%":3 + (21' + HI.

electron acceptor for dechlorinators which use sulfate,

 

 

Growth
allowing for growth of this population. Aﬁer the _ _ 1
s. a;-
sulfate is depleted (after several weeks), this much 1‘92+1
. . o- donor
larger population can dechlorInate PCBs at a faster
rate. The electron donors for both sulfate reduction as

and PCB dechlorination are presumed to be

endogenous carbon sources in the sediment. Second, £3813: sizinzlilati 01:10:: I Pg];
dechlorination. 1. Sulfate is
an alternate electron acceptor
for growth of the
dechlorinating population. 2.
This larger bacterial population
is able to dechlorinate PCBs at
a faster rate. 3. Sulﬁde from
sulfate reduction is detoxiﬁed
by co—precipitation with ferrous
iron.

ferrous iron co-precipitates with sulﬁde, the toxic
product of sulfate reduction. Sulﬁde is inhibitory to
many organisms including dechlorinating bacteria.
Speciﬁcally, ferrous sulfate appears to stimulate
dechlorinating populations that remove both para and
meta chlorines. Individual additions of sulfate, as Na2SO4, or F e”, as F eCl2, also

signiﬁcantly enhanced PCB dechlorination, but the combined effect was much greater.

135

Thus, the addition of ferrous sulfate as an alternate electron acceptor and sulﬁde co-
precipitate has proven an effective PCB dechlorination stimulant, but this process occurs

after a lag period and does not stimulate removal of ortho—substituted PCBs.

Presently research on the bioremediation of PCBs focuses on sequential anaerobic
and aerobic treatment [3]. Such systems take advantage of the ability of the anaerobic
process to attack highly chlorinated congeners and the ability of the aerobic process to
mineralize lesser-chlorinated congers. Research is now centered on the difﬁcult task of
engineering and implementing such systems that can effectively alternate between these
two redox states. Of course, a more desirable remediation approach would be a purely
anaerobic system that completely dechlorinates PCBs avoiding costly aeration procedures

and allowing for a more passive, less expensive, remediation.

Several studies have shown that zero-valent iron (Feo, F e(0), or ZVI) can
effectively catalyze the dechlorination of a variety of chlorinated aliphatic and aromatic
compounds. To date, most research has focused on chlorinated aliphatics such as
tetrachloroethene, trichloroethene, carbon tetrachloride, and chloroform, which are
readily reduced at room temperatures and pressures [7]. Tetrachloroethene, for example,
is completely dechlorinated and detoxiﬁed to ethene. Recently DDT, DDD, and DDE
were also shown to be dechlorinated by zero-valent iron [8]. Like PCBs, these
compounds also contain two aromatic rings and have low water solubilites. The rate of
transformation of these highly insoluble chlorinated aromatics was found to be limited by

dissolution and could be increased through the addition of a surfactant (Triton X-114).

136

 

Many zero-valent iron remediation technologies are currently being employed, including
permeable reactive barriers, slurry walls, ground covers, and canisters [7]. These
treatment approaches are becoming popular as they are relatively inexpensive, and once
implemented, treatment is passive. Although zero-valent iron has been shown to
completely dechlorinate a variety of chlorinated organics, few studies have examined
Fe(O)-mediated dechlorination of PCBs.

Whereas studies on chlorinated aliphatics and DDT analogs were performed at room
temperatures and pressures, most of the studies on zero-valent iron mediated
dechlorination of PCBs were performed at high temperatures and pressure. For example,
one study on the degradation of PCBs showed complete dechlorination to biphenyl using
zero-valent iron but only at temperatures above 200 °C [9]. Another study using nano-
scale iron particles showed about 25% dechlorination of PCBs after 17 h of incubation at
room temperature and pressure [10]. By coating these nano-scale iron particles with
palladium, the authors were able to completely degrade the PC88 to biphenyl in that time
period. Because of the greater success with the palladium-coated iron, little attention was
given to the promising results with uncoated iron. A 25% decrease in PCBs in only 17 h
should be considered impressive, as many microbial PCB dechlorination experiments are
incubated for 4-8 months. The iron experiments were conducted with specially prepared
nano-scale iron; the more commonly used commercial iron would certainly react slower
due to nature of the surface catalyzed reactions. We propose that although F e(0)-
catalyzed dechlorination of PCBs may be slow compared to dechlorination rates of
chlorinated solvents or incineration, it will be at least as fast as microbial degradation and

perhaps less speciﬁc.

