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This is to certify that the
dissertation entitled

DEVELOPMENT AND CHARACTERIZATION
OF A RAPID DECHLORINATING ENRICHMENT
STIMULATED FROM AN AQUIFER
CONTAMINATED WITH CHLORINATED
ETHENES AND ETHANES

presented by
HAEKYUNG KIM

has been accepted towards fulfillment
of the requirements for the

 

 

 

 

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Engineering
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Date

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DEVELOPMENT AND CHARACTERIZATION OF A RAPID DECHLORINATING
ENRICHMENT STIMULATED FROM AN AQUEER CONTAMINATED WITH
CHLORINATED ETHENES AND ETHANES

By

Haekyung Kim

A DISSERTATION

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

DOCTOR OF PHILOSOPHY

Department of Civil and Environmental Engineering

2004

ABSTRACT

DEVELOPMENT AND CHARACTERIZATION OF A RAPID
DECHLORINATING EN RICHMENT STINIULATED FROM AN AQUIFER
CONTAMINATED WITH CHLORINATED ETHENES AND ETHANES

By

Haekyung Kim

Biostimulation has become an accepted potential remedial alternative for treating
contaminated aquifers. This study evaluates a novel approach to enrich for rapid
dechlorinating microbes by pulse feedings in a batch system over an extended period. It
was demonstrated that by varying the electron donor to acceptor ratio during pulse feedings
of chlorinated compounds, two different types of activity could be developed. Cometabolic
dechlorination was dominant under high lactate conditions, and dehalorespiration activity
was enriched under low lactate conditions. The high lactate/PCB enrichments exhibited
slow and incomplete dechlorination of PCB with cDCE as the major degradation product.
The low lactate/PCB enrichments completely dechlorinated PCE to ethene, and further
reduction of ethene to ethane was observed. Dehalococcoides 16S rRNA gene targeted
real —time PCR confirmed that two orders of magnitude higher amounts of
Dehalococcoides DNA was present in low lactate enrichments compared to high lactate
enrichments. This demonstrated that pulse feedings with low-level electron donor and a

chlorinated ethenes (at a 1:10 electron acceptor/donor ratio) favors stimulation of

halorespirers. Using metabolic inhibitors revealed that the dechlorination observed in low
lactate enrichments was not impacted by molybdate (inhibitor of sulfate-reducer), but was
inhibited by high levels of sulfate, indicating sulfate reducers are not involved in the
observed complete and rapid dechlorination. Low levels of BES (methanogen inhibitors)
specifically inhibited methane productions, but the cultures retained dechlorination activity,
indicating that the responsible microorganism(s) is (are) non-methanogenic. High levels of
vancomycin (100 mg/L; acetogens inhibitors) successfiilly inhibited acetate production
from lactate. The reduction in acetate production also inhibited dechlorination, flirther
suggesting that acetate-utilizing dechlorinating microbes are responsible for the observed
dechlorination.

The development of methodology for the stimulation of dehalorespiring
microorganisms presented in this dissertation represents a potential strategy for the
design of successful biostirnulation systems for chlorinated solvents. It also has

significant design implications for in—situ field systems.

Copyright by
HAEKYUNG KIM
2004

DEDICATION

This dissertation is dedicated to my family and friends: my parents, NamYeun
Chang and KiTae Kim; my sister and brother, SoAh Kim and ByungSoo Kim; and my
biggest supporters, Y.A, TL, and AJN. No matter where I go in life, you all are in my

heart and prayers, now and for always. I could not have done this without you.

ACKNOWLEDGEMENTS

I would like to express my appreciation to my advisor, Michael Dybas, for his
encouragement and support. I am also grateful for my committee, Syed Hashsham, James
Tiedje, and David Wiggert for their comments and guidance during this process. Also to,
Davis Mckenzie, my mentor, for providing unconditional support and advise whenever I

needed.

Special thanks are extended to Lesile Dybas, for introducing me to the DNA extraction,

Yan Pan and Xianda Zhao for answering every question I tossed at them.

This work was fiinded by MDEQ Contract. Y403 86.

TABLE OF CONTENTS

LIST OF TABLES ........................................................................................................ x

LIST OF FIGURES .................................................................................................... xiii

LIST OF ABBREVIATIONS ..................................................................................... xvi

CHAPTER 1

PROBLEM STATEMENT AND OBJECTIVES ............................................................ l
1.1 Hypothesis and research objective .................................................................... 2
1.2 Literature cited ................................................................................................. 4

CHAPTER 2

BACKGROUND AND THEORY ................................................................................... 5
2.1 Energetic Considerations ................................................................................. 5
2.2 Microbial Competition for Hydrogen ................................................................ 7
2.3 Design of Experiments ................................................................................... 11
2.4 Literature cited .............................................................................................. 13

CHAPTER 3

LITERATURE REVIEW ............................................................................................. 15
3.1 Chlorinated Solvents ...................................................................................... 15
3.2 Biotransformation of Chlorinated Ethenes ..................................................... 17
3.3 Mechanism of Chloroethenes Transformation ............................................... 17
3.4 Kinetics of Dechlorination ............................................................................. 20
3.5 Electron Donors for Reductive Dechlorination ............................................... 22
3.6 Methanogens vs. Sulfate Reducers ................................................................. 23
3.7 Abiotic Degradation of Chlorinated Compounds ............................................ 25
3.8 Literature cited .............................................................................................. 37

CHAPTER 4

MATERIALS AND METHODS ................................................................................. 43
4.1 Chemicals and Reagents ............................................................................... 43
4.2 Description of Site and Sediments ................................................................. 43
4.3 Microcosms .................................................................................................. 45
4.4 Analysis of Volatile Organics ........................................................................ 48
4.5 Analysis of Volatile Fatty Acids .................................................................... 49

vii

4.6 Anion Analysis ............................................................................................. 50

4.7 Hydrogen Analysis ........................................................................................ 50
4.8 Hydrogen Sulfide Analysis ............................................................................ 51
4.9 Data Reduction ............................................................................................ 51
4.10 DNA Analysis .............................................................................................. 53
4.11 Literature cited ............................................................................................. 56

CHAPTER 5

EVALUATION OF EFFECT OF ELECTRON DONORS ON

DECHLORINATION ................................................................................................... 58
5.1 Abstract ........................................................................................................ 58
5.2 Introduction .................................................................................................. 59
5.3 Materials and Methods ................................................................................... 62
5.4 Results ........................................................................................................... 65
5.5 Conclusions and Discussions ......................................................................... 78
5.6 Literature cited .............................................................................................. 82

CHAPTER 6

STIMULATION OF HIGH-RATE COMPLETE
DECHLORINATION ACTIVITY BY PULSE FEEDINGS OF

CHLOROETHENES ..................................................................................................... 85
6.1 Abstract ........................................................................................................ 85
6.2 Introduction .................................................................................................. 86
6.3 Materials and Methods ................................................................................... 88
6.4 Results ............................................................................................................ 92
6.5 Conclusions and Discussions ....................................................................... 116
6.6 Literature cited ............................................................................................ 120

CHAPTER 7

CHARACTERIZATION OF DECHLORINATING

COMMUNITY BY METABOLIC INHIBITORS ..................................................... 122
7.1 Abstract ...................................................................................................... 122
7.2 Introduction ................................................................................................ 123
7.3 Materials and Methods ................................................................................. 125
7.4 Results .......................................................................................................... 128
7.5 Conclusions and Discussions ....................................................................... 145
7.6 Literature cited ............................................................................................ 149

CHAPTER 8

ENGINEERING APPLICATIONS ............................................................................ 151

8.1 Enrichment of Rapid and Complete
Dechlorinating Activity ............................................................................... 151

viii

8.2 Field Application ............................................................................................ 154

8.3 Literature cited ................................................................................................. 160
CHAPTER 9
CONCLUSIONS AND FUTURE RESEARCH ........................................................... 161
9.1 Conclusions ..................................................................................................... 161
9.2 Future Research .............................................................................................. 162

APPENDIX A — ELECTRON DONOR AMOUNT

CALCULATIONS .............................................................................. l 64
APPENDIX B — EFFECT OF ADDED SULFATE ON THE

POWDER X-RAY DIFFRACTION PATTERN ................................. 168
APPENDIX C — DESIGN OF EXPERIMENT ............................................................. 169

APPENDIX D -— ADDITIONS AND CONSUMPTION FOR THE
STIMULATIONS OF ENRICHMENTS FOR

EXPERIMENT II ................................................................................ 172
APPENDIX E — T-RFLP PATTERNS FROM CHAPTER 6 ...................................... 174
APPENDIX F — T-RFLP PATTERNS FROM CHAPTER 7 ....................................... 176

LIST OF TABLES

Table 2.1. Reduction potentials of redox couples found in
anaerobic dechlorinating environments ........................................................ 6

Table 2.2. Maximum free energies (AG°’) for electron donors
commonly used in dechlorination of chlorinated
compounds ................................................................................................... 7

Table 2.3. Values of half-velocity constants K8 and maximum
specific utilization rate k with respect to hydrogen, for

hydrogenotrophic PCE dechlorinators and methanogens ................................ 9
Table 2.4. Threshold H2 concentrations and energetics of H2

utilizing Reactions ...................................................................................... 10
Table 2.5. Sign table for 2k factorial design ................................................................... 12
Table 3.1. Properties of common chlorinated organic compounds ................................ 16

Table 3.2. Rate coefficients for various mixed and pure cultures in
dechlorination ............................................................................................. 21

Table 3.3. Summary of research efforts regarding the degradation
of chlorinated ethenes and ethanes in biotic and abiotic

conditions. ................................................................................................ 33
Table 4.1. Typical contaminants found from the study area ......................................... 44
Table 4.2. Typical chemical composition of G plume site ........................................... 45
Table 4.3. Basal medium for microcosm ...................................................................... 46

Table 5.1. Effect of substrates on transformation of Chloroethenes
afier 240 days of incubation in sulfate-reducing
conditions .................................................................................................... 69

Table 5.2. Reduction of TCE depending on presence of reductant
after 180 days of incubation at 15°C ............................................................ 72

Table 5.3. The byproducts and mass recoveries for c-DCE and

Table 5.4.

Table 6.1.

Table 6.2.

Table 6.3.

Table 6.4.

Table 6.5.

Table 6.6.

Table 6.7.

Table 6.8.

Table 7.1.

Table 7.2.

Table 7.3.

Table 7 .4.

TCE transformation after 180 days ............................................................... 74

Estimated transformation rate and half-time of TCE and
c-DCE biodegradation in Plume G .............................................................. 77

Experiment arrangements for respective conditions.

Microcosms were exposed to pulse feedings of PCB

over 1.5 years for Experiment I and pulse feedings of

TCE, c-DCE, and TCA over 1 year for Experiment 11. ................................ 90

Effects of prolonged exposure of high versus low
lactate on enhancing transformation activity (after 12
month of incubation). ................................................................................ 93

Electron balance for a single, PCE degrading
microcosm (15 umol PCE /bottle, 12.5 mM (2500
umol) lactate; data obtained at day 10) . ..................................................... 95

Electron balance for a single, PCE degrading
microcosm (15.2 umol PCE /bottle, 0.5 mM, lactate;
data obtained at day 10) ............................................................................. 99

Summary of dechlorination rate observed by low
lactate/PCB, low lactate/T CE, and low lactate/c-DCE
enrichments (/day). ................................................................................... 105

Effects of prolonged exposure of TCA on stimulation of
transformation activity . ............................................................................ 107

Mass balance determined for PCE transformation to
ethane after 60 days .................................................................................. 109

Quantitative estimation on total DNA and
Dehalococcoides populations in pre-enrichment,
Low/PCB, High/PCB, c-DCE, TCE, and TCA enriched

bottles using PicoGreen and RTm-PCR ..................................................... l 15
The list of inhibitors and their concentrations used in

this chapter ................................................................................................ 127
Effect of inhibitors on reductive dechlorination of PCB

after 18 days .............................................................................................. 138
Influences of inhibitors on electron balance after 7 days ............................ 139
Quantitative Estimation of Total DNA and

xi

Dehalococcoides populations in the BES, molybdate,
sulfate, and vancomycin impacted microcosms using
PicoGreen and RTm-PCR ......................................................................... 144

Table 8.1. Enhancement of first order rate k (1/day) on
disappearance of parent products by pulse feedings over
time ........................................................................................................... 152

Table 8.2. Comparisons of maximum rate between the pulse fed
and once-fed enrichments in enhancing dechlorination

activity observed fi'om each condition. .................................................... 154

Table 8.3. Levels of Dehalococcoides DNA in the biostirnuation
and bioaugmentation test sites ................................................................... 157

xii

LIST OF FIGURES

Figure 3.1. Pathway of microbial PCE transformation under
anaerobic conditions ................................................................................... 18

Figure 4.1. Location of study area ................................................................................ 47

Figure 5.]. Powder X-ray diffraction pattern by
experimentally-prepared FeS .................................................................... 64

Figure 5.2. Differences in (a) slow vs (b) rapid T CE
transformation due to electron donors during 240 days. ............................. 66

Figure 5.3. The differences in dechlorination of TCE after 240
days expressed as percentage removal of abiotic soil
control ......................................................................................................... 67

Figure 5.4. The differences in dechlorination of c-DCE after 240
days expressed as percent removal of abiotic soil
control. ...................................................................................................... 68

Figure 5.5. Anion degradation in CMB fed microcosm ................................................ 70

Figure 5.6. Effect of sulfate on transformation of TCE in CMB fed
microcosm. ................................................................................................. 71

Figure 5.7. Powder x-ray diffraction pattern by iron sulfide
produced in microcosm. ............................................................................. 75

Figure 5.8. Possible pathways for the reductive dechlorination of
chlorinated ethenes by iron sulfide ............................................................ 81

Figure 6.1. The overview and sequence of the experiments
performed during enrichments .................................................................. 91

Figure 6.2. PCB additions and consumption by high lactate/PCB
treatment over 1.5 years ........................................................................... 92

Figure 6.3. Effects of prolonged exposure of high lactate (12.5

leI) on enhancing transformation activity (after 12
month of incubation). ............................................................................. 93

xiii

Figure 6.4.

Figure 6.5.

Figure 6.6.

Figure 6.7.

Figure 6.8.

Figure 6.9.

Figure 6.10.

Figure 7.1.

Figure 7 .2.

Figure 7.3.

Figure 7.4.

Figure 7.5.

PCB additions and consumption by low lactate/PCB
treatment during 450 days ..........................................

Transformation of PCB and intermediate productions
from lactate/PCB enrichments received high dosage
of PCB (80 umol). ......................................................

Conversion of chlorinated ethenes to ethene by
lactate/PCB enrichment. This experiment was done
with the microcosms showing highest activity.
Approximately 15 umol of chlorinated ethenes and
0.5 mM of lactate were used: (a) PCE; (b) TCE; (c) c-
DCE; (d) t-DCE; (e) 1,1-DCE; and (f) VC. ...............

Conversion of chlorinated ethanes by low
lactate/TCA enrichment. Individual serum bottles
have received 15 umol of chlorinated ethanes and
lactate of 0.5 mM: (a) TCA and (b) 1,1-DCA. ..........

Temperature effects on PCE transformation on low
lactate/PCB enrichment. Individual serum bottles
have received 6 pmol of PCB and lactate of 0.5 mM.

Final PCE transformation byproducts by low
lactate/PCB enrichments Temperature effects on PCE
transformation on low lactate/PCB enrichment ..........

T-RFLP profiles generated using pre-enrichment, low
lactate/PCB, and high lactate/PCB enrichments using
HhaI. ........................................................................

Chemical Structures of inhibitors used in this study.
The PCE transformation by control. ...........................

The effects of PCB transformation by 0.5 mM 5 mM,
and 50 mM of BES. Microcosms were amended at
0.5 mM, 5 mM, and 50 mM of BES at zero time only

The effects of PCB transformation by 0.5 mM 2 mM,
and 6 mM of molybdate. Microcosms were amended
at 0.5 mM, 2 mM, and 6 mM of molybdate at zero
time only ....................................................................

The effects of PCB transformation by 0.6 mM and 2
mM of sulfate. Microcosms were amended with

xiv

.............................. 97

............................ 101

............................ 102

............................ 106

........................... 108

............................ 110

............................ 112

. ............................ 124

............................ 129

............................ 130

............................ 133

sulfate at time zero only ......................................................................... 134

Figure 7.6. Sulfate reduction and effect of molybdate on residual
sulfate in microcosms. The reduction of high sulfate
(2 mM) is shown on the bigger scale. ................................................... 135

Figure 7.7. The effects of PCB transformation by 25 mg/L and
100 mg/L of vancomycin. Microcosms were amended
with vancomycin at time zero only ......................................................... 137

Figure 7.8. T-RFLP profiles generated by using dechlorinating
enrichments treated with inhibitors. Each fragment
size indicates percent fragment area: (a) BES; (b)
molybdate; (c) sulfate; and (d) vancomycin. .......................................... 141

Figure 7.9. Model based on DiStefano et al (2), for carbon and
electron flow in a lactate-PCB anaerobic enrichment .............................. 148

Figure 8.1. A Schematic of the relationship between fermentators
and Dehalococcoides ............................................................................. 153

Figure 8.2 Layout of the tracer injection system ....................................................... 158
Figure 8.3. Comparison on the observed reduction in chlorinated

compounds in the biostimuation and bioaugmentation
test sites ................................................................................................. 159

ATP
BES

CA
c-DCE
t-DCE

1, l-DCE
1, l, l-TCA
1 , l-DCA
1 ,2-DCA
DNA
EPA
PCE
PCR
rRNA
TCE
TOC
T-RFLP
VC

VOC

LIST OF ABBREVIATIONS

Adenosine Triphosphate
2-Bromoethanesulfonic acid
Chloroethane

cis-l, 2 Dichloroethene

trans-1,2 Dichloroethene
Dichloroethene

1, 1, l-Trichloroethane

1 , l-Dichloroethane

1 ,2-Dichloroethane
Deoxyribonucleic acid
Environmental Protection Agency
Perchloroethene

Polymerase chain reaction
Ribosomal Ribonucleic acid
Trichloroethene

Total Organic Carbon
Terminal-Restriction Fragment Length Polymorphism
Vinyl Chloride

Volatile Organic Compound

Chapter One: PROBLEM STATEMENT AND OBJECTIVES

Due to past agricultural practices, accidental chemical spills, and inappropriate
disposal of industrial and commercial wastes, groundwater contamination is a significant
national and international problem. In the United States, for example, groundwater
contamination can be found in 1200 Superfund sites and an estimated 7000 Department
of Defense sites (EPA, 1993). The most ubiquitous class of organic compounds polluting
groundwater is the chlorinated solvents (EPA, 1998). Traditional approaches to water
containing these contaminants have relied on groundwater extraction, followed by a
physical/chemical process (e. g., air stripping and carbon adsorption). This approach has
some major disadvantages such as high cost, inefficiencies in removing contaminants
sorbed to the aquifer material, and requirement of pumping, treating, and ultimately
disposal of large volumes of water. Accordingly, there has been a great deal of interest in
in-situ alternative treatment strategies, many of which incorporate biological remediation
of contaminants.

The three basic bioremediation approaches are natural attenuation, biostirnulation,
and bioaugmentation. Natural attenuation refers to unenhanced “natural” processes,
which may occur at contaminated sites if sufficient nutrients and degrading microbes are
present. This can include both abiotic and biotic mechanisms. For chlorinated organics,
natural attenuation may be insufficient, and electron donor addition is often needed to
accelerate reductive dechlorination or stimulate cometabolic activities. Biostimulation

refers to enhancing the metabolic activity of indigenous microflora to transform target

compounds and usually requires amendments (electron donors and/or acceptors) to
enhance the microbial populations already present.

In some cases, indigenous bacteria may be unable to mediate the desired
transformation (Mayotte et al. 1996; Dybas et al. 1998; Ellis et al. 1999) and specific
microorganisms with the desired traits need to be provided (bioaugmentation). Although
bioaugmentation offers the promise of increased control over transformation of a specific
compound, competition between indigenous microflora and the introduced organisms
usually presents a challenge. Factors such as the nutritional requirements of
dehalogenating microbes and electron donor competition between dechlorinating and
non-dechlorinating microorganisms have presented challenges to implementation of this
promising technology in the field. The research presented in this thesis provides insight
into the issue of selective biostimulation: that is, how to enrich microbial populations that

are capable of complete and rapid degradation of contaminant.

1.1. Hypothesis and objectives
The hypotheses of this research are:
l. Palmitate, a slow releasing carbon source, will enhance the rate of dechlorination
compared to simple, readily available carbon sources.
2. Relative abundance of electron donor with respect to PCE will influence the level
of dehalogenation populations such as Dehalococcoides ethenogenes, will impact
rates, and will determine the extent of transformation. A lower abundance of

electron donor with respect to PCE will produce a higher yield of

Dehalococcoides, which results in higher degradation rates and complete

transformation.

The specific objectives were:

1. Evaluate the effect of various types of nutrient on stimulation of dechlorination.

2. Evaluate the effect of various chlorinated compounds and 2 different levels of
lactate on stimulation of dechlorinating enrichments.

3. Characterize the mixed dechlorinating community by the inhibitors and hydrogen
effects on dechlorination and growth.

4. Characterize the mixed dechlorinating community by using the molecular
techniques such as RTm-PCR and T-RF LP.

5. Evaluate the abiotic degradation of chlorinated compounds.

The format of this thesis is as follows: theory supporting the aforementioned hypotheses
are outlined in Chapter 2; Chapter 3 contains review of related researches; materials and
methods are illustrated in Chapter 4; results from the screening of the various electron
donors and byproduct analysis presented in Chapter 5; development strategies and
complete dechlorination by enriched sediments are presented in Chapter 6; and effects of
metabolic inhibitors on dechlorinating communities in Chapter 7; engineering

applications in Chapter 8; conclusions and future research are summarized in Chapter 9.

1.2. Literature cited

US. Environmental Protection Agency. 1993. Cleaning Up the Nation’s Waste Sites:
Markets and Technology Trends. Office of Solid Waste and Emergency Response,
EPA 542-R—92-012.

US. Environmental Protection Agency. 1998. National Water Quality. Inventory Report
to Congress.

Mayotte, T., M. Dybas, and C. S. Criddle. 1996. Bench-scale evaluation of
bioaugmentation to remediate carbon tetrachloride-contaminated aquifer materials.
Ground Water 34 (2), 358-367.

Dybas, M. 1., Barcelona, M., Bezborodnikov, 8., Davies, S., Fomey, L., Heuer, H.,
Kawka, O., Mayotte, T., Sepulveda-Torres, L., Smalla, K., Sneathen, M., Tiedje, J .,
Voice, T., Wiggert, D. C., Witt, M. E., and C. S. Criddle. 1998. Pilot-Scale Evaluation of
Bioaugmentation for In-Situ Remediation of a Carbon Tetrachloride-Contaminated
Aquifer. Environ. Sci. Technol. 32(22): 3598-3611.

Ellis, D.E., E.J. Lutz, R.J. Buchanan, M.D. Lee, C.L., Bartlett, M. Harkness, and K.
DeWeerd. 1999. Bioaugmentation for accelerated in situ anaerobic bioremediation. In
Platform Abstracts of the Fifth International Symposium on In-Situ and On-Site
Bioremediation, Session D2, San Diego, California, April 19-22. Columbus, Ohio:
Battelle

Chapter Two: BACKGROUND AND THEORY

2.1 Energetic Considerations

Microorganisms that are capable of utilizing the most energy rich electron
acceptor/donor pair (the half reactions exhibiting the most positive electron EH°) will be
dominant in a given environment. These organisms then in turn determine the redox
conditions of the environment. The maximum free energy (AG°') microorganisms may
gain is commonly presented by reducing-oxidation reactions. Table 2.1 contains standard
reducing potentials and free energies (AG°') of redox couples and chlorinated compounds
that are commonly found in contaminated groundwater at 25 °C. Thermodynamic
consideration indicates that reductive dechlorination is exergonic and chlorinated
compounds can be used as electron acceptors. Table 2.2 contains maximum free energies
of selected organic acids and alcohols commonly used as electron donors in
dechlorinating experiments.

McCarty (1969) proposed stoichiometric methodology to evaluate the energy

requirement for cell synthesis from various carbon and nitrogen sources.

AG /k’"+AGc+AGn/k
A=—— P (2.1)
kAG

f

 

where: A = electron equivalent of electron donor to energy/ electron equivalent of cells
synthesized : AGp represents the energy required to reduce carbon source to pyruvate : k
is cell efficiency of energy transfer : AGn represents energy required to convert N source

to NH] : AGr presents energy available energy per electron equivalent of substrate

converted for energy : AGc presents energy of conversion of one electron equivalent of

intermediate to one electron equivalent of cells.

Table 2.1 Reduction Potentials of redox couples found in anaerobic dechlorinating

 

 

environments.
Eo’+ AG°’
Half-Reaction (V) (kJ /mol e)

Fe3+ + e' = Fe2+ 0.77 -114
$0424 9H+ + 8e° = HS' + 4 H20 022 -146.7
C020,) + SPF + 8e' = CH4 (9 + 2 H20 025 -131.2
C2C14 + H*+ 2e‘ = C2HCI3 + Cl' 0.58 -191.0
C2HC13 + H+ + 2e‘ = C2H2C12+ Cl' 0.54 -183.0
C2H2C12 + H+ + 2e‘ = C2H3CI + Cl' 0.37 -151.1
C2H3Cl + I—F+ 2e‘ = C2H4 + Cl’ 0.49 -173.7
C2H4+ 2H+ + 2e' = C2H6 0.1 -98.9
C2H3Cl3 + H+ + 2e‘ = C2H4C12 + Cl' 0.57 -189.1
C2H4C12 + H+ + 2e' = C2H5Cl + Cl' 0.51 -177.5

 

I Redox-potentials were obtained from Criddle et al. (1991), Dolfing and Janssen (1994), and
Vogel et al. (1987).

In heterotrophic reactions, the value of A is by definition feis so that the fraction

of organic substrate utilized for energy is

A

=_ 2.2
A+1 ( )

fe

where: fs is fraction of electron donor synthesized and fe is fraction of electron donors

converted to energy.

Table 2.2. Maximum free energies (AG°’) for electron donors commonly used in

dechlorination of chlorinated compoundsa

 

 

AG°’ kcal
. per mole
Half-Reactions electrons
1/92CH3(CH2)14COO‘ + 31/92 H20 —+ 15/92 C02+1/92 HCO3‘+ IF+ e' -6657
Palmitate
1/12 C3H503' + 1/3 H209 1/6 C02 + 1/12 HCOg' + I-F + e' -7.873
Lactate
1/14 C3H502'+ 5/14 H20 —> 1/7 C02 + 1/14 HC03'+ W + e‘ -6.664
Propionate
1/6 CH40 +1/6 H20 —-> 1/6 C02 + W+e‘ -8.965
Methanol

 

aEnergetics and Bacterial Growth. Perry L. McCarty. Fifth Rudolf Research
Conference, Rutgers. The State University, New Jersey, 1969. July 2.
2.2 Microbial Competition for Hydrogen

The ability of certain microorganisms to out compete others for the same grth
nutrient is dependent upon both the kinetics and energetics of the competing reactions.
When microbial strains compete for the same limiting nutrient, resource-based
competition theory predicts that only one group of organisms will survive and all other
will die out (Hansen et a1, 1980). Experimental results have shown that two species
were observed to coexist only if either (1) each was limited by a different resource and
met the theoretical criteria for coexistence, or (2) the species were limited by the same

resource and did not differ significantly in their resource requirements.

The rate of bacterial growth (rg) and the rate of substrate consumption (rsc) can be
used to determine whether one organism has a kinetic advantage over another. The rate

of bacterial grth is described mathematically by Monod’s equations.

XS
eta-:3- <2.”

Where; um (hr'l) is the organism’s maximum specific growth rate; X represents the
concentration of biomass (mg cells L'l) ;S is the concentration of the growth-limiting
substrate (mgL'l) ; Ks is the half velocity constant which is the substrate concentration at
half the maximum specific growth rate. The rate of substrate consumption, rsc is related

to the rate of bacterial growth by

r. = — (2.4)

The organism’s maximum yield coefficient (Y) represents the maximum mass of
cells that can be produced per mass of substrate consumed. The Y and mm define the
maximum specific substrate utilization rate k (mg substrate consumed mg cells"1 h'l )
according to the relationship k = umax/Y . Combining equation 2.3 into equation 2.4

yield the following expression for rSC

r50 = kXS (2.5)
K, + S

 

Examination of the Monod’s expression reveals that the rates of bacterial growth
and substrate consumption are dependent upon values of Ks, Y, and k. Half-velocity rate
constants represent the affinity of an organism for a substrate, and lower values of K8 are
associated with higher affinities.