137

The general mechanism of dechlorination by iron is understood. The low redox
potential of F eO/F e2+ half reaction drives the reduction of redox-reactive species including
chlorinated organics (eq 1), and both water in anoxic systems (eq 2) and oxygen under

aerobic conditions (eq 3) [7]:

Fe°+RCl+H+—-Fe2++RH+CI' (1)
Fe0 + 21120 —I Fe” + H2 + 20H' (2)
2Fe° + 02 + 2H2O —I 21=e2+ + 40H’ (3)

138

Recent work has shown that the rates of dechlorination of chlorinated solvents and of
reduction of nitroaromatic compounds by zero-valent iron are proportional to the speciﬁc
surface area (surface area per unit reactor volume)[7]. These results imply that the
reduction reactions occur by electron transfer at the iron surface. In an anoxic system in
addition to directly dechlorinating chlorinated compounds, FeO oxidation results in the
production of ferrous iron and hydrogen, which could be utilized by dechlorinating
bacteria (Figure AI.3). Hydrogen derived from the oxidation of elemental iron, termed
cathodic hydrogen, has been shown to support sulﬁdogenesis, acetogenesis,
methanogenesis, denitriﬁcation, and a dechlorinating methanogenic enrichment. In
addition to directly dechlorinating chlorinated organics, oxidation of zero-valent iron
results in the production of hydrogen, a potential electron donor for microbial reductive

dechlorination of PCBs.
Figure AI.3. Proposed

stimulation of PCB by F e(0).
1. Direct dechlorination of

3 PCBs by Fe(0). 2. Fe(O)

4 "3 "3+ C" t l" reduction of water to form

Poi”... er I 0.5 PCB AW 0,5 I.“ ferrous iron and hydrogen. 3.
V Cathodic hydrogen is an

electron donor for microbial

2\2 j/l dechlorination of PCBs. 4,

Ferrous iron is also a

 

 

I . F90 thermodynamically favorable

/’ 1 electron donor for abiotic or

, , microbial reductive
PCB+H F02 PCB+¢I' . .

5 dechlonnatlon. 5. Ferrous

IrIIs<—z s- iron co-precipitates

endogenous sulﬁde.

I further proposed the combined addition of FeSO4 and Fe(O) would have
synergistic effects leading to rapid dechlorination, with little or no lag, and resulting in

complete dechlorination of PCBs to biphenyl. In effect, this approach involves the

139

addition of both an alternate electron acceptor and an effective electron donor (Figure
AI.4). Such a treatment would rapidly increase the sulfate reducing population, consume
sulfate, precipitate sulﬁde, and provide the increased dechlorinating population with
abundant electron donor. Additionally, the Fe(O) itself may dechlorinate the PCBs
directly in rates comparable to microbial dechlorination, but perhaps not hindered by the
strict conformational requirements of enzyme-catalyzed degradation. Production of
cathodic hydrogen produces an abundance of electron donor to complement the added

electron acceptor, sulfate.

4 Pee PCB + 01' + H+

F835, or + 0.5 PCB

  

Figure AI.4. Proposed model for the synergism between FeSO4 and Fe(0) amended

PCB reductive dechlorination.

140

Research Design and Methods

Research Design. Experiments tested the efﬁcacy of zero-valent iron amended
remediation in sediment microcosms. The Hudson River sediments used were the same
as those used to study the effects of ferrous sulfate amendment on dechlorination [6].
Microbial reductive dechlorination of the PCB mixture was compared in the presence and
absence of Fe(0). Additionally, reductive dechlorination of PCBs was compared among

Fe(O), FeSO4, Fe(O) + FeSO4, formate, formate + FeSO4 and unamended sediments.

Sediment microcosms. Microcosms were established in 20 m1 glass vials with Teﬂon-
lined septum caps using sterile anaerobic RAMM media as previously described [6].
Amendments were made according to Table 1. The sediment-RAMM slurry (7 ml) was
added to vials under a stream of 20% C02 in N2. Selected vials were then autoclaved for
60 min at 121° C on three consecutive days to serve as abiotic controls. Arochlor 1242
(200 pg) was added to the reactors dissolved in 7 ul of acetone. An F e(0) suspension was
prepared by adding iron powder to a nitrogen ﬂushed serum bottle, sealing it, and
autoclaving it (Chapter 5). Sterile, anaerobic 30 mM bicarbonate buffer (pH 7.3) was then
transferred into the bottle. The iron powder suspension was then immediately shaken and
dispensed into selected treatments. Sterile, anaerobic F eSO4 was added to selected
treatments. Hydrogen (3 ml) was also added to selected treatments. Live treatments were
conducted in quadruplicate, and abiotic controls were performed in triplicate. Samples
were sacriﬁced and analyzed over a l6-week incubation period. Each sample was

analyzed for headspace hydrogen and methane prior to replacing septa and freezing the

141

samples for extraction. In addition to analysis of speciﬁc PCB congeners, Fe(II), and pH

 

were determined.

Extraction and analysis of PCBs. The protocol for extraction and analysis of PCBs was
performed as previously described[6]. The entire sample vials were extracted with hexane
and acetone by Dr. M. Zwiemik as previously described[6]. PCBs were analyzed by GC-

ECD by Dr. J. Quensen.

Headspace analysis. Hydrogen concentrations in the headspace were determined by
direct injection of a headspace sample into a GC equipped with a reducing gas detector.
Methane concentrations in the headspace were determined by GC-FID headspace

analysis.

Iron, sulfate and sulﬁde analysis. Acid extractable Fe(II)alq was measured by ﬁltering a
0.1 —0.5 ml aliquot of sample through a 0.22pm nylon syringe ﬁlter into 0.5 M HCl.
After an 1 h incubation an aliquot of the acidic iron solution was added ferrozine buffer
and the A562 recorded. The pH was determined using pH test strips. Volatile fatty acids

were determined by HPLC as previously described [11].

142

RESULTS
W- Methane was produced in all live treatments (Figure AI.5). The F e(0) only
treatment accumulated 4.2 (1.75 SD) % headspace methane by day 114. Within this time
period the three live treatments amended with FeSO4 accumulated between 16-18%
headspace methane. Formate and the no-amendment treatments accumulated 38 (10.4
SD) % and 31 (4.2 SD) %, respectively.
Hydrogen. F e(0) amendment, with or without F eSO4, resulted in very high headspace
hydrogen, > 35% H2 by day 20 (Figure A16). The hydrogen in the Fe(O) + F eSO4
treatment decreased by a sigmoidal function to 710 (210 SD) ppmv by day 114, where as,
in the Fe(O) treatment H2 continued to increase reaching 47 (7.8 SD) %. The no
amendment and F e804 treatments showed high variation in the ﬁrst two or three time
periods. In each case, the variation was due to a single replicate that was higher than the
others. In both treatments the trend showed an increase in H2 before decreasing to 3.3
(1.2 SD) ppm for no amendment and 0.8 (0.1 SD) ppm for FeSO4 treatment. The
hydrogen in the formate treatment decreased to 2.2 (1.2 SD) ppm, and in the FeSO4 +
formate treatment, it remained constant throughout the time course reaching 0.9 (0.4 SD)
ppm at day 114.
Volaﬁ fagv acids. Volatile fatty acids were only detected in the two live Fe(O) amended
treatments at week 16 (Figure AI.7). The Fe(O) treatment had 5.1 mM acetate and 160

uM propionate; Fe(O) + FeSO4 had 650 MW acetate and 60 pm propionate.

143

 

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A. Fe(O)

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pH. The pH was 8.75 for both Fe(0) and Fe(O) + FeSOI, and for the no amendment and
format treatment were 7.25 and 7.5, respectively (Figure A18). The pH was not measured
for theFeSO4 or F eSO4 + formate treatments.

Acid extractable aqueous Fe(II). Acid extractable Fe(II)aq was lowest for the Fe(O)
treatments with Fe(O) at 0.07 (0.01) and F e(0) + FeSO4 below the detection limit (Figure
A19). The other treatments were substantially higher in F e(II) aq mM (SD): None 1.17
(0.38), formate 0.96 (0.16), FeSO4 1.27, (0), F eSO4 + formate 0.54, (0.05).