Many organisms like acetogens, methanogens, and sulfidogens use hydrogen
produced from anaerobic metabolism of organic matters (Brock et al. 1999). Since
hydrogen is also needed for the hydrogenotrophic organisms (such as hydrogen-utilizing
PCE dechlorinators) to survive, the microbial competition for hydrogen can play an
important role in the transformation of the chlorinated compounds. Values of half-
velocity constants with respect to hydrogen, Ks (Hg), for hydrogenotrophic dechlorinators
and methanogens and measurements of the maximum specific utilization rate reported are

summarized are listed in Table 2.3. For hydrogenotrophic dechlorinators, half-velocity

Table 2.3. Values of half-velocity constants K8 and maximum specific utilization rate k

with respect to hydrogen, for hydrogenotrophic PCE dechlorinators and methanogens.

 

 

K5 (H2) k
Methanogens 2,500 —l3,000 nMa loo-2,600Pmol-mgprotein 'lhr" c
960 i 180 M
1339111011113me 100 i 50 an 20 umol-mgprotein 'lhr'”
9-21 nM°

 

’ Robinson and Tiedje, 1984
" Smatlak et al. 1996
° Ballapragada et al. 1997

constants were found to be one-tenth that of the methanogenic organisms (Smatlak et a1.
1996). The use of fermentable substrates that maintain low levels hydrogen was
suggested to offer a competitive advantage for hydrogen utilizing dechlorinators.

The energetics of competing reactions can also be used to determine whether the
competition will emerge between organisms utilizing the same electron donor. Cord-
Ruwisch et al. (1988) observed that the hydrogen threshold is inversely correlated with
changes in Gibbs free energy (AG°'). As can be seen in Table 2.4, since dechlorination has
lower free energy compared to other processes, dechlorinators are expected to grow at
lower hydrogen partial pressure. Supporting observations had been made in many studies

(Smatlak et al. 1996; Ballapragada et al. 1997).

Table 2.4. Threshold H2 Concentrations and Energetics of H2 Utilizing Reactions
(adopted from Loffler et al. 1999).

 

 

Biological . Threshold H2 AG°’
Equation
process (nM)' (KJ/mol H2)
Dehalogenation PCE + 4H2 —) ethene +I—I+ + Cl’ < 0.3 -130 to -187
Methanogenesis C02 + 4H2 —> CH4 + 2H20 5 — 95 -33.9
Acetogenesis 2CO2 + 4H2 —) CH3COO' + H+ + 2H20 336- 3640 -26.1
Sulfate reduction 8042' + 4H2 +I-I+ —> HS' + H2O 1-15 -3 8.0

 

’ In aqueous phase concentration

10

2.3 Design of Experiments

Significant factors

Many factors contribute to transformation of chemicals and it can be difficult to
separate the contribution of any one factor to the overall degradation. Reported data on
PCB and TCE transformation often does not include consideration of any role of abiotic
mechanisms, assuming transformation occurred only biologically with microorganisms.
To attempt to determine which factors are most significant, a statistical design approach
(factorial design) is adopted which allows separation of significant factors from
insignificant factors and thus the effects of interactions between each factor can be

evaluated (Box et al. 1978).

Design Matrix

The main objective of factorial design is to see the effects of two or more
independent variables (Wadsworth, 1990). It is usually more efficient to manipulate
these variables in one experiment than run a separate experiment for each variable. Also
it is possible to test for interactions among variables using factorial design. 2" factorial
designs need k parameters and helps in sorting out parameters of impact. The method

uses a nonlinear regression model and assuming there are only 2 parameters, x1, x2, gives:

y=q0+q1*XI+QZ*X2+(112*X1*X2 (2.4)

11

Here, y is the system performance metric; qo is the average performance; q,- is the effect
on the performance of parameter x,; q12 is the effect of the interaction of x1 and x2; The

value of x,- is defined as:-1 if x, assumes the minimum value in the experiment, otherwise,

 

 

 

 

 

+1.
Expenment ........... .x.l. x2 xlxé y
. 0-1 -1 1.y.0.
............... 1........... . . 1_1’_'1 yfll
2 -1W1._1“ yz
3 1 1 1. y3

 

 

 

 

 

 

 

Table 2.5. Sign table for 2k factorial designs

In this example, 22 experiments with 4 different combination of x1,x2 are needed. Then,
by substituting x, and y into the above performance equation, 4 equations with 4

unknowns qo, q1, q2, q12 are obtained. Then, we can solve for qo, q1, q2, q12 unknowns.

Analysis of Variance (AN OVA)

In order to further verify the degradation of Chloroethenes from these preliminary
results, collected data will be analyzed using (1) Qalysis gf ygriance (ANOVA), which
tests the difference between 2 or more means by examing the ratio of variability between
two conditions and variability within each condition; and (2) factorial design, which

identifies positive and negative factors on degradation of Chloroethenes. Values of the F

12

statistics and and p-value will be computed to determine the significance of main and

interaction effects.

2.4. Literature cited

McCarty, Perry L. 1969. Energetics and Bacterial Growth. The Fifth Rudolf Research
Conference, Rutgers, New Brunswick, New Jersey

Criddle, CS. and PL. McCarty. 1991. Electrolytic model system for reductive
dehalogenation in aqueous environments. Environ. Sci. T echnol. 25:973 -978

Dolfmg J. and DB. Janssen. 1994. Estimates of Gibbs free energies of formation
of chlorinated aliphatic compounds. Biodegradation 5: 21-28

Vogel, T.M., C.S. Criddle, and PL. McCarty. 1987. Transformation of halogenated
aliphatic compounds. Environ. Sci. T echnol. 21: 722-736.

Hansen, S. R., and S. P. Hubbell. 1980. Single-nutrient microbial competition: qualitative
agreement between experimental and theoretically forecast outcomes. Science. 207: 1491—
1493.

Brock, T. D., Madigan, M. T., Martinko, J. M., and Parker, J. 1999. Biology of
Microorganisms, Prentice-Hall, Inc, Englewood Cliffs.

Smatlak, C. R., Gossett, J. M., and Zinder, S. H. 1996. Comparative kinetics of
hydrogen utilization for reductive dechlorination of tetrachloroethene and
methanogenesis in an anaerobic enrichment culture. Environ. Sci. Techno]. 30 (9): 2850-
2858.

Ballapragada, B. S., Stensel, H. D., Puhakka, J. A., and Ferguson, J. F. 1997. Effect of
hydrogen on reductive dechlorination of chlorinated ethenes. Environ. Sci.
Technol. 31(6): 1728-1734.

Robinson, J .A., and Tiedje, J.M. 1984. Competition between sulfate-reducing and
methanogenic bacterial for H2 under resting and growing conditions. Archives of
Microbiology. 137 : 26-32.

Cord-Ruwisch, R, H.-J. Seitz, and R. Conrad. 1988. The capacity of hydrogenotrophic
anaerobic bacteria to compete for traces of hydrogen depends on the redox potential of the
terminal electron acceptor. Arch. Microbiol. 149:3 50-357 .

Loffler F. E., James M. Tiedje, and Robert A. Sanford. 1999. Appl. Environ. Microbiol.
65(9): 4049-4056

13

Distefano, T. D., Gossett, J. M., and Zinder, S. H. 1992. Hydrogen as an electron-donor
for dechlorination of Tetrachloroethene by an anaerobic mixed culture. App]. Environ.
Microbiol. 58 (11): 3622-3629.

Monod, J. 1949. ”The growth of bacterial cultures.” Annual Review of Microbiology.

Box, GE. 1978. Statistics for Experimenters: An introduction to design, data analysis,
and model building. John Wiley & Sons.

Harrison M. Wadswroth. 1990. Handbook of statistical methods for engineers and scientist.
McCraw-Hill Publishing Company.

14

Chapter Three: LITERATURE REVIEW

This chapter provides previous research efforts on biotransformation of
chlorinated ethenes under anaerobic conditions. The review begins with a general
overview of chlorinated ethenes properties. The remaining review is divided into two
sections. The first contains discussions on mechanisms and transformation kinetics of
chlorinated solvents. A summary of previous research efforts on biotransformation of

chlorinated ethenes follows shortly after.

3.1. Chlorinated Solvents

Due to the similarities in the electrical charges (or electronegativites) of oxygen
and chlorine, chlorinated solvents are considered highly oxidized. Perchloroethylene
(PCE), the most highly chlorinated ethylene molecule, is environmentally persistent
partly because it only degrades anaerobically. PCE degradation proceeds through
trichloroethylene (TCE), the TCE is then reduced to isomers of dichloroethylene.
Although the cis-isomer (c-DCE) appears to be the predominantly produced
dechlorination byproduct, trans-isomer (t-DCE) is also reported to be produced by D.
ethenogenes 195 (Maymo-Gatell et al. 1999). A third isomer, 1, l-dichloroethylene (1,1-
DCE) is chemically possible, however, there is little evidence that it is biologically
produced from TCE dechlorination. The next step of reductive dechlorination is the
production of vinyl chloride (VC) from the dichloroethylene isomers. Removal of the
final chlorine produces the innocuous ethylene molecule making this product the end

point of the reductive dechlorination process, which satisfies the treatment purpose. Note

15

that because of the double bond in the ethylene molecule can be further reduced to

ethane, which contains only a single bond between the carbons with three hydrogens on

each carbon. All of the chlorinated ethylene are listed as priority pollutants under the

Safe Drinking Water Act Amendments of 1986. Typical chemical and physical

properties of these compounds are summarized in Table 3.1.

Table 3.1. Properties of Common Chlorinated Organic Compounds

 

 

Chemical Molecular Density Solubility in Henry’s
Compound Formula Weight (m g/m’)’ water b Constant Hcc
(g/mol) (mM at 25 °C) (dimensronless)
Tricthlrblethane CHCl3 133.41 1350 8.5 0.703
Tetrachloroethene C2Cl4 165.83 1626 0.91 0.723
Trichloroethene C2HC13 13 1.39 1460 8 .4 03”
c-Dichloroethene C2H2Cl2 96.94 1284 3 6 O. 167
Vinyl Chloride C2H3Cl 62.50 91 1 43 1.137
Ethene C2H4 28.04 2085 4.7 8.5b

 

‘ Data taken from Sawyer and McCarty (1994).
bYaws, C. L. (1999)
° Gossett, J. M. (1987) at 24.8 °C

16

3.2. Biotransformation of Chlorinated Ethenes
Reductive Dechlorination

Reductive dechlorination or hydrogenolysis is a term to describe reactions where
sequential electron and hydrogen additions replace chlorine atoms. Chlorinated ethenes
are susceptible to biologically mediated reductive dechlorination reactions under
anaerobic conditions. For example, PCE is sequentially reduced to TCE, c-DCE or t-
DCE, VC, and then to ethene (Figure 3.1). The extent of reaction varies for different
compounds and with differing environmental conditions. Generally more chlorinated
compounds are more susceptible to rapid reductive dechlorination reactions while the
more toxic intermediates such as DCE, VC are less readily degraded (Haston & McCarty,
1999; Vogel et al. 1987). These less chlorinated intermediates are susceptible to
oxidation under aerobic and iron reducing conditions but PCE is only removed by

anaerobically (Freedman and Gossett, 1989).

3.3. Mechanism of Chloroethenes Transformation: Halorespiration vs. Cometabolism

Most of these transformations were explained as cometabolism, which are
fortuitous transformation of chemicals by enzymes or cofactors produced by organisms
for other purposes. During the cometabolic process, microorganisms that are responsible
for catalyzing such reactions do not obtain nutritional or energetic benefits. Researchers
showed that the majority of chlorinated compounds could undergo cometabolic
transformation to less chlorinated products and in some instances complete
dechlorinations to benign products in environments. Reductive dechlorination has been

reported to occur cometabolically under various redox environments including

17

denitrifying, sulfate reducing, fermentating, acetogenic, and methanogenic conditions

(Fetzner & Lingens, 1994; Mohn & Tiedje, 1992).

C1 C1
PCE
Cl L C1
C1 C1
>:.< TCE
Cl H
A/ \
C1 C1 H Cl
H H Cl H
cis-DCE trans-DCE
\ A/
H H

H: :H Ethene
H H
H H
M Ethane
H H
H H

Figure 3.1 Pathway of microbial PCE transformation under anaerobic conditions.

18

However, recent research has indicated that in addition to cometabolic processes,
anaerobic bacteria may degrade chlorinated ethenes by using them as an electron
acceptor, which couples chlorinated ethenes directly to energy production (McCarty,
1997). In the absence of 02, many anaerobes use electron acceptors such as nitrate,
ferric iron, sulfate, protons, and CO2. Reductive dechlorination has been shown to occur
in all of the above mentioned alternative electron accepting environments, although it is
more frequent in the environments with the lower redox potentials (F etzner & Lingens,
1994; Mohn & Tiedje, 1992; Vogel et a]. 1987). Due to similarity in redox potentials
between chlorinated compounds to NO3-/NO2- redox couple, chlorinated compounds
have the thermodynamic potential to serve as terminal electron acceptors in anaerobic
microorganisms. Currently, only a few bacteria have been isolated that use chlorinated
compounds as electron acceptors and conserve energy for growth.

Holliger et al. (1993) have described Dehalobacter restrictus, a highly purified
enrichment culture that is able to grow by the reduction of PCB to c-DCE using hydrogen
as the electron donor. This bacterium was able to grow on a limited number of electron
donors, especially; hydrogen and formate, and could use only PCB and TCE as electron
acceptors with carbon dioxide and yeast extract acting as carbon sources.

Five additional isolates, all capable of utilizing PCE as their terminal electron
acceptors, have since been isolated: Dehalaspirillum multivorans, strain TEA,
Desulfirromonas chloroethenica, Desulfitobacterium sp. strain PCB], and
Dehalococcoides ethenogens strain 195. Dehalosporillium multivorans was able to grow
on a range of electron donors including pyruvate, formate, hydrogen, lactate, ethanol and

glycerol (Neumann et a], 1994). It was further demonstrated that only PCB and fiimarate

l9

could act as electron acceptors for this organism. Strain MS-l has also been shown to
dechlorinate PCE in pure culture, but PCE dependent growth has not been confirmed.
Another organism capable of the incomplete dechlorination of PCB to c-DCE was the
freshwater anaerobe TT4B, which used acetate or pyruvate as electron donor and only
PCE or fumarate as an electron acceptor (Krumholz eta]. 1996).

Of these isolates, only Dehalacoccodes ethenogenes strain 195 catalyzes the
complete dechlorination of PCB to ethene. Dehalococcoides ethenogenes was extracted
from anaerobic digester supernatant and grown on hydrogen and PCB using carbon
dioxide and yeast extract as carbon sources (Maymo-Gatell et al. 1997). It is reported
that the growth of strain 195 was resistant to ampicillin and vancomycin and its cell wall
did not react with a peptidoglycan —specific lectin and its ultra structure resembled S-
layers of Archaea (Maymo-Gatell et al. 1997). Based on its 16S ribosomal DNA
sequence, strain 195 clustered phylogenetically with the eubacteria but did not fall in any
of the previously described eubacterial branches (Maymo-Gatell et a]. 1997). The final
step of VC transformation to ethene was found to occur only after PCE was completely
degraded and cometabolic. No other known microorganisms capable of dechlorination

beyond c-DCE have been reported so far.

3.4. Kinetics of Dechlorination

As reductive dechlorination proceeds, successive steps are increasingly slow and
the conversion of VC to ethene has been observed to be a rate—limiting step in complete
degradation (Freedman & Gossett, 1989). Studies of the kinetics of chlorinated ethene

biodegradation have shown that DCE and VC have higher half velocity coefficients and

20

lower maximum transformation rates than their parent compounds so that their rate of
degradation will be slower, especially at low concentrations (Haston and McCarty, 1999).
The kinetics of dechlorination has been determined in both mixed and pure cultures,
although inconsistencies in the manner the dechlorination rates were determined have
made comparisons between studies difficult. Table 3.2 lists rate coefficients for various

cultures and electron donors used to maintain the dechlorination activity.

Table 3.2. Rate coefficients for various mixed and pure cultures in dechlorination

 

Electron

Rate of dechlorination Reference
Donor (5)

Culture Dechlorination

 

Methanasarcina PCE —) TCE Methanol 3.5 pmol-mgprotein ' Fathepure et
1 .
hr

sp. and acetate al. (1987)
Sulfate-reducing TCE —) c-DCE Sodium 213 p.rnol-L"d‘l Pavlstathis &
enrichment lactate and Zhuang

acetate (1991)
Enrichment PCE —) ethene ' Methanol 275 umol-L'Id'l Distefano et
culture a].
(1991)
Fixed-bed PCE -—> ethane Lactate 3,7 pmolL'Ihr'l DeBruin et al.
column (1992)
Anaerobic C-DCE —) Glucose 4.2 timolg'l VSSd'l Komatsu et al.
sewage sludge — ethene, ethane (1994)
Batch
Aquifer solids PCE —-) TCE Mixture of 0,3 nMd'1 Gibson et al.
from polluted TCE _> c-DCE lactate, (1994)
sediment — Batch acetate and
propionate
Laboratory PCE —) VC Methanol 4.6 i 0.4 pmolg“ Tandoi er al.
culture VSSd" (1994)
Anaerobic TCE —> DCE Glucose 149 ter'1 Wild et al.
Reactor treating (1995)
dechloromethane

 

21

 

Dehalaspirillum PCE—> c-DCE
multivarans

Strain MS-l PCE —) c-DCE

Enriched PCE ——) c-DCE
contaminated soil
sample

Enrichment from c-DCE —)

polluted sediment ethene
- Batch

Pyruvate 50 nmolmin'l
(mg cell protein ‘1)

Glucose, 0.5 umol'mgCell 'lhr'1
pyruvate,

formate,

lactate,

acetate,

yeast
extract,
amino acid

Citrate, 0.4 umol-mg VSS'lhr'1
pyruvate,
succinate,
formate,
and acetate
with
hydrogen

H2 l6uMd"

Scholz et a].
(1995)

Shanna &
McCarty
(1996)

Lee et al.
(1997)

Luijten et al.
(1997)

 

3.5. Electron Donors for Reductive Dechlorination

A wide variety of fatty acids, alcohols, sugars and other compounds have been

successfully demonstrated to support reductive dechlorination, and current researches

focus on finding a particular electron donor that best support mixed dechlorinating

anaerobes. Researchers reported that hydrogen is a key electron donor and organic

compounds are mainly precursors to supply the needed hydrogen via fermentation

(Distefano et a]. 1992; Fennell et al. 1997). However, several studies have suggested the

ability to enhance and sustain in situ dechlorination can be influenced by the selection of

electron donor. Accordingly, many different types of electron donors were screened to

stimulate and sustain dechlorination activity. Some researchers focused on hydrogen at

22

high partial pressures, where methanogens and other hydrogenotrophs may compete with
dechlorinating microorganisms for hydrogen, resulting in incomplete dechlorination
activity or its exclusion entirely (Ballapragada et a]. 1997; Yang & McCarty, 1998).
Meanwhile other researcher reported that the difference in the hydrogen partial pressure

fail to represent long-term dechlorination activity (F ennell et al. 1997).

3.6. Methanogens vs. Sulfate Reducers

Anaerobic conditions encompass a range of redox potentials, depending on the
terminal electron acceptors. Reductive dechlorination has been observed repeatedly in
both natural and laboratory methanogenic environments (F athepure et a]. 1987;
Fathepure & Boyd, 1988; DiStefano et a]. 1992), under nitrate reducing conditions
(Bouwer & McCarty, 1983; Jafvert & Wolfe, 1987), and in sulfate reducing conditions
(Egli et al. 1987; Bagley & Gossett, 1990). Many reports on dechlorination in sulfate
reducing conditions followed (Kohring et a]. 1989; Stevens et al. 1988), but for most
compounds, dechlorination occurred more readily when methanogenesis predominates
(Bouwer & Wright, 1988; Gibson & Suflita, 1986). Therefore, early works on
transformation of chloroethene were dominated in methanogenic mixed cultures or pure
cultures of specific methanogens. However, many cases reported interactions/inhibitions
in dechlorinations between microorganisms. Bacterial such as Methanosarcina sp. and
Methanosarcina mazei could convert PCE to TCE with acetate as an electron donor and
this conversion was found to be faster in mixed cultures and was inhibited when

methanogensis was inhibited (Fathepure et al. 1987).

23

Freedman and Gossett (1989) reported a mixed culture capable of completely
reducing PCE to ethene did experience inhibition when methanogenesis was inhibited,
however, PCE degradation was only temporarily halted and did resume even with
repeated additions of a methanogenesis inhibitor bromoethanesulfonic acid (BES). The
same inhibitor was also found to prevent ethane production, but have no effect on other
steps in reductive dechlorination of PCE to ethene in column studies (deBruin et
al.1992). In other experiments with a methanol enrichment culture, high concentrations
of PCB were reductively dechlorinated to ethene and methanogenic activity decreased as
PCE loading and subsequent dechlorination increased (DiStefano et al. 1991).

While these studies suggested that methanogens are still involved in partial PCE
or TCE reductive dechlorination and may be necessary members of bacterial populations
of complete reductive dechlorination, other works found that PCB and TCE could be
degraded in the complete absence of methanogenesis. Bagley and Gossett (1990),
reported that PCE degradation to TCE and c-DCE was enhanced in sulfate reducing
cultures when methanogenesis was inhibited. There are still no published reports of
complete degradation of PCB to ethene under sulfate reducing conditions in mixed
cultures. But pure cultures of Desulfovibrio desulfirricans are reported to degrade PCE
without any other productions of dichloroethene (Fathepure et a]. 1987). In addition,
sulfidogenic bacterium DCB-l (Shelton & Tiedje, 1984; Stevens et al. 1988) has been
shown to dechlorinate PCE to TCE, and Desulfobacterium autotrophicum reduces PCE
quantitatively to TCE and DCE (Egli et al. 1987), indicating the existence of reductive
dechlorination capabilities in sufidogenic and sulfate-reducing bacteria. The research

efforts on biotransformation of chlorinated ethenes are summarized in Table 3 .3.

24

3.7. Abiotic degradation of chlorinated compounds

The biological reductive dechlorination of PCB, TCE to ethene and ethane has
been demonstrated at a number of sites. However, biological activity often leads to
incomplete or slow degradation of a specific compound. This is generally due to a
deficiency in an electron donor or acceptor, other nutrients required for the dechlorinating
microbial population, and/or absence of appropriate dechlorinating microbial populations.

FeS has been identified as a soil precipitate in sulfate-reducing environments
where it is formed through the biologically mediated reduction of sulfate to sulfide and
subsequent reaction of sulfide with available iron species (Rickard 1969; Freney 1979).
Several forms of ferrous sulfide can be produced; amorphous iron sulfide, makinawite
(F eo,995-1,o23S), greigite (Fe3S4), pyrrohotitie (F es”), and pyrite (F eS2). These soil
minerals are reported to transform halogenated organic compounds. (Magnetite: Sivavec
and Horney 1997; McCormick et al. 1998), pyrite (Kriegman-King & Reinhard, 1991
and 1994). Butler and Hayes (1999) have demonstrated that ferric sulfides and ferric
disulfides such as mackinwite and pyrite can promote the abiotic dechlorination of
Chloroethenes. The major byproduct of was acetylene after 120 days. Lee and Batchelor
(2002) also reported that pyrite degraded TCE to 3.3% aDCE, 43% acetylene, 2.2%
ethene, and 50 % residual TCE after 32 days. The research efforts on abiotic

transformation of chlorinated ethenes are also summarized in Table 3.3.

25

Table 3.3. Summary of Research Efforts regarding the degradation of chlorinated

ethenes and ethanes in biotic and abiotic conditions.

(a) Electron donors and source of inoculums in methanogenic mixed and pure cultures.

 

Source of

Electron Donor (s)

 

Summary Reference
Inoculum and Dechlorination
Methanogenic None, except Degradation of PCB and TCE in anaerobic Bouwer and
mixed culture from methanol used in and aerobic cultures. Both compounds were McCarty
laboratory PCB and TCE recalcitrant under aerobic conditions and (1981)
digester solutions appreciable degradation under anaerobic
conditions was not observed within 16
PCE —> DCE weeks.
TCE —> DCE
Same as above Acetate Under methanogenic conditions, PCB (100 Bouwer and
< ug/L) was reductively dechlorinated to McCarty
PCE —> TCE TCE. Addition of BES causes a reduction (1983)
in acetate utilization, but did not affect the
extent of PCB dechlorination The role of
acetoclasts in PCE dechlorination remained
uncertain
Soil from a TCE Soybean meal With radiolabeled TCE, authors proved that Kleopfer er
contaminated site it reduced to 1, 2-DCE isomers in sail a]. (1985)
PCE —> DCE microcosms.
Laboratory culture Acetate PCE dechlorinate to TCE, DCEs, and VC in Vogel and
fixed film methanogenic columns. 24 % of McCarty
PCE -—> DCE the initial PCE added was reported to be (1985)
mineralized to CO2
Pure cultures of Methanol and Pure cultures of anaerobic bacteria capable F athepure,
acetoclastic acetate 0f dechlorinating PCE were identified. Nengu, and
methanogens DCB-l, a chlorobenzoate fed anaerobe, and Boyd
PCE —> TCE two strains of Methanosarcina dechlorinated (1987)

PCE. DCB-l stoichiometrically converted
PCE to TCE. When DCB-l was mixed
with the two Methanosarcina strains, higher
rates of PCB dechlorination were achieved
and TCE degraded.

 

26

 

Sewer sludge and
acetoclastic
methanogens in
pure culture

Purecultureof

Methanosarcina sp.

Strain DCM

Laboratory
anaerobic digester
seeded with
digested sludge
from NY
wastewater
treatment plant

Aquifer solids

Laboratory culture
used by Freedman
and Gossett

Methanol and
acetate

PCE —) TCE

Methanol, acetate,
methylamine, and
trirnethylamine

PCE —) TCE

Methanol,
hydrogen, formate,
acetate, and
glucose

PCE —> ethene

Toluene

PCE -> DCE

Methanol

PCE —> DCE

BES significantly inhibited PCE
dechlorination and methanogensis.
Acetoclastic methanogens, Methanosarcina
sp. and Ms. mazei were shown to
dechlorinate PCE in pure cultures.
Relationships may exist between PCE
dechlorination and methanogesis.

Methanosarcina sp. Strain DCM as shown
to dechlorinate PCE to TCE while growing
on methanol, acetate, methylamine and
trimethylamine. PCE dechlorination was
observed only during mathanogenesis, and
was found to be contingent upon methanol
utilization. Authors suggest that a reduced
electron carrier diverted electrons generated
during methane synthesis to PCE, and that
stimulating methanogenesis could enhance
PCE dechlorination.

First publication of the complete
dechlorination of PCB to ethene in mixed
cultures. Under methanogenic conditions,
the reduction of VC to ethene was found to
be the rate-limiting step in the
dechlorination process. Hydrogen, formate,
acetate, and glucose were found to sustain
dechlorination, although methanol was the
most effective. TCE dechlorination and
methanogenesis were inhibited in cultures
amended with BES. BES did not
immediately inhibit PCE dechlorination, but
resulted in the accumulation of TCE and
1,2-DCEs. Authors concluded that
methanogens may have played a key role in
PCE dechlorination, yet emphasized that
complete dechlorination was only occurring
in mixed cultures.

Metabolism of toluene was found to create
an initial source of reducing equivalents for
the dechlorination of PCE. Either benzoate
or acetate, both toluene metabolites, served
as the final electron donor.

PCE (55mg/L) was degraded to ethene
(<l% remaining as VC) within 4 days by a
methanol-fed anaerobic enrichment culture.
Thirty-one percent of the methanol added
was used for dechlorination, and the
remainder was utilized for acetate
production Dechlorination occurred in the

27

Fathepure
and Boyd
(1988)

Fathepure
and Boyd
(1988)

Frwdman
and Gossett
(1989)

Sewell and
Gibson
(1991)

Distefano,
Gossett, and
Zinder.
(1991)

 

Aquifer Solids Simple organics
from Traverse City, and alcohols
MI.