PCB analysis. Transformation of PCBs are commonly assessed by comparing the average
meta + para and ortho chlorines per biphenyl to the same in the starting PCB mixture, in
this case Arochlor 1242. After 16 weeks there was no change in ortho chlorines in any
treatment with all remaining in the range of 1.38 -1.42 chlorines per biphenyl (Figure
AI.10). Like the autoclaved control, no transformation of meta + para chlorines was
observed in either Fe(O) or Fe(O) + F eSO4 treatments with the average meta + para
chlorines per biphenyl remain at 1.79 (0.01 SD) for each treatment. FeSO4 and FeSO4 +
formate showed more dechlorination at the meta and para positions with 1.27 (0.07 SD)
and 1.38 (0.29 SD) respectively. Dechlorination in the formate treatment was 1.03 (0.55
SD) meta + para chlorines per biphenyl. The greatest dechlorination was detected in the

no amendment treatment with average meta + para chlorines per biphenyl of 0.87 (0.16

so).

148

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osms aﬁer 16 weeks of incubation.

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150

DISCUSSION

Two hypotheses were tested in this study. (1) That Fe(0) amendment would
enhance microbial PCB dechlorination and (2) that a combined Fe(0) + F e804 treatment
would have synergistic effects ﬁsrther stimulating microbial PCB dechlorination. The
experiments presented here clearly refute both of these hypotheses, as F e(0) and Fe(0) +
FeSO4 both showed no PCB dechlorination after 16 weeks. In fact the most
dechlorination was observed in the no amendment treatments, suggesting that the Fe(0)
addition inhibited microbial PCB dechlorination. Although no F e(0)-enhanced PCB
dechlorination was found in this study, insights into F e(0) dechlorination and microbial
PCB dechlorination were gained.

First, the high loading of Fe(0) used here exceeded the buffering capacity of the
system upon OH' release from anaerobic corrosion. By 16 weeks, the pH in F e(0)
amended microcosms reached 8.75. This is likely beyond the optima for PCB
dechlorinators, which appear to already be slow growers.

Second, despite the high iron loading, there was no abiotic PCB dechlorination. A
high fraction of Fe(0) was added to the system to ensure abiotic dechlorination, since
these sediments were known to support microbial dechlorination. There are two likely
reasons for the absence of F e(0) mediated dechlorination in the sediment microcosms
even though substantial activity was reported for sediment-free abiotic systems. (1) Low
solubility of PCBs and their subsequent partitioning in to organic matter may have
greatly reduced the PCB sorption to catalytic sites on the iron surface. (2) The Fe(0)

surface may be vulnerable to passivation by components in the sediment resulting in a

151

 

 

corrosion rate faster than that for dechlorination (but slower than for hydrogen
production).

To minimize these potential limits to abiotic dechlorination, use of a bimetallic
form of Fe(0) such as Pd coated iron, which has been shown to readily dechlorinate PCBs
[10] may resist corrosion from sediments. Studies with surfactants may show increased
PCB reactivity as was seen in sediment- free system with Fe(0) and DDT [8].

In general the analysis of hydrogen and volatile fatty acids in these long term
incubations showed that electron donors were extremely limited in all treatments not
amended with F e(0). Even addition of 10 mM formate did not sustain hydrogen levels for
more than a few weeks. Since reductive dechlorination is an electron consuming process,
electron donors are likely limiting in this system. Since dechlorinators may not compete
well kinetically with methanogens for high concentrations of electron donor (i.e.
hydrogen), sustained low-level ﬂuxes of hydrogen would be desirable to capitalize on the
thermodynamic advantage of reductive dechlorination. In fact, the focus of much
research on the reductive dechlorination of chlorinated solvents in groundwater has been
the development of slow release hydrogen-producing compounds. Such an approach may

prove useﬁIl for PCB remediation also.

152

10.

11.

References

Wiegel, J .and.Q.Wu., Microbial reductive dehalogenation of polychlorinated
biphenyls. FEMS Microbiology Ecology, 2000. 32: 1-15.

Bedard, D.L.and .J.F.Quensen., III, Microbial reductive dechlorination of
polychlorinated biphenyls, in Microbial Transformation and Degradation of
Toxic Organic Chemicals, L.Y.a.C.C. Young, Editor. 1995, Wiley-Liss Division:
New York. p. 127-216.

Tiedje, J.M., et al., Microbial reductive dechlorination of PCBs. Biodegradation,
1993. 4: 231-40.

Alder, A.C., M. M. Haggblom, S. Oppenheimer, and L. Y. Young, Reductive
dechlorination of polychlorinated biphenyls in anaerobic sediments. Environ. Sci.
Technol., 1993. 27: 530-538.

Sokol, R.C., C. M. Bethoney, and G. Y. Rhee, Effect of hydrogen on the pathway
and products of PCB dechlorination. Chemosphere, 1994. 29: 1743-1753.

Zwiemik, M.J., J. F. Quenson III, and S. A. Boyd, F eSO4 Amendements Stimulate
Extensive Anaerobic PCB Dechlorination. Environ. Sci. Technol., 1998. 32:
3360-3365.

Tratnyek, P.G., Remediation with iron. 1998.

Sayles, G.D., G. You, M. Wang, and M.J. Kupferle, DDT, DDD, and DDE
Dechlorination by Zero- Valent Iron. Environ. Sci. Technol., 1997. 31: 3448-3454.

Yak, H.K., B. W. Wenclawiak, 1. F. Cheng, J. G. Doyle, and C. M. Wai,
Reductive Dechlorination of Polychlorinated Biphenyls by Zerovalent Iron in
Subcrtical Water. Environ. Sci. Technol., 1999.33: 1307-1310.

Wang, CB. and W.X. Zhang, Synthesizing Nanoscale Iron Particles for Rapid
and Complete Dechlorination of TCE and PCBs. Environ. Sci. Technol., 1997.
31: 2154-2156.

Ldfﬂer, F.E., J .M. Tiedje, and RA. Sanford, Fraction of electrons consumed in

electron acceptor reduction and hydrogen thresholds as indicators of
halorespiratory physiology. Appl Environ Microbiol, 1999. 65: 4049-56.

153

APPENDIX II
MICROBIAL DEHALORESPIRATION WITH 1,1,l-TRICHLOROETHANE

Sun, B., B.M. Grifﬁn, H. Ayala-del-Rio, S.A. Hashsham, and J.M. Tiedje. 2002.
Science. 298: 1023-1025.

The primary author of the work presented in Appendix II is Dr. Baolin Sun. As second
author, I guided the design and execution of the enrichments cultures, isolation
procedures, and analyses of chlorinated compounds. I participated directly in the
preparation of the manuscript.

154

 

Microbial Dehalorespiration
with 1,1,1-Trichloroethane ~

Baolin Sun,1 Benjamin M. Griffin,"2 Hector 1.. Ayala-del-Rio,”
Syed A. Hashsham,” james M. Tiedjel-z"

1,1,1-Trichloroethane (TCA) is a ubiquitous environmental pollutant because of
its widespread use as an industrial solvent, its improper disposal, and its
substantial emission to the atmosphere. We report the isolation of an anaerobic
bacterium. strain TCA1, that reductively dechlorinates TCA to 1,1-dichloro-
ethane and Chloroethane. Strain TCA1 required HZ as an electron donor and
TCA as an electron acceptor for growth, indicating that dechlorination is a
respiratory process. Phylogenetic analysis indicated that strain TCA1 is
related to gram-positive bacteria with low DNA G+C content and that its
closest relative is Dehalobacter restrictus, an obligate Hz-oxidizing, chlo-

roethene-respiring bacterium.