PCE —) DCE

Laboratory culture Methanol and
used by Freedman Hydrogen

and Gossett (1989)

PCE —> DCE
Rhine River Lactate
sediment and

anaerobic granular PCE —> Ethane
sludge

absence of methanogenesis. This
observation was inconsistent with previous
work, which indicated that methanogens
participated in PCE dechlorination
Authors concluded that other organisms,
potentially acetogens, played a role in PCE

dechlorination

Lactate- and ethanol-fed cultures Gibson and
dechlorinated PCE (5 mg/L) with smaller Sewell

lag times than did cultures fed butyrate, (1992)

crotonate, or propionate. Dechlorination
activity was not enriched using acetate,
isopropanol, or methanol. Unlike the other
substrates tested, methanol and acetate
usually do not produce large amormts of
hydrogen during their anaerobic
metabolism. Therefore, the authors
concluded that hydrogen, generated through
alcohol and fatty acid metabolism, was the
final electron donor in PCE dechlorination.

Hydrogen has been proposed as the direct DiStefano,
electron donor, derived from the Gossett, and
fermentation or breakdown of other more Zinder
complex compounds. Hydrogen was found (1992)
to sustain dechlorination in an anaerobic

enrichment culture for a period of 14 to 40

days. Dechlorination beyond 40 days could

not be sustained without the addition if

nutritional factors from a culture containing

the same inoculum but fed methanol.

Acetogenesis was inhibited in both the

methanol- and hydrogen-fed cultures by

vancomycin, and eubacterial inhibitor of

cell wall synthesis. In the presence of

vancomycin, dechlorination was inhibited in

only the methanol-fed cultures. Authors

concluded that dechlorination in the

methanol-fed cultures was sustained by

hydrogen produced during acetogenesis of

methanol. BES was found to inhibit

dechlorination in cultures fed methanol and

hydrogen, suggesting that hydrogen-

utilizing methanogens, not acetogens, were

possibly the dechlorination microorganism.

PCE (1.5 ppm) was dechlorinate to ethene DeBruin er
and ethane in a fixed-bed column. When al. (1992)
BES inhibited methanogenesis, ethene was

not reduced to ethane. Both river sediment

and sludge were need for complete

dechlorination

 

28

 

PER-K23
(Dehalobacter
restrictus)

Aquifer solids from
Traverse City, MI

Enrichment culture
seeded with D.
tiedjei DCB-l

Laboratory culture
used by Freedman
and Gossett

Anaerobic digested
sewage sludge

Hydrogen and
forrnate

PCE -—> Ethane (?)

Mixtures of lactate,
acetate, and
propionate

PCE —> DCE

3-chlorobenzoate

PCE —> DCE

Methanol

PCE —> VC

Glucose, yeast
extract, propionate,
hydrogen,
methanol and
acetate

A gram-negative bacterium (PER-K23) was
insolated from the inoculum from DeBruin
et a1 (1992). First isolate to couple the
dechlorination of PCE to c-DCE to growth.
Growth on PCE or TCE could only be
sustained using hydrogen or formate.
Biomass and dechlorination were found
account for all of the electrons generated
from hydrogen and forrnate utilization

Hollinger,
Schraa,
Stams, and
Zehnder
(1993)

A mixture of fatty acids, at three different
concentrations, was fed in conjunction with
5 mg/L, PCE to microcosms containing
aquifer solids. The amount of PE
dechlorinated was similar regardless of fatty
acid concentration, although higher
substrate concentrations led to shorter lag
periods. A zero order rate constant of 0.3
uM-day“ was calculated for PCB and TCE
dechlorination TCE and the DCE isomers
were the reduced end products formed. The
authors concluded that butyrate oxidation
supported dechlorination activity since it
preceded the onset of CE reduction to TCE.

Gibson,
Robertson,
Russell, and

Sewell
(1994)

F athepure
and Tiedje
(1994)

Biofilm reactor packed with an enrichment
culture containing Desulfomonile tiedjei
DCB-l was shown to dechlorinate PCE to
cis- and trans-DCE. Acetate, methanol,
glucose, and benzoate could not replace 3-
chlorobenzoate (3-CB) as electron donor.
Dependence of the culture on 3-CB
indicated that strain DCB-l of a similar
microorganism was involved in
dechlorination. The maximum PCE
dechlorination rte was 10.3 umol-L“.hr".

Tandoi et al.
(1994)

Anaerobic enrichment culture degraded 53
mg/L PCE to VC within 20 hours. The
PCE degradation rate was calculated to be
4.6 i 0.4 mol PCEmg VSS'loday'l. VC
dechlorination was inhibited by the
presence of PCE. Zero order kinetics were
used to describe PCE, TCE, cis-DCE and
1,1-DCE reductions to VC. Trans-DCE
conversion to VC was modeled using first
order kinetics.

Cis-DCE dechlorination in anaerobic Komatsu er
culture was supported by glucose, yeast al.

extract, propionate, and to a lesser extent by ( 1994)
hydrogen The rate of cis- dechlorination

was slower in methanol-fed cultures than in

29

 

Dehalaspirillum
multivorans

D. multivarans

Laboratory culture
used by Freedman
and Gossett

Butyrate
enrichment culture
seeded from
Distefano et al.
(1991)

Strain MS-l

c-DCE—a ethene,
ethane

Pyruvate,
hydrogen, and
forrnate

PCE —> c-DCE

Pyruvate and PCB

PCE —> DCE
TCE —> DCE

Hydrogen

PCE —+ VC, ethene

Hydrogen and
formate

PCE -> VC

Glucose, pyruvate,
formate, lactate,
acetate, yeast
extract, amino
acids

cultures in which no electron donor was
added. This implied that methanol might
have had an inhibitory effect on
dechlorination

Dehalaspirillum multivorans, an anaerobe
capable of utilizing PCE as its terminal
electron acceptor, dechlorinated PCE to cis-
DCE. Pyruvate, formate, and hydrogen
could serve as electron donors. Furnarate
was found to be an alternative terminal
electron acceptor. The enzyme PCE
dehalogenase was recovered in cell free
extracts.

D. multivarans grew in defined medium
with PCB and H2 as sole energy sources and
acetate as carbon source. Alternatively to
PCE, fumarate and nitrate could serve as
electron acceptors: sulfate could not. The
organism utilized a variety of electron
donors for dechlorination (pyruvate, lactate,
ethanol, formate, glycerol, H2).

Using the anaerobic enrichment culture
developed by (Freedman and Gossett), a
hydrogen-utilizing culture capable of
growing on PCE was isolated The
hydrogen-PCE culture was capable of
dechlorinating PCE to VC and ethene, but
required supplements of vitamin 812,
supematant from an anaerobic digester
sludge, and acetate as a source of carbon.
Methanol and acetate could not replace
hydrogen as an electron donor, and the
culture did not produce methane or acetate.
This suggests that the hydrogen-PCE
culture did not contain methanogens or
acetogens.

Dechlorinating organisms are able to utilize
lower concentrations of hydrogen than
competing bacteria such as methanogens.

A facultative aerobe capable of
dechlorinating PCE at 0.5 umol PCE.hr'
l.mg (dry weight) cell“1 was isolated from a
site in Victoria, TX, contaminated with
PCE. PCE was dechlorinated to cis-DCE.

3O

Neumarm er
al. (1994)

Scholz-
Muramatsu-
Heidrun et
al. ( 1 995)

Maymo-
Gatell,
Tandoi,
Gossett, and
Zinder
(1995)

Smatlak,
Gossett, and
Zinder
(1996)

Sharma and
McCarty
(1996)

 

Strain TF4B

Laboratory
methanogenic
consortium

Laboratory culture

Contaminated soil
sample

PCE —-> c-DCE

Acetate or pyruvate

PCE —> c-DCE

Lactate, acetate,
hydrogen, and
propionate

PCE —> TCE
VC, ethene

Butyric acid,
ethanol, lactic acid,
propionic acid

PCE —> DCE

Citrate, pyruvate,
succinate, formate,
acetate, and acetate
with hydrogen

PCE —> DCE

The presence of oxygen and nitrate
inhibited dechlorination, suggesting that
strain MS-l would only dechlorinate when
more thermodynamically favorable terminal
electron acceptors were absent. Numerous
electron donors could sustain
dechlorination, although high
concentrations of fermentable compounds
were inhibitory. Characteristics of strain
MS-l were found to be very similar to the
Enterobacteriaceae family. Enterobacter
agglomerans was also shown to

dechlorinate PCE to c-DCE.
Strain TT4B was isolated from stream Krumholz,
sediments contaminated with TCE and sharp, and

toluene. The isolate dechlorinated PCE to Fishbain
c-DCE, and was found to use other terminal (1996)
electron acceptors (TCE, fumarate, and

ferric nitroacetate).

Higher dechlorination rates were observed Ballapragada
to correspond to higher hydrogen partial et al.
pressures. Dechlorinators were found to (1997)

have an advantage in competing with

methanogens for hydrogen (K values for

hydrogen uptake by dechlorinators were

reported to be 12-28 ppm).

Four electron donors were compared for F ennell er
their ability to stimulate and sustain al. (1997)
dechlorination in short and long-term

studies. It was found that for butyrate and

propionate, which are fermented at low

hydrogen partial pressures of 10-3.5 and 10-

4.4 at respectively and ethanol and lactic

acid fermented at partial pressures orders of

magnitude higher, long term differences

were negligible. In all cases, increased

electron donor levels resulted in greater

ethene production

Demonstrated that dechlorinating bacteria Lee at al.
are relatively abundant in nature by (1997)
showing that cultures from both

contaminated and uncontaminated sites

could dechlorinate PCE (nominal

concentration of 10 mg/L) within two

weeks. PCE dechlorinating activity

increased with increasing PCE

concentration up to 150 mg/L nominal

concentration (58 rug/L in aqueous phase).

31

 

Sediments taken
from several
contaminated sites

Dehalococcoides
ethenogenes strain
195

Laboratory
enrichment culture
seeded from an
anaerobic upflow
sludge blanket;
sediments from a
contaminated site
in Texas

Laboratory culture

Methanol, lactate,
acetate, and
sucrose

PCE —> TCE

Hydrogen

PCE —+ ethene

Methanol, lactate,
and hydrogen

PCE ->
VC, ethene

Butyric acid,
ethanol, lactic acid,
and propionic acid

PCE —>
c-DCE, VC

The optimum values for pH and
temperature were 7.0 and 30°C,
respectively. Yeast extract was required to
maintain activity, and all electron donors
tested supported dechlorination.
Dechlorinating rate was determined to be
0.4 umol PCE-mg VSS'I-hr'1

Tested the ability of several electron donors
to enrich PCE dechlorination in
contaminated sediments. Long lag-times
were observed before PCE dechlorination
commenced in all systems. Lactate was the
only electron donor that was able to support
dechlorination in more than one of the
sediments tested. Lactate also supported
greater extents of dechlorination.

Gao, Skeen,
Hooker, and

Quesenberry
(1997)

An anaerobic bacterium was isolated that
had the capability to dechlorinate PCE to
ethene. Strain 195 coupled the reduction of
PCB with growth on hydrogen, and required
the addition of anaerobic digester sludge
supernatant for growth. This is the first
report of an isolate with the capability of
reducing PCE completely to ethene.

Maymo-
Gatell et al.
(1997)

Carr and
Hughes
(1998)

A high-rate PCE dechlorinating culture was
enriched from a culture with no previous
exposure to chlorinated ethenes. The ability
to enrich and sustain dechlorination activity
with various electron donors was tested
with this culture either alone or with a (1:1)
mixture containing a second culture derived
from contaminated aquifer sediments. Over
extended periods of time (approx. 430
days), it was demonstrated that similar rates
and extents of PCB dechlorination could be
achieved regardless of electron donor fed or
inoculum used. Later studies demonstrated
that PCE dechlorination could be sustained
at high hydrogen partial pressures despite
the presence of an actively methanogenic
community.

Fennell and
Gossett
(1998)

Using data obtained from a laboratory, a
model was developed to predict the
formation of hydrogen from the
fermentation of various substrates coupled
with the competition for hydrogen among
methanogens and dechlorinating bacteria
Model simulations suggested that

32

 

Aquifer material Benzoate and
from a PCE- propionate
contaminated

groundwater site in c-DCE —> ETH
Victoria, TX

Aquifer material Benzoate and
from a PCE- propionate
contaminated

grormdwater site in

Victoria, TX

TCE-contaminated Glucose
aquifer
PCE —> VC

compensating for competition of electron
donor by adding excess donor eventually
led to failure of dechlorination and the

development of a predominantly

methanogenic population.

Demonstrated that in batch studies Yang and
containing benzoate and c-DCE, McCarty

dechlorinators maintained hydrogen (1998)
concentrations at levels below that which
support methanogenesis. Cultures fed
propionate, which is fermented to hydrogen
at slower rates than benzoate, had 100%
conversion to ethene rather than 73% in
benzoate-fed cultures. Dechlorinators had
higher hydrogen utilization efficiencies in
continuous flow, completely mixed systems
than in batch reactors, indicating that
different approaches may be used to favor
dechlorination over other competing
microbial processes.

The maximum degradation rates and half- Haston and
velocity coefficients for PCE, TCE, c-DCE, McCarty
and VC dechlorination were determined in (1999)
batch cultures. Degradation rates were

highest with PCB (77 uM/day), and

degradation rates for c-DCE and VC were

similar (14 uM/day, 13 uM/day,

respectively). Half-velocity coefficients for

PCE, TCE, c-DCE, and VC were 0.11, 1.4,

3.3, and 2.6 uM, respectively. Authors

concluded that the common observation of

slow or incomplete dechlorination to the

level of c-DCE and/or VC could be partially

explained by kinetics.
Rates of PCE and TCE dechlorination were Nielsen and
shown to increase with increasing PCB and Keasing

TCE concentrations, including (1999)
concentrations near the solubility limit.

Under saturating conditions (i.e., DNAPL

present), PCB and TCE dechlorination did

not result in the buildup of VC, potentially

due to less competition for electron an

electron donor. PCE dechlorination was

shown to follow Monod kinetics, and TCE

and VC dechlorination were first-order.

 

33

(b) Electron Donors and source of inoculums in sulfate-reducing mixed and pure cultures.

 

 

Source of Electron Donor (5) Summary Reference
Inoculum
Soil sediments Acetate Trace halogenated aliphatics were Bouwer
transformed under denitrifying, sulfate— and Wright
reducrng, and methanogemc COIldlthIlS. (1 98 8)
Laboratory culture Lactate PCE dechlorinated to TCE and c-DCE Bagley and
under sulfate-reducing conditions. Extent Gossett
PCE —+ DCE and rate of PCB dechlorination was less (1990)
TCE _) DCE than that observed in mixed methanogenic
cultures. Neither lactate, acetate, methanol,
isobutyric acid, valeric acid, isovaleric acid,
hexanoic acid, succinic acid, nor hydrogen
appeared directly to support
tetrachloroethene dechlorination, although
lactate-fed inocula demonstrated longer-
term dechlorinating capability.
Enrichments from Sodium lactate and TCE reductively dechlorinated to c-DCE by Pavlostathis
contaminated acetate sulfate-reducing. The highest observed & Zhuang
subsurface soil transformation rate of TCE was 213 (1991)
TCE —> c-DCE micromoles 1'1 per day at 35°C. No further
dechlorination of c-DCE. A decrease in the
rate and extent of TCE transformation was
observed with an increase in the
concentration of bromoethanesulfonate up
to 50 mM.
Desulfitabacreriu Lactate, pyruvate, Anaerobic bacterium was isolated from a Genitse er
m sp. Strain PCEl butyrate, formate, PCE dechlorinating enrichment culture a]. (1996)
succinate, and containing sulfate reducers and acetogens.
ethanol Desulfitobacterium could utilize PCE, 2-
chlorophenol, 2, 4, 6-trichlorophenol,
PCE —> DCE fumarate, sulfite, thiosulfate, and 3-chloro-
4-hydroxy-phenylactetate as terminal
electron acceptors. PCE was dechlorinated
to TCE, and c-DCE and t-DCE. Hydrogen
was found to inhibit dechlorination.
Laboratory Acetate (1 mM) The rate of transformation and the by- DeBest et
Enriclunent products depended on the concentrations of al. (1997)
1,1,1-TCA —9 TCA, acetate and sulfate. Both packed-bed
DC A and reactor studies and batch experiments with
DC A _) unknown BEA and molybdate demonstrated the
products involvement of methanogens and sulfate-

reducing bacteria in transformation of TCA.

 

34

 

Subsurface soils

PCE-enrichment

PCE
contaminated soil

Methanol, lactate,
acetate and sucrose

PCE —> TCE

Yeast extract,
sodium lactate

PCE —> TCE

PCE —-> c-DCE,
TCE

Various levels of sulfate-reducing,
acetogenic, ferrnentative, and methanogenic
activity were observed in all sediments.
PCE dechlorination was detected in all
microcosms, but the amount of
dehalogenation varied by several orders of
magnitude. Lactate-amended microcosms
showed large amounts of dehalogenation
but elevated levels of dehalogenation were
not consistently associated with any
observable anaerobic metabolisms.

A methanogenic and sulfate-reducing
consortium showed that dehalogenation was
due to the direct activity of methanogens.

In the presence of sulfate, methanogenesis
and dechlorination decreased because of
interspecific competition, probably between
the H2-oxidizing methanogenic and sulfate-
reducing bacteria in batch conditions.

PCE was transformed to c-DCE and TCE
under sulfate-reducing conditions using
acetate, lactate, and methanol as electron
donors. Biotransformation of PCB was
dependent upon the type of electron donor
used.

Gao-
J ianwei et
al(1997)

Cabnoler
al(1998)

Ndon et a].
(2000)

 

35

(c) Summary of dechlorination by iron-reducing or iron related abiotic mechanism.

 

 

Source Dechlorination Summary Reference
Anaerobic aquifer Vinyl chloride First demonstration of vinyl chloride can be Bradley
sediments added as oxidized to CO2 under iron (ED-reducing and

substrate conditions. Transformation was dependent Chapelle
upon the bioavailability of Fe (III), which (1996)
was added as Fe-EDTA.
VC —> C02
Creek bed c-DCE and VC c-DCE and VC mineralization were observed Bradley
sediments obtained added as under both methanogenic and Fe (III)- and
near discharge of substrates reducing conditions. In the methanogenic Chapelle
contaminated microcosms, 5% to 44% of 1“VC and 4% to (1997)
groundwater C-DCE, VC_> 14% of 14DCE were recovered as 1"C02.
C02 Under iron reducing conditions, the recovery
of 14co2 from labeled vc was twice that of
the methanogenic microcosms, and similar
for labeled c-DCE. The kinetics of DCE and
VC mineralization varied between the two
compounds: DCE was modeled using first
order kinetics, and VC was modeled with
Michaelis-Menten kinetics.
1.1 M Na2S added PCE —> TCE and PCB were transformed by 10 g/L Butler and
to 0.57 M F€C12 in acetylene F eS in aqueous solution at pH 8.3. Hayes
anaerobic glove TCE _) Acetylene was the major reaction product for (1999)
bag acetylene both TCE and PCB. 1,1-DCE was persistent
over 120 days.
Zero-valent iron Reductive B-elimination accounts for 87 % William
of PCB, 97% of TCE, 94 % of c-DCE, and and
99 % of t-DCE. Reaction of 1,1-DCE gives Roberts
rise to ethylene, consistent with a reductive (2000)
u-elimination pathway.
Pyrite and Abiotic reductive dechlorination of Lee and
magnetite chlorinated ethenes by pyrite and magnetite Batchelor
was monitored in batch system. Target (2002)

organics were mainly transformed to
acetylene and small amount of chlorinated
intermediates, which suggests that B-
elimination, was the main dechlorination
pathway.

 

36

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concentrations of tetrachloroethene to ethene by an anaerobic enrichment culture in the
absence of methanogenesis. Appl. Environ. Microbiol. 57: 2287-2292.

Pavlostathis, 8G. and P. Zhuang. 1991. Transformation of trichloroethylene by sulfate-
reducing cultures enriched from a contaminated subsurface soil. Appl. Microbial.
Biotech. 36 (3): 416-420.

Komatsu, T, Momonoi, K., Matsuo, T., K. Hanaki. 1994. Biotransformation of c-DCE to
ethylene and ethane under anaerobic conditions. Water Science and Technology.
Kidlington, Oxford, UK; Elsevier Science Ltd. v. 30 (7): 75-84.

Gibson, S.A., Roberson, D.S., Russell, H.H., and G.W. Sewell. 1994. Effects of three
concentrations of mixed fatty acids on dechlorination of tetrachloroethene in aquifer
microcosms. Environmental Toxicology and Chemistry, 13 (3): 453-460.

Tandoi, V., Distefano, T. D., Bowser, P. A., Gossett, J. M., and S. H. Zinder. 1994.
Reductive dehalogenation of chlorinated ethenes and halogenated ethanes by a high-rate
anaerobic enrichment culture. Environ. Sci. T echnol. 28(5): 973-979.

Wild, A. P., Winkelbauer, W., and T. Leisinger. 1995. Anaerobic dechlorination of
trichloroethene, tetrachloroethene and 1,2-dichloroethane by an acetogenic mixed culture
in a fixed-bed reactor. Biadegradatian 6(4): 309-318.

Scholz. Muramatsu Heidrun, Neumann. Anke, Messmer. Michael, Moore. Edward, and
Diekert. Gabriele. 1995. Isolation and characterization of Dehalaspirillum multivorans

gen. nov., sp. nov., a tetrachloroethene-utilizing, strictly anaerobic bacterium. Archives
ofMicrobialogy; 163 (1): 48-56.

39

Sharma. P. K. and PL. McCarty. 1996. Isolation and characterization of a facultatively
aerobic bacterium that reductively dehalogenates tetrachloroethene to cis—l,2-
dichloroethene. App]. Environ. Microbial. 62 (3): 761-765.

Lee, T. H., Yoshimi, M., Ike, M., and M. Fujita. 1997. Characterization of an anaerobic
soil enrichment capable of dechlorinating high concentrations of tetrachloroethylene.
Water Science and Technology. 36(6/7): 117-124.

Luijten. M.L.G.C., S.W.M. Kengen, G. Schraa, and AIM. Stams. 1997. Complete
reductive dehloagnation of chlorinated ethenes. In: H.Verachtert and W.Verstraete (Ed),
Intern. Symposium. Environ. Biotechnol, pp. 24-25.

Shelton, D.R., Boyd, SA. and J .M. Tiedje. 1984. Anaerobic biodegradation of phthalic
acid esters in sludge. Environ. Sci. T echnol. 18 (2) p. 93-97.

Rickard, D. T. 1969. The microbiological formation of iron sulphides, Stockholm Cantr,
Geol. (26): 49-6

Freney, J. R. 1979. Sulfur Transformations, in R. W. Fairbridge, and C. W. Fink], Jr.,
Eds, The Encyclopedia of Soil Science Part 1, in Encyclopedia of Earth Science Volume
XII, Dowden, Hutchison and Ross, Inc., Stroudsburg, PA, pp. 536-544.

Sivavec, TM. and D. P. Horney. 1997. Reduction of chlorinated solvents by Fe(II)
minerals. Preprints of papers presented at the 213 th ACS National Meeting, April 13-17,
1997, San Francisco, CA, V0137 (1), pp. 115-117

McCormick, M.L., Kim, H.S., Bouwer, El, and P. Adriaens. 1998. Abiotic
transformation of chlorinated solvents as a consequence of microbial iron reduction: An
investigation of the roles of biogenic magnetite in mediating reductive dechlorination,
Proceedings of the 13 th Mid-Atlantic Industrial and Hazardous Waste Conference, Suri,
R.P.S. and Christensen, G.L., eds, Villanova University, Philadelphia, PA, July 12-15,
1998, pp. 339-348.

Kriegman-King, M. R. and M. Reinhard. 1991. Reduction of hexachloroethane and
carbon tetrachloride at surfaces of Biotite, Vermiculite, Pyrite, and Marcasite, in R
Baker, Ed., Organic Substances and Sediments in Water, Vol. 2., Processes and
Analytical, Lewis Publishers, Inc. Chelsea, Michigan, pp. 349-3 64.

Kriegman-King, M. R. and M. Reinhard. 1994. Transformation of carbon tetrachloride by
Pyrite in aqueous solution, Environ. Sci. T echnol (28), 692-700.

Butler, E., and Kim F. Hayes. 1999. Kinetics of the transformation of Trichloroethylene
and Tetrachloroethylene by iron sulfide. Environ. Sci. Technol. 33(12): 2021-2027

40

Lee, W., and B. Batchelor. 2002. Abiotic reductive dechlorination of chlorinated

ethylenes by iron-bearing soil minerals. 1. Pyrite and Magnetite, Environ. Sci. T echnol,
36(23): 5147-5154

Bouwer, E. J., B. E. Rittmann, and P. L. McCarty 1981. Anaerobic degradation of
halogenated 1- and 2-carbon organic compounds. Environ. Sci. Techno]. 15: 40-49.

Kleopfer, R.D., Easley. D. M. , Haas, B.B., Deihl, T.G., Jackson, DE, and OJ. Wurrey.
(1985). Anaerobic degradation of trichloroethylene in soil. Environ. Sci. T echnol, 19(3):
277-280.

Vogel T. M. and P. L. McCarty. 1985. Biotransformation of tetrachloroethylene to
trichloroethylene, dichloroethylene, vinyl chloride, and carbon dioxide under
methanogenic conditions. Appl. Environ. Microbiol, 49(5): 1080-1083.

Fathepure, B. Z. and S. A. Boyd. 1988. Dependence of tetrachloroethylene
dechlorination on methanogenic substrate consumption by Methanosarcina Sp. Strain
Dcm. Appl. Environ. Microbial. 54(12): 2976-2980.

Sewell, G. W., and S. A. Gibson. 1991. Stimulation of the reductive dechlorination of
tetrachloroethene in anaerobic aquifer microcosms by the addition of toluene. Environ.
Sci. T echnol., 25(5): 982-984

Gibson, SA, and G.W. Sewell. 1992. Stimulation of reductive dechlorination of

tetrachloroethene in anaerobic aquifer microcosms by addition of short-chain organic
acids or alcohols. Appl. Environ. Microbial, 58 (4): 1392-1393.

Fathepure, El. and J.M. Tiedje. 1994. Reductive dechlorination of tetrachloroethylene
by a chlorobenzoate enriched biofilm reactor. Environ. Sci. Technology. 28:746-752.

Maymo-Gatell, X., Tandoi, V., Gossett, J. M., and S. H. Zinder. 1995. Characterization
of an H2-utilizing enrichment culture that reductively dechlorinates tetrachloroethene to
vinyl chloride and ethene in the absence of methanogenesis and acetogenesis. Appl.
Environ. Microbiol. 61(11): 3928-3 933.

Smatlak, C. R., Gossett, J. M., and S. H. Zinder. 1996. Comparative kinetics of hydrogen
utilization for reductive dechlorination of tetrachloroethene and methanogenesis in an
anaerobic enrichment culture. Environ. Sci. T echnol. 30(9): 2850-2858.

Gao, J. W., Skeen, R.S., Hooker, B. S., and R. D. Quesenberry. 1997. Effects of several
electron donors on tetrachloroethylene dechlorination in anaerobic soil microcosms.
Water Research. 3 1(10): 2479-2486.

Carr, C. S. and J. B. Hughes. 1998. Enrichment of high rate PCE dechlorination and
comparative study of lactate, methanol, and hydrogen as electron dopers to sustain

41

activity. Environ. Sci. Technol. 32(12): 1817-1824.

Nielsen, R B. and J. D. Keasing 1999. Reductive dechlorination of chlorinated ethene
DNAPLs by a culture enriched from contaminated groundwater. Biotechnol. Bioeng. 62
(2): 160-165.

Gerritse, J., Renard, V., Games, T. M. P., Lawson, P. A., Collins, M. D., and J. C.
Gottschal. 1996. Desulfitabacterium sp. strain PCE], an anaerobic bacterium that can

grow by reductive dechlorination of tetrachloroethene or ortho-chlorinated phenols.
Archives of Microbiology 165(2): 132-140.

DeBest]. H., Jongema. H, Weijling. A, Doddema. H.J., Janssen. D. B., and W. Harder.
1997. Transformation of 1,1,1-trichloroethane in an anaerobic packed-bed reactor at
various concentrations of 1,1,1-trichloroethane, acetate and sulfate. Applied Microbiolog/
and Biotechnology. 48 (3): 417-423.

Gao, J. W., Skeen, R.S., Hooker, B. S., and R. D. Quesenberry. 1997. Effects of several
electron donors on tetrachloroethylene dechlorination in anaerobic soil microcosms.
Water Research. 31(10): 2479—2486.

Cabirol. Nathalie, Jacob. Francois, Perrier. Joseph, Fouillet. Bruno, and P. Chambon.
1998. Interaction between methanogenic and sulfate-reducing microorganisms during
dechlorination of a high concentration of tetrachloroethylene. Journal of General and
Applied Microbiology. 44 (4): 297-301.

Ndon, U. J ., Randall, A. A., and T. Z. Khouri. 2000. Reductive dehalogenation of
tetrachloroethylene by soil sulfate reducing microbes under various electron donor
conditions. Journal of Environmental Monitoring and Assessment, 60: 329-3 36.

Bradley, P. M., and F. H. Chapelle. 1996. Anaerobic mineralization of vinyl chloride in
Fe (RD-reducing, aquifer sediments. Environ. Sci. T echno]. 30 (6): 2084 -2086

Bradley, P.M., and Chapelle, F .H. 1997. Kinetics of DCE and VC mineralization under
methanogenic and Fe(III)-reducing conditions. Environ. Sci. T echnal. 31 :2692-2696.

Arnold, W. A.and A. L. Roberts. 2000. Pathways and kinetics of chlorinated ethylene
and chlorinated acetylene reaction with F e(0) particles. Environ. Sci. T echnol, 34(9):
1794-1805.

42

Chapter Four: MATERIALS AND METHODS

4.1. Chemicals and Reagents.

All chemicals used in this thesis were ACS reagent grade or higher. The
chlorinated compounds used as substrates in microcosms studies were obtained in neat
liquid forms from Sigma-Aldrich (Milwaukee, WI): PCE (HPLC grade, 99.9+%); TCE
(certified ACS grade); c-DCE (certified ACS grade); 1,1, l,-TCA, trans-1,2-
dichloroethene (t-DCE), 1,1-dichloroethene (1,1-DCE), 1,1-dichloroethane (1,1-DCA).
Analytical standards such as PCE, TCE, c-DCE, t-DEC, 1,1-DCE, 1,1,1-TCA, 1,1-DCA,
1,2-DCA, VC, and CA (Chloroethane) were purchased from Supelco (Bellefonete, PA) in
methanol-dissolved forms.

Gas standards including vinyl chloride (1000 ppm), methane (100%), hydrogen
(100%), ethene (100 ppm), acetylene (l %), carbon dioxide (100 %), and ethane (100%)

were purchased from Scott Specialty Gases.

4.2. Description of Site and Sediments.

The study site is an aquifer located approximately 16 km south of Kalamazoo, MI
(Figure 4.1). This aquifer (designated by Michigan Department of Environmental
Quality as Plume G) has been contaminated with mixed volatile organic compounds
(VOCs) and metal compounds due to PCE use and improper disposal of wastes by a
plastic and rubber manufacturer, which impacts 1.7x107 cubic yards of sediments
(Mayotte, 1994). Immediately down gradient of the Plume G source, groundwater and

sediments are contaminated with chromium and arsenic from a lumber preserving

43

operation (Plume F). Migration of these two plumes to the southeast has resulted in a
zone of significant hexavalent chromium/chlorinated organic contamination. Data
collected from 1988 on has revealed evidence of intrinsic reductive dechlorination that
the daughter products (i.e.TCE and c-DCE), not the parent compound (PCE) prevail. The
background chlorinated compounds and their levels in plume G site are summarized in

Table 4.1.

Table 4.1. Typical Contaminants found from The Study Area“

 

 

Contaminants Concentration (pg/L)

TCE 1015

1, l, l-TCA 232

c-DCE 1518
1,1-DCA 249
VC 19
Chromium 7O

Arsenic 152b

 

aVOC concentrations fiom location MPS-S. No PCE was detected in MPS-5 vicinity.
bMaximum concentration detected in MDEQ monitoring wells. No significant arsenic is present in
MPS-S vicinity

The Plume G aquifer is comprised of relatively homogeneous glacial outwash
sands. This glacial outwash consisted of mostly coarse to medium sand with two gravel
and cobble layers located just above the confining clay layer (depth 80-83 ft). Sulfate,

nitrate, dissolved iron concentrations and total organic carbon analysis indicated that the

44

site groundwater was predominantly anaerobic and low in electron donors (Table 4.2).
The mean annual temperature of this plume is about 12°C. The sediment used in this
study was collected from the mixed waste impacted section of the plume G site, depth
between 50 ft and 75 it, which were mainly composed of fine to medium sand. Batch
experiments were designed to imitate field environment (temperature, electron acceptor

levels, and chloroethene concentrations) as much as possible.

Table 4.2. Typical Chemical Composition of G plume site“:

 

 

Parameter Value Parameter Value
Dissolved iron (aqueous) 12-20 mg/L Nitrate 45 mg/L
Total iron 2.5-8.2 mg/g Sulfate 38 mg/L
pH 7 .2 Total Organic Carbon < 0.04 %b

 

aAnion concentrations from groundwater (MW-MSUI). Iron concentrations fi'om core sediment

samples.
”Provided by Xianda Zhao.

4.3. Microcosms

All experiments were conducted either in 250—ml serum bottles containing 200 ml
of groundwater or 40-ml vials containing 35 ml of anaerobically collected groundwater
from the same aquifer location (MW-MSUI) at 65 it below groundwater surface (bgs).
Bottles were sealed with Mini-inert® (teflon) valves or sealed with screw cap containing
22-mm teflon-lined butyl rubber septa. Groundwater was supplemented to create a sulfate
reducing media as described in Atlas (2). The detailed description of the amended media is
shown in Table 4.3. The groundwater medium was kept under 02 free N2-CO2 (80:20)

prior to addition of NaHC03 and vitamin solutions. The vitamin solutions were filter

45

sterilized. The pH of the medium was adjusted to 7.3 to 7.5 with a sterile HCl or Na2C03
solution. Approximately 20g of soil with large particles removed by sieving with 3.35 mm
screen were added to 250-ml bottle. Abiotic (pasteurized) controls were included in all
experiments. For abiotic controls, the medium and sediments were pasteurized for 8 hours
in 70°C water bath and cooled to room temperature under 02 free N2-CO2 (80:20) prior to
addition of NaHCO3 and vitamin solutions. Sorption controls were added to account for
potential partitioning associated with large-chain hydrophobic carbon sources such as CMB
and palmitate.

The pH of cultures was routinely checked with a pH meter and probe (Orion). A
sub-sample of the culture was removed and pH was determined both before and after

adjustment additions. The pH probe was calibrated before each use (Fisher Scientific).

Table 4.3. Basal Medium for Microcosm"

 

 

Component Trace metal solution Vitamin solution
NI-14C1 0.16 g CoCl2 061-120 012 g p-Aminobenzoic acid
KH2P04 23 m8 MnCl2 -4H20 0.1 g 5 m8
Resazurin 10 mg 21102 0.07 g Vitamin Bu 5 mg
NaHCO; 5.0 g H3B03 0.06 g Biotin 1 mg
Trace metal solution 1 mL Na2M004 02H2O 0025 g Groundwater 100 mL
Vitannn solution 1 mL NiCl2 06H2O 0.025 g

Groundwater 1000 mL CuCl2- 2H2O 0.015 g

HCl (25 %) 6.5 mL
Groundwater 1000 mL

 

a Modified from Atlas (1993)

46

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a 5. 2am

 

 

 

WEBB: Z— NA<UM
000 com o

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47

4.4. Analysis of Volatile Organics
4.4.1. Acetylene, Methane, Ethane, Ethene, and C02

Concentrations of acetylene, ethane, ethene, and methane were routinely
determined by gas chromatographic (GC) analysis utilizing PerkinElmer (Autosystem)
equipped with a flame-ionization detector and a l % SP-1000 on 60/80 Carbopack-B (3.2
mm x 2.44 mm) column from Supelco. Standards and identification were accomplished by
matching the retention times of certified gas standards with peaks detected in headspace
samples. The oven program was as follows: 35 °C, hold 2 min, 35 °C/min to 180 °C, no
hold, flow rate was at 20 ml/min. The detector and injection port temperatures were
200°C. Standards were prepared in serum bottles closed with aluminum crimp cats and
teflon lined butyl rubber stoppers. The detection limit of ethene was 65 nmol/L.

In case of high concentrations of methane and CO2, these were quantified by
direct injection of headspace samples (500 pL) into GC equipped with a thermal
conductivity detector (TCD) and a molecular sieve 5A column (15 ft x 1/8 in. stainless
steel 60/80 Carboxen-lOOO: Supelco) using helium (30 ml/min) as carrier gas. The GC
oven was operated isothermally at 200°C. Standards were prepared by volumetric

dilutions of certified gas standards.

4.4.2. Chlorinated compound
Liquid samples (0.5 mL) were collected with a gas-tight syringe and immediately
dispensed to 40 mL EPA/VOA vials (0.125” TFE/Silicone Liner: I-Chem), which contains

4.5 mL of 2 % sodium bisulfate as a preservative. Vials were sealed immediately with

48

Teflon lined septa and kept in the refrigerator until final analysis. All samples were
analyzed within a 14-day holding period.

Identification and quantification of PCE, TCE, c-DCE, t-DCE, 1,1-DCE, and VC
was made by gas chromatography/mass spectrometry (Agilent 5973 Mass Selective
Detector) combined with a purge-and-trap system (Tekmar/Dohrmann Precept II and
Tekmar/Dohrmann 3100 Sample Concentrator). GC calibration was performed by adding
known masses of PCB, TCE, c-DCE, trans-DCE, 1,1-DCE, and VC directly to a 40-mL
EPANOA vial containing 5-ml chilled distilled deionized water. Quantification of
chlorinated ethenes and ethene was corrected for partitioning between the aqueous and gas
phases. Dimensionless Henry’s Gas Law constants were obtained from Gossett (1987)
except for gases. These values are as follows: TCE, 0.392; c-DCE, 0.154; VC, 1.077;
ethane, 20.4 (Schwarzenbach et al. 1993); ethylene, 8.5 (Y aws, 1999); methane, 33.1
(Freedman & Gossett, 1989), and acetylene, 0.887 (Butler & Kim, 1999). Throughout this
thesis, all reported aqueous and gas-phase concentrations have been corrected for

partitioning between aqueous and gas phase.

4.5. Analysis of Volatile Fatty Acids

Liquid samples were dispensed into Eppendolfs and clarified by centrifirge at 18,
000 rpm for 5 minutes. The supernatant was the acidified with 50 pl of 1.25 N H2SO4, and
dispensed into a 1.5 ml glass vial (Sun Microsystems). The samples were kept frozen until
analysis. Acetic, propionic, and lactic acids were measured by high performance liquid
chromatography using a Supelco Discovery C8® column (ultraviolet absorbance at 210 nm)

and the flow rate was 0.6 mL/min. Acetonitrile was used to regenerate the column between

49

samples and 3.4 g/L of KH2P04 was used as an eluent. Chromatograms were recorded and

data integrated using Turbochrom R 4 software (Perkin Elmer Corp).

4.6. Anion Analysis

Anions were assayed by ion chromatography with suppressed conductivity
detection on a Dionex model 2000i/ SP ion chromatography equipped with a Dionex AS4A
IonPac column and utilizing a 1.8 mM biocarbonate—1.7 mM carbonate mobile phase
(3mL/min). Chromatograms were recorded and data integrated using Turbochrom R 4
software (Perkin Elmer Corp.) Five—point calibration curves were prepared by diluting
primary anion standards into secondary deionized water standards. 1 mL liquid samples
were taken from bottles and filtered through a 0.22 um nylon filter (Scientific Resources
Inc.) 120 microliter of filtered samples were diluted with 480 pl of deionized water, and
dispensed into a polypropylene sample vial (Alcott Chromatography). Samples were
stored at -20 °C until analysis was performed. Each sample was analyzed for acetate,

bromide, chloride, nitrate, nitrite, phosphate, and sulfate.

4.7. Hydrogen Analysis

Hydrogen levels were monitored by injecting 0.5 mL of headspace gas into a
mercury vapor reduction gas analyzer (RGA2 Reduction Gas Analyzer;Trace Analytical,
Menlo Park, CA). Data was collected with Turbochrom R4 software. Hydrogen partial
pressures were measured in the headspaces of bottles at room temperature. The detection

limit for hydrogen in headspace was 10 ppb (0.4 nM).

50

Standards were prepared in serum bottles purged with N2 and capped with butyl
septa prior to adding H2. Hydrogen values are expressed in parts per million by volume
(1 ppmv = 0.1 pa = 10*5 atm) according to the convention adapted by Conrad (1996).
Since most of the values published in other studies are expressed as hydrogen dissolved
in aqueous, dissolved H2 concentrations were calculated by using the following

equations:
LP
H dissolved = — 4.1
.( ) RT ( )

Where, H2 (dissolved) is the aqueous concentration in males per liter; L is the Ostwald
coefficient for H2 solubility, which is 0.01913 at 25°C (Wilhelm et a]. 1977); R is the
universal gas constant (0.0821 liter atm K'lmol'l); P is the absolute pressure (in
atmospheres); and T is the Kelvin temperature (Lovley et a]. 1994). Therefore, 1 ppmv =

ca. 0.78 nM at 25 °C.

4.8. Hydrogen Sulfide Analysis
Hydrogen sulfide was colormetrically measured using sulfide kit (Chemtrics, Inc.)

based on Methylene blue method.

4.9. Data Reduction
Transformation of chlorinated compounds was decided using first-order kinetics.
First-order rate constant k can be obtained by plotting natural logarithm of measured

aqueous concentrations of chlorinated compound as a function of time. Evidence for fit of

51

the data to such a rate law includes the good agreement between experimental data and rate
laws. First order reaction rate (r) in the liquid phase:
r = (dCaq /dt) = —k,,,,sCaq (4.2)

Due to the existence of headspace in batch systems, a fraction of the chlorinated
ethenes partitioned to the headspace and was effectively withdrawn from the reaction phase
(aqueous phase). As a result, the measured rates of transformation can be impeded. The
inherent rates of transformation occurring in aqueous phase can be determined fiom
measured rates using the volumes of the gaseous and aqueous phases, assuming

equilibrium partitioning occurring between these phases at all times. The inherent rate is

the product of the measured rate, corrected by factor f (Burris eta]. 1996)

V
f=1+HV—g (4.3)

Where H is the dimensionless Henry’s law constant (molL"gas / mol L'1 alqmam), and V3 and
Va,l are the gas and aqueous volume respectively. Values of k were corrected to account for

the effects of partitioning of the reactant between the aqueous and gas phases

(kobs =f*k)-

52

4.10. DNA Analysis
4.10.1. DNA Extraction and RTm-PCR

Microcosm DNA samples were extracted from vigorously mixed suspended
sediment slurries (3 ml). After collection, sediment slurries were centrifuged (10,000
rpm) for 6 minutes to remove supernatant and DNA was extracted from sediment pellets
using DNA extraction kit (MO BIO Laboratories. Inc. Solana Beach, CA.) following
manufacturer’s guidelines.

Total DNA quantification was accomplished by detecting and quantifying double-
stranded DNA (dsDNA) in sediment using the PicoGreen® method (Molecular Probes co.
Eugene, OR). PicoGreen® quantification reagent is an ultra sensitive fluorescent nucleic
acid stain for quantitating double-stranded DNA (dsDNA) in solutions. Quantification
assays were performed using 1-10 % dilutions of sample extracts resuspended in 50 11L
of 10 mM TRIS-HCl [pH 7.5]. The samples were excited at 485 nm and the
fluoresecence emission intensity was measured at 538 nm using a spectroflurometer.
Fluoresecence emission intensity was then plotted versus DNA concentrations. Sample
quantitations were accomplished by using MFX Microliter Plate Flourometer and
integration software.

Real-time PCR (TaqMan-PCR) was used to quantify Dehalococcoides
populations and conducted by following the outlines by Gruntzig, et al. (Gruntzvig eta].
2001). PCR amplification was performed in 50 )1] reaction volumes. The reaction
mixture for real-time PCR consisted of IX TagMan Universal Master Mix (containing
AmpliTaq Gold DNA polymerase, AmpErase Uracil-N-glycosylase, which degrades

PCR carryover products from previous reactions, deoxynucleoside triphosphates with

53

dUTP, a passive reference [6 carboxy-X-rhodamine], and optimized buffer components)
(PE Applied Biosystems), 50 nM forward primer and 50 nM reverse primer. MicroAmp
optical caps and tubes were used for the final reactions. The 168 rRNA gene forward
and reverse primers were 5’ CTGGAGCTAATCCCCAAAGCT 3’; 5’
CAACTTCATGCAGGCGGG 3 ’(He et a]. 2002)’. PCR conditions were as follows: 2
minutes at 50 °C, 10 minutes at 95 °C, then 40 cycles of 15 seconds at 95 °C and 1
minutes at 60 °C. Negative controls with no template DNA or no probe were run in each
reaction.

The increase in fluorescence emission, due to the degradation of the probe by the
DNA polymerase in each elongation step, was monitored during PCR amplification using
the 7700 Sequence Detector (PE Applied Biosystems). The fluorescence signal was
normalized by dividing the emission of the reported dye (6-carboxyflourscene) by the
emission of the passive reference dye, 6-carboxy-X-rhodamine. The parameter threshold
cycle (Tc) is the fractional cycle number at which the fluorescence emission crosses an
arbitrarily defined threshold within the logarithmic increase phase (0.1 in our reactions).
The higher the amount of initial template DNA, the earlier the fluorescence will cross the
threshold and the smaller will be Tc. The TC values obtained for each sample were

compared with a standard curve to determine the initial copy number of the target gene.

4.10.2. T-RFLP analysis
T-RFLP analysis was performed as previously described by Liu, et al. (1997).
The amplification was performed using the pair of universal bacterial primers, 8F [5’

AGAGTTTGATCCTGGCTCAG 3 ’] labeled with hexachlorofluorescene (Hex) at the 5’

54

end (synthesized by Operon Technologies, Inc., Alameda, CA). PCR mixtures included
Taq DNA Polymerase (Gibco BRL, Gaithersburg, MD) and were done according to
manufacturer’s recommendations with the addition of 0.2 mg/ml of bovine serum
albumin (Sigma Chemical Co., St. Louis, MO). PCR amplification was conducted in a
GeneAmp PCR system 9600 thermal cycler (Perkin-Elmer Cetus, Norwalk, CT). The
protocol consisted of an initial denaturnation step at 94 °C for 30 sec, annulling at 57 °C
for 45 sec, elongation at 72 °C for 1 min 30 sec and one extension cycle at 72 °C for 7
min to finalize the PCR. Negative controls included tubes that received no template
DNA, as well as positive controls containing pure culture genomic DNA. Aliquots (10
pl) of the PCR products were separated by electrophoresis in a 1.5 % agarose gel using 1
X TAE buffer. The gel was stained with ethidium bromide (500 ng/ 1.11) and visualized by
ultraviolet excitation (Helton, 2000).

Amplified PCR products were purified using the Wizard PCR purification kit
(Promega, Madison, WI) and separately digested with the restriction endonucleases HhaI
overnight at 37 °C. Resulting fragments were resolved on an ABI 373A sequencer in a 6
% urea-containing polyacrylamide gel (PE Applied Biosystems Sequencer), running in
gene scan mode. The resulting electropherograms were analyzed for similarities using

GeneScan software version 3.1 (PE Applied Biosystems).

55

4.11. Literature cited

Mayotte, Timothy. J. 1994. Field Trips Guidebook. The geological society of America.
North Central Section.

Ronald, M. Atlas. 1993. Handbook of Microbiological Media. CRC Press, Florida

Jaime A. Graulau-Santiago. 2003. Development and Application of a Methodology to
Evaluate Natural Attenuation for Chlorinated Solvents using Conceptual and Numerical
Models. A Doctoral Thesis. Civil and Environmental Engineering. Michigan State
University.

Gossett, J. M. 1987. Measurement of Henrys Law Constants For C1 And C2
Chlorinated Hydrocarbons. Environ. Sci. T echnol. 21(2): 202-208.

Schwarzenbach R.P., P.M. Gschwend, and D.M. Imboden. 1993. Environmental Organic
Chemistry. Wiley and Sons, Inc.

Yaws, CL. 1999. Chemical Properties Handbook. McGraw-Hill.

Freedman, D. L. and J. M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions. .
Appl. Environ. Microbiol. 5 5(9): 2144-21 5 l .

Butler, E., and Kim F. Hayes. 1999. Kinetics of the transformation of trichloroethene and
tetrachloroethene by Iron Sulfide. Environ. Sci. T echnal. 33: 2021-2027

Conrad, R 1996. Soil microorganisms as controllers of atmospheric trace gases (H2, CO2,
CI-Ia, OCS, N20, and NO). Microbiol. Rev. 60: 609-640

Wilhelm, E., R. Battino, and R]. Wilcock. 1977. Low pressure solubility of gases in
liquid water. Chem. Rev. 77:219-262

Lovley, D.R., F.H.Chapelle, and J .C. Woodward. 1994. Use of dissolved H2
concentrations to determine distribution of microbially catalyzed redox reactions in
anaerobic groundwater. Environ. Sci. T echnol. 28: 1205-1210

Burris, D.R., C.A. Delcomyn, M.H. Smith, and AL. Roberts. 1996. Reductive
dechlorination of tetrachloroethylene and trichloroethylene catalyzed by vitamin B-12 in
homogeneous and heterogeneous systems. Environ. Sci. Technol. 30:3047-3 052.

Gruntzvig, V., Nold, S.C., Zhou, J., and J .M. Tiedje. 2001. Pseudomonas stutzeri Nitrite
Reductase gene abundance in environmental samples measured by RT-PCR App].
Environ. Microbial. 67, 760-768.

56

He, J., Sung, Y., Dollhopf, ME, Fathepure, B.Z., Tiedje, J. M., and FE. Loffler. 2002.
Acetate versus hydrogen as direct electron donors to stimulate the microbial reductive
dechlorination process at chloroethene-contaminated sites. Environ. Sci. T echnol. 36,
3945-3952.

Liu, W., T.L. Marsh, H. Cheng, and LI. Fomey. 1997. Characterization of microbial
diversity by determining Terminal Restriction Fragment Length Polymorphisms of genes
encoding 16S rRNA. Appl. Environ. Microbiol. 63:4516-4522

Helton, Rebekah R. 2000. Assessment of the microbial communities derived from two
tetrachloroethene contaminated aquifers. MS. Thesis. Michigan State University.

57

Chapter Five: EVALUATION OF ELECTRON DONOR EFFECTS ON THE
TRANSFORMATION OF TCE AND c-DCE USING ANAEROBIC SEDIMENTS

5.1. Abstract

Anaerobic microcosms prepared with sediment and groundwater from an aquifer
contaminated with mixed chlorinated compounds (perchloroethene (PCE), 1,1,1-
trichloethane (1,1,1-TCA), and daughter products (c-DCE and vinyl chloride», and
chromium were used to investigate the effects of electron donors on dechlorination.
Screened substrates were ethanol, methanol, lactate, palmitate, propionate, and
commercial food grade palmitate (Crisco®) delivered via Microporous Beads (CMB).
Substrates were added at levels calculated to stoichiometrically reduce all available
anions in the system. Hydrogen levels were ranged from 0.3 to 1.1 mM in biotic bottles.
But no clear relationship was observed between extent of dechlorination and hydrogen
levels. In microcosms spiked with TCE, the greatest transformation was observed with
CMB (44 %) or lactate (36 %) as carbon sources. For all electron donors tested, the
predominant biotransformation daughter product was c-DCE, with little further
dechlorination beyond c-DCE over a 240-day period. Productions of acetylene have

indicated that abiotic degradation by iron sulfide also participated in dechlorination of

TCE.

58

5.2. Introduction

Chlorinated volatile organic compounds (VOCs) are the most frequently
occurring type of contaminant in groundwater at Superfund sites in US (EPA, 1994). It
has been estimated by EPA that cleanup of these sites will cost more than $45 billion
(EPA, 2000). Biological remediation of these compounds offers the potential for
economical and destructive removal of these hazardous compounds from contaminated
aquifers. Biological treatments may be passive (natural attenuation) or engineered.
Naturally occurring biological activity often leads to incomplete or slow attenuation or
incomplete degradation of contaminants. This is generally due to a deficiency in an
electron donor or acceptor, other nutrients required for the dechlorinating microbial
population, and/or absence of appropriate dechlorinating microbial populations. In these
cases, biostimulation or bioaugmentation may be successful. Biostimulation can be a less
complicated and costly alternative compared to bioaugmentation, if the appropriate

population(s) are present.

Many substrates have been shown to stimulate dechlorination of chlorinated
solvents (Gibson et a]. 1994; Gibson and Sewell, 1992; Gao et a]. 1997). These are
generally organic acids and alcohols. Four fatty acids (lactate, palmitate, propionate, and
CMB) and two alcohols (methanol and ethanol) were used as electron donors to support
dechlorination in this study. Palmitate and CMB were evaluated as potential slow
degrading/persistent electron donors for dechlorination. Propionate is a common
fermentation intermediate that could be produced during anaerobic metabolism.

Alcohols such as ethanol and methanol were used as easily degradable substrates, and

59

methanol has been often used to support methanogens (F ennell et a]. 1997; Carr and
Hughes, 1998; Fathepure and Boyd, 1988 (a)(b)). Furthermore, fermentation of these
substrates is known to produce different levels of hydrogen (F ennell er a]. 1997; Carr and
Hughes, 1998). Ethanol and lactate are known to generate relatively high hydrogen
levels, whereas propionate is known to generate lower hydrogen levels. Since differences
in dechlorination might be due to the possible competition between dechlorinators and
other hydrogen-utilizing microorganisms for hydrogen, hydrogen production was
monitored.

Much of the work on biodegradation of VOCs has evaluated degradation under
methanogenic conditions (Fathepure and Boyd, 1988 (a)(b); Freedman and Gossett,
1989). Fewer studies have been conducted on the transformation of chlorinated
compounds under sulfate reducing conditions (Pavlostathis and Zhuang, 1991) in spite of
the dechlorinating potential of sulfate-reducing bacteria and the relative abundance of
sulfate in subsurface environments.

FeS has been identified as a soil precipitate in anaerobic environments where it is
formed through the biologically mediated reduction of sulfate to sulfide and subsequent
reaction of sulfide with available iron species (Richard, 1969; Freney, 1979). Several
forms of ferrous sulfide can be produced; amorphous iron sulfide, mackinawite (F e29”-
1_o23S), greigite (F e384), and pyrrohotitie (F eSLl). Formation of amorphous iron sulfide
predominates at low pH, while pyrite (F eS2) can be developed by further reaction with
elemental sulfur.

Butler and Hayes (1999) have demonstrated that ferric sulfides and ferric

disulfrdes such as mackinwite and pyrite can promote the abiotic dechlorination of

60

Chloroethenes. Trichloroethene was transformed to 65% acetylene, 6% c-DCE, and 9%
residual TCE (at pH 8.3) after 120 days. Lee and Batchelor (2002) reported that pyrite
degraded TCE to 3.3% c-DCE, 43% acetylene, 2.2% ethene, and 50% residual TCE after
32 days. However, reports by Hassan et a]. (1995) and Doong and Wu (1992) indicated
that FeS is unreactive in the transformation of chlorinated compounds. Discrepancies in
experimental conditions among these different studies may be responsible for these
disparate results.

The purpose of the studies reported here were to screen for the biostimulation
potential of the native site microbes when augmented with different electron donors. I
sought to demonstrate that by adjusting the levels and types of substrates, I could poise
the systems reducing potential above methanogenesis and thus selectively influence
which indigenous populations were enriched. I also attempted to evaluate the production
and reactivity of FeS since it is naturally occurring in sulfate-reducing environments and
known to be involved in dechlorination of chlorinated compounds (Butler and Hayes,
1999; Lee and Batchelor, 2002; Hassan et a]. 1995; Doong and Wu, 1992). Here I report
on comparative studies in TCE/c-DCE degradation by both biostimulation and biogenic

iron sulfide.

61

5.3. Materials and Methods

All experiments were conducted in 250—ml serum bottles containing 200 ml of
groundwater anaerobically collected from the same aquifer location (MW -MSU1) at 65 it
below groundwater. The bottles were sealed with mini-inert® teflon valves, and
incubated in the dark, at aquifer temperature (15°C) and inverted to minimize volatile
loss. The bottles were shaken periodically (twice a week). Reagent-grade chemicals
were used in nutrient medium preparation. Groundwater was supplemented to create a
sulfate reducing media as described in previously in Chapter 4. In order to study the
effect of reductant (N a2S) on dechlorination of TCE and c-DCE, some bottles were
spiked with 0.15 or 1.5 mM of sodium sulfide. Not only abiotic (pasteurized) controls,
but also sorption controls were added to account for potential partitioning associated with
hydrophobic carbon sources in all experiments. To prevent the formation of siderite
(F eC03), and mimic in-situ conditions, the pH was maintained between 7 .0-7 .5 by
addition of NaOH.