TCA is a synthetic organic solvent widely used
in industrial processes and is a major environ-
mental pollutant commonly found in soil (1),
groundwater (2), and the atmosphere (3 ). TCA
is present in at least 696 of the 1430 National
Priorities List sites identiﬁed by the US. Envi-
ronmental Protection Agency (EPA) (I). Be-
cause of TCA’s adverse effects on human
health, the EPA has set a maximum contami-
nant level of 200 ug/ liter in drinking water (4 ).
TCA is also listed as an ozone-depleting sub-
stance by the United Nations Environment Pro-
gramme (5). Even when released to soil or
leached to groundwater. the primary environ-
mental fate of TCA is volatilization to the
atmosphere, where it interacts with ozone and
contributes to the erosion of the ozone layer
(I. 5). TCA is often a co-contaminant in

 

‘Center for Microbial Ecology, 2 Department of Micro-
biology and Molecular Genetics, 3Department of Civil
and Environmental Engineering, Michigan State Uni-
versity, East Lansing, MI 48824—1325. USA.

‘To whom correspondence should be addressed. E-
mail: tiedjej©msu.edu

aquifers with chlorinated ethenes, especially
tetrachloroethene (PCB) and trichloroethene
(TCE), because they have similar industrial
uses. Although in situ bioremediation pro-
cesses for the chloroethenes are known (6),
TCA remediation remains problematic and
can prevent site restoration.

TCA undergoes slow abiotic degradation
to acetic acid and 1,1-dichloroethene. an EPA
priority pollutant (7). Biotransforrnation of
TCA has been observed under aerobic and
anaerobic conditions only in cometabolic
processes (8—11). A growth-linked, or deha-
lorespiratory, process would be more effec-
tive for in situ bioremediation of TCA-
contaminated sites, because reaction rates
would be faster and natural selection would
ensure growth in situ.

Although bacterial growth by dehalorespi-
ration of chloroethenes, chlorobenzenes,
3-chlorobenzoate. and 2-chlorophenol has
been well documented (12—15), bacterial
growth by reductive dechlorination of TCA
has not been reported until now. We describe
the isolation of a bacterium capable of energy

155

 

conservation for growth through the reduc-
tive dechlorination of TCA.

Isolation was initiated from a sediment mi-
crocosm that reductively dechlorinated TCA
(16). Single colonies were subcultured and de-
chlorination activity was maintained in deep
agarose shake cultures until a pure culture was
obtained. The isolate is designated strain TCA1
and is a motile. short rod with a diameter of 0.4
to 0.6 pm and a length of l .0 to 2.0 um (Fig. 1).
Cells stain gram-negative. No spores were ob-
served in starved cultures. Although strain
TCA1 did not grow or dechlorinate in the pres-
ence of oxygen, exposure of the active culture
to aerobic conditions for up to 3 days did not
result in the loss of dechlorinating activity ([6).
Dechlorination of TCA was sequential with the
accumulation of 1,1-dichloroethane (DCA) be-
fore conversion to Chloroethane (CA) (Fig. 2).
Dechlorination of TCA occurred at a faster rate
than that of DCA. The temperature range for

 

Fig. 1. Scanning electron micrograph of strain
TCA1. The rod-shaped morphology and dividing
cells are shown. Scale bar, 1 pm.

156

reductive dechlorination was from 12° to 30°C
with the optimum at 25°C. No dechlorination
occurred at 37°C. Growth yield from reductive
dechlorination was 5.60 i 1.26 g (dry weight)
(mean : SD, n = 3 cultures) of cells per mole
of chloride released. TCA, H2, and acetate were
essential for growth of strain TCA1. Formate
could replace H2 as an electron donor, which
resulted in a similar dechlorination rate. Acetate
alone did not support TCA dechlorination, sug-
gesting that it was used only as a carbon source
and not as an electron donor for reductive de-
chlorination. Because H2 oxidation does not
support substrate-level phosphorylation, growth
in a deﬁned medium with TCA as the electron
acceptor and H2 as the electron donor indicated
that strain TC Al conserves energy in a respira-
tory process.