The concentrations of TCE and c-DCE in the bottles were 15 11M and 20 11M,
respectively. To study the effect of different substrates in dechlorination, soluble
nutrients including ethanol, 2 mM; lactate, 2 mM; methanol, 4.25 mM; palmitate, 0.25
mM; propionate, 1.5 mM; and slow release nutrients (CMB, 400 mg/L each) were added,
respectively. CMB carbon beads were prepared by loading molten Crisco® into clay
hydrophobic plant growth media. Substrate concentrations were calculated using
McCarty’s stoichiometric approach at standard conditions to poise the reducing potential
to sulfate reducing environments to prevent methanogensis due to excess carbon (See

Appendix A). The equations for the heterotrophic sulfate reducing of various electron

62

donors were calculated using cell efficiency of energy transfer as 0.6 and ammonia as N-
source (McCarty, 1969).

The identification of iron sulfide’s crystalline structure produced in the
microcosms was made by powder x-ray diffraction (XRD) analysis (McCormick et al.
2002). Figure 5.1 shows a X-ray diffraction pattern for the experimentally-prepared FeS
match with the mackinawite (The mackinawite power diffraction pattern is from the
Powder Diffraction File, International Center for Diffraction Data, Swarthmore, PA).
Poorly-crystalline mackinawite and amorphose FeS are found in natural systems (Butler,
1998). X—ray diffraction analyses were conducted using a Rigaku Rotaflex rotating
anode XRD apparatus (Cu K01 radiation, 40 kV, 100 mA). XRD samples were prepared
by freeze-drying and back-filling under nitrogen to potential oxidation. Samples were
mounted on glass slides in the anaerobic glovebox and then sealed under tape to prevent
sample oxidation during analysis.

Headspace samples (acetylene, ethane, ethene, and methane) were determined by

gas chromatographic (GC) analysis (PerkinElmer) as described in Chapter 4.

63

Figure 5.1. Powder X-ray diffraction pattern by experimentally prepared FeS.

(The mackinawite power diffraction pattern is from the Powder Diffraction File,
International Center for Diffraction Data, Swarthmore, PA).

 

"“l

 

 

Intensity

 

 

 

 

 

 

 

PI'I'I‘I'flSTTI I" I'I'TYT J'T'T'T‘ T‘T‘Trl

 

I'I'T‘ ' r'r'r‘r' I'Ivr'r

 

TYT'IrI I'I'I‘I'

20

64

 

5.4. Results

5.4.1. Effect of carbon sources on transformation of c-DCE and TCE

1. TCE transformation

In order to determine which substrates support enhanced dechlorination activity,
multiple carbon sources were screened, including both alcohols and fatty acids. Long-
term incubation cultures were provided with carbon sources including methanol, ethanol,
lactate, palmitate, CMB, and propionate at the beginning of incubation only. Within a
week of nutrient addition, nitrate presented in the site groundwater (average
concentration of 35 ppm) was consumed. All biotic bottles blackened within two weeks
and the presence of sulfide was confrrmed by Chemetrics kit (methylene blue method).
Anion levels, turbidity, and color remained unchanged in pasteurized abiotic controls
throughout the 240-day period.

All bioactive (non-control) microcosms exhibited some dechlorination of TCE over
a 240-day period. The activity generally initiated after approximately 60 days of
incubation, however the onset of degradation and 240 day removal amounts varied among
replicates. Lactate, CMB, and palmitate were effective in stimulating TCE degradation
(Figure 5.2). Propionate, methanol, and ethanol stimulated slight dechlorination of TCE.
The dominant byproduct of TCE degradation was c-DCE, but small amount of acetylene
was detected in most replicates. The percent removal of TCE by fatty acids and their
sorption controls are shown in Figure 5.3. Sorption controls are included for large
molecular weight (C-16) carbon sources to accounts for partitions. After 240 days of
incubation, CMB and lactate showed the most significant removal (about 44 % and 36 %

removal respectively, corrected for partition). An approximately 15 % decrease in TCE

65

.998 ovm macaw meadow sebum—o 8 26 mega—Emcee“ MOB Eng 3v .m> 32m 3 a 80:80me .N.m eBwE

 

 

Beam 3 32m 8
3.550 m5 ESE—am IT 8393 IT 2:538.— Eiucm 111 .8282 II... 353nm 19...
are a: 2.3 a:
a: a: a: 8 c as a: as 8 e
. _ P _ Q — p p p o

 

r
In

 

r
G
‘—

(wt) °ou03 snoanbv

 

. . . .
.V“\\\\\\\>{\§s\\)\s\\:0‘r
.. . .. (r.

 

 

 

E

66

level was observed in the CMB specific abiotic control bottles, most likely due to
partitioning to the hydrophobic Crisco® in the CMB beads (Figure 5.2). This indicated the
removal in the CMB treated microcosms were not due to partitioning. However, without
additional amendments, c-DCE appeared to resist firrther dechlorination under the

conditions of this study.

 

 

80 -
60 -
"<3 40 —
O
8
d)
04
°\°
20 -
O I l '
Coiuol Lactate Palmitate Plamitate CMB CMB
control control

 

-20 - Electron Donors

Figure 5.3. The differences in dechlorination of TCE after 240 days expressed as
percentage removal of abiotic soil control. This experiment was conducted in the presence
of Na2S (1.5 mM). Sorption controls are included for large molecular weight (C-16)
carbon sources. Values are averages of 4 replicates and the error bars represent the

standard deviation.

67

2. Cis-DCE transformation

Unlike TCE, only slight degradation of c-DCE was observed over a 240-day period.
Palmitate and CMB amended microcosms demonstrated some degradation of c-DCE
(Figure 5.4). For c-DCE-amended microcosms, minimal degradation daughter products
were detected. As can be seen on Table 5.2, about 25 % and 12 % degradation of the initial
c-DCE were observed in the CMB and CMB specific abiotic control bottles, respectively,
after 240 days of incubation. This again indicated partitioning did not account for c-DCE

removal in the CMB fed microcosms.

80-

60-

% Removal

 

Cool Lactate Palmitate Palmitate CMB CMB
control control

'20 I Electron Donors

 

Figure 5.4. The differences in dechlorination of c-DCE after 240 days expressed as percent
removal of abiotic soil control. This experiment was conducted in the presence of Na2S
(1.5 mM). Sorption controls are included for large molecular weight (C-16) carbon

sources. Values are averages of 4 replicates and the error bars represent the standard

deviation.

68

5.4.2. Oxidation-Reducing Potential and pH

The pH of the bottles dropped slowly during the course of incubation and remained
stable around 6.8, while the oxidation-reduction potential (ORP) changed rapidly at first
(from an initial value of -102 mV to about -230 mV on day 30), then remained relatively
constant. The transformation of Chloroethenes, pH/ORP final values after 240 days of
incubation, and hydrogen production after 25 days of incubation at 15°C was summarized

in Table 5.1.

Table 5.1. Effect of substrates on transformation of chloroethenesa after 240 days of

incubation in sulfate-reducing conditionsb.

 

 

 

Electron H dro en Chlorinate ethene
Reactant donors p HIGH) 3(1th Reduction (as %)°
TCE Controls 7.2/-138 O 8.6 i 2.6
No nutrient 7.2/~189 0 12.8 i 2.6
CMB 6.8/—206 0.4 i 0.03 43.7 i 4.7
Ethanol 7.2/-168 0.4 i 0.12 14.2 :t 3.3
Lactate 6.9/-192 0.3 i 0.07 35.7 i 4.1
Methanol 7.1/-211 1.1i0.18 17,0;th
Palmitate 6.9/-182 0.3 i 0.09 32, 5 i 29
Propionate 7 .0/-202 0.5 i 0.03 14_1 i 3.6
c-DCE Controls 7. l/-123 0 12.1 i 2.7
No nutrients 7.2/-147 0 11.6 i 2.7
CMB 6.8/-248 0.5 i: 0.08 25.3 i 5.6
Ethanol 7.1/-180 0.6 i 0.12 9.1 i 2.8
Lactate 6.9/-217 0.4 i 0.09 11.8 i 3.8
Methanol 7.0/-199 1.0 i 0.28 13.0 i 4.4
Palmitate 7 .0/-192 0.4 i 0.09 10.9 i 32
Propionate 7.0/-185 0.5 i 0.01 13.2 i 3.5

 

“ Values represent average of triplicate. i indicates standard deviation

bAll experiments are done in the presence of reductant NaZS of 1.5 mM; Methane levels were
below 0.1 uM at the end of the incubation; At end of incubation H2 levels were below 0.01 mM.
°Based on initial moles of Chloroethenes individually.

69

5.4.3. Anion Utilization

Within a week of carbon addition, complete removal of available nitrate/nitrite was
observed. When Nags was not added as a reductant, all sulfates present in the
groundwater (65 to 92 ppm) were removed within 12 days (Figure 5.5). In order to see
the effect of sulfate on TCE transformation, a similar amount of sulfate was respiked on
llO-th day for CMB fed microcosms (Figure 5.6). Even though complete removal of

sulfate occurred again, no significant enhancement on TCE transformation was observed

during this period.
90
-I— Sulfate
~6- Nitrite
65 : "air-“Nitrate

  
 

 

Conc. (ppm)
A
O

p...
kit

 

- 10 5 10 1 5
Time (days)

Figure 5.5. Anion degradation in CMB fed microcosm. Values represent average of

triplicate. The error bars indicate stande deviation

70

100 - —~ 16

 

 

 

 

Sulfate respiked at 65 ppm
75 ~ J + 12
A .- T r 8
g was . m
3 50 - “L .. i —- s 9
5 ‘2‘?
fi v
.2 25 __ 4
c:
<
0 2.29““. 0
0 45 90 135 180
Time (days)

+Nitrate %Nitrite + Sulfate %TCE

Figure 5.6. Effect of sulfate on transformation of TCE in CMB fed microcosm.

5.4.4. Hydrogen production

High levels of hydrogen were produced by fermentation of substrates at first (Table
5.2). Ethanol fed microcosms exhibited the highest levels of hydrogen immediately
following the electron donor addition. But no clear relationship was observed between
extent of dechlorination and hydrogen levels. Methane production was minimum in all
biotic bottles. All these observations indicate that amount of added substrate was enough

to sustain sulfate reducers but not for methanogens.

71

5.4.5. Effect of sodium sulfide on TCE transformation

To determine the effect of reductant on transformation of TCE, additional
microcosms were prepared with 2 different levels of sodium sulfide (0.15 and 1.5 mM).
Approximately 32 % of TCE were transformed with and without reductant, respectively
(lactate as substrate). Addition of reductant did not lead to a significant increased in TCE
transformation (Table 5.2). Sterile controls amended with 1.5 mM of sodium sulfide
showed no significant degradation of TCE (i.e. < 7 %), and no c-DCE or other daughter
products were detected. Microcosms with and without added Na2S, demonstrated similar
ORP values at the same substrate levels (ranging from -182 to -248). Thus, not only
microcosms with reductant, but also those without reductant sustained desirable reducing
conditions, indicating the biological processes, as well as added reductants were capable
of generating the drop in reducing potential. Therefore, carbon loading (not use of
reductant) may be sufficient to control ORP in the field. Table 5.2 shows transformation

of TCE in the presence and absence of reductant using CMB and lactate as substrates.

Table 5.2. Reduction of TCE depending on presence of reductant after 180 days of

incubation at 15°C.

 

 

 

 

 

 

 

Substrate Reductants (mM) Reduction (as %)a ORP
CMB 0 37.8 i 5.8 -187 i 7.3
0.15 40.0 i 4.4 -209 i' 5.9
1.5 41.4:72 -211¢4.5
Lactate 0 32.4 i 1.9 -201 i- 5.7
0.15 3l.4i2.1 -l96i4.5
1.5 32.7:t3.7 -213 123.1

 

“ Values represent average of triplicate. i indicates standard deviation.

72

 

5.4.6. Byproduct Analysis

Table 5.3 shows the byproduct investigation after the dechlorination of TCE after
180-day period. Productions of trace amount of ethene and acetylene initiated
approximately 90 days after incubation, however the onset of degradation and 180-day
removal amounts varied among replicates. The principal reaction byproduct for TCE
transformation using CMB as substrate was c-DCE and acetylene; c-DCE accounted 65
% of the transformed TCE, while acetylene accounted for 27 %. Even though trace
amounts of ethene were detected in some bottles, neither t-DCE nor 1,1-DCE was
observed above the method detection limit.

Unlike TCE, c-DCE treated microcosms exhibited little degradation of c-DCE.
However, trace amount of acetylene was observed above the method detection limit
(acetylene: 50 nM) using CMB as substrate. No 1,1-DCE or VC was observed above the
method detection limit. Mass balance of c-DCE shown in Table 5.3 shows more than 17
percent discrepancy, which may have resulted from VOCs loss from handling and
sampling.

In order to see the effect of sulfate on TCE transformation, a similar amount of
sulfate was respiked on 110-th day for CMB fed microcosms. Even though complete
removal of sulfate was observed again, no significant enhancement in acetylene
production was observed during this period.

The crystalline structure of microcosms is illustrated in Figure 5.7. The figures
show wide peaks and relatively small intensities. These diffraction patterns by the
microcosm sediment shows poor degree of crystallinity and thus, indicating that the iron

sulfide produced was amorphous. A crystal structure of microcosm treated with added

73

sulfate can be seen in Appendix B (no difference in crystal structure was observed due to

added sulfate).

Table 5.3. The byproducts and mass recoveries for transformation of c-DCE and

TCE after 180 days“.

 

 

 

 

 

 

 

 

Reactant Amendment ijroduct Mass recovery (%)b

TCE CMB Acetylene 7.3 i 4.4
Ethene 2.2 i 3.7

eDCE 175i49

TCE remaining 62.2 3: 5.8

Tm“ 892:68

Ethanol Acetylene 5.3 i 2.0
C-DCE 1.8 i 0.5

TCE remaining 87.7 i 7.3

Tad 948:83

Lactate Acetylene 8.2 i 3.7
Ethene 2.1 i 1.3

eDCE 152:24

TCE remaining 67.6 i- 1.9

Tad 931:89

Methanol Acetylene 4.7 i 1.2
04303 1.5 i 0.9

TCE remaining 84.9 i- 5.8

Tmfl 911:6?

Palmitate Acetylene 6.3 i 2.1
Ethene 2.2 i 0.7

eDCE 15:39

TCE remaining 68.2 i 9.3

de 842:128

Propionate Acetylene 1.9 i 1.4
c-DCE 2.0 d: 1.1

TCE remaining 88.2 i 5.3

TNfl 922:68

c-DCE CMB Acetylene 7.3 i 1.0
c-DCE remaining 755 i 49

Tmm 828:58

 

 

 

 

 

 

“The pH of microcosms was maintained and nearly constant between 7 -7.5.
Reductant (N a28) was not added. Only trace amounts of VC were detected.
Values represent average of triplicates. i indicates standard deviation
bBased on initial moles of chloroethenes individually.

74

 

0.250

0.200 -

0.150 d

0.100 .

CPS norm

0.050 ~

 

1
0a» . . . L 11.4.1.1 1...; -l. 1. link... .‘J 11‘.“ A . .LL L.
10 20 30 40 50 60 70 80 9C

 

 

 

 

 

-0.050
29

Figure 5.7 . Powder x-ray diffraction pattern by iron sulfide produced in microcosm.

75

5.4.7. Field Estimation

The extrapolation of laboratory result to field-scale processes is an uncertain
endeavor, particularly in environmental microbiology (Madsen, 1996). However, studies
of several sites contaminated with chlorinated ethenes suggest that rates of reductive
dechlorination estimated from laboratory microcosms are similar to those estimated from
field studies (Wilson et al. 1996). Half-life prediction of chloroethene in the groundwater
was made using first order rate equation to obtain treatment estimation. As can be seen in
Table 5.4, remediation of TCE and especially c-DCE in groundwater will results in
prolonged operation time and high cost without electron donor addition Natural
attenuation occurring in the aquifer seems not only incomplete, but also insufficient. The
effects of sulfate, reductant (N azs) and nutrient and their interactions on the dechlorinations

are summarized on Appendix C.

76

Table 5.4. Estimated transformation rate and half-time of TCE and c-DCE biodegradation

 

 

 

in Plume G

Chloroethene Electron donor k (l/month)a 2311:3133

TCE CMB 0.13 5.6

Ethanol 0.09 8.0

Lactate 0.11 6.5

Methanol 0.08 8.3

Palmitate 0.08 9.0

Propionate 0. 1 6.6

No nutrient 0.006 108

c-DCE CMB 0.03 20.3

Ethanol 0.01 62.4

Lactate 0.03 24.5

Methanol 0.02 37.5

Palmitate 0.01 47.1

Propionate 0.01 53.7

No nutrient 0.003 256

 

 

 

 

 

 

“ Based on first-order rate equation C=Coe‘k‘.
b t1/2 = 111 Z/k.

77

5.5. Conclusion and Discussion

Several researchers have demonstrated the stimulation of reductive dechlorination

of chlorinated ethene by addition of organic supplements or hydrogen (Gibson et al. 1994;
Fathepure and Boyd, 1988(a); Freedman and Gossett, 1989). These studies were
conducted in soils that were rich in organic material (Gibson et a1. 1994; Freedman and
Gossett, 1989) or a pure culture derived from anaerobic sludge (Fathepure and Boyd,
l988(a)). In contrast, our study used anaerobic materials from an oligotrophic aquifer
(maximum total organic carbon < 0.04 %) impacted with mixed chlorinated compounds
(PCE, 1,1,1-TCA, and daughter products), and chromium to investigate the effects of
nutrients on dechlorination. TCE degradation was achieved by selectively stimulating
indigenous microflora using fatty acids such as CMB and lactate. The predominant
transformation daughter product was c-DCE and acetylene.

Previous studies on the transformation of chloroethenes under sulfate-reducing
conditions have also resulted in the production and accumulation of c-DCE (Pavlostathis
and Zhuang, 1991; Bagley and Gossett, 1990). Even though TCE degradation continued
afier depletion of sulfate, it occurred at a much lower rate in our study. This suggests these
conditions led to enrichment of sulfate reducers that the degradation observed might have
been linked to sulfate reducing populations such as Desulfomonile tieay'ei or Desulfovibrio
fiuctosivorans (Drzyzga and Gottschal, 2002; Mohn and Tiedje, 1991).

Hydrogen is an important intermediate in the anaerobic degradation of organic
matter (Conrad, 1999). Microbial competition for H2 plays an important role in the natural
attenuation of chloroethenes, and recent studies have focused on the role of H2 as a key

electron donor for the reductive transformation of these compounds. In our study, we

78

designed a sulfate-reducing environment where sulfate is depleted with minimal methane
production. Even though high H2 production was observed in some cases, no clear
relationship with transformation rate was observed.

Even though the predominant transformation byproduct was c-DCE, small amounts
of acetylene, known as an abiotic dechlorination indicator (Butler and Hayes, 1999; Lee
and Batchelor, 2002), were also detected. This observation suggested that the abiotic
degradation by iron sulfide have participated in dechlorination of TCE in microcosm.

A plausible scheme for the reduction of the TCE to acetylene by abiotic degradation
(hydrogenolysis (replacement of a halogen by hydrogen), reductive elimination
(dihaloelimination), and hydrogenation (reduction of multiple bonds)) has been proposed
by Lee et al (2002). Because acetylene was one of the main transformation products, a
reductive elimination pathway would appear occur (Figure 5.8). The removal of TCE
without observing stoichiometric production of a chlorinated intermediate suggests that the
acetylene formed during the TCE dechlorination is produced via an unstable intermediate
(chloroacetylene) that quickly decays.

The diffraction pattern by XRD indicated poor-crystalline (amorphous) formation
of FeS in microcosm. According to Bemer (1964), the poorly-crystalline mackinawite
and amorphous FeS are often found in natural systems and their crystallinity is similar to
the materials precipitated from sodium sulfide and ferrous sulfate.

Additional sulfate, which would have provided added source of the reductive
power in microcosm, did not coincide with rapid enhanced dechlorination. No immediate

increase on the TCE degradation by the added sulfate indicated that the dechlorinations

79

by indigenous sulfate reducers and iron sulfide are very slow. This suggests that the

pulse feeding strategy is required to achieve sufficient dechlorinating activity.

80

Figure 5.8. Possible pathways for the reductive dechlorination of chlorinated ethenes by
iron sulfide (modified from Lee et al. (2002)). Chemical compounds in italic characters

were detected in this experiment

C1 C1
/ Cl H \
C1 C1 TCE
H—m— Cl
H H
Chloroa et 1
cis-DCE \ _ c y we
Reductive
l B-elrmmatron l Hydrogenolysis

H>=<H \ H—m—H

C1 H Acetylene
VC

\ H H A/Hydrogenation

>=<

H H
Ethene

l Hydrogenation

RH»

H H

Ethane

81

5.6. Literature cited

EPA. 1994. “Common Chemicals found at Superfund Sites”, Office of Emergency and
Remedial Response, Washington, DC.

EPA. 2000. Engineered Approaches to In Situ Bioremediation of Chlorinated Solvents:
Fundamentals and Field Applications.

Gibson, S.A., Roberson, D.S., Russell, H.H., and G.W. Sewell. 1994. Effects of three
concentrations of mixed fatty acids on dechlorination of tetrachloroethene in aquifer
microcosms. Environmental Toxicology and Chemistry, 13 (3): 453-460.

Gibson, SA, and G.W. Sewell. 1992. Stimulation of reductive dechlorination of
tetrachloroethene in anaerobic aquifer microcosms by addition of short-chain organic
acids or alcohols. App] Environ. Microbial. 58 (4): 1392-1393.

Gao, J. W., Skeen, R.S., Hooker, B. S., and R. D. Quesenberry. 1997. Effects of several
electron donors on tetrachloroethylene dechlorination in anaerobic soil microcosms.
Water Research, 3 1(10): 2479-2486.

Fennell, D. E., Gossett, J. M., and S. H.Zinder. 1997. Comparison of butyric acid,
ethanol, lactic acid, and propionic acid as hydrogen donors for the reductive
dechlorination of tetrachloroethene. Environ. Sci. T echnol., 31(3): 918-926.

Carr, C. S. and J. B. Hughes. 1998. Enrichment of high rate PCE dechlorination and
comparative study of lactate, methanol, and hydrogen as electron donors to sustain
activity. Environ. Sci. T echnol., 32(12): 1817-1824.

(a) Fathepure, B. Z. and S. A. Boyd. 1988. Dependence of tetrachloroethylene
dechlorination on methanogenic substrate consumption by Methanosarcina sp. Strain
Dcm.. Appl. Environ. Microbiol. 54(12): 2976-2980.

(b) Fathepure, B. Z. and S. A. Boyd. 1988. Reductive dechlorination of
perchloroethylene and the role of methanogens. FEMS Microbiology Letters 49(2): 149-
156.

Freedman, D. L. and J. M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions.
App]. Environ. Microbiol. 55(9): 2144-215 1.

Pavlostathis, S. G. and Ping Zhuang. 1991. Transformation of trichloroethylene by

sufate-reducing cultures enriched from a contaminated subsurface soil. Appl. Microbial.
Biotech, 36(3): 416-420.

82

Rickard, D. T. 1969. The microbiological formation of iron sulphides, Stockholm Contr,
Geo]. (26): 49-6

Freney, J. R 1979. Sulfur Transformations, in R. W. Fairbridge, and C. W. Fink], Jr.,
Eds, The Encyclopedia of Soil Science Part 1, in Encyclopedia of Earth Science Volume
XII, Dowden, Hutchison and Ross, Inc., Stroudsburg, PA, pp. 536-544.

Butler, E., and Kim F. Hayes. 1999. Kinetics of the Transformation of Trichloroethene
and Tetrachloroethene by Iron Sulfide. Environ. Sci. Techno]. 33: 2021-2027

Lee, W., and B. Batchelor. 2002. Abiotic reductive dechlorination of chlorinated
ethylenes by iron-bearing soil minerals. 1. Pyrite and Magnetite, Environ. Sci. T echno],
36(23): 5147-5154

Hassan, S. M., Wolfe, N.L., and MG. Cippolone. 1995. Reaction mechanisms involved
in the reduction of halogenated hydrocarbons with sulfated iron, “209'” ACS National
Meeting, April 2-7, Anaheim, CA, Vol. 35 (1), pp.73-737.

Doong, R. A. and S. C. Wu. 1992. Reductive dechlorination of chlorinated hydrocarbons
in aqueous solutions containing ferrous and sulfide ions, Chemasphere 24, 1063-107 5.

McCarty, P. L. 1969. Energetics and Bacterial Growth. 5'” Rudolf Research Conf.
Rutgers.

McCormick, M.L. E.J. Bouwer, and P. Adriaens. 2002. Carbon tetrachloride
transformation in a model iron-reducing culture: relative kinetics of biotic and abiotic
reactions. Environ. Sci. T echnol. 36(3); 403-410

Butler, Elizabeth Constance. 1998. Transformation of haolgenated organic compounds by
iron sulfide. A Ph.D. Thesis. pp. 28. Univeristy of Michigan.

Madsen, E.L. In Soil Biochemistry. Vol.9; Stotzky, G., Bollag, J ., M. Eds; Marcel
Dekker: New York, 1996. pp.287-370

Wilson, J .T.; Kampbell, DH. and J .W. Weaver. In US. Environmental Protection
Agency Symposium in natural attenuation of chlorinated organics in ground waters. Sept.
11-13, 1996; Dallas, TX; EPA/540/R-96/509; 1996;pp.124-127

Bagley D. M. and J. M. Gossett. 1990. Tetrachloroethene transformation to

trichloroethene and cis-1,2- dichloroethene by sulfate-reducing enrichment cultures.
Appl. Environ. Micro]. 56(8): 251 1-2516.

83

Oliver Drzyzga and Jan C. Gottschal. 2002. Tetrachloroethene dehalorespiration and
growth of Desulfitobacteriumfrappieri TCE] in strict dependence on the activity
of Desulfavibrio fructosivorans. Appl. Environ. Microbiol, p. 642-649, Vol. 68, No. 2

Mohn, W.W., and J .M. Tiedje. 1991. Evidence for chemiosmotic coupling of reductive
dechlorination and ATP synthesis in Desulfomonile tieay'ei. Arch. Microbiol. 157:1-6.

Conrad, R. 1999. Contribution of hydrogen to methane production and control of
hydrogen concentrations in methanogenic soils and sediments. FEMS Microbiol. Ecol.
28, 193-202.

Bemer, R A. 1964. Iron sulfides formed from aqueous solution at low temperatures and
atmospheric pressure. .1. Geol. 72, 293-306.

84

Chapter Six: STINIULATION OF HIGH-RATE COMPLETE
DECHLORINATION ACTIVITY BY PULSE FEEDIN GS OF
CHLOROETHEN ES

6.1. Abstract

Once-fed microcosms that demonstrated initial positive dechlorinating activity
were further characterized for chloroethene/ethane transformation potential by pulse-
feedings of VOCs, lactate, and native levels of sulfate using plume G groundwater.
Experiments were performed in batch microcosms receiving two levels of lactate; low
(0.5 mM) or high (12.5 mM). PCE was fed approximately 15 umol/ 2 month for the first
year followed by 15-umol/ month in the second year. High lactate enriched bottles
exhibited limited transformation of PCB with TCE or c-DCE as byproduct (0.03 6/day).
Low lactate enrichments demonstrated high rate of degradation; PCE, TCE, c-DCE, t-
DCE, 1,1-DCE, and VC to ethenes (rates ranged from 0.19 to 0.82/day). Further
reduction to ethane was observed in many cases. Highest rate of PCE transformation was
observed at 25 °C in low lactate/PCE enrichments. 16S rRNA gene-based RTm-PCR
identified that the pre—enrichments contain no quantifiable amounts of Dehalococcoides
populations. Significant levels of Dehalococcoides were detected in all of the post-
enrichment samples. The low lactate/PCE microcosms contained Dehalococcoides at
least two orders of magnitude higher than high lactate/PCE enriched microcosms. This
demonstrated that pulse feedings of chlorinated ethenes, and low levels of electron donor

effectively stimulated Dehalococcoides populations.