Strain TCA1 did not dechlorinate 1.1.1.2-

tetrachloroethane, 1,1,2-trichloroethane, l ,2-
dichloroethane, 1,2-dichloropropane, PCE, or
TCE when they were added as potential elec-
tron acceptors. No growth occurred in liquid
media amended with 4 mM H2 as an electron
donor, 5 mM acetate as a carbon source, and
either 5 mM sulfate, sulﬁte, thiosulfate, ni-
trate, or fumarate as an electron acceptor.
Strain TCA1 did not use lactate, pyruvate,
propionate, fumarate, butyrate, benzoate.
phenol, glucose, ethanol. or methanol as elec-
tron donors. No fermentative growth was
observed. Sulﬁte or thiosulfate at 5 mM con-
centration completely inhibited TCA dechlo-
rination, whereas 5 mM sulfate, nitrate, or
fumarate had no effect. Because these com-
pounds did not serve as electron acceptors for
strain TCA1, inhibition was not due to com-
petition for electron donors. Instead, the re-

 

°°°l

§

Chloroethane: (pM)
N
8 §

100-

 

 

Fig. 2. Stoichiometry of
the dechlorination of
TCA to DCA and CA by
strain TCAI. Samples
were incubated at 25°C
and analyzed every 3
days over a 2-month
period for depletion of
TCA and production of
DCA and CA. Error bars
(shown if larger than
the symbols) represent
standard deviations of
triplicate cultures.

 

 

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158

active sulfur oxyanions may directly inhibit
the components of the electron transport
chain or even the reductive dehalogenase, as
observed in Desulfomonile tiedjei (I 7).

Comparative analysis of the nearly full-
length 165 ribosomal DNA (rDNA) sequence
of strain TCA] and available 165 rDNA se-
quences revealed that strain TCA1 is related
to gram-positive bacteria with low G+C con-
tent in their DNA. Strain TCA1, Deha-
lobacter restrictus (l8, l9), and three clones
from trichlorobenzene- and l,2-dichloropro-
pane—dechlorinating consortia (20, 2]) form
a phylogenetic cluster (Fig. 3) deﬁned by 165
rDNA sequence similarities of 97% and high-
er. The closest relative of strain TCA1 is D.
restrictus strain TEA, with a sequence simi-
larity of 99%. D. restrictus strains PER-K23
and TEA are strict anaerobes capable of cou-
pling PCB and TCE dechlorination to H2
oxidation for growth in a respiratory process.
However, we did not detect TCA dechlorina-
tion by strain PER-K23. These bacterial
strains seem to be obligate ill-oxidizing de-
chlorinators that only grow by reductive de-
chlorination of speciﬁc chlorinated ethanes or
ethenes in anaerobic respiration. The physi-
ology. morphology, and 16S rDNA sequence
of strain TCA1 suggest that it is a Deha-
lobacter and perhaps represents a new spe-
cies based on its unique features of TCA
dechlorination and formate oxidation. The
isolation of strain TCA1 further suggests an
important role for Dehalobacter species in
polluted anoxic environments.

To determine the potential of strain TCA1
to attenuate TCA in the natural environment,
we bioaugmented anoxic aquifer sediments
from Bachman and Schoolcraft (both in
Michigan, USA) contaminant plumes. Both
sites are contaminated with PCE and daugh-
ter products. and the Schoolcraft plume G site
is also contaminated with TCA, DCA, and
chromium. TCA was completely converted to
CA within 2 months in both aquifer sediment
samples amended with strain TCA1, whereas
no dechlorination was observed in samples
without the inoculum. These results suggest
that bioaugmentation with strain TCA1 could
ensure and speed the degradation of TCA,
especially if naturally occurring populations
are patchy or absent.

CA, rather than ethane. appears to be the
terminal TCA product from our culture, and
studies have shown that both DCA and CA
can be degraded under aerobic conditions
(22. 23). Because the aerobic transformation
of DCA is much slower than that of CA (24 ),
complete conversion to CA would result in
more reliable removal of chloroethanes on
the aerobic fringes of a plume.

The discovery of an anaerobic dehalore-
spiring Dehalobacter that couples reductive
dechlorination of TCA to growth not only
may lead to a better understanding of the

159

 

physiology, phylogeny, and biochemistry of
dchalorespiring bacteria, but also suggests a
strategy for bioremediation of TC A in soils
and ground water, thereby aiding in the at-
tenuation of this ozone-depleting compound.

LII-#01

13.

14.

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Supported by the Michigan Department of Environ-
mental Quality and the Center for Microbial Ecology
at Michigan State University.

Supporting Online Material
www.5ciencemagorg/cgiIcontent/iull/298/ 5 595/102 3/
DC1

Materials and Methods

4 June 2002; accepted 19 August 2002

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