85

6.2. Introduction

Methanogens are known to dechlorinate PCE to TCE and c-DCE (F athepure and
Boyd, 1988(a)(b); Freedman and Gossett, 1989). However, the dechlorination by
methanogens is known to be incomplete and proceed at low rates. So far only
Dehalococcoides ethenogenes is known to fully dechlorinate PCE to ethene. It uses H2
as a sole electron donor, grows slowly in pure culture, and is fastidious in its growth
conditions (Maymo-Gatell et a]. 1997). In pure culture, Dehalococcoides ethenogenes
requires the extract of an anaerobic sludge community for sustained growth, suggesting
that it relies on biochemical collaboration with other microorganisms and that it benefits
significantly by growing in a mixed community.

Several molecular approaches are currently available as means of exploring
microbial communities including T-RFLP and RTm—PCR (Hendrickson et a]. 2002; Liu
et a]. 1997). This research discusses the characterization of highly enriched mixed
cultures that reductively dechlorinates PCE to ethene/ethane with lactate as an electron
donor using two molecular techniques. Effects on the community structure and activity
on the original community by prolonged exposure to high versus low lactate/PCE
feedings were studied. Terminal restriction length polymorphism (T-RFLP) analyses
were employed to identify community structure in the mixed cultures and to track
population dynamics after enrichment by PCE. Additionally, specific primers for
detecting Dehalococcoides species were used to determine the presence and enumeration

of such Dehalococcoides populations.

86

Studies presented here focus on the enrichment of a high-rate PCE dechlorinating
culture through pulse feedings of PCB, lactate, and native levels of groundwater sulfate
over 2-year period. Microcosms which demonstrated initial positive dechlorinating
activities, were further enriched using 2 different ratios of lactate and pulse feedings of
chlorinated ethenes to evaluate the effect of excess electron donor on stimulating
dechlorinating activity. I have evaluated the effect of lowering the ratio between
substrate to PCE, along with pulse feedings on the substantial enrichment of
Dehalococcoides. Characterization of these enrichments and enumeration of
Dehalococcoides are presented later in this chapter. Finally, I conducted daughter
product analysis to evaluate the completeness of PCB transformation by rapidly

dechlorinating enrichments.

87

6.3. Materials and Methods

Enrichment Procedure

Anaerobic microcosms assembled from plume G sediments initially exhibited
neither rapid nor complete transformation of parent compounds (TCE and c-DCE). In
order to enhance dechlorinating activity, enrichments that demonstrated the highest
dechlorination activity were subcultured and incubated with three different
concentrations of lactate (0, 0.5 mM, and 12.5 mM) and equal amount of PCE (15
umol/bottle). Table 6.1 shows the additions of chlorinated compounds and lactate
corresponding to the appropriate experiment. Chlorinated compounds were added neat
using SuL microsyringe except for VC, which was added as a gas using airtight syringe.
The Roman numeral corresponds to the chronological order in which the experiments
were performed. Experiment I indicates two conditions that favor one or the other
process based on electron donor/acceptor ratio.

Experiment 11 was designed to screen for the substrate specificity of PCE over
TCE, and c-DCE. Experiment [I was set up using c-DCE long-term microcosms from
chapter 5. Only low levels of lactate (see Table 6.1 for exact concentration) were fed
once every 2 months for 6 months and once a month thereafter for c-DCE and TCE.
TCA microcosms were spiked once every 2 months, as the TCA microcosms never
exhibited high rates of dechlorination. Therefore, all chloroethene enrichments
consumed similar amounts (12 cycles of 15 umol) before the rate experiments except
TCA. TCA enrichments had only 6 cycles of 15 umol over 360 days of enrichment. The

additions and consumptions of chlorinated compounds for experiment II are depicted in

88

Appendix D. The overview of enrichments and experiments performed according to their

time sequence is described in Figure 6.1.

Sampling and Analysis

Over the initial period of enrichment, complete transformation of PCB was not
observed; however, periodic respiking with PCE was conducted. During enrichment, one-
quarter of the bottle (liquid volume, ZOO-mL), contents were replaced with fresh
groundwater media (typically once/2 month), to replenish vitamin, phosphate, and lactate
(0.5 or 12.5 mM). Groundwater media was made by adding phosphate (KH2PO4, 0.2g/L),
sodium bicarbonate (NaHCOg, 5g/L), and vitamin supplements (see Table 4.3) to plume G
groundwater, which contained native amounts of sulfate (range between 35 to 60 ppm) and
nitrate (range between 25 to 40 ppm). After lactate addition, PCE was added at 15
umol/bottle. Throughout these studies, bottles were purged thoroughly afier sampling but
prior to each PCE addition to remove VC and ethene, and thus to prevent potential toxicity
(Schink, 1985). The pH was monitored and maintained between 6.8-7.5 by addition of
small amounts of 1 M NaHC03 solution. Chlorinated compounds, headspace, anion, and
fatty acid analyses were done as described previously in Chapter 4. Terminal restriction
length polymorphism (T-RFLP) and Real time PCR (RTm-PCR) analyses were performed

as described in Chapter 4.

Effect of Temperature
All bottles were kept in a constant room temperature during 2-year enrichment

period. Nine low lactate/PCE enriched bottles were incubated at three different

89

temperatures (15, 25, and 35 °C) for 3 weeks to evaluate temperature effects on

transformation of PCE.

Table 6.1. Experiment arrangements for respective conditions. Microcosms were
exposed to pulse feedings of PCB over 1.5 years for Experiment I and pulse feedings of

TCE, c-DCE, and TCA over 1 year for Experiment 11“.

 

Chloroethene
added Lactate/PCE uequiv of umol of
Conditions (umol) (equiv/ equiv) lactate lactate

 

Experiment I: Cometabolic reductive dechlorination favoring versus chlororespiration
favoring conditions

(using TCE exposed bottles from Chapter 5)

PCE 1 15 0/0 0 0
PCE 2 15 10/1 1200 100
PCE 3 15 250/ 1 30000 2500

Experiment 11: Chlororespiration favoring conditions

(using c-DCE exposed bottles from Chapter 5)

TCE 15 10/1 900 75
c-DCE 15 10/1 600 50
TCA 15 10/1 600 50

aThe electron donor ratio was calculated assuming: PCE 8 equiv/mol; TCE 6 equiv/mol;
c-DCE 4 equiv/mol; and TCA 4 equiv/mol. Lactate was assumed to provide 12
equiv/mol (on CO2 basis).

90

 

 

 

 

 

 

 

 

 

0.5 mM Lactate, 12 mM Lactate,

 

 

 

 

 

 

 

 

 

15 mol PCE/2 mo. 15 mol PCE/2 mo.
f """""" '1
¢ : 1 st year 5
0.5 mM Lactate, 12 mM Lactate,
15 umol PCE/l mo 15 umol PCE/l mo.

 

 

 

 

 

 

 

 

 

 

 

 

 

Figure 6.1. The overview and sequence of the experiments performed during

enrichments.

91

6.4. Results
6.4.1. High lactate conditions
6.4.1.1. Transformation of PCE

After one year of repeated feedings, transformation rates became faster (once a
month feeding of PCE) and c-DCE was produced as byproduct for some bottles. Figure
6.2. depicts repeated additions and consumptions of PCB at the early stages of
enrichment period. However, transformation beyond TCE was limited, with only slight
c-DCE degradation during 2 years. Figure 6.3 illustrates PCE transformation by high
lactate/PCE enrichments and the first order rate achieved by high and low lactate/PCE

enrichments.

 

 

 

 

 

 

O I T T 1 I Y

0 60 120 180 240 300 360 420 480 540
Time (days)

Figure 6.2. PCB additions and consumption by high lactate/PCE treatment over 1.5 years.
12 cycles of PCB was added.

92

 

 

60

20 40
Time (days)
+ VC + c-DCE +TCE + PCE

Figure 6.3. Effects of prolonged exposure of high lactate (12.5 mM) on enhancing

transformation activity (after 12 month of incubation).

Table 6.2. Effects of prolonged exposure of high versus low lactate on enhancing

transformation activity (after 12 month of incubation).

 

 

Transformation rate Half life Final products
(l/day) a (day) b
Low lactate/PCE 0.71 0.98 ethene and ethane
High lactate/PCE 0.04 19.4 c-DCE and TCE

 

3Based on first-order rate equation C=Coe'1“.
b rm = ln 2/k.

93

6.4.1.2. Fermentation balance

An electron balance was determined after 10 days of fermentation. To
accomplish this, fatty acid transformation, methane, and hydrogen were monitored 10
days after the spiking of lactate and PCE. Using lactate and PCE as starting compounds,
microequivalents of lactate consumed were balanced with the sum of the
microequivalents of reduced products formed. The electron balance for one of the six
replicate bottles is shown in Table 6.3. It is apparent that nearly 76.5 % of the reducing
equivalents from lactate were used to produce acetate and propionate and the other 10 %
being used in methane production. Dechlorination and hydrogen production were
insignificant sinks (< 1.5 %) for electrons on 10th day electron balance.

The other five replicate bottles exhibited similar results. The fraction of lactate
equivalents used in dechlorination varied from 0.7-1.5 % and averaged 1.2 % among 5
bottles. The electron equivalents for dehalogenation were calculated assuming complete
dechlorination from PCE to ethene requires 8 equiv/mol and lactate breaks down to
propionate and acetate. Biomass increment was ignored. However, CO2 change during
this 10 day was not negligible. The missing 8 % of reducing equivalents produced from
lactate consumption may have used in nitrate or sulfate reduction. Negative and positive

values represent consumption and production during 10 days.

94

 

Table 6.3. Electron balance for a single, PCE degrading microcosm (15 umol PCE
/bottle, 12.5 mM (2500 umol) lactate; data obtained at day 10)

 

 

Amount Reducing Electron Electron Percenta
present Equivalents equiv. equiv. (%)
Compound
(umol) (eq/mol) consumed formed
(ueq) (ueq)
Lactate -476 4 - 1 904
Propionate 52 6 3 12 16.4
Acetate 129 8 1032 54.2
CH4 23 8 184 9.7
H2 0.27 2 0. 54 0 . 03
PCE - l 3 0
TCE l 3 2 26 l .4
DCE isomers 0 4
VC 0 6
Ethene 0 8

 

aPercent calculated by dividing reducing equivalents produced over total reducing

equivalents consumed.

95

 

6.4.2. Low lactate conditions

6.4.2.1. Enrichment of high rate activity

Low lactate/PCE bottles were respiked with PCB and replenished with fi'esh
groundwater media just like high lactate/PCE bottles. However, lactate was supplied 25
times less than high lactate/PCE bottles (100 mol). Reducing potential for low
lactate/PCE reached approximately -200 and pH was monitored and maintained between
6.8-7 .5. After 4 cycles of repeated feedings, transformation rates became faster (once a
month feeding of PCE) and more complete. For example, primarily VC was produced as
byproduct. By 9 cycles of repeated feedings (12 month), bottles treated with low lactate
ones exhibited complete transformation of PCE to ethene and 20 times faster than high
lactate treated ones (Figure 6.4).

Since complete transformation of PCB was only observed under the low lactate
conditions, the ranges of VOCs, which could be degraded under these conditions, were
determined. As parent products, TCE, c-DCE, and TCA were evaluated since these
compounds already present in plume G. TCE and c-DCE were completely transformed
to ethene after 6 month of repeated feedings. However, limited number of TCA fed
microcosms (5 out of 12) transformed to Chloroethane (CA). Ethane production from CA

was not observed over 60 days.

96

 

 

 

 

 

 

 

 

 

 

 

 

 

0 l 1
0 90 180 270 360 450

Time (days)

Figure 6.4. PCE additions and consumption by low lactate/PCE treatment during 450
days (12 cycles).

97

6.4.2.2. Fermentation balance

An electron balance was calculated after 10 days of fermentation for low lactate
treated microcosms. To accomplish this, fatty acid transformation, methane, and
hydrogen were monitored 10 days after the spiking of lactate and PCE. Using lactate and
PCE as starting compounds, microequivalents of lactate consumed were balanced with
the sum of the microequivalents of reduced products formed. The electron balance for
one of the six replicate bottles is shown in Table 6.4. It is apparent that about nearly half
of the reducing equivalents from lactate were used to produce acetate (43%) and the other
20 percent being used in reductive dechlorination. Methanogensis and hydrogen
production were insignificant sinks (4 %) for electrons on 10th day electron balance.

The other five replicate bottles exhibited similar results. The fraction of lactate
equivalents used in dechlorination varied from 12-28 % and averaged 24 % among 5
bottles. The electron equivalents for dehalogenation were calculated assuming complete
dechlorination from PCE to ethene requires 8 equiv/mol and lactate breaks down to
propionate and acetate. Biomass increment was ignored and CO2 change during this 10
day was negligible. Negative and positive values represent consumption and production

during 10 days.

98

Table 6.4. Electron balance for a single, PCE degrading microcosm (15.2 mol PCE

/bottle, 0.5 mM, lactate; data obtained at day 10)

 

 

Amount Reducing Electron Electron
present Equivalents equiv. equiv. Percenta
Compound
(umol) (eq/mol) consumed formed (%)
(ueq) (ueq)
Lactate -102 4 -408
Propionate 13 6 78 19
Acetate 22 8 17 6 43
CH4 0.2 8 1.6 0.4
H2 < 0.01 2 0.02 < 0.005
PCE ~14 0
TCE 0 2
DCE isomers 3.2 4 12.8 3
VC 11.5 6 69 17
Ethene 0 8

 

aPercent calculated by dividing reducing equivalents produced over total reducing
equivalents consumed.

99

 

6.4.2.3 . Utilization of Chlorinated Ethenes

Low lactate/PCE enrichments which showed complete degradation of PCB were
employed to determine the ability to degrade other chlorinated compounds. In addition,
degradation rates were compared with bottles enriched with TCE and c-DCE. The
transformation rates observed in these experiments are summarized in Table 6.4.

The enrichments usually required 5-15 days to completely degrade PCE depending on the
bottles and concentration of parent products. There was stoichiometric conversion of
PCE to VC, and VC dechlorination occurred only when all other chlorinated
intermediates were nearly consumed. Figure 6.5 shows the sequence and duration of
intermediates produced by high concentration of PCE dechlorination. The productions of
TCE and c-DCE were observed as PCE degraded. Cis-dichloroethene was completely
dechlorinated to VC in approximately 20 days.

When high concentrations of PCE or TCE were received, some bottles produced
t-DCE as intermediate, which persisted longer than c-DCE. When t-DCE alone was
added in the bottles, its rate of degradation was again slower than c-DCE. VC production
from t-DCE was slow and transformation of VC was substantially slower than any other
chloroethenes tested. Utilization of 1,1-DCE by the culture was similar to c-DCE in that
conversion from VC to ethene only occurred after depletion of 1,1-DCE (Figure 6.6).
When VC was added alone, it was converted to ethene with no lag period.

All these tests were done in the bottles that were previously enriched with
repeated PCE feedings over a 2-year period. However, the disappearance of TCE and c-
DCE from TCE, c-DCE incubated bottles did not show any increase in rate of

transformation compared to PCE enriched bottles. For example, the rates of

100

transformation of c-DCE in PCE enriched bottles were not significantly different from

that of c-DCE enriched bottles (Table 6.5).

Figure 6.5. Transformation of PCB and intermediate productions from lactate/PCE
enrichments received high dosage of PCE (80 umol).

100 e
80 —

604

  
 

amol/bottle
A
O

20-

 

Time (days)

+ PCE -u- TCE -------- c-DCE -I- t-DCE ~>K~ VC ~0— Ethene

101

Figure 6.6. Conversion of chlorinated ethenes to ethene by low lactate/PCE enrichment.
This experiment was done with the microcosms showing highest activity. Approximately
15 umol of chlorinated ethenes and 0.5 mM of lactate were used: (a) PCE; (b) TCE; (c)
c-DCE; (d) t-DCE; (e) 1,1-DCE; and (0 VC. Error bars represent the standard deviation
of triplicates.

 

 

(a) PCE

d)

E

0

fl

0

E

3;

-5 Tune (days)
-0- PCE mi»— TCE + c-DCEwir/iw VC 916- Ethene

(b) TCE

 

 

0 7 14 21
Time (days)

+ TCE W“: c-DCE '1'- VC ”We”: Ethene

 

102

(c) c-DCE

    

0 6 12 l 8

'5 ‘ Time (days)

 

+ c-DCE -~I-- VC Ethene

(d) t-DCE

“mol/bottle

 

 

0 1 5 3 0 45
Time (days)

+t-DCE +VC - Ethene

103

 

(e) 1,1-DCE

20 ‘ +1,1-DCE
wan-vc
15 - dwEthene

 

 

 

 

0 7 14 21
Time (days)
20
(f) VC 1
15
0
g 1
Q10 —
E t
:r --o— VC
5 _ -I-- Ethene
0 r - 1 3—9
0 6 12 18

Time (days)

104

 

Table 6.5. Summary of dechlorination rate observed by low lactate/PCE, low
lactate/T CE, and low lactate/c-DCE enrichments (/day)“’b

 

Compound PCE Enriched TCE Enriched c-DCE Enriched

 

PCE 0.82 a: 0.07 N/Ac N/A°
TCE 0.69 a 0.08 0.76 a: 0.02 N/A°
c-DCE 0.66 2: 0.04 0.70 a. 0.04d 0.67 :2 0.03
t-DCE 0.19 i 0.06 N/P° N/Ac
1,1-DCE 0.45 :t 0.04 N/P° N/Ac
VC 0.29 d: 0.03 0.292 0.03 d 0.31 :t 0.04‘1

 

“ Based on first-order rate equation C=Coe"“. Microcosms had 15 cycles of respking.
b Experiments were conducted in triplicates (i indicates standard deviation from
triplicates).

cN/A = Not Available.

dSome compounds like c-DCE and VC were produced as daughter products.
°Some compounds are NP (Not Present or Not Produced).

6.4.2.4. Metabolism of chlorinated ethanes.

The enrichments (pulse fed TCA for 12 month) also transformed 1,1-DCA and
TCA in a rather different way. Some of the bottles received TCA did not show any signs
of degradation. Some bottles showed rapid transformation after a l-month lag phase.
Only bottles that showed TCA transformation degraded 1,1-DCA. 1,1-DCA had been
found to be the main product of TCA biotransformation. Chloroethane was the final

product and no ethane was observed over 2-month period (Table 6.6).

105

Figure 6.7. Conversion of chlorinated ethanes by low lactate/T CA enrichment.
Individual serum bottles have received 15 pmol of chlorinated ethanes and lactate of 0.5

mM: (a) TCA and (b) 1,1-DCA. Error bars represent the standard deviation of triplicates.

 

 

 

      
  

 

 

(a) TCA
20
+TCA
+ 1,1-DCA
15 +CA
O
8 x-
%10 - Fax“
8 N»-
a
5 — f.
/rr"’/
0 K I I F r .
0 9 18 27 36 45
Time (days)
(b) 1,1-DCA
20 ~
-+— 1,1-DCA
15 if .4
.243- ---------
a) _ flflflflflflfl %fl,...
.1310 — ”””” £1
% /
8
:3.
5 — ,.
0 r“ 1 . . . Y
0 9 18 27 36 45

106

Table 6.6. Effects of prolonged exposure of TCA on stimulation of transformation

activity (after 12 month of incubation).

 

 

Transformation rate Half life Final product
(l/day)ab (daY)°
TCA 0.23 i 0.03 3.0 1,1-DCA and CA
1,1-DCA 0.12 at 0.02 3.7 CA

 

“ Based on first-order rate equation C=Coe'k‘.

bExperiments were conducted in quadruplicates (i indicates standard deviation from
quadruplicates).

c11/2 = ln Z/k.

6.4.2.5. Effect of temperature

The microcosms spiked with 6 umol/bottle of PCE were kept in stationary and dark place
at different temperatures (15 °C, 25 °C, and 35 °C). The effects of temperature on PCE
transformation were monitored over 3 -week periods. As can be seen from Figure 6.8,
all microcosms exhibited similar rates of PCB transformation. However, the byproduct
disappearance was difference among treatments. The ethene production from 25 °C
started at day 10, when others started at day 14. In additions, cis-DCE persisted longer at

low temperature (15 °C) than others.

107

Figure 6.8. Temperature effects on PCE transformation on low lactate/PCE enrichment.
Individual serum bottles have received 6 umol of PCB and lactate of 0.5 mM.

Error bars represent standard deviation of triplicates.

15 °C
+ PCE
"I" TCE
+ c-DCE

”)6“ VC
--)1(--- Ethene

umollbottle

 

 

25 °C

+ PCE
«aw TCE
~A-~ c-DCE
mam VC
-)K- Ethene

 

 

35 °C
+ PCE
-I— TCE
~*- c-DCE
--\-- VC
--)K.- Ethene

 

 

 

108

6.4.2.6. Did it go all the way to ethane?

Reduction of ethene to ethane has been previously demonstrated under
methanogenic condition by enrichment (deBruin et a]. 1992). In order to examine
whether the enriched mixed community can produce ethane from ethene, 8 bottles were
injected with PCB and sampled every 10 days. Ethane production was observed from
ethene after 60 days. Utilization of ethene by the culture was similar to the rest of the
chloroethenes, in that the conversion from ethene to ethane only occurred after depletion
of all chlorinated parent and intermediate products (Figure 6.9). However, some bottles
exhibited very slow reduction of ethene to ethane. As can be seen from high standard
deviation of total ethane produced, the degree of reduction from ethene to ethane varied

between microcosms. Mass balance of PCE to ethane is shown in Table 6.7.

Table 6.7 . Mass balance determined for PCE transformation to ethane after 60 days

 

Total PCE added Total Ethene produced Total Ethane produced Conversion
(umol i SD)a (umol i SD) (umol :1: SD) (%)

 

13.15:l:1.15 0.69i1.24 10.10i1.36 82%

 

“ SD — standard deviation from 6 out of 8 microcosms tested. The rest 2 bottles
had very slow ethane production that they were not included in the calculations.

109

 

r l '.. ‘F‘

Figure 6.9. Final PCE transformation byproducts by low lactate/PCE enrichments.
6 out of 8 microcosms tested. The rest 2 bottles had very slow ethane production that
they were not included in this figure. Individual serum bottles have received 15 umol of

PCB and lactate of 0.5 mM.

“mol/bottle

 

 

 

110

6.4.3. Microbial Community Analysis

After extracting DNA from 3—ml samples of sediment groundwater slurry of pre-
enriched, low, and high lactate enrichment, PCR is amplified for T-RFLP analysis and
Real time PCR (RTm-PCR) for the quantification of dechlorinating populations. T-RF LP
was used to determine if the microbial community had been changed over prolonged
exposure to chlorinated compound loadings and lactate. Community shifts were
indicated by changing peak patterns, i.e. the terminal restriction fragments (TRFs), on the
electropherograms. Since it is possible for different organisms to share a common
restriction site in the 168 genes, only a single dominant peak on the electropherogram
was considered for that fragment.

Using digestions enzyme HhaI, pre-enrichment and enrichments showed
differences in fragments numbers. The patterns show that the numbers of terminal
restriction fragments (TRFs) detected in the enriched ones have significantly increased
from those in the pre—enrichment samples. All low lactate chloroethene enrichment
(PCE, TCE, and c-DCE) had similar fragments and overall bacterial community
structure, indicating that the dominant indigenous populations were similar to each other
(Figure 6.10). The TCA and high lactate enrichments had different patterns from low
chloroethene enrichments in terms of terminal restriction fragments (TRFs) sizes
indicating change in overall bacterial community structure.

T-RFLP profile fail to show a single fragment that could represent as
Dehalococcoides specific fragment (List of the T-RFLP fragments and peak areas using

HhaI are shown on Appendix D).

111

Figure 6.10. T-RFLP profiles generated using pre—enrichment, low lactate/PCE, and high
lactate/PCE enrichments using HhaI. Each fragment size indicates percent fragment
area: (a) pre-enrichment; (b) high lactate/PCE; (c) low lactate/PCE; (d) low lactate/c-
DCE; (e) low lactate/TCE; and (0 low lactate/1‘ CA

(a) Pre-enrichment

25-

% Total Area

 

 

207 650 700 900 950 999
Fragment Size (bp)

(b) High Lactate/PCE Enrichment

% Total Area

 

   

       

-.- :1
- ->u5 W5 a

450 475 500 550 600 650 700 750 801 850 899 950 1000
Fragment size (bp)

112

(0) Low Lactate/PCE Enrichment

 

 

228 400 450 475 500 550 600 651 700 751 800 850 900 950
Fragment size (bp)

((1) Low Lactate/T CE Enrichment

12-

% Total Area

 

126 219 250 351 450 475 501 551
Fragment size (bp)

 

113

(e) Low Lactate/c-DCE Enrichment

% Total Area

 

 

650 701 750 800 901 950 1000

475 551 600
Fragment size (bp)

(f) Low Lactate/T CA Enrichment

12-

% Total Area

   

 

 

369 450 475 500 550 600 648 650 700 751 850 900 951

Fragment size (bp)

114

Compared to pre-enrichment bottles which exhibited non-quantifiable amounts of

Dehalococcoides, greater numbers of Dehalococcoides were observed in enriched

bottles. As described in Table 6.7, the enrichment by low lactate/PCE treatment

contained at least 160 times higher numbers of Dehalococcoides populations than the

high lactate/PCE enrichments. The differences in total DNA for low chlorinated

compounds were small. However, the differences in Dehalococcoides concentration

were substantial. TCA enrichment did not have any detectable Dehalococcoides. The

quantitative estimates on total DNA and Dehalococcoides 16S rRNA gene copies are

summarized in Table 6.8.

Table 6.8. Quantitative Estimation on Total DNA and Dehalococcoides populations in
pre—enrichment, Low/PCE, High/PCE, c-DCE, TCE, and TCA enriched bottles using

 

 

PicoGreen and RTm—PCRa
Total DNA Dehalococcoides l6S rRNA gene
(ng/ml) (pg/ml) (copies/ml) b

Pre—enrichment 85.33 at 16.42 NQ° NRd
High lactate/PCE 186.28 i 64.25 17° [1.0 a 0.78] x 103
Low lactate/PCE 129.60 :1: 77.41 256.8 [1.6 a 0.11] x 105
Low lactate/c-DCE 130.57 at 47.33 195.3 [1.2 a 0.57] x 105
Low lactate/TCE 78.76 a 4.25 118.1 [7.2 :1: 0.40] x 104

Low lactate/T CA 90.59 i 15.70 NDf NRd

aTriplicate samples of slurry extracted DNA were used except for the high lactate/PCE
enrichment. Three PCR reactions were run for each sample point.
"16 s rRNA gene copies/ml = DNA (ug/ml)x6.023x1023/(1.5x106x660x106) from He et al. (16)
“N Q = detectable but Not Quantifiable by RTm-PCR.

NR = Not Reported due to the absence of data

c4 Out of 8 results were used in calculated due to the high variation of DNA between samples.

‘ND = Not Detectable by RTm-PCR.

115

6.5. Conclusion and Discussion

After 2 years of pulse feedings, once-fed microcosms (microcosms from chapter
5), which initially exhibited slow degradation, showed much higher rates of PCB
dechlorination. The high lactate/PCE treatments exhibited improved rates of PCE
dechlorination and produced TCE and c-DCE as final byproducts. The low lactate/PCE
treatments were capable of sequentially dechlorinating chlorinated ethenes to ethene, and
then ethane as final byproduct. These results provide evidence that the ratio between
electron acceptor and donor is important. Even though 16S rRNA gene targeted primers
revealed the presence of PCE-dechlorinating Dehalococcoides in high lactate
enrichments, none of the laboratory microcosms completely reduced PCE to ethene.
Possible reasons for incomplete dechlorination could be that the high nutrient
environment resulted in elevated H2 levels, which stimulated methanogenic populations,
and resulted in limited growth of halorespirating populations.

In this study, 100-umol lactate (electron donor) and 15 umol of PCE were used.
This is 10 times the minimum reducing equivalents necessary for a complete reduction of
the 15 umol PCE provided to ethene, and is similar to the amounts supplied in other
studies (Freedman and Gossett, 1989; Bagley and Gossett, 1990; Vogel and McCarty,
1985). It is not clear from our study whether a smaller amount of lactate or other electron
donors would have been a better candidate for even faster and complete dechlorination.

The highest rate of transformation was 0.82/day using PCE as parent product.
VC, 1,1-DCE, and even t-DCE were dechlorinated by the PCB enrichment, but

degradation rates were slow compared to PCE. Transformation of chlorinated ethenes

116

did not show specificity over what compound it was enrich with. TCE dechlorination by
TCE enrichment have similar rate as TCE dechlorination by PCE enrichments.

Lowering the ratio between substrate to PCE, along with pulse feedings has
resulted in substantial enrichment of Dehalococcoides. Dehalococcoides populations
were detected in all of the enriched samples by RTm-PCR On the contrary, the once fed
long-term microcosms (from chapter 5) and TCA enrichments did not show detectable
amounts of Dehalococcoides. Distinct differences in Dehalococcoides DNA levels were
observed in low lactate and high lactate enrichments. The low lactate/PCE contained at
least 160 times higher Dehalococcoides numbers than high lactate/PCE bottles,
supporting the hypothesis that the Dehalococcoides cells compete better in a low rather
than high nutrient environment.

T-RFLP was used to generate profiles of the bacterial communities in chlorinated
compound enrichments and pre-enrichment. The number of fragments was consistently
higher after enrichment. The detection of greater number of terminal fragments in
chlorinated enrichment is an indication that a relatively higher bacterial diversity exists
after enrichment on lactate/PCE.

The profiles generated by T-RF LP can vary in two ways. First, there can be
variation in the number and size (in basepairs) of terminal restriction fragments (T-RFs)
present in profile. Secondly, variation can be found in the height (and consequently the

area) of any particular peak (Osborn et a]. 2000). However, T-RFLP does not resolve
closely related organisms or organisms that happen to share the same restriction sites and

hence yielding the same TRF. Therefore, a single peak in the electropherogram may

r epr esent more than one organism.

117

A single common fragment present in all of the enriched microcosms (not found
in pre-enrichment) by T-RF LP was not found. This may have resulted since these
populations are a small percentage of the total community and only small sample amount
(3 ml) was used to extract the DNA that it is not sufficient to represent whole microbial
system in the bottle.

Even though CO2 and methane have also been observed as dechlorination
products under methanogenic conditions (V ogel and McCarty, 1985; Bradley and
Chapelle, 1999), most biotransformation results in the accumulation of partially
dechlorinated ethenes such as c-DCE or VC. Complete dechlorinations of chloroethene
are known to produce ethene as a frnal transformation product (Maymo-Gatell, 1995).
Despite the reports of ethene persistence in the subsurface (Oremland, 1981; Schink,

1985), ethene reduction was occurred predominant in low lactate enriched conditions.
DeBruin (1992) also reported that ethane could be the final product of dechlorination in
mixed community using lactate as an electron donor. These often-observed diverse
transformations byproducts emphasize the need for better understanding of the
underlying microbiological process involved.

Low temperature (15 °C) had no significant effects on transformation of PCE.
Transformation of PCE was highest at room temperature instead of 35 °C. This might
have resulted since the bottles have been enriched in room temperature over 2 years. The
high and complete transformation of chlorinated compound, and the absence of
chlorinated end products like VC, makes bioremediation an attractive method for the

removal of chlorinated compounds at the plume G site.

118

It is not known which organisms are responsible for the reduction of ethene to
ethane in our study. It is also not clear whether that same organisms are also responsible
for TCA transformation since ethane production from CA was not observed. Since they
were pulse fed with a different compound over an extended period, we suspect both
communities are very different. More research will be required in the fiJture to backup

this assumption.

119

6.6. Literature cited

(a) Fathepure, B. Z. and Boyd, S. A. 1988. Dependence of tetrachloroethylene
dechlorination on methanogenic substrate consumption by Methanosarcina Sp. Strain
Dem. Appl. Environ. Microbiol. 54(12): 2976-2980.

(b) Fathepure, B. Z. and Boyd, S. A. 1988. Reductive dechlorination of
perchloroethylene and the role of methanogens. FEMS Microbiology Letters. 49(2): 149-
156.

Freedman, D. L. and J. M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions.
Appl. Environ. Microbiol. 55(9): 2144-215 1.

Maymo-Gatell, X., Y-T, Chien, J.Gossett, and SH. Zinder. 1997. Isolation of a bacterium
that reductively dechlorinates tetrachloroethene to ethene. Science 276;]568-1571.

Hendrickson, E.R. et a]. 2002. Molecular analysis of Dehalococcoides 16S Ribosomal
DNA from chloroethene-contaminated sites throughout North America and Europe. Appl
Environ. Microbiol. 68 (2) 485-495.

Liu, Wen-Tso, Marsh T.L., Hans Cheng, and Larry J. Forney. 1997. Characterization of
microbial diversity by determining Terminal Restriction Fragment Length
Polymorphisms of genes encoding l6S rRNA. Appl. Environ. A/Iicrobiol. 63 (l 1) 4516-

4522.

Schink, B. 1985. Inhibition of methanogensis by ethylene and other unsaturated
hydrocarbons. FEMS Microbiol. Ecol. 31:63-69

Debruin, W. P., Kotterman, M. J. J ., Posthumus, M. A., Schraa, G., and A. J. B. Zehnder,
1992. Complete biological reductive transformation tetrachloroethene to ethane. Appl

Environ. Microbiol. 58(6): 1996-2000.

Bagley, D. M. and J. M. Gossett. 1990. Tetrachloroethene transformation to
trichloroethene and cis-l,2- dichloroethene by sulfate-reducing enrichment cultures.
Appl. Environ. Micro]. 56(8): 251 1-2516.

Vogel T. M. and P. L. McCarty. 1985. Biotransformation of tetrachloroethylene to
trichloroethylene, dichloroethylene, vinyl chloride, and carbon dioxide under
methanogenic conditions. Appl. Environ. Microbiol. 49(5): 1080-1083.

Osborn, A. M., E. R. B. Moore and K. N. Timmis. 2000. An evaluation of terminal-
restriction fragment length polymorphism (T-RF LP) analysis for the study of microbial
community structure and dynamics. Environm. Microbiol. 2: 39-50.

120

Bradley, P. M. and F. H. Chapelle. 1999. Methane as a product of chloroethene
biodegradation under methanogenic conditions. Environ. Sci. T echno]. 33(4): 653-656

Maymo-Gatell, X., Tandoi, V., Gossett, J. M., and, S. H. Zinder. 1995. Characterization
of an H2 utilizing enrichment culture that reductively dechlorinates tetrachloroethene to
vinyl chloride and ethene in the absence of methanogenesis and acetogenesis. Appl.
Environ. Microbiol 61(1 1): 3 92 8-3 933 .

Oremland, RS. 1981. Microbial formation of ethane in anoxic estuarine sediments. Appl.
Environ. Microbiol 42: 122-129.

He, Jianzhong, Kirsti M. Ritalahti, Michael R Aiello, and Frank Loffler. 2003. Complete
detoxicification of vinyl chloride by an anaerobic enrichment culture and identification of
the reductively dechlorinating population as a Dehalococcoides species, Appl. Environ.
Microbiol 69(2): 996- 1003 .

121

Chapter Seven: CHARACTERIZATION OF DECHLORINATIN G
COWUNITY BY METABOLIC INHIBITORS

7.1. Abstract

Metabolic inhibitors were used to characterize the anaerobic enrichment, which
dechlorinates chloroethene to ethene/ethane using lactate as an electron donor. The
effects of BES, molybdate, vancomycin, and sulfate on these enrichments were evaluated
over 18 days. While the inhibitory effects depended on experimental concentrations,
dechlorination was not hindered by molybdate (inhibitor of sulfate-reducer), but was
inhibited by high levels of sulfate. High levels of sulfate stimulated PCE to c-DCE
reduction, but not the complete and rapid dechlorination. Therefore, sulfate reducers
appear to be not involved in the observed rapid and complete dechlorination. Low levels
of BES (5 mM; methanogen inhibitor) specifically inhibited methane production, but the
cultures retained dechlorination activity, indicating that the responsible microorganism(s)
is (are) non-methanogenic. High concentration of BES (50 mM) inhibited the on-set of
dechlorination and dechlorination stopped at c-DCE. The decrease in Dehalococcoides
DNA in the presence of high levels of molybdate and BES suggests non-specific toxic
effects. High levels of vancomycin (100 mg/L; acetogens inhibitors) successfully
inhibited acetate production from lactate. The reduction in acetate production also
inhibited dechlorination, further suggesting that acetate-utilizing dechlorinating microbes

are responsible for the observed dechlorination.

122

7.2. Introduction

One of the main topics under investigation with regard to the process of reductive
dechlorination is the interaction between microorganisms (methanogens, sulfate reducers,
and acetogens, etc.) inhabiting contaminated groundwater and aquifer solids. To clarify
the: se interactions and the roles of specific microorganisms in dechlorination, specific
inhibitors were used to block the activity of potentially degradative bacteria.
Molybdate is often used as an inhibitor of sulfate reduction (Smith and Klug,
19 81) and vancomycin is used to inhibit acetogenic bacteria since it is an inhibitor of cell
wall synthesis in gram-positive eubacteria (Distefano et al 1992). The 2-
bro moethanesulfonic acid (BES) is frequently used to inhibit of methanogenesis
(D istefano et al 1992). Unlike vancomycin, molybdate, and BES, sulfate is not a
sp ecific inhibitor of any microbial physiological group. Sulfate addition usually
stimulates grth of sulfate-reducing bacteria and thus inhibits metabolic activity of

methanogenic bacteria.

The effects of molybdate, BES, vancomycin, and sulfate on PCE dechlorinating

community enriched from mixed chlorinated solvents contaminated sediments was
investigated. The main objective of this study was to characterize the anaerobic
dechlorination activity observed in low lactate/PCE enrichment stimulated by pulse
feedings of chlorinated compounds and lactate over 2 years. High rates of dechlorination

rndic at ed that the transformation observed in the lactate/PCE enrichment was occurring

V13 h'C‘I-lorespiration.

123

“‘0‘ MO

(a) BES (b) Molybdate
HO
me
i 0 0°“ .
O 0 9
F0... 0 H
5042. 0% 0&0 o :11,
1...; .. $.47 .a’p
; 0
We" "1’
m 0"

(c) Sulfate (d) Vancomycin

Figure 7.1. Chemical Structures of inhibitors used in this study. Vancomycin structure
obtained from Doug et a]. (2002).

Two molecular studies were conducted to characterize the effect of inhibitors on
enrichment. Terminal restriction length polymorphism (T-RFLP) analyses were
employ ed to identify community shifts due to the addition of inhibitors. Specific primers

for dete cting Dehalococcoides species were used to determine the effects of inhibitors on

SUCh P O pulations.

124

7.3- Materials and Methods

Cl: etnicals

Bromoethanesfulfonic acid (sodium salt, 98 %), vancomycine hydrochloride (4 %
water content), and molybdate (sodium salt, 99+%) were purchased from Sigma-Aldrich

(1V1 ilwaukee, WI).

In hibitors Studies
Inhibitors were applied to low lactate/PCE vials to investigate the roles of
ac etogens, methanogens, and sulfate-reducers on the onset of dechlorination. The
itlhibitors were fed only once at the beginning of the experiment and their effects on
transformation were monitored for 18 days. Approximately 4 g (wet weight) of
sediments were anaerobically transferred to 40-ml serum vials each and sealed with
screw cap containing 22-mm teflon—lined butyl rubber septa. Only low lactate/PCE
rni crocosms were used in this study. The 30-mL of growth medium was amended with
lactate (200 umol), vitamin solutions, phosphate, and groundwater, as described
previously (Chapter 4).
After transferred to their new 40-ml vials, vials were spiked with 3 umol of
PCE/month for 3 month. Thirty-three vials out of 45, which showed similar
tr 3118f O rmations rates to each other, were selected for this experiment in order to reduce
the variation of PCB transformation between vials. A total of 33 vials were employed: 3
were prepared with no inhibitor, 9 had 3 different levels of BES (0.5, 5, 50 mM), and 6

had two different levels of vancomycin (25 mg and 100 mg), 9 had 3 different levels of

125

molybdate (0.5, 2, 6 mM), and 6 had 2 levels of sulfate (0.6 mM, 2 mM). The inhibitor
concentrations were similar to other reported studies using dechlorinating enrichments for
comparisons (Distefano et a]. 1992; Pavlostathis and Zhuang, 1991;,DeBest et al 1997;

Bagely and Gossett, 1990; dean et al. 1992). Table 7.] depicts the list of inhibitors and

their concentrations used in this chapter.

Prior to each electron donor and PCE addition, the vials were thoroughly purged
with 80% N2- 20% C02. The intent of purging was to prevent accumulations of VC
and/or ethene, which is a known inhibitor of methanogenesis (Schink. 1985). All vials
had 200 umol (5.72 mM) lactate as electron donor before 3 umol (85 11M) PCE addition.
Vancomycin was added directly after purging the vials. Therest of inhibitors (BES,
no lybdate, and sulfate) were bubbled with N2, autoclaved, and then used for inhibition

experiments. Vials were kept in a constant room temperature throughout the course of

th i 5 experiment.

126

Tab le 7.1. The list of inhibitors and their concentrations used in this chapter

 

 

Samples Inhibitor Concentrations Number of
(mM)a bottles
Controlb 0 3
BES 0.5 3
5 3
50 3
Molybdate 0.5 3
2 3
8 3
Sulfate 0.6 3
2 3
Vancomycin 25 mg/L 3
100 mg/L 3

 

aConcentrations were in mM otherwise noted.
bControl had no inhibitor addition.

127

7.4- Results

7.4- 1. Effects of BES

The effects of three different levels of BES on the dechlorination of PCE by low

lactate/PCE enrichments was examined. Earlier microcosm results (Chapter 6) indicated
that methane production was low, suggesting that methanogens do not play significant
role in dechlorination. However, this observation does not completely rule out a role for
methanogens in PCE dechlorination.

The effects of three levels of BES on PCE transformation during 18 days are
depicted in Figure 7.3. Low levels of BES (0.5 mM) amendment showed no immediate
eff‘ects on dechlorination, but 5 mM amendments exhibited delayed and partially
inhibited dechlorination that DCEs and VC accumulated. Even though, 50 mM BES
inhibited PCE dechlorination significantly, it did not stop PCE degradation completely.
P CE transformation was delayed by 50 mM BES and did cause the production and

accumulation of c-DCEs. This indicates that repeated dose of BES may be need to
completely step PCE dechlorination.

Ethene production was observed only by 0.5 mM BES treated vials during this

18-day period and both 0.5 mM BES treated vials and control vials reduced PCE in 6
days that VC and ethene were final byproduct after 18 days (Figure 7 .2 and 7.3). This
result suggests that small dose of BES had no effect on dechlorination.
Methane formation was inhibited by all doses of BES (Table 7 .3). Methane
produCtions from 0.5 mM and 5 mM BES were only 43 and 18 percent of the methane
pr Oduced by control microcosms. This result shows that the residual PCE were probably

110i su‘ITT‘icient to suppress methanogenesis and BES was the one, which inhibited

128

methanogenesis. Hydrogen sulfide, lactate, acetate, and propionate concentrations were
measured for all vials, where possible (Table 7 .3). Even though BES inhibited PCE
transformation, no apparent negative or positive effects on acetogenesis was observed

(Table 7 .3). Methane production in the BES 50 mM was less than 1 % of total
equivalents were measured as methane, however non-specific effects were observed at
this concentration. Sulfide production was not measured in the BES inhibited

mi crocosms due to interference by BES.

+ Ethene
100 — WVC
-—*-- cDCE
ween TCE
% PCE

 

 

0 6 1 2 1 8
Time (days)

Figure 7 .2. The PCE transformation by control.

129

.bao 08w Son 3 mmm mo 28 cm 98 .28 m .28 md we Began 0.53 358832
.mmm he 28 on e5 .28 m 28 no 3 8888382 mofie 38% By .2. 63mm

Ea 8 ma 9 Es m mmm 3v Ea 3 ma 3
a
A33 baa. E63 bag a eg new
M: S e o 2 S e o M: S e o

 

 

 

mop. {kmvmtza a. . m5
mph man—to 1‘1
89.2.4.1. to f U> +
mUQIO E I OO~ m8 Ofiufiwm + f COM

 

o>1..-T t o2
U>i§t

130

7.4- 2. Effects of Molybdate and Sulfate

Two levels of sulfate and three level of molybdate were added to observe their
potential effects on PCE transformation. The effects of molybdate and sulfate on PCE

dechlorination during 18-day period are shown on Figure 7.4 and Figure 7 .5.

Dechlorination was neither inhibited nor improved by low levels of molybdate
ad ditions (0.5 and 2 mM). Both molybdate treated vials and control vials reduced PCE in
3 — 6 days and VC and ethene were final byproducts after 18 days (Table 7.2). High levels
of molybdate (6 mM) amendments showed delayed and partially inhibited dechlorination

that DCEs and VC accumulated. This suggests non-specific toxic effect of molydate on

d e chlorination.
The PCE dechlorination was neither delayed nor inhibited by 0.6 mM sulfate, but

high levels of sulfate (2 mM) significantly reduced dechlorination of PCE that c—DCE

were major byproduct after 18 days.

Molybdate did inhibit sulfate reduction (Figure 7 .6). After 18 days of incubation,

the residual sulfate from groundwater persisted in the microcosms amended with

molydate. In sulfate-amended vials, low levels of sulfate (0.6 mM) were all reduced

within 7 days, but high levels of sulfate (2 mM) persisted over 18 days.

Hydrogen sulfide, lactate, acetate, and propionate concentrations were measured
for all Vials, where possible (Table 7.3). Essentially stoichiometric sulfide production
was observed in the 0.6 mM and 2 mM sulfate treated vials. Methane production was

very low, accounting for only 1 % of the electron equivalents measured.

13]

Both high sulfate and high molybdate inhibited methane production. Lesser
amount of methane was observed by sulfate (0.6 mM and 2 mM) amendments compare
to the molybdates and controls. However, high levels of molybdate (6 mM) did exhibit
inhibitory effect on methanogenesis. As shown in Table 7 .3, low levels of molybdate
(O -5 and 2 mM) did not inhibit methane production while high level inhibited

tn ethanogensis up to 40 percent.

132

.380 08% Son 3 339308 mo 28 o 98 .28 N .28 md as 3258a 803 m8moo€oflz
.3828: we Ea e .8 .E8 N Ea no 3 8886.88 mom .8 38% 65 .I. 28E

E8 e eceBreE 3 Ea N eaBbeE 3V Ee no eeBEoE 3
88V 98. Ages «a: was a:

M: S o o M: NM o o E S o o

 

 

 

 

mom lyml. .
881T - .
m8. . t . mom o 111 mon if:
r U> 1 OS U> o3
mooé ldl o3 285m 101 985m Iol

o>lal

133

Figure 7.5. The effects of PCE transformation by 0.6 mM and 2 mM of sulfate.

Microcosms were amended with sulfate at time zero only.

100 a -O— Ethene
-*- c-DCE
mew» TCE

 

 

(a) 0.6 mM Sulfate
0 6 12 18
Time (days)

100 — "*"VC
+c-DCE
we» TCE
+PCE

(b) 2 mM Sulfate

 

 

0 6 12 18

134

300 -

+ 2 mM
200 sulfate

100 .

 

 

\]
(It

 

 

 

Sulfate (ppm)

 

 

O l T 7‘ [TR

Tune (days)

+6mMmolybdate will~-~2mMmolybdate
- ~ 0.5 mM molybdate ->(- 0.6 mM sulfate

Figure 7 .6. Sulfate reduction and effect of molybdate on residual sulfate in microcosms.

The reduction of high sulfate (2 mM) is shown on the bigger scale.

135

7.4.3. Effect of Vancomycin

Two levels of vancomycin were added to lactate/PCE enrichments to observe its
potential effects on PCE transformation. On the basis of the electron balance from
Chapter 6, acetogenesis from lactate was the predominant pathway, and the role of
acetogens in the PCB dechlorination must be considered. Results for vancomycin
(acetogenesis inhibitor)-amended bottles are shown in Figure 7.6. Figure 7.7 indicates
that 25 mg/L of vancomycin had no significant effect on transformation while 100 mg/L
of vancomycin delayed the transformation of PCE that VC and c-DCE was observed as
final byproduct during 18 days.

Table 7 .3. shows only low acetate production, demonstrating the inhibitory effect
of vancomycin on acetogens. Methane productions from vancomycin treated vials were
also smaller than controls.

The decline in ethene production and the detection of VC and DCEs by higher
concentration of vancomycin indicated that acetogens are indirectly involved in PCE
dechlorination. Lactate metabolism and the resultant products are shown in Table 7 .3.
Acetogenesis accounted for almost 50 percent of electron donor used for controls, and

reduced products from PCE dechlorination represented nearly 10 percent.

136

100 - +Ethene
w“WVC

(a) 25 mg/L + c-DCE

 

 

0 6 12 18
Time (days)
100 - + VC
«qt-- c-DCE
"Mic“ TCE
75 2 "7'16— PCE
(b) 100 mg/L >

 

 

0 6 12 18

Figure 7.7. The effects of PCE transformation by 25 mg/L and 100 mg/L of vancomycin.

Microcosms were amended with vancomycin at time zero only.

137

Table 7.2. Effect of inhibitors on reductive dechlorination of PCE after 18 days3
(Initial PCE load was between 2.33 and 2.63 umol/bottle).

 

Amt of product formed (pmol)

 

Inh’b‘tor PCE TCE c-DCE vc Ethene
BES
0.5 mM 0 0 0 0.7 1.8
5 mM 0 0 0.5 1.9 0
50 mM 0 0 2.2 0.3 0
Sulfate
0.6 mM 0 0 0 0.9 1.4
2 mM 0 0 2.3 0.2 0
Vancomycin
25 mg/L 0 0 0 0.1 2.4
100 mg/L 0 0 0.5 1.8 0
Molybdate
0.5 mM 0 0 0 0.6 1.6
2 mM 0 0 0 1.0 1.5
6 mM 0 0 0 0.5 2.0
No
inhibitors 0 0 0 1. 1 l .2

 

aLactate (200 mol) was supplied as electron donor. The data are means obtained for
triplicates samples after 18 days. The standard deviation was < i 0.2.

138

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139

7.4.4. Microbial Community Analysis

After extracting DNA from 3-ml sediment slurry of microcosms treated with
inhibitors for 18 days, PCR is amplified for T-RFLP analysis and Real time PCR was
conducted for quantification of dechlorinating populations. T-RFLP was used to
determine if the microbial community had been changed by the exposure to inhibitors.
Changes in peak patterns and number of fragments can indicate community shifts. Since
it is possible for different organisms to share a common restriction site in the 16S genes,
only a single dominant peak on the electropherogram was considered for that fragment.

Not all T-RFLP produced usable patterns for inhibitor study. Many of them were
over digested or not digested enough to represent the effect of inhibitors on microbial
communities. Using digestions with the enzyme Hhal, inhibitors produced different
patterns compare with controls. Each inhibit treated microcosms had different patterns
from control in terms of terminal restriction fragments (TRFs) sizes indicating change in
overall bacterial community structure (Figure 7.8). List of the T-RFLP fragments and

peak areas using HhaI are shown on Appendix F.

140

Figure 7 .8. T-RFLP profiles generated by using dechlorinating enrichments treated with
inhibitors. Each fragment size indicates percent fragment area: (a) BES (0.5 mM); (b)
molybdate (2 mM); (c) sulfate (0.6 mg/L); and (d) vancomycin (25 mg/L).

(a) BES

._- N N
Ur 0 tn
1 J

% Total Area
8

 

 

176 268 348 393 475 476 600 651 685 950
Fragment size (bp)

(b) Molybdate

% Total Area

 

 

334 476 500 600 651 700 751 850 900 950 999
Fragment size (bp)

00

15

141

(c) Sulfate

16 -

% Total Area
00 ii?

A

 

 

185 217 464 475 500 600 649 850 900 951
Fragment size (bp)

((1) Vancomycin

12]

oo
1

% Total Area

 

 

207 450 476 492 501 550 601 650 700 750 800 850 900 950
Fragment size (bp)

142

Microcosms amended with inhibitors such as BES showed significant decay in
the total populations and Dehalococcoides populations at high levels (Table 7 .4). These
high levels (50 mM BES) exhibited non-specific toxic effect, which led to a 60 %
decrease in total DNA. The 16S rRNA demonstrated a decrease in Dehalococcoides
populations due to the high levels of BES. High levels of molybdate (6 mM) also
exhibited non-specific toxic effects in that both total DNA and Dehalococcoides levels
declined.

The low levels of sulfate and vancomycin did not significantly alter the total DNA
levels and Dehalococcoides populations. While 25 mg/L vancomycin exhibited no
effects on both DNA and Dehalococcoides yield, less total DNA and Dehalococcoides

were observed at high levels (100 mg/L).

143

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144

7.5. Conclusion and Discussion

Characterization of the dechlorination activities in an undefined microbial
community is an important step in the identification of the responsible microbial
populations. Information from these studies can provide a basis for further enrichment,
and may facilitate the identification and isolation of microorganisms capable of
dechlorination. The microbial group specific inhibitors can sometimes be used to
distinguish which group is responsible for a specific activity (Oremland and Capone,
1988)

To investigate the role of methanogenic, acetogenic and sulfate-reducing
microorganisms on the transformation of PCB, 4 inhibitors; BES, molybdate, sulfate, and
vancomycin were used. Although methane production was low in lactate/PCE
enrichments, this fact alone does not rule out a role for methanogens in PCE reduction.
BES, considered a selective inhibitor of methyl coenzyme M reductase, the enzyme that
catalyzes the final step in methanogensis (Gunsalus et a1. 1978), was therefore used to
inhibit methanogenesis in the enrichments. Adding BES eliminated methanogenic
activity, which plays an important role in anaerobic microbial processes (Bhatnagar et al.
1991; Zeikus, 1977), but a partial dechlorination activity was still remained. Repeated
additions of BES might have been required in order to completely stop dechlorination,
possibly because microbes can degrade BES. Schink (1985) has noted this is a draw back
to the use of BES.

Many puzzling (diverse) results have been published regarding the use of BES.
Previous works have demonstrated that methanogens can dechlorinate PCE (F athepure et

al. 1987 ; Fathepure and Boyd, 1988). However, when methanogenesis stops, PCE

145

dechlorination by methanogens also stops. In previous research conducted by Freedman
and Gossett (Freedman and Gossett, 1989), 5 mM BES was found to be sufficient for the
complete cessation of methane production and TCE dechlorination by mixed
methanogenic cultures.

In the case of mixed sulfate-reducing cultures spiked with 50 mM BES, methane
was not produced and PCB dechlorination declined by approximately 25 % as compared
to non-inhibited cultures (Bagley and Gossett, 1990). Lofiler et al. (1997) showed that
BES (2 mM) inhibited dechlorination in the absence of methanogens, indicating BES
results should be interpreted with caution.

From our study, BES inhibited dechlorination beyond DCEs when added at 50
mM. Methane production in the presence of 50 mM BES was less than 1 % of the total
measured electron equivalents, indicating very little methanogenic activity. However,
PCE dechlorination was not completely stopped. These observations deviate
substantially from the pattern of PCE dechlorination observed with methanogenic
systems. Therefore, it appears probable that dechlorinators are responsible for the
observed dechlorination.

Lactate/PCE enrichment also had high acetate production that a majority of
reducing equivalents was accounted for acetate production with the remainder being
accounted for by PCE reduction (refer to Chapter 6). Vancomycin effectively inhibited
acetogenesis in lactate fed bottles in every concentration. Vancomycin is a eubacterial
peptidoglycan synthesis inhibitor, (is bacteriocidal,) and is more effective against gram-
positive than gram-negative eubacteria (Bock and Kandler, 1985; J oklik et al. 1984). It is

unlikely that vancomycin directly inhibited acetogenesis in the microcosms but rather the

146

 

- I" .—m-—_

inhibition was probably due to the bacteriocidal effects of vancomycin on acetogens,
most of which are gram positive (Ljungdahl, 1986). In contrast to 100 mg/L—amended
bottles, 25 mg/L amended bottles continued to dechlorinate well after the inhibition of
acetogenesis.

The inhibition of PCB dechlorination by the high levels of vancomycin (100
mg/L) suggests that PCE dechlorinator could not use lactate directly but rather required
lactate metabolism, most likely acetate (or H2) as the electron donor for reductive
dechlorination. Figure 7.9 shows a hypothetical model of the lactate metabolism and
PCB transformation by the enrichment, which includes the hypothesis that acetate and H2
as the actual electron donor for PCE dechlorination.

Control (no inhibitors) performed similarly with respect to electron donor use;
acetate production accounted for a majority of reducing equivalents, with the remainder

being accounted for by PCE reduction and methane production.

Molybdate is generally regarded as a specific inhibitor of sulfate-reducing
bacteria (SRB) (Taylor and Oremland, 1979) and thus widely used in microbiology
studies due to its specificity (Oremland and Capone, 1988). Molybdate is an analog that
compete with sulfate for the active site of ATP sulfurylase, resulting in formation of an
unstable analog-AMP complex, which readily hydrolyzes to AMP and the sulfate analog;
the latter is then available to again react with ATP sulfierlase (Wilson and Bandurski,
195 8). Repetition of these events depletes intracellular ATP, thereby halting growth of

the bacteria and, as a result, inhibiting sulfate reduction.

Even though sulfate reduction was efficiently prevented by small amounts of

molybdate (0.5 and 2 mM), no inhibition on PCE dechlorination was observed. The

147

inhibition of PCB dechlorination by high molybdate concentrations might have been
resulted by a toxic effect since it can bind free sulfide ions to form a molybdosulfide
complex (Tonsager and Averill, 1980), and this could influence microorganisms
requiring sulfide ions (Oremland and Capone, 1988) e. g. methanogens (Bhatnagar et al.
1991; Zeikus, 1977). The methane production from 6 mM molybdate treated vials was

less than 0.5 mM or 2 mM, which confirms the toxic effects on methanogens.

T-RF LP was used to discern the microbial communities’s response to the
inhibitor treatments. The restriction enzyme digest using the restriction enzymes Hhal
was used to obtain a T-RF LP community fingerprint from the extracted DNA of the
microcosm sample. Profiles of the bacterial communities showed extensive shifis due to

inhibitors in terms of terminal restriction fiagments (TRFs) numbers/sizes.

Figure 7 .9. Hypothetical model based on DiStefano et al. (1992), for carbon and electron

flow in a lactate-PCE enrichment

Lactate

  

Acetogens

  

 

 

Dechlorinating organisms 4 HCl
Acetate
,1 / \
/,/ PCE Ethene
/'/
o/.
H2 / ’ CH4

Methano gens

148

7.6. Literature cited

Smith, LR. and M.J. Klug. 1981. Electron donors utilized by sulfate-reducing bacteria in
eutrophic lake sediment. Appl. Environ. Microbial. 42:116-121

Distefano, T. D., Gossett, J. M., and S. H. Zinder. 1992. Hydrogen as an electron donor
for dechlorination of tetrachloroethene by an anaerobic mixed culture. Appl. Environ.
Microbiol. 58(11): 3622-3629.

Steven D. Dong, Markus Oberthur, Heather C. Losey, John W. Anderson, Ulrike S.
Eggert, Mark W. Peczuh, Christopher T. Walsh, and Daniel Kahne. 2002. The structural
basis for induction of vanB resistance. Am. Chem. Soc., 124 (31), 9064 -9065

Pavlostathis, S.G. and P. Zhuang. 1991. Transformation of trichloroethylene by sulfate-
reducing cultures enriched from a contaminated subsurface soil. Appl. Microbial.
Biotech. 36 (3): 416-420.

DeBest, J. H., Jongema, H.,Weij1ing. A., Doddema, H. J., Janssen.,D. B. and W. Harder.
1997. Transformation of 1,1,1-trichloroethane in an anaerobic packed-bed reactor at
various concentrations of 1,1,1-trichloroethane, acetate and sulfate. Applied Microbiology
and Biotechnology. 48 (3 ): 417-423 .

Bagley, D. M. and J. M. Gossett. 1990. Tetrachloroethene transformation to
trichloroethene and cis-l,2-dichloroethene by sulfate-reducing enrichment cultures. Appl.
Environ. Microbial. 56(8): 2511-2516.

DeBruin, W. P., Kotterman, M.J. J, Posthumus, M.A., Schraa, G., and A.J.B. Zehnder
1992. Complete biological reductive transformation of tetrachloroethene to ethane. Appl.
Environ. Microbiol. 58(6) 1996-2000.

Schink, B. 1985. Inhibiton of methanogensis by ethylene and other unsaturated
hydrocarbons. FEMS Microbiol. Ecol. 31:63-69

Ballapragada, B. S., Stensel, H. D., Puhakka, J. A., and J. F. Ferguson. 1997. Effect of
hydrogen on reductive dechlorination of chlorinated ethenes. Environ. Sci. T echnal.
31(6): 1728-1734.

Fennell, D. E., Gossett, J. M., and, S. H. Zinder. 1997. Comparison of butyric acid,
ethanol, lactic acid, and propionic acid as hydrogen donors for the reductive
dechlorination of tetrachloroethene. Environ. Sci. T echnal. 31(3): 918-926.

He, Jianzhong, Kirsti M. Ritalahti, Michael R. Aiello, and Frank Loffler. 2003. Complete
detoxicification of vinyl chloride by an anaerobic enrichment culture and identification of
the reductively dechlorinating population as a Dehalococcoides species, Appl. Environ.
Microbiol. 69(2): 996- 1 003 .

149

Oremland, RS, and D.G. Capone. 1988. Use of specific inhibitors in biogeochemistry
and microbial ecology. In K.C.Marshall (ed.), Advances in microbial ecology. Vol. 10.
Plenum Publishing Co.

Gunsalus, R. P., J. A. Romesser, and R. S. Wolfe. 1978. Preparation of coenzyme M
analogues and their activity in the methyl coenzyme M reductase system of
Methanabacterium thermotrophicum. Biochemistry 17:23 74-23 7 6

Bhatnagar, L., M.K.Jain, and J .G.Zeikus. 1991. p. 251-270. Methanogenic bacteria. In
variations in autotrophic life. Academic Press, Orlando, FL.

Zeikus, J .G. 1977. The biology of methanogenic bacteria. Bacterial. Rev. 411514-541

Fathepure, B., Nengu, J. P. and SA. Boyd. 1987 . Anaerobic bacteria that dechlorinate
perchloroethene. Appl. Environ. Microbiol. 53(11): 2671-2674.

Fathepure, B. Z. and S. A. Boyd. 1988. Reductive dechlorination of perchloroethylene
and the role of methanogens. FEMS Microbiology Letters. 49(2): 149-156.

Freedman, D. L. and J. M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions.
Appl. Environ. Microbiol. 5 5 (9): 2144-21 5 1 .

Loffler, Frank E., K. M. Ritalahtiand James M. Tiedje. 1997. Dechlorination of
chlorethenes is inhibited by 2 —Bromoethanesulfonate in the absence of methanogens.
Appl. Environ. Microbiol. 63(12): 4982-4985

Back, A. and O. Kandler. 1985. Antibiotic sensitivity of Archaebacteria, p. 525-544. In
C. R Woese and RS. Wolfe (ed.), The bacteria: a treatise on structure and function, vol.
VIII. Archaebacteria. Academic Press, Inc., New York.

Joklik, W.K., H.P. Willett, and D. B Amos. 1984. Zinsser microbiology. Appleton-
Century-Crofts, East Norwalk, Conn.

Ljungdahl, LG. 1986. The autotrophic pathwa of acetate synthesis in acetogenic bacteria.
Annu. Rev. Microbiol. 40: 415-450.

Taylor, B. F., and R. S. Oremland. 1979. Depletion of adenosine triphosphate in
Desulfovibrio by oxyanionsof group VI elements. Curr. Microbiol. 3: 101-103

Wilson, L. G., and R. Bandurski. 1958. Enzymatic reactions involving sulfate, sulfite,
selenate, and molybdate. J. Biol. Chem. 233:975-981

Tonsager, SR, and B. A. Averill. 1980. Difficulties in the analysis of acid-labile sulfide
in Mo-S and Mo-Fe-S systems, Anal. Biochem. 102213-15.

150

Chapter Eight: ENGINEERING APPLICATIONS

8.1. Enrichment of Rapid and Complete Dechlorinating Activity

The research presented in this thesis provides insights into the practical
application of selective biostimulation: how to enrich microbial populations that are
capable of complete and rapid degradation of a contaminant. For oligotrophic aquifers
such as Plume G, the native dechlorinating microflora are present at such small numbers
that selective stimulation of the degradative indigenous soil microorganisms is required.

However, the success of stimulating dechlorinating activity may be severely
limited by biological factors such as competition among dechlorinating and non-
dechlorinating microorganisms, and the nutritional requirements of dechlorinating
bacteria. Although the complete transformation of PCE to ethene have been observed
(DiStefano et al. 1991; Freedman and Gossett, 1989), the PCE transformation often
stopped at TCE (F athepure and Boyd, 1988), or c-DCE (Sewell and Gibson, 1991) or VC
(Vogel and McCarty, 1985). These incomplete reductive dechlorination reactions pose
environmental risks because the intermediates c-DCE and VC are themselves hazardous
environmental pollutants.

One of the most significant findings from this research has been the achievement
of complete transformation of chlorinated ethenes by mixed cultures, which originally
exhibited only limited and incomplete dechlorination. By pulse feedings of lactate and
chlorinated compounds over extended periods, rapid and complete dechlorinating activity

has been enriched. However, when enrichments were stimulated in high nutrient

151

conditions, slow and incomplete dechlorinating activity was expressed. The time
sequences of the enrichments are summarized in Table 8.1. After 1 year of enrichment,
36 out of 36 bottles from low lactate/PCE enrichment showed complete activity
compared to 0 out of 28 bottles from high lactate/PCE enrichments (Table 8.2). In
addition, the transformation rate of PCE became 15 times faster in low lactate/PCE
enrichments than high lactate/PCE enrichment after approximately same pulse feedings

(12 cycles).

Table 8.1. Enhancement of first order rate k (1/day) on disappearance of parent products by

pulse feedings overtime.

 

 

 

 

Number of High Low Low Low
Pulse feedings Lactate/PCE Lactate/PCE Lactate/TCE Lactate/c-DCE

3 0.008 0.012 0.01 0.01

6 0.04 0.38 0.42 0.33

12 0.06 0.85 0.72 0.62

 

 

 

 

 

 

 

Higher levels of Dehalococcoides populations were present in low lactate
enrichments compare to high lactate enrichments. This demonstrated that pulse feedings
strategy of chlorinated compounds and low electron donor could be a preferred method
for the stimulation of halorespiring microbes. The mechanisms by which nutrient pulse
feeding stimulates Dehalococcoides has two main components: (1) pulse feeding allows
sustained Hz/acetate production and cofactors generation by fermentators, and (2) the
relative low nutrient/high chlorinated ethene levels favor microbes that can exploit the
energy (AG°') present in the oxidized Cl-C bonds. Another examples of cofactors known

to affect bioremediation are iron and copper for Pseudomonas KC (Tatara et al. 1993;

152

Dybas et al. 1995; Kim, 1998; Dybas et al. 1998). Schematics showing these

relationships are included in Figure 8.1.

Hz and Acetate

 

 

 

 
   
 

Lactate Fermentators Dehalococcoides

 

 

 

J
Cofactors and B 1 2

 

Figure 8.1. A schematic of the relationship between fermentators and Dehalococcoides.

153

Table 8.2. Comparisons of maximum rate between the pulse fed and once-fed enrichments
in enhancing dechlorination activity observed from each condition. First order rates are

obtained after the end of the enrichment experiment.

 

 

 

 

 

 

 

Method Chloroethene Electron donor Maxrmun: k Success RatesCl
(l/day)

Once-fed TCE" High Lactate 0.004 0/220
Pulse-fed PCEb High Lactate 0.03 0/28

Pulse-fed PCE” Low Lactate 0.82 36/36
Pulse-fed TCEb Low Lactate 0.73 36/36
Pulse-fed cDCE" Low Lactate 0.66 36/36
Pulse-fed TCAc Low Lactate 0.25 12/36

 

 

 

 

 

 

 

‘ Based on first-order rate equation C=Coe"“.
b Final byproduct to ethene or ethane.
“ Final byproduct to Chloroethane.

d Occurrence of complete and rapid degradation.

8.2 Field Applications

The remediation of groundwater containing chlorinated compounds is
challenging. Traditional approaches for groundwater remediation have relied an
extraction, followed by a physical/chemical process (e. g., air stripping and carbon
adsorption). This approach has some major disadvantages such as high cost,
inefficiencies in removing contaminants absorbed to the aquifer material, and
requirement of pumping, treating, and ultimately disposal of large volumes of water.

In situ bioremediation is an alternative approaches to these more traditional
methods due to the discovery of bacterial populations that grow on chlorinated

compounds, thus efficiently reducing and detoxifying these compounds. Major

154

 

advantages of employing such a method are potentially lower operation costs and
minimum ecological disturbances. In addition, in-situ bioremedation does not generate
waste solids for disposal.

Biostimulation refers to enhancing the metabolic activity of indigenous microflora
to transform target compounds and usually requires amendments (electron donors and/or
acceptors) to enhance the microbial populations already present. Factors such as the
delicate nutritional requirements of dehalogenating microbes and electron donor
competition between dechlorinating and non-dechlorinating microorganisms have
presented challenges to implementation of this promising technology in the field.

Much of the work presented in this thesis has been directed toward feasibility of
biostimulation in Schoolcrafi Plume G site. For oligotrophic aquifers such as Plume G,
natural attenuation may be insufficient, and electron donor addition is often needed to
accelerate reductive dechlorination or stimulate cometabolic activities. Although
bioaugmentation offers the promise of increased control over transformation of a specific
compound, competition between indigenous microflora and the introduced organisms
usually presents a challenge.

Laboratory and field investigations were used to develop cost effective methods
of biologically generated reducing conditions to support reductive dehalogenation of
PCE/TCE/T CA mixtures. We have evaluated the two levels of enhanced remediation in
the Plume G: biostimulation and bioaugmentation: and compared their results over three-
year period. Biostimulation could be part of a remediation strategy for a VOC plume
(PCE and daughter products, 1,1,1-TCA and daughter products) co-mingled with a

chromium and arsenic plume.

155

 

A series of periodic lactate additions were conducted in the field to test the L. .
hypothesis that pulse feedings of nutrient influence the stimulation of dechlorinating
activity in the study region. Figure 8.1 depicts the delivery setups employed in the
Schoolcrafi Plume G. Groundwater was pumped from the flux control well to a mixing
tank where lactate was added once a week using an in-tank recirculation system.
Solution from the mixing tank was injected into the ground in delivery wells 1 and 2.
The screened interval for these wells is between 18.9 and 25.0m bgs.

Biostimulation by weekly nutrient feeding resulted in TCE, c-DCE, and VC
degradation (Figure 8.2). TCE and daughter products were degraded with a different
motif than occurred under bioaugmentation conditions, with slower initial TCE removal,
but more rapid c-DCE and VC removal. 1,1,1-TCA was more rapidly degraded in the
biostimulation than the bioaugmentation test cell (data not shown).

Bioaugmentation with halorespiring microbes in a flow through system can be
accomplished with a one-time inoculation of Bachman Road halorespiring consortium
and weekly nutrient feeding. No pre-reducing of the aquifer was required. Degradation
of TCE, c-DCE and VC has been observed, with aqueous phase reductions of up to 99%
for TCE within 60 days. Degradation of TCE and daughter products has continued for
over 15 months following inoculation. Both laboratory results and field evaluations have
successfully demonstrated the importance of periodic feedings in stimulation of
indigenous dechlorinating populations.

Observed enhanced dechlorination and enumeration of Dehalococcoides
populations are summarized in the Table 8.3. Compared to a test grid inoculated with

Bachman Consortium (0.2 %), biostimulation shows similar levels of Dehalococcoides.

156

 

Overall, in situ biostimulation using native flora to remove chlorinated compounds shows
promise for cleaning up contaminated groundwater and sediments. The levels of
Dehalococcoides stimulated in the biostimulation grid was comparable to the
bioaugmentation system, suggesting that stimulation using indigenous flora shows
promise for cleaning up chlorinated compounds contaminated aquifers in a cost-effective

manner.

Table 8.3. Levels of Dehalococcoides DNA in the biostimuation and bioaugmentation

test sites.

 

Dehalococcoide DNA (%)

 

 

 

Days 36 Days 235
Bioaugmentation 2.08 0.024
Biostimulation 2.86 0.025

 

 

 

 

 

157

 

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158

 

Figure 8.3. Comparison on the Observed Reduction in Chlorinated Compounds in the

biostimuation and bioaugmentation test sites.

(a) Biostimulation System

 

1500

pt
N
c
Q
l

900 _ c-DCE VC

 

 

 

Chlorinated Ethenes (ppb)

 

 

Weeks from Initial Feed

(b) Bioaugmentation System

 

1500

1200 -

900— g TCE
It:
i / X
600- g ‘
: ,
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Chlorinated Ethenes (ppb)

 

 

0 10 20 30 40 50 60 70

Weeks from Inoculation

159

 

8.3. Literature cited

DiStefano, T. D., J. M. Gossett and S. H. Zinder. 1991. Reductive dechlorination of high
concentrations of Tetrachloroethene to ethene by an anaerobic enrichment culture in the
absence of methanogenesis. Appl. Environ. Microbiol. 57: 2287-2292.

Freedman, D. L. and J. M. Gossett. 1989. Biological reductive dechlorination of

Tetrachloroethylene and Trichloroethylene to ethylene under methanogenic conditions.
Appl. Environ. Microbiol. 5 5(9): 2144-215 1.

F athepure, B. Z. and S. A. Boyd. 1988. Reductive dechlorination of perchloroethylene
and the role of methanogens. FEMS Microbiology Letters 49(2): 149-156.

Sewell, G. W., and S. A. Gibson. 1991. Stimulation of the reductive dechlorination of
tetrachloroethene in anaerobic aquifer microcosms by the addition of toluene. Environ.
Sci. T echno]., 25(5): 982-984

Vogel T. M. and P. L. McCarty. 1985. Biotransformation of tetrachloroethylene to
trichloroethylene, dichloroethylene, vinyl chloride, and carbon dioxide under
methanogenic conditions. Appl. Environ. Microbiol, 49(5): 1080-1083.

Tatara, G. M., Dybas, M. J ., and C. S. Criddle. 1993. Effects of medium and trace
metals on kinetics of carbon tetrachloride transformation by Pseudomonas sp. strain KC.
Appl. Environ. Microbiol. 59(7), 2126-2131.

Dybas, M. J ., Tatara, G. M., and C. S. Criddle. 1995. Localization and
characterization of the carbon tetrachloride transformation activity of Pseudomonas sp.
Strain KC. Appl. Environ. Microbiol. 61(2): 758-762.

Kim, HaeKyung. 1998. The role of trace copper in the transformation of carbon
tetrachloride by Pseudomonas stutzeri KC. A MS Thesis, Department of Civil and
Environmental Engineering, Michigans State University.

Dybas, M. J ., Barcelona, M., Bezborodnikov, 8, Davies, S., Forney, L., Heuer, H.,
Kawka, O., Mayotte, T., Sepulveda-Torres, L., Smalla, K., Sneathen, M., Tiedje, J.,
Voice, T., Wiggert, D. C., Witt, M. E., and C. S. Criddle. 1998. Pilot scale evaluation of
Bioaugmentation for in-situ remediation of a carbon tetrachloride-contaminated aquifer.
Environ. Sci. T echnal. 32(22): 3598-3611.

Jaime A. Graulau-Santiago. 2003. Development and Application of a Methodology to
Evaluate Natural Attenuation for Chlorinated Solvents using Conceptual and Numerical
Models. A Doctoral Thesis. Civil and Environmental Engineering. Michigan State
University.

160

Chapter Nine: CONCLUSIONS AND FUTURE STUDY

9.1. Conclusions
The following specific and general conclusions have been drawn regarding the
development and characterization of dechlorinating microorganisms based on research

performed using Plume G sediment and its enrichment:

o Incomplete and slow dechlorination is stimulated in Plume G by microcosms by
addition of nutrients such as CMB, lactate, and palmitate.

o Cometabolic dechlorination was dominant in high lactate and PCE pulse fed
conditions

0 Halorespiration was dominant in low lactate and PCB pulse fed conditions

0 Rates and extent of dechlorination increased with repeated feeding of chlorinated
compounds

0 High hydrogen concentration did not favor dechlorination over methanogenesis.
Enhanced dechlorination activity under these conditions was not observed, which
agrees with predictions by other researchers based on half-velocity constants with
respect to hydrogen.

0 Dehalococcoides 16S rRNA gene targeted real—time PCR confirmed that two
order of magnitude higher amounts of Dehalococcoides DNA present in low
lactate enrichments than high lactate enrichment.

0 Metabolic inhibitors such as molybdate did not impact the low PCE/lactate
enrichments, but high levels of sulfate inhibited dechlorination. This suggested

that the sulfate-reducers are not involved in the observed dechlorination.

161

0 Low levels of BES specifically inhibited methane production, but retained
dechlorinating activity, indicating that the responsible microorganism(s) is (are)
non-methanogenic.

o The reduction of acetate production was observed in vancomycin treated
enrichment and inhibited dechlorination. This suggests that acetate-utilizing

microbes are responsible for the observed dechlorination.

9.2. Future Study

Results from research presented in the previous chapters indicated that fithher
research is required in specific areas. m, no attempts have been made to isolate PCE,
TCE, and c-DCE dechlorinating microorganisms. The isolation of these dechlorinators
from the enrichment appears to be necessary to better understand its nutritional
dependencies on other organisms present in the culture. A better understanding of the
growth requirements of TCA dechlorinating microorganisms can also be obtained
through research using highly purified or pure cultures. Sew—11d, abiotic dechlorination
observed in the microcosms has to be explored further due to the prevalence of high
sulfate area (such as CA and CO) contaminated with chlorinated compounds. I‘m, the
organism(s) responsible for the reduction of ethene to ethane has to be explored. This
can be achieved again by purifying or isolating microorganisms from the enrichment.
Egyfih, the role of sulfate reducing dechlorinators such as Desulfomonile tieay'ei in
enrichment has to be addressed clearly. Enumeration of such microorganisms using
specific primers would further dissect the enrichment community and their interactions

with each other.

162

APPENDICIES

163

APPENDIX A. Electron Donor Calculations

(Adapted from McCarty McCarty, Perry L. 1969. Energetics and Bacterial Growth. The
Fifth Rudolf Research Conference, Rutgers, New Brunswick, New Jersey)

Simple Models of Electron Donor Requirement Calculations. Microorganisms
obtain energy for growth and maintenance by removing electrons from electron donors
and transferring them via macromolecules to electron —deficient compounds, such as
nitrate or sulfate, the terminal electron acceptors. A fraction of the electrons removed
from the electron donor may also be used to reduce oxidized forms of carbon and
nitrogen in the creation of new biomass. These two pathways for the consumption of
reducing power are summarized in this equation where fs is the fraction of electrons
diverted for synthesis and the fraction used for energy generation is termed fe.

R = Rd + feRa + stc
Where: Rd = half reaction for the oxidation of an electron donor normalized by the

moles of electrons removed fi'om the donor,

Ra = half reaction for the reduction of an electron acceptor used for energy

normalized by males of electrons added to the acceptor,

Rc = half reaction for the reduction of an electron accept used for synthesis

normalized by the males of electrons added to the acceptor, and fs + fe = 1

 

 

AG°'
Electron Half reaction (Kl/mol)
donor
Methanol 0.167 methanol + 0.167 water => 0.167 C0; + H+ + e' -37.53
Lactate 1/ 12 lactate + 1/3 water => 1/6 C02 +1/12HC03'+ l-F + e' -32.96
Ethanol l/12 ethanol + 3/12 water => 2/12 C02 +1/12HCO3‘+ IF +e' -31.13
Propionate 1/14 propionate + 5/14 H20 => 1/7 C02 +1/14HC03‘+ H)“ + e' -27.9
Acetate 1/8acetate + 3/8 water => 1/8C02 + 1/8HC03‘+ I-F + e' -27.98
Palmitate 1/92 C16HglOz' + 31/92 water => 15/92 C02 + 1/92HC03'+ H* + e' -27.87

 

164

E0 I AGO I

Electron Half reaction
acceptor (V) (KJ/mol)

Nitrate 1/5 N03" + 6/5H+ + e’ => 1/10N2+2/5 H20 0.74 71.71

reduction
1/8 SO42' + 5/4l-F + e' => 1/8 HS’ + 1/2 H20 -0.22 20.79

Sulfate
reduction
Cell % C0; + 1/20 NH; + H+ + 6 =>
synthesis
“20 C5H702N + 2/5 H20

 

The appropriate value for fs depends upon the type of microorganism or enrichment
involved, its electron donors and acceptors, and the amount of decay a culture

 

 

experiences.
-kAGr
(film. =
—kAGr+188+A0n+ AGp
k k’"

where: AGs = free energy per e- mole of the electron donor used for the creation of

biomass = AGp + 18.8 + AGn
AGr = free energy per e- mole released by oxidation of the electron donor

AGp =free energy per e- mole of the electron donor— energy need for pyruvate

AGn = free energy need for N assimilation
k = the efficiency which energy is converted into new chemical

m=-1 if AGp<0
m=1ifAGp>0

In our system:
Residual amounts of nitrate and sulfate in the groundwater; 65 ppm and 80 ppm each

Assume ammonia is N-source
K is 0.6

165

 

 

 

 

 

 

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167

APPENDIX B.

Effect of Added Sulfate on the Powder X-ray Diffraction Pattern by Iron Sulfide

produced in Microcosm.

 

 

 

 

 

 

 

 

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168

APPENDIX C. Design of Experiment

A statistical design of experiment was implemented to evaluate the effects of
sulfate, reductant (Na2S), and nutrient in the stimulation of dechlorination (Table 1).
Replications were performed to gain a sufficient level of confidence in the estimated
parameters. The main and interaction effects were analyzed by AN OVA and presented in
Table 2. The two main effects such as sulfate and Na2S exhibited as insignificant
components on dechlorination. The nutrient alone had significant effect. According to
the P-value, the two selected factors sulfate and reductant were insignificant at the 0.01
levels. The interactions between each factor were insignificant. In statistical analyses a

P value of <0.01 was considered the criterion for significance.

Table 1. Experimental Design Matrix for Dechlorination

 

 

Run Sulfate Reductant Nutrient Run label
(A) (B) (C)

1 - - - (1)

2 + - -

3 - + - b

4 + + - ab

5 - - + c

6 + - + ac

7 - + + bc

8 + + + abc

 

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APPENDIX D. The Additions and Consumptions of Chlorinated Compounds for
Experiment II

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% Total Area

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400.64
450.82
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600.76
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351 401 451 475 501 551 601 650 701 750 801

Fragment size (bp)

Terminal fragment

174

Peak area

1335
1234
1355
1690
1532
1462
1845
1820
1784
1890
1558
1599
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2088

APPENDIX E. T-RFLP Patterns from Chapter 6

 

850 900 951

2. T-RFLP PATTERS SHOWN IN CHAPTER 6 (Figure 6.7).

(a) Pre-enrichment (b) High Lactate/PCE (c) Low Lactate/PCE
Terminal Peak area Terminal Peak area Terminal Peak area
fragment (bp) fragment (bp) fragment (bp)
206.92 518 449.68 1168 228.09 1115
650.26 1144 474.55 1352 400.3 1211
700 1102 499.58 1257 450.34 1243
900.19 1055 550.23 1170 475.43 1427
949.8 1341 600.12 1348 500.44 1301
999.33 1416 650.13 1455 550 1473
699.86 1358 600.48 1775
750.29 1272 650.64 1654
800.62 1292 700.4 1515
849.83 1308 750.58 1664
899.46 1515 800.31 1321
949.8 1503 850.33 1512
999.56 1956 899.82 1602
949.6 1853
(d) Low Lactate/TCE (e) Low lactate/cDCE (1) Low lactate/T CA
Terminal Peak area Terminal Peak area Terminal Peak area
fragment (bp) fragment (bp) fragment (bp)
126.02 475 475.29 1379 368.61 252
218.56 385 550.8 1208 450.11 1045
250.25 765 600.31 1495 475.3 1227
350.92 649 650.33 1625 500.41 1128
450.21 1223 700.92 683 550 1207
475.2 1532 750.49 1441 600.44 1457
500.72 1388 799.74 1221 648.47 610
550.53 1579 900.6 1476 650 1561
600.33 1670 950.32 1851 699.87 1414
650 1764 1000 1970 750.67 1403
700.25 1713 850.15 1251
750.4 1691 900 1517
800 1446 950.72 1675

850.45 1735
900.33 1752
950.36 2084

175

 

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Fragment size (bp)

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Fragment size (bp)

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(c) Sulfate before

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% Total Area

 

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Fragment size (bp)

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% Total Area

 

 

450 475 500 550 600 650 700 750 801 850 899 950 1000
Fragment size (bp)

182

